Environ. Sci. Technol. 1996, 30, 1358-1364
Photostability of Nitro-Polycyclic Aromatic Hydrocarbons on Combustion Soot Particles in Sunlight ZHIHUA FAN,† RICHARD M. KAMENS,* JIANXIN HU,‡ JIANBO ZHANG,‡ AND STEPHEN MCDOW Department of Environmental Sciences and Engineering, University of North Carolina, Chapel Hill, North Carolina 27599-7400
Little is known about the stability of nitro-polycyclic aromatic hydrocarbons (NPAH) on atmospheric aerosols. In this study, the photostability of particleassociated NPAH was investigated under natural sunlight. Deuterated and native NPAH along with diesel exhaust or wood smoke particles were added to a 190m3 outdoor smog chamber and permitted to age under sunlight in cold and warm temperatures. Ozone (O3), nitrogen oxides (NOx), and volatile hydrocarbons in the gas phase were monitored. A sampling train consisting of an annular denuder filter plus another denuder was used for the collection of gas- and particlephase PAH and NPAH. Rapid degradation of deuterated and native NPAH was observed in sunlight, over a temperature range of -19 to +38 °C. Deuterated 1-nitropyrene (d9-1NP) displayed the same behavior as native 1-nitropyrene (1NP), which indicated that it was reasonable to use deuterated NPAH as substitutes for native NPAH. The photolysis rate of NPAH was referenced to the NO2 photolysis rate in order to relate the observed decay of NPAH to the changing solar radiation. To model the decay of NPAH on diesel particles, an average rate constant of kNPAH ) (0.04 ( 0.01) × kNO2 was used for nitropyrenes (NPs), and a average rate of kNPAH ) (0.025 ( 0.005) × kNO2 was needed to model the behavior of nitrofluoranthenes (NFs). A higher rate, kNPAH ) (0.050 ( 0.005) × kNO2, was needed to model the decay of NFs and NPs decay on wood smoke. A photolysis rate of NO2 (kNO2 ) 8.3 × 10-3 s-1) at noon on June 15, 1994, gave half-lives of 0.8 h for 1NP and 2-nitropyrene (2NP) and 1.2 h for 2-nitrofluoranthene (2NF), d9-3-, and d98-nitrofluoranthene (d9-3NF and d9-8NF) on diesel soot particles. The half-life was 0.5 h for d9-1NP, d93NF, and d9-8NF on wood soot particles. These results showed that photodecay was the main loss pathway for NPAH on diesel soot and wood smoke and that photodecomposition of NPAH was dependent on the solar radiation and the chemical and physical
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properties of the substrates. The effect of temperature was not significant.
Introduction Nitro-polycyclic aromatic hydrocarbons (NPAH) are widespread pollutants in both primary combustion emissions and ambient air particulate matter (1-9). NPAH can be formed during combustion processes or by the reactions of polycyclic aromatic hydrocarbons (PAH) with hydroxyl radicals (OH), nitrate (NO3) radicals, and dinitrogen pentoxide (N2O5) present in the polluted atmosphere (3-5, 7-10). The nitroarene isomers from atmospheric reactions, such as 2-nitrofluoranthene (2NF) and 2-nitropyrene (2NP), are different from those emitted directly, such as 1-nitropyrene (1NP) (3, 8-10). A number of studies have suggested that the reaction of OH radicals with gaseous PAH and the subsequent addition of NO2 to the PAH-OH adduct is the major formation process for atmospheric NPAH (3-5, 7-10). Many NPAH are more mutagenic than their parent PAH in the Salmonella bacterial mutagenicity assay, and several have shown carcinogenic potential in animals as well (1114). Compared to our knowledge of PAH reactions, the chemical transformations of NPAH have only been studied to a limited extent. One probable environmental fate of certain NPAH is their photodecomposition. The photolysis of gaseous nitronaphthalenes was studied by Atkinson et al. (15), and they concluded that the photolysis pathway was the dominant loss process for gaseous NPAH in the atmosphere. A number of studies of NPAH photodegradation in solution or coated on substrates have been conducted in the past several years (16-20), and it has been shown that some NPAH can be readily decomposed when exposed to light both in solution or on particles. The stability of NPAH on fresh diesel soot particles has also been investigated in an outdoor smog chamber (10, 21). Particle-bound 6-nitrobenzo[a]pyrene (6N-BaP), 1NP, and 2NP decayed rapidly in sunlight. In contrast, NPAH associated with aerosolized diesel soot (U.S. National Institute of Standards and Technology SRM 1650) did not decay when exposed to sunlight (21); 1NP was resistant to photodecomposition when adsorbed on coal fly ash (22); and 2NP in methanol was very stable when exposed to 6 Dura Test Vita lights (>300 nm) (19). In the studies described above, compounds were dissolved in organic solvents(16, 18-20) or coated onto a variety of substrates such as silica (19) and glass plates (17) or indigenously associated with particles. These inconsistencies regarding the photostability of NPAH have suggested that photodecomposition of NPAH was strongly influenced by the physical and chemical nature of the substrates in which NPAH were sorbed. * Author to whom correspondence should be addressed; telephone: (919) 966-5452; fax: (919) 966-7911; e-mail address: Kamens@ sophia.sph.unc.edu. † Present address: Research Triangle Institute, P.O. Box 12194, 3040 Cornwallis Road, Research Triangle Park, NC 27709. ‡ Present address: Center of Environmental Science, Beijing University, 100871 People’s Republic of China.
0013-936X/96/0930-1358$12.00/0
1996 American Chemical Society
SCHEME 1
Photochemical Pathway of 1-Nitropyrene (20)
In previous NPAH studies with fresh diesel soot particles (10), the behavior of NPAH was modeled (10, 23) by adding PAH and NPAH reactions to a photochemical smog mechanism (24). Since no data were available for the photodegradation rates of 2NF and 2NP on diesel soot particles, the photolysis rates of 2NF and 2NP were based on the photodecay rate of 1NP (10, 23), and an uncertainty of at least a factor of 2 existed. In this work, we conducted chamber experiments to investigate NPAH photodegradation on diesel soot and wood smoke particles in the presence of sunlight with different photochemical conditions and temperatures and reduced the uncertainty associated with NPAH rate constants.
Experimental Section Outdoor chamber experiments were conducted in a 190m3 Teflon film smog chamber located in Pittsboro, NC, (10, 23). Both diesel exhaust and wood smoke were used as the particle source for the studies of NPAH photodegradation. Diesel exhaust from a 1967 Mercedes sedan (200D) engine was added to the smog chamber for 1-2 min. Wood emissions were added directly from the chimney of a wood stove to the large chamber for 3-4 min in a manner described in previous studies (25-27). After the addition of diesel emissions or wood smoke, deuterated NPAH were vaporized to the chamber via a hot injector for a period of 5-10 min. The resulting system was permitted to age in the presence of sunlight for 5-7 h. The diesel car was operated in the heavy load condition during the chamber injection process to create a large amount of particles. The particle concentration that resulted in the chamber was in the range of 500-1000 µg/m3, and about 0.4 ppm nitrogen monoxide (NO) and 0.1-0.2 ppm nitrogen dioxide (NO2) were generated from the combustion. The measurements of NO, NO2, ozone (O3), and other parameters such as particle size, temperature, solar radiation, etc. are described in detail elsewhere (25-27). The sampling train consisted of an upstream 40-cm fivechannel annular denuder to collect gas-phase semivolatiles (i.d. ) ∼2.4 cm with each annulus separated by a space of 0.1 cm, University Research Glassware, Carrboro, NC) followed by a 47-mm Teflon-impregnated glass fiber filter (type T60A20, Pallflex Products Corp., Putnam, CT) and another 40-cm five-channel denuder. The denuder walls were coated with ground XAD-4 particles (Supelco, Bellefonte, PA) with diameters smaller than 40 µm (10, 23, 27). After taking the samples, the denuder was extracted with 30-50 mL of 2:1 (v/v) hexane and acetone (Fisher optima grade) four times in the field. The denuder was then dried
with pure nitrogen gas and reused. Filter samples were Soxhlet extracted in dichloromethane (MeCl2) for 12-16 h and fractionated via normal-phase high-pressure liquid chromatography (25-27). The fractions between 5-19 min (PAH) and 23-32 min (NPAH) were collected, concentrated to 10-100 µL with rotary evaporation and a nitrogen stream, and then analyzed on a Hewlett Packard 5891A/5890II gas chromatography/mass spectrometer (GC/MS). Identities were referenced to authentic standards of PAH (SRM 1647b) and NPAH (SRM 1587) available from the U.S. National Institute of Standards and Technology. The precision for PAH was generally less than (12.5% (10), and for NPAH it was less than (15% ((relative standard deviation from nine observations) through the entire workup procedure (23). The detailed coating technique by Gundel and co-workers (28), denuder characterization, and description of the sample workup procedure are reported in previous studies (10, 25-27). Synthesis of Deuterated NPAH. To investigate the photodecay of NPAH, a technique was needed that could distinguish the production and formation processes of NPAH. NPAH can be formed by gas- and particle-phase PAH reactions from diesel emissions under photochemical conditions; thus, the use of deuterated NPAH was explored to investigate the decay process. Deuterated nitropyrenes (d9-NPs) and nitrofluoranthenes (d9-NFs) were prepared (29) by direct nitration of deuterated PY (d10-PY) and deuterated FL (d10-FL) with ammonium nitrate (NH4NO3) and trifluoroacetic anhydride (TFAA) in acetonitrile. Based on electrophilic substitution, a nitro group was introduced to d10-PY or d10-FL, and d9-1NP and d9-1-, d9-3-, d9-7-, and d9-8-NF were obtained. Scheme 1 shows the diagram of the synthesis of d9-8NF. After synthesis, the mixture was separated by an open bed silica column with a 9:1 hexane/ MeCl2 mixture and then separated with a 97:3 petroleum ether/MeCl2 mixture. The purity of these compounds was determined by GC/MS (Figure 1), and >98% purity was obtained for d9-1NP and d9-1-, -3-, -7-, and -8-NF. d9-1NP, d9-3NF, and d9-8NF were used to investigate photodecay of NPAH; d9-7NF was used as an internal standard for NPAH quantitation. Introduction of Compounds to the Chamber. NPAH with four or more fused-aromatic rings are generally associated with the particle phase under ambient conditions (2-10); therefore, we attempted to introduce NPAH to the chamber by vaporizing pure NPAH with a hot injector and have them condense on diesel soot particles. This simulates their formation in the gas phase and rapid migration to particles. Their decay in sunlight was then monitored.
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FIGURE 2. Migration of d12-BaA and BaP from gas phase to particles.
FIGURE 1. Chromatogram of d9-8NF and 2NP and mass spectrum of d9-8NF. Analysis was on a J&W 30 m, 0.25 µm film thickness, 0.32 mm i.d. DB5, GC oven started at 110 °C, held for 0.1 min, and programmed to 250 °C at 10 °C/min and then at 6 °C/min to 300 °C and held for 2 min.
Because NPAH are more mutagenic than PAH, the feasibility of the hot injector technique was first tested with PAH. Deuterated pyrene (d10-PY), benzo[a]pyrene (BaP), and deuterated benz[a]anthracene (d12-BaA) were used as target compounds to investigate the migration of PAH from the gas phase to diesel soot particles. Two chamber experiments were conducted under darkness on July 1 and September 6, 1993. On the evening of July 1, 1993, 1 mg of d10-PY (the pure solid vapor pressure of PY is 4.5 × 10-6 Torr at 25 °C; see ref 30) was placed in a U-tube, vaporized via a hot injector (200 °C), and carried into the chamber with an air stream. The gas phase d10-PY in the chamber was 344 ng/m3. The wall loss of gaseous d10-PY was monitored for 2 h before the addition of diesel emissions. Diesel exhaust was then added to the chamber for 5 min, and the resulting particle concentration in the chamber was 929 µg/m3. Temperatures in the chamber ranged from 24 to 22 °C. Rapid migration of d10-PY from the gas to the particle phase was observed, and 120 ng/m3 d10-PY was resulted in the particle phase. Migration to the particle phase was further tested with heavy molecular weight PAH. To prevent the rapid loss of gas-phase PAH to the chamber walls, in the September 1993 experiment, diesel exhaust was added to the chamber for 5 min before the addition of the NPAH, and 810 µg/m3 of particulate matter appeared. After the addition of diesel emissions, 1 mg each of solid BaP and d12-BaA (vapor pressures are 5.6 × 10-9 and 2.1 × 10-7 Torr at 25 °C, respectively) was vaporized via a hot injector (200 °C) by the method described above; the injection time took 5 min. During this period, the diesel emissions and the PAH were mixed with the three chamber fans and two extra floor fans located inside the chamber. The volumetric flow rate of the fan blowing over the hot injector was 120 m3/min. Hence, during injection approximately 3 chamber volumes
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crossed over the top of the injector, providing more than ample opportunity for even coating of the particles. The chamber temperature ranged between 20 and 22 °C during the experiment. The initial chamber particle concentration of BaP was 635 ng/m3 and 455 ng/m3 for d12-BaA in the chamber. d12-BaA and -BaP transferred from the hot flowing air steam of the hot injection to chamber diesel particles rapidly and reached equilibrium immediately (Figure 2). In order to further understand the mechanism of NPAH photooxidation, irradiation of NPAH was conducted in a photochemical turntable reactor (ACE Glassware, Vineland, NJ) in our laboratory (31-33). A 450-W medium pressure mercury arc lamp placed in a borosilicate immersion well was used as a light source. Under these conditions, only light above 300 nm was transmitted through the borosilicate. The temperature was maintained at 16.5 ( 0.5 °C by circulating water through exterior copper coils in an ice bath (31-33). A 0.2-0.4-mg sample of each compound described above was dissolved in 4 mL of hexane (Fisher optima grade) in 13 mm × 100 mm quartz reaction tubes. The concentration was ∼2 × 10-4 M, and the solution was irradiated for 4 h. A total of 10 µL of solution was removed every 15 min during the first hour, and 1 µL of internal standard was added prior to the analysis. Later, aliquots were removed every 30 min. Quantitative analysis was carried out by the GC/MS procedure described earlier.
Results and Discussion Photodecomposition of Particle-Associated NPAH under Low Photochemical Conditions. Outdoor chamber experiments of NPAH photodecomposition on diesel particles were conducted under cold and warm temperatures. A mixture of d9-1NP, d9-3NF, and d9-8NF (0.5 mg of each compound) was vaporized into the chamber for 5-7 min with the hot injector after the addition of diesel exhaust; about 100-300 ng/m3 d9-NPAH appeared on particles in the chamber. These compounds were then aged in sunlight for 5-7 h. Temperatures ranged from -18.6 to -5 °C in the January 1994 experiment and from 13 to 38 °C in the April 1994 experiment. The NOx concentration was 0.83 ppm and the O3 was 0.021 ppm in the first experiment, and 1.38 ppm NOx and 0.001 ppm of O3 in the second experiment. Little O3 formed in the chamber because of the low amount of volatile hydrocarbons in the diesel emissions (Figure 3). Rapid decay of d9-1NP, d9-3NF and d9-8NF was observed in sunlight even during cold temperatures (temperatures on January 16, 1994, ranged between -18.6 and -5.6 °C).
FIGURE 3. O3, NO, and NO2 model simulation results and data for daytime diesel soot experiment conducted on January 16, 1994. Dashed lines are data, and solid lines are model simulations.
The half-life of NFs and NPs observed from this experiment was 2 h, and the photodegradation rate followed pseudofirst-order kinetics. d9-1NP displayed the same behavior as the native 1NP. This suggests that it is reasonable to use deuterated NPAH as substitutes for native NPAH. Like NO2, NPAH absorb in the UV and the visible region. The photolysis rate of NPAH was referenced to the NO2 photolysis rate to relate the observed decay of NPAH to the changing solar radiation. The degradation of NPAH in the cold chamber experiment was modeled with the rate constant used in previous studies (10, 23; kNPAH ) 0.03kNO2). The photolysis rate constant of NO2 (kNO2) can be calculated using
kNO2 (s-1) )
∫
λi
λ)290nm
σ(λ)φ(λ)J(λ) dλ
FIGURE 4. Simulation (solid line) and data (symbols) of 1NP and d9-1NP decay on diesel soot in sunlight (k1NP ) 0.030 ( 0.005kNO2, April 27, 1994). The photolysis rate of PAH was referenced to that of NO2 in order to relate the observed decay of NPAH to the changing solar irradiation.
(1)
where σ(λ) is the absorption cross section base e of NO2 in cm2 molecule-1 averaged over a wavelength interval ∆λ, centered at λ; φ(λ) is the primary quantum yield for the loss of NO2 averaged over a wavelength interval ∆λ, centered at λ; J(λ) is the actinic flux in photons cm2 s-1 summed over the wavelength interval ∆λ, centered at λ, at solar zenith angle θ, corrected for season and latitude. If desired, corrections for altitude and surface albedo can be included. The cross section, primary quantum yield, and actinic flux of NO2 are available in the scientific literature (34, 35). The photolysis rate at a given time for NO2 was calculated from the solar radiation intensity inside the chamber using an in-chamber actinic flux simulation program (36). Reasonable fits were obtained for both 1NP and d9-1NP (not shown here). To get the best fits, model rate constants of kNPAH ) (0.025 ( 0.005)kNO2 were used to predict the loss of d9-3NF and d9-8NF (not shown here). In the second experiment, 2NF and 2NP were vaporized into the chamber along with the other deuterated NPAH. The interfering atmospheric formation processes of 2NF and 2NP were minimized by the low initial concentrations of volatile hydrocarbons in the chamber. By adding a high concentration of 2NF and 2NP to the chamber, it was therefore possible to overwhelm the small confounding formation process of 2NF and 2NP under relatively inactive photochemical conditions. As shown in Figure 3, almost no O3 was formed in the chamber during the reaction process, and the predicted maximum concentration of OH radical was less than 10-7 ppm. Under these conditions, the concentration of 2NF and 2NP predicted by modeling
FIGURE 5. Simulation (lines) and data (symbols) of 2NF and 2NP decay on diesel soot in sunlight (April 27, 1994).
the chamber system (10, 23) was only 10-7 ppm, which was 3 orders of magnitude lower than the 2NF and the 2NP added to the chamber. As observed in the experiment above, a rapid degradation of d9-1NP, d9-3NF, and d9-8NF occurred (Figure 4), and more than half of these deuterated NPAH photodecayed within 1.5 h. 2NF and 2NP also decayed rapidly; the decay rate (k ) 0.025kNO2 for 2NF and 0.03kNO2 for 2NP) was not significantly different from that of 1NP on diesel particles (Figure 5). This is close to values used in previous studies (10, 23) of (0.03 ( 0.01)kNO2 for 2NF, but these were based on model fits with an assigned uncertainty of a factor of 2. In the past, we have used a value of 0.045kNO2 for 2NP based on modeling studies (10, 23). This appears too high when compared to the direct measurements in this study. In both experiments reported here, rapid decay of particle FL and NP was observed. Since O3 was very low in both experiments, photodegradation was the major factor influencing the decay of particle PAH as well. To model the decay of particle PY and FL for all the chamber experiments conducted, the photooxidation rate constants were (0.04 ( 0.01)kNO2 and (0.020 ( 0.005)kNO2 for particle PY and particle FL, respectively. Photodecomposition of Particle-Associated NPAH under Active Photochemical Conditions. In order to further test NPAH photodecomposition under different conditions, two daytime outdoor smog chamber experiments were
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FIGURE 6. Simulation (lines) and data (symbols) of d9-1NP, d9-3NF, and d9-8NF decay on wood smoke in sunlight (kNP ) 0.050 ( 0.005kNO2, September 30, 1994).
conducted with diluted diesel exhaust on June 15, 1994, and wood smoke on September 30, 1994. Photochemical conditions were created by adding extra NO and propylene to the chamber after the addition of diesel soot or wood soot. The temperature ranged from 15 to 38 °C and from 10 to 32 °C in the June and September experiments, respectively. The particle concentration started at 1720 and 1530 µg/m3. The initial NOx concentration was 1.48 ppm, and the maximum O3 concentration was 0.99 ppm at 14:35 P.M. in the June experiment; and the initial NOx concentration was 1.25 ppm, and the maximum O3 concentration was 0.95 ppm at 14:07 P.M. in the September experiment. d9-NPAH decayed rapidly in both chamber experiments (deuterated NPAH decay in the September 30, 1994, experiment is given in Figure 6). The decay rate constants on diesel soot in the June 15 experiment under active photochemical smog conditions (high O3 concentration) were not significantly different from those under low photochemical conditions (low O3 concentration) (Table 1). However, a more rapid degradation of NPAH on wood soot compared to diesel particles was observed; approximately 90% of the NPAH degraded after 1.5 h. This required a higher rate constant (kNPAH ) (0.050 ( 0.005) × kNO2) to fit the decay rate of NPAH on wood smoke compared to kNPAH ) (0.02 to 0.035)kNO2 on diesel soot. Again, the same decay rate was observed for all NPAH compounds on wood smoke. These results strongly suggest that different substrates can strongly influence the photostability of NPAH. Wood smoke tends to have less elemental carbon and more extractables than diesel exhaust particles (26, 31, 37). In the September 30 experiment, the organic extractable portion from the wood smoke samples was approximately 85-100% while the June 15 experiment with diesel soot had only 30-40% extractables. Compared to diesel soot or coal fly ash (10, 22), NPAH on wood soot particles decayed more rapidly. The organic material associated with wood smoke is much more polar than diesel soot organic matter (31); this results from the high alkane content of diesel soot vs the abundance of methoxyphenols and methoxybenzaldehydes associated with wood smoke (31, 38). Previous studies in our group have shown that methoxyphenols promoted PAH photodecay (31-33); therefore, it is possible that the organics initiate or participate in reactions leading to the decay of NPAH or that NPAH are more mobile in
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wood smoke and less shielded from sunlight, and direct photolysis of NPAH is possible. A linear relationship was obtained when we plotted observed decay rate constants of NPAH versus the average solar radiation from all of our experiments. The results have shown that photodecay was the main loss pathway for NPAH associated with diesel soot particles and wood smoke. A photolysis rate of NO2 (kNO2 ) 8.3 × 10-3 s-1) at noon on June 15, 1994, gave half-lives of 0.8 h for 1NP and 2NP, and 1.2 h for 2NF, d9-3NF, and d9-8NF for diesel soot particles. It was about 0.5 h for d9-1NP, d9-3NF, and d98NF for wood soot particles. The above results indicated that the photodecomposition of NPAH was dependent on the solar radiation and the chemical and physical properties of the substrates, and there was no significant effect of temperature on NPAH photodegradation (Table 1). Influence of Matrix and Substrate. Alkanes constitute a significant portion of the organic extractables of diesel exhaust particles (39). To test the reactivity of NPAH in an alkane solution, irradiation of 3NF, 8NF, 2NP, 2NF, and d9-1NP in hexane (16.5 °C) was carried out in a photochemical turntable reactor (31-33). Fast decay of d9-1NP occurred, slight decay of 3NF and 2NF was observed, and almost no degradation was observed for 8NF or 2NP (Figure 7) even though each of these compounds has strong absorption at the intense emission lines of 313 and 334 of a medium Hg arc lamp. As mentioned above, all the compounds photodecayed at almost the same rate on wood soot particles and at slightly different rates on diesel soot. These results suggest that the solution phase and the particle matrix can influence NPAH decay. Other studies have shown that the photostability of PAH and NPAH depends on the physical and chemical properties of the substrate (such as the carbon content), the color of the substrate, the sorptivity of the ash particle size fraction, and the mineral contents (which can quench the excited states of adsorbate molecules) such as metal ions of the sorbent (22, 40-46). It has been suggested that the relative quantity of the carbon substrate is the main factor determining the photoreactivity of the compounds adsorbed on it (44-46). Carbonaceous coal fly ash particles have a great affinity for PAH (43-46) and an even greater affinity for NPAH (22). Adsorbed compounds are possibly deposited in the pores of the substrate and shielded from incident light by the opaque particulate matrix (43-46). Highly colored coal fly ash can also absorb incident light, which reduces the amount received by adsorbed compounds (43, 44). In contrast, diesel exhaust consists of more extractables and less carbon than coal fly ash, and it is very likely that there is a surrounding liquid layer on diesel soot (31). Thus, the affinity of diesel soot to an adsorbed compound may not be as strong as that of coal fly ash. Furthermore, the geometric mean count diameter of the diesel soot from our chamber experiments was 0.15 µm, which was much smaller than that of coal fly ash (22, 43). This provided diesel soot with much more exposed reaction surface area than coal fly. NPAH photooxidation in liquid hexane represents photooxidation in a “simple” photochemical environment. Under these conditions, the structure of the compounds, as will be discussed in the next section, is the main factor determining the photoreactivity of adsorbed compounds. However, diesel soot particles, with its large portion of organic extractables (30-40%) involves a much more
TABLE 1
NPAH Photolysis Rate Constants on Diesel and Wood Soot Particlesa date
temp (°C)
1-16-94 4-27-94 6-15-94 9-30-94c
-19 to -6 °C 13 to 38 °C 24 to 38 °C 10 to 32 °C
1NP 0.03 0.03 0.03
d9-1NP
2NP
2NF
d9-3NF
d9-8NF
0.03 0.03 0.03 0.05
NDn
ND 0.025 0.022 0.045
0.022 0.03 0.022 0.05
0.022 0.03 0.022 0.05
0.03 0.035 ND
a NPAH photolysis rate is referred to as a multiplier (A) on the NO photolysis rate constant (k 2 NPAH ) AkNO2). smoke chamber experiment. The others were diesel soot experiments.
FIGURE 7. Comparison of NPAH decay in hexane after irradiation to medium mercury lamp as light source.
complex reaction system (39). Aliphatic and aromatic hydrocarbons can constitute more than half of the extractable organic matter associated with diesel soot (37, 39, 47). Also present are known photosensitizers such as anthraquinone that promote PAH decay. In wood smoke, methoxyphenols perform a similar role (31, 32). Hence, as with PAH (31-33), it is possible that chemical composition controls the mechanism of NPAH photodecay in particulate matter. NPAH Photolysis Mechanisms. Only a few studies are available on NPAH photolysis mechanisms (20, 48-51). The photodecay of 9-nitroanthracene was investigated by Chapman (48), and later the detailed mechanism of 1-, 2-, and 4-nitropyrene photodecay in solution was discussed by Van den Braken-van Leersum (20). Previous studies have suggested that the photoreaction described in Scheme 1 proceeded via the electronically excited singlet state and that both C-N bond scission and intramolecular rearrangement may occur (20, 48, 49). The study from Chapman (48) suggested that the major photochemical transformation of NPAH was an intramolecular nitro to nitrite rearrangement following an n-π* transition. Detection of several types of oxyl radicals produced by photolysis of nitro compounds (50) strongly supports this mechanism. Pitts has suggested (51) that the different photostabilities of NPAH are dependent upon the number of hydrogens peri to the nitro group. It has been shown that 6N-BaP with two peri-hydrogens is photochemically less stable than 1- or 3N-BaP with 1 perihydrogen (17, 18, 51). In addition, the nitro group in 2NP is not hindered by a peri proton as in 1NP, and the nitro group is expected to be in the plane of the aromatic π-system (19, 20). The lack of electronic interaction between the nitro group and the aromatic π-system makes 2NP photochemically stable toward light (20). The final products detected are quinones, phenols, and nitrohydroxy compounds from previous studies (16-20). In our solution
b
Not detected (ND). c Daytime wood
studies, 1NP and d9-1NP decayed faster than 2NP, which was consistent with the above theory and with the idea that the structure of the compound is the dominant factor for NPAH decay. On diesel particles, there probably were some other species acting as radical transfer media, and the radical pair formed by either C-N or N-O bond scission may recombine in solution but might not do so on particles. Also, some other reactions possibly occurred on the particles, which all eventually caused NPAH decay. In this study, we were unable to determine which mechanism resulted in NPAH decay on particles. We attempted to identify the photoreaction products of NPAH on diesel soot and wood smoke in order to gain insight into the photodecay mechanism, but the concentrations of the products were too low to be identified by GC/MS. If the photolysis mechanism follows Chapman and Pitts’ theory, d8-1,6- or 1,8-pyrenedione, d8-1,6- or 1,8-fluoranthenedione, or hydroxy PAH may be detected in polar fractions of the samples. Additional studies are needed to identify the photodegradation products of NPAH to further understand their photoreaction mechanism. NPAH, in spite of the structural differences, all showed similar decay rate on diesel soot particles or wood smoke. The results indicate that other chemicals associated with the diesel particles have an important effect on the photooxidation of NPAH in sunlight and that the effect of NPAH structure is not a dominant factor when NPAH are adsorbed on diesel soot or wood smoke.
Acknowledgments We thank Dr. Ramiah Sangaiah for his guidance of the deuterated NPAH synthesis. This work has been funded by a gift in 1995 to the University of North Carolina from the Ford Motor Company. The authors would like to thank Myoseon Jang for help with the photoreactor experiments.
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Received for review August 11, 1995. Revised manuscript received November 6, 1995. Accepted November 7, 1995.X ES9505964 X
Abstract published in Advance ACS Abstracts, January 15, 1996.
