Environ. Sci. Technol. 1994, 28, 1437-1443
Effect of Combustion Temperature on the Atmospheric Stability of Polybrominated Dibenzo-pdioxins and Dibenzofurans Parag Blrlat and Richard M. Kamens'
Department of Environmental Sciences and Engineering, University of North Carolina, Chapel Hill, North Carolina 27599-7400 Polybrominated dibenzo-p-dioxins and dibenzofurans (PBDDs and PBDFs) produced from the combustion of polyurethane foam (PUF) containing 4.4 % ' polybrominated diphenyl ethers (PBDPEs) were injected into outdoor Teflon film chambers and aged in the presence of sunlight under typical atmospheric conditions. Experiments with combustion temperatures in the range of 400470 "C were categorized as "low-temperature" experiments and those in the range of 670-780 "C as "high-temperature" experiments. Production of PBDFs, namely, tetrabrominated dibenzofuran (TBDF), and pentabrominated dibenzofuran (PeBDF) and decay of tetrabrominated-pdioxin (TBDD) were observed in low-temperature experiments. Production of TBDF and PeBDF is believed to occur from the photolysis of unburned PBDPEs. TBDF, PeBDF, and TBDD emissions from high-temperature experiments were stable. Particle-bound PAHs from lowtemperature experiments degraded while corresponding PAHs from high-temperature experiments were stable. Experimental observations suggest that under incinerator conditions at combustion temperatures of 800 "C, particulate-bound emissions of PBDDs, PBDFs, and PAHs will be transported over long distances due to long halflives. Combustion temperatures around 450 "C can lead to unstable emissions of these compounds with atmospheric half-lives of the order of 1-6 h.
Introduction Incineration has been projected as an attractive waste disposal technology. This projection is underscored by the fact that since 1979, 3500 landfills have been closed and within the next 5 years landfill capacity will exist for only 20 million t of the anticipated 160 million t of waste generated (1). Concomitant with the likely increased use of incineration is the growing production of brominated organics, particularly with respect to polybrominated diphenyl ethers (PBDPEs), which are used as fire retardants in textiles, plastics, carpets, and other materials (2). Incomplete combustion of these brominated organics, present in municipal waste, can lead to the formation of polybrominated dibenzo-p-dioxins (PBDDs) and dibenzofurans (PBDFs) (3). Therefore, it is expected that atmospheric concentrations of PBDDs and PBDFs will rise in the future in light of the increasing production of brominated compounds. While toxicological studies have primarily focused on polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDDs and PCDFs), a study conducted on guinea pigs has shown that PBDDs and PBDFs are at least as toxic as their chlorinated analogs (2). Central to biological exposure and uptake of PBDDs and PBDFs is their atmospheric fate and transport. Operating
* To whom correspondence should be addressed. t Present address: Alpha-Gamma Technologies, Inc., 900 Ridgefield Dr., Suite 350, Raleigh, NC 27609.
0013-936X194/0926-1437$04.50/0
0 1994 American Chemical Society
combustion temperatures in incinerators have been found to vary from 300 to 1300 "C ( 4 ) . Current designs target temperatures of 1100 "C and a minimum of 2-9 residence time in the combustion zone for the incinerated material. However, nonideal conditions can result from fluctuations in processconditions such as waste feed rate, waste content, inadequate turbulence in the combustion zone, and other factors. These factors may lead to lower combustion temperatures in the older or conventional incinerators ( 4 ) . Thus, a need arises to study the effect of combustion temperature on the atmospheric stability of incinerationgenerated pollutants. The need for research in this area assumes added importance in light of a recent study that has cited savings worth thousands of dollars and health benefits as a result of lower incinerator operating temperatures (5). The present study addresses the question of the effect of combustion temperature on the atmospheric stability of incineration-generated pollutants by focusing on PBDDs and PBDFs generated from the incineration of brominated organics in a prototype incinerator. An important consideration in determining the atmospheric stability of semivolatile organics like PBDDs and PBDFs is their gas-particle partitioning. Theoretical predictions using the partitioning theory developed by Junge (6) and experimental observations by Lutes et al. (7) revealed that more than 95% of PBDDs and PBDFs occur on the particulate phase (8). Thus, the atmospheric stability of PBDDs and PBDFs is primarily relevant to their occurrence on atmospheric aerosols as opposed to their presence in the gas phase. Currently, there exists a paucity of data on the photodegradation of PBDDs and PBDFs under realistic atmospheric conditions. Laboratory studies have been carried out in organic solutions or on solid substrates. An important observation regarding these experiments is the rapid decay of 2,3,7,8-tetrabrominated dibenzo-p-dioxin (TBDD) and 2,3,7,8-tetrabrominated dibenzofuran (TBDF) in isooctane (half-lives of 0.8 and 0.7 min, respectively) and relative stability on quartz surface (half-lives of 32 and 35 h, respectively) (9). Given such disparity in halflives it becomes all the more imperative to arrive at representative values for atmospheric half-lives of PBDDs and PBDFs using outdoor chambers. Preliminary work in this direction revealed that thermal destruction of polyurethane foam- (PUF-) containing PBDPEs at combustion temperatures ranging from 640 to 760 "C yielded particulate-bound PBDDs and PBDFs, which were relatively stable as in the solid-phase laboratory experiments (7). In the present study, experiments were carried out with PUF-containing PBDPEs at combustion temperatures ranging from 400 to 780 "C to determine the effect of combustion temperature on the atmospheric stability of PBDDs and PBDFs. Particulate-bound polycyclic aromatic hydrocarbons (PAHs) generated from these experiments were also quantified to determine if their behavior was similar to or different from the behavior of PBDDs and PBDFs. Envlron. Sci. Technol., Vol. 28. No. 8, 1994
1437
_l__ll._
Table 1. Comparison of Experimental Conditions date of expt chamber type of expt combustion temp ( O C ) chamber temp ("C) relative humidity ( % ) av TSRa max [NO] (ppm) max [ N o d (ppm) max [031 (ppm) PUF burned (9) initial particle concn (pg/m3) injection time (a.m.)
08/21/91 east low 400 32.5-40.5 76-83 1.02 0.002 0.016 0.07 1.48 9935 1001
08/21/91 west high 745 30-42 73-76 0.85 0.32 0.16 0.24 1.69 2025 1012
11/05/91 east high 780 19.5-22.2 10-13 0.72 0.074 0.061 0.06 0.48 942 11:oo
11/05/91 west high 780 20.2-21.8 11-13 0.72 0.094 0.074 0.05 0.63 3154 11:Ol
05/21/92 aerosol low 470 26.3-26.7 45-46 0.99b 0.002 0.005 0.045 3.91 1234 9:57
10/03/92 east low 470 23.9-27.4 39-42 0.96 0.016 0.026 0.041 2.08 6043 10:15
10/03/92 west low 470 24.1-27.1 39-42 0.96 0.013 0.021 0.014 1.96 14159 1025
*
TSR (total solar radiation) in cal cm-2 min-l; av of first 4 h of reaction. Av for 3 h.
Experimental Section Incineration experiments were carried out in an ignition vessel (5-in. i.d., 15-in. high) described in a previous study (7). Liquid petroleum gas was used as the combustion fuel. The combustion material consisted of PUF containing 4.4% w/w of industrial grade DE-71 mixture (40% tetrabromodiphenyl ether, TBDPE, 57 % pentabromodiphenyl ether, PeBDPE, and 3 % hexabromodiphenyl ether, HxBDPE). Experiments carried out with combustion temperatures ranging from 400 to 470 OC were categorized as low-temperature experiments and those in the range of 650 to 780 "C as high-temperature experiments. Although the present study refers to the lowtemperature experiments as incineration experiments, ideally such experiments would not reflect incineration but simply thermal destruction. Emissions from the prototype incinerator were directed into outdoor smog chambers located in Pittsboro, NC (7). Three chambers were used in this study. Two identical 25-m3chambers, designated as east and west, allowed for two experiments to be performed under similar conditions (7). In addition, on one occasion a 190-m3 chamber called the aerosol chamber was used as it facilitated the collection of two sets of samples from the same experiment (10). The emissions were allowed to age within the chamber atmosphere in the presence of sunlight. During this time, particulate samples were collected by passing chamber air through a sampling train containing a 47-mm T60 A20 Teflon-impregnated glass fiber filter, followed by a 4 in. X 1.5 in. PUF cartridge for the collection of vapor-phase species. PUF used for vapor collection did not contain PBDPEs and was soxhlet extracted in toluene prior to sample collection. Chamber ozone and oxides of nitrogen were monitored during the course of an experiment with chemiluminescent monitors (Bendix Model 8101-B and 8002 analyzers). Relative humidity and chamber temperature were measured by a thermohygrometer (Hanna Model 8564). Total solar radiation was measured with an Eppley black and white pyrometer (Newport, RI), and dilution rate was measured by a gas chromatograph/ electron capture detector using SFe as a tracer. An electrical aerosol analyzer (EAA Model 3030) and a laser aerosol spectrometer (LAS-CRT)were used to gather data on the size distribution of the combustion-generated particles. In addition, nuclepore (0.1 rm) filter samples were collected and analyzed with a scanning electron microscope (SEM, Cambridge Model S200). SEM photographs were used to describe particle morphology. Table 1438 Envlron. Scl. Technol., Vol. 28, No. 8, 1994
1presents the experimental conditions for the combustion experiments carried out in this study. Prior to extraction and enrichment, filter samples from chamber experiments were weighed on a three-place milligram balance (Sartorious microbalance, Model 4503MP6) to determine the particulate mass collected. Particulate samples were analyzed for PBDDs and PBDFs in accordance with EPA Method 8290 (11). Briefly, this involved extraction of the samples in toluene and enrichment using three sets of columns: silica gel, florisil, and carbodcelite. All laboratory work was carried out under artificial lights and shielded against UV radiation, and samples were kept wrapped in foil to preclude photodegradation of the analytes in the laboratory. Quantification of PBDDs and PBDFs was achieved by internal standardization using 13Clz-labeledcompounds. The extracted samples were analyzed by high-resolution gas chromatography/high resolution mass spectrometry (HRGC/ HRMS) using a Hewlett-Packard 8290 gas chromatograph interfaced to a VG 70-250SEQ hybrid mass spectrometer. Gas chromatographic separation of the PBDDs and PBDFs was achieved through a 30-m, 0.32-mm i.d. DB-5 column (J&W, Folsom, CAI. The analysis was carried out in the selected ion monitoring (SIM) mode at a resolution of 10000. Precision and accuracy checks on the method employed were performed in an earlier study. An analysis of similar samples (7) taken from the chambers during periods of darkness revealed a repeatability (one standard deviation) of f 6-13 % . Filter samples collected from the low-temperature experiment on May 21, 1992, were also analyzed for the presence of PBDPEs. The procedure for PBDPE analysis was similar to that for PBDDs and PBDFs, except that the carbon column was excluded in the chromatographic cleanup process, since the carbon column retains PeCDPEs (12). Therefore, sample extracts were split into two following the florisil column. One part was used for brominated dioxin and furan analysis and the other was quantified for PBDPEs. l3C1z-labeledPBDFs were used as internal standards for the quantification of PBDPEs. Samples for PAH analysis were extracted in dichloromethane (DCM) and analyzed by HRGC/HRMS in the SIM mode a t a resolution of 10 000. 0,P-Binaphthyl was used as the internal standard for quantifying PAHs. Filter samples for PAH analysis were collected under experimental conditions identical to that for PBDD and PBDF samples, thereby facilitating a direct comparison of the behavior of PAHs versus PBDDs and PBDFs.
1
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Flgure 1. TBDD, TBDF, and PeBDF behavior in sunlight on particles derived from hlgh-temperature combustion (Nov 5, 1991).
Analysis for particulate organic mass (POM)was carried out by extraction of particulate samples in DCM. The resultant extracts were filtered through a syringe equipped with a filter paper to exclude particles that were carried over into the extract during soxhlet extraction. The filtered extracts were then spiked on preweighed filters. The spiked filters were left uncovered for approximately 30 min to permit evaporation of DCM. Following this, the filters were reweighed to determine the change in mass. This change in mass was attributed to POM.
Results and Discussion High-Temperature Experiments. The compounds detected were TBDD, TBDF, and PeBDF. Detection of individual isomers was not possible due to the lack of appropriate standards. Hence, results have been presented in terms of the total concentration of a compound and not in terms of specific isomers. Two high-temperature experiments were carried out (Aug 21, 1991, and Nov 5, 1991). Figure 1 illustrates typical TBDD, TBDF, and PeBDF behavior from these experiments. For these experiments, plots of chamber concentrations of TBDD, TBDF and PeBDF versus chamber aging time had insignificant slopes at a 90% level of confidence. Therefore, it can be concluded that there is little evidence of decay for PBDF and PBDD emissions from hightemperature experiments. These results are similar to those obtained by Lutes et al. (7) and are somewhat comparable to the laboratory solid-phase photodegradation experiments carried out by Buser et al. (9). Examination of Table 1 reveals that environmental conditions with respect to relative humidity and ambient temperature are very different for the two high-temperature experiments carried out in this study. However, the stable behavior of the analytes in both cases renders these parameters inconsequential in assessment of their atmospheric stability. Two high-temperature experiments were carried out on November 5, 1991. Particulate samples from one of the experiments were analyzed for PAHs. Results from this analysis are presented in Figure 2. PAHs like phenanthrene, fluoranthene, pyrene, and chrysene had insignificant slopes at a confidence level of 90% , implying that these compounds were stable in the chamber atmosphere. This behavior is similar to that of TBDD, TBDF, and PeBDF and suggests that particulate-bound emissions from our high-temperature experiments may have long
0
~
I I
1
2
3
4
5
TIME SINCE INJECTION (h)
Flgure 2. PAH behavior in sunlight on particles derived from hightemperature combustion (Nov 5, 1991).
half-lives under atmospheric conditions, regardless of their chemical structures. However, one of the PAHs, benz[alanthracene (BaA), had a significant negative slope at a confidence level of 90%. The assumption of first-order decay yielded a half-life of 8 h for BaA. Previous investigations have found BaA as one of the most reactive PAHs on atmospheric aerosols with half-lives as low as 0.7 h (13). In the context of this finding, even BaA appears to exhibit a stable behavior on combustion particles generated from our high-temperature experiments, in spite of the significant negative slope. In the past, PAHs have been found to decay rapidly on spark ignition and on wood soot aerosols (13,14). However, high-temperature experiments conducted in this study yielded particles which do not permit rapid PAH photodegradation. Low-Temperature Experiments. Low-temperature experiments were conducted on August 21,1991; May 21, 1992; and October 3, 1992. Curves depicting concentrations of TBDD, TBDF, and PeBDF with chamber aging times are presented in Figures 3 and 4. Significant negative slopes for TBDD were obtained at a confidence level of 90%. Calculations using the first four points and firstorder kinetics yielded TBDD half-lives of 0.79 and 0.98 h for the August 21,1991, and October 3,1992 experiments. Both days had similar levels of solar radiation (Table l ) , although the October day had lower humidity and temperature conditions, hence a slightly lower rate constant. This rapid decay of TBDD is a departure from the stable TBDD behavior observed in the high-temperature experiments (Figure 1)in this study and in the Lutes et al. study (7). It is believed that the major decomposition pathway is reductive debromination, leading to lower halogenated species (9). In contrast to TBDD degradation, the formation of TBDF and PeBDF over time yielded significant positive slopes in low-temperature experiments (Figure 3). We believe that the formation of these compounds was the result of photolysis of unburned PBDPEs (15). The major reaction pathway (Scheme 1)for the formation of PBDFs from photolysis of PBDPEs involves dehalogenation and ring closure (15). Therefore, in our system TBDF can be formed from the photolysis of PeBDPE or HxBDPE, and PeBDF can be formed from the photolysis of HxBDPE only. The scope for formation of these compounds in hightemperature experiments is negligible as it has been found that thermolysis of PBDPEs at temperatures around 800 Environ. Sci. Technol.. Vol. 28, No. 8, 1994
1430
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Flgure 3. TBDD, TBDF, and PeBDF behavior in sunlight on particles derived from low-temperature combustion (top, Aug 8, 1991, east chamber; bottom Oct 3, 1992). I
t
0
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0.8 1.2
1.6
Hours after
2
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injection
A A
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Flgure 4. Data and model of TBDF, PeBDF, PeBDPE and HxBDPE from a low-temperature burn experiment (May 21, 1992).
"C causes 96-97 % destruction of the ethers. In contrast, thermolysis at 510 OC causes only 10% destruction (16). Hence, in our low-temperature experiments photolysis of 1440
Envlron. Scl. Technol., Vol. 28, No. 8, 1994
unburned PBDPEs may play a major role and confound any photodegradation of TBDF and PeBDF. An investigation of this hypothesis resulted in the quantification of PBDPEs. A portion of the extract from the May 21, 1992, experiment was used for PBDPE analysis in accordance with the criteria specified in the preceding section. Results obtained from PBDPE analysis are shown in Figure 4. Also plotted in the same figure are TBDF and PeBDF concentrations over time. Decay was observed in TBDPE, PeBDPE, and HxBDPE concentrations over time. An estimation of TBDF and PeBDF decay by eliminating their formation from ether photolysis would require the knowledge of yield rates. It has been found by Choudhry et al. that the irradiation of 2,2',4,4',5-pentachlorodiphenyl ether (PeCDPE) dissolved in cyclohexane yielded 14% TCDF over a 4 h period (15). Wantanabe and Tatsukawa (17 ) irradiated DeBDPE in hexane with sunlight and observed a 2 % yield of TBDF in 2 h. These yields represent formation as well as photooxidation losses. Moreover, Buser observed sunlight (midday, cloudless, latitude 47O, 400 m above sea level) photoinduced half lives of TBDFs and TBDD on the order of minutes in an isooctane mixture. In our system (Figure 3), we observed TBDD half-lives on the order of 50 min. These inconsistencies can be explained by differences in the nature of the reaction liquid or particle substrate. McDow and coworkers have reported that different solvents can dramatically influence rates of photooxidation reactions of PAH (18, 19). Hence, extrapolation of the laboratory results to our system with real soot particles has major limitations. In a study performed in our laboratory by Lutes et al. (7), glass fiber filters were spiked with the DE-71 mixture and allowed to photodegrade in the presence of sunlight on the filter surface. A yield rate of TBDF of 0.06% over a 4-h period was reported. Again, this rate includes both synthesis as well as destruction reactions for TBDF. These experiments most closely addressed photodecay on a surface and do not represent the particle surface generated in our low-temperature experiments. In addition, pores present on the filter surface may allow diffusion of the PBDPEs into the filter material. This phenomenon will shield some of the PBDPEs from sunlight, thereby preventing their photolysis. In light of the limitations associated with the above photooxidation studies, we explored a kinetic treatment of our TBDF and PeBDF data to gain insights into the confounding reactions of TBDF loss via photoinduced processes and TBDF production from PBDPEs. We started with HxBDPE, which debrominates in sunlight to form PeBDPE and PeBDF (see reaction 1 in Table 2). PeBDPE reacts in light (reaction 2) to form TBDF and PeBDPE products. PeBDF photodecays to TBDF in reaction 3, and TBDF further photodecays in reaction 4. Reactions were assumed to be first order, and the units of analyte concentration were in ng/mg of particles. These units implicitly account for particle losses in the chamber due to dilution or chamber walls. Similar results are obtained when units of ng/m3 or molecules/m3 are used. A particle loss rate constant, however, developed from the particle concentration decay over time must be used in the case when concentration is in the units of ng/m3. Brominated dioxin, furan, and ether loss to the vapor phase are negligible because of the extremely low vapor pressures
-
Scheme 1
Slow
Fast Br
OtherPBDFs
Further Debromination Table 2. Reaction Mechanism Used
+
---
To Model PeBDF a n d T B D F Formation
+ +
kl = 0.003 min-l TSRa kz = 0.0075 min-1 TSRa ks = 0.002 min-l TSRa k d = 0.002 min-' TSRa
HxBDPE light 0.9PeBDF O.1PeBDPE PeBDPE + light 0.3TBDF products2 PeBDF + light TBDF TBDF + light products4 a
(1) (2) (3) (4)
TSR in cal cm-2 min-1.
of these compounds. A chemical kinetics solver developed by Jeffries (20)was used to simultaneously solve the simple kinetic differential equations which result from reactions 1-4. Each of the rate constants was keyed to the measured total solar radiation (TSR). The exact value of each rate constant (kl-k4) at a given time was obtained in the model by multiplying its value given above by the solar radiation (in cal cm-2min-l) measured at that time in the experiment. Rate constants for reactions 1and 2 were obtained from the May 21,1992, experiment for which complete time vs concentration data were obtained for PeBDPE and HxBDPE (see Figure 4). Estimates of the yield of TBDF from PeBDPE (reaction 2) and the photodecay of PeBDF and TBDF (reactions 3 and 4) were based on best-fit simulations of the May 21, 1992, experiment. A comparison between the data and model-fit is shown in Figure 4. A muchlower rate constant for the photodecay of TBDF and PeBDF was used than experimentally observed for TBDD (Figure 3). This was needed to model the observed buildup of TBDF and PeBDF in Figure 4. In addition, a significant photolysis yield of TBDF (30%) from PeDBPEs was necessary to fit the data. If a higher TBDF photodecay was used than in reaction 4, an even greater yield from PeBDPE would be needed. To fit the PeBDF data, we had to assume that almost all of the PeBDPE decomposed to PeBDF. These modeling assumptions highlight the need for further characterization of the reactions involving the decay of PBDPEs into PBDFs. In spite of these limitations, it appears that as low temperature combustion particles age in sunlight and PBDPEs are consumed, a net photodecay process of TBDF and PeBDF will ultimately follow. PAHs were also quantified from one of the lowtemperature experiments conducted on October 3,1992. Unlike the high-temperature experiment, BaA and chrysene were not detected and the concentrations of phenanthrene, fluoranthene, and pyrene were lower. This observation highlights the importance of combustion conditions on the formation of PAHs. It has been found
0
1
2 TIME SINCE INJECTION (h)
3
4
-T- PHENANTHRENE -W FLUORANTHENE t- PYRENE
Figure 5. PAH behavior in sunlight on particles derived from iowtemperature combustion (Oct 3, 199 1 experiment).
that the optimum reaction temperature for PAH formation from wood combustion is around 800 "C (21). This fact can account for the higher PAH concentrations in hightemperature experiments and the absence of BaA and chrysene in low-temperature experiments. Plots of phenanthrene, fluoranthene, and pyrene are included in Figure 5. Significant negative slopes were obtained at a 90% level of confidencefrom these curves. Half-lives were in the range of 1.5-2 h. The similar values of half-lives for TBDD and PAHs indicate that photodegradation may be very dependent on the physical and chemical nature of the substrate. The substrate in question is the surface of combustion-generated particles. The following paragraphs delve into measurements related to particle characteristics. Comparison of Particle Characteristics. Characterization of particle size distributions was done by two different instruments, EAA and LAS. Table 3 summarizes the data gathered from these measurements. Calculations Environ. Sci. Technol., Vol. 28, No. 8. 1994 1441
_I_-_______-
Table 3. Comparison of Particle Size Distributions= experiment device TPM (mg) GMD (pm) GSD MMD (rm) SMD (pm) SSA (m2/s) Nov 5 (H) east EAA 23.55 0.084 1.756 0.256 0.186 12.43 Nov 5 (H) east LAS 23.55 0.230 1.448 0.372 0.324 12.26 May 21 (L) EAA 234.46 0.095 1.708 0.259 0.194 13.06 May 21 (L) aerosol LAS 234.46 0.215 1.398 0.319 0.285 15.03 Oct 3 (L) west EAA 353.97 0.127 1.515 0.233 0.196 18.23 Feb 17 (H) west EAA 103.24 0.067 1.845 0.250 0.172 11.33 Feb 17 (L) east EAA 78.49 0.097 1.741 0.285 0.210 11.38 a Abbreviations: H, high-temperature experiment; L, low-temperature experiment; TPM, total particulate mass present in the chamber at the start of the experiment; GMD, geometric mean diameter (d,) = MMD/exp(3.5 In ug);GSD, geometric standard deviation (ug);MMD, mass mean diameter; SMD, surface mean diameter = d, exp(2.5 ln2 ug);SSA, specific surface area = total surface area/TPM.
performed on the data gathered suggested that the particles were log-normally distributed, as least-square regression analysis revealed R2 values greater than 0.95 for a lognormal fit. Table 3 also includes data from lowtemperature and a high-temperature experiments conducted on February 17, 1993. These experiments were conducted solely for the characterization of particle characteristics. Filter samples collected from these experiments were utilized for POM analysis. From Table 3 it is evident that there is not a noticeable difference in particle size distribution obtained from the two kinds of experiments. Higher specific surface area (SSA) has been cited as one of the reasons to explain different rates of photodegradation for PAWS adsorbed on surfaces like carbon black, fly ash, silica gel, and alumina (22). Significantly higher rates of degradation were observed for PAHs like BaA and benzo[alpyrene adsorbed on silica gel or alumina when compared to those adsorbed on carbon black or fly ash. Silica gel and alumina have significantly higher SSAs than fly ash or carbon black. However, in our experiments there was not a significant difference in particles generated from high-temperature and low-temperature experiments (Table 3). Thus, a higher SSA can be precluded as a plausible explanation for higher rates of photodegradation observed in lowtemperature experiments. However,examination of §EM photographs indicated a discernible difference in particle morphology for the low-temperature and high-temperature experiments. The former is characterized by singular and spherical particles, while the latter distinguished itself by the presence of coagulated masses. POM analysis indicated the presence of more organic matter on low-temperatureparticles. The averagepercent extractable from low-temperature experiment particulate samples was 50% as opposed to only 30% from hightemperature filter samples. This observation can be attributed to more efficient combustion at higher temperatures leading to a lower concentration of products of incomplete combustion. At this stage, an interesting parallel can be drawn between solid-phase versus solutionphase laboratory experiments and high-temperature versus low-temperature chamber experiments. The analogy from the laboratory experiments suggests that the presence of an organic layer around low-temperature particles may expedite photodegradation as in the case of solution-phase laboratory experiments (19). This hypothesis is illustrated in Figure 6. The organic layer may be comprised of unburned PUF that manifested itself in the form of a solution around the particles. The analytes may be dissolved in this layer, thus mimicking the conditions of a solution-phase laboratory experiment. Research on wood soot and diesel soot particles has revealed that POM 1442
Environ. Sci. Technol., Vol. 20, No. 8, 1994
"HIGH TEMPERATURE" PARTICLE PBDD, PBDF etc, molecules adsorbed on the aerosol core Aerosol Core
"LOW TEMPERATURE" PARTICLE Or anic la er with P A D S , PbDFs,etc, dissolved Aerosol Core PBDD, PBDF,etc, molecules
FlyAsh ;e '. Growth by Condensation Reaction Products
Ooo
00 Z C O
Growth by Coagulation
Figure 7. Schematic of the burner showing particle formation and growth mechanisms. (SR is the saturation ratio.)
comprises 80-90% of the former compared to only 3040% of the latter (14, 18). PAHs have been found to photodegrade faster on wood soot particles as compared to diesel soot particles (10). Our hypothesis is also espoused by the particle morphology observed from SEM photographs. SEM photographs in correlation with a model for dynamics of particle formation in incinerators provided a valuable insight into the mechanisms for particle formation and growth in our combustion experiments (23). According to this model (Figure 7), particle formation at high temperatures is dominated by coagulation of fly ash and reaction products. However, at low combustion temperatures, particle formation and growth are controlled by nucleation and condensation. The mechanisms discussed above are consistent with the particle morphology observed for high-temperature and low-temperature experiments. This suggeststhat products from PUF combustion form condensation nuclei and may
account for the presence of an organic layer around lowtemperature particles. In high-temperature experiments, condensation of unburned PUF may not be possible as it is expected that most of the PUF is oxidized in the flame zone. Summary and Conclusions
The current study has demonstrated the effect of combustion temperature on the atmospheric stability of PBDDs and PBDFs produced from the combustion of polyurethane foam (PUF) containing PBDPEs. Experiments with combustion temperatures in the range of 400470 "C were categorized as low-temperature experiments and those in the range of 670-780 OC as high-temperature experiments. Particulate samples were collected over time to ascertain the atmospheric stability of PBDDs and PBDFs, since these compounds have been found to occur predominantly in the particulate phase. Particulatebound PAHs were also monitored to determine if their behavior was similar to or different from the behavior of PBDDs and PBDFs. The high-temperature experiments evinced no photodegradation of the compounds monitored. Production of PBDFs, namely, tetrabrominated dibenzofuran (TBDF) and pentabrominated dibenzofuran (PeBDF),and decay of tetrabrominatedp-dioxin (TBDD) and certain PAHs were observed in low-temperature experiments. Production of TBDF and PeBDF appears to have occurredfrom the photolysis of unburned PBDPEs. It is believed that the observed behavior of TBDD and PAHs may be attributable to the differences in physical and chemical properties of the particles generated from high-temperature and low-temperature experiments, respectively. The present study demonstrates that, under ideal operating conditions, incinerators may lead to particulatebound air emissions, which can be transported over long distances due to the stable nature of the emissions. It was observed that lower combustion temperatures can lead to the formation of particulate-bound toxic air pollutants like PBDDs, PBDFs and PAHs that may be relatively unstable when exposed to sunlight. Previous work has shown these same classes of compounds to be photodecayresistant on particles derived from high-combustion temperatures. Particle characteristics can play an important role in photodegradation of the compounds adsorbed on them. The dependence of atmospheric rates of production and decay of incineration-generated air pollutants on the combustion temperature should be taken into account for accurate prediction of biological exposure. Acknowledgments
We want to thank Tom Merz for his help in conducting combustion experiments, G. Dean Marbury for assistance in mass spectrometry analyses, and Dr. Bob Bagnell for assistance in scanning electron microscopy. We also wish to thank Dr. M. Judith Charles for her helpful discussions. Funding for this project was received from the Office of
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Abstract published in Advance ACS Abstracts, May 15, 1994.
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