Environ. Sci. Technol. 2002, 36, 4364-4369
Aqueous Photodegradation of Polycyclic Aromatic Hydrocarbons MATTHEW P. FASNACHT AND NEIL V. BLOUGH* Department of Chemistry and Biochemistry, University of Maryland, College Park, Maryland 20742
Photodegradation of 12 polycyclic aromatic hydrocarbons was studied in aerated pure water, solutions of Suwannee River fulvic acid, and natural waters using polychromatic light (>290 nm). Quantum yields in pure water varied from 3.2 × 10-5 to 9.2 × 10-3. No obvious relationships were evident among the quantum yields and molecular properties. Photodegradation rate constants in solutions of Suwannee River fulvic acid or natural waters were largely unchanged compared to rate constants in pure water. Estimates of PAH photodegradation rates in natural waters can thus be obtained employing the quantum yields in pure water, PAH absorption, and solar irradiance. Calculated rate constants for photodegradation in surface waters during the summertime at mid-latitude varied from 3.2 × 10-3 to 7.6 h-1.
Introduction Due to incomplete combustion of fossil fuels, large amounts of polycyclic aromatic hydrocarbons (PAHs) are introduced into the atmosphere daily (1). These compounds can enter the aquatic environment through both atmospheric deposition and watershed runoff. Because PAHs can be carcinogenic as well as toxic to many organisms (1), an understanding of the pathways and rates of their removal from the environment is important. Although several past studies indicated that PAHs could be degraded photochemically in natural waters (2, 3), the mechanisms and magnitude of this degradation are still not thoroughly understood. PAHs may be degraded through either direct or sensitized photochemical reactions (2, 4, 5). Most PAHs can absorb surface solar radiation, allowing for the possibility of direct photodegradation. Indeed, several studies have shown that a number of PAHs are destroyed when irradiated with 313 and 366 nm light in pure water (2, 3). This degradation could be enhanced in natural waters through reactions with intermediates produced photochemically from chromophoric dissolved organic matter (CDOM) (6-9) or other constituents in natural waters (10, 11). Moreover, the binding of some PAHs to dissolved organic matter (DOM) in natural waters is known to cause fluorescence quenching (12-14), which could inhibit or enhance PAH photodegradation depending on the quenching mechanism. Inhibition would occur if the binding to DOM increased the rate of relaxation of the PAH excited singlet state to the ground state without the formation of a reactive intermediate, whereas enhanced degradation could occur if the quenching produced a reactive PAH intermediate. How PAH photodegradation may be affected by constituents of natural waters such as DOM is still largely unknown. * Corresponding author phone: (301)405-0051; fax: (301)314-9121; e-mail:
[email protected]. 4364
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 36, NO. 20, 2002
Although PAH photochemistry in both pure and natural waters has been studied for the last 35 years (2, 3, 15-19), much of this past work either employed irradiation wavelengths outside the surface solar spectrum (17-19) or ignored the possible effects of DOM on PAH photodegradation (15, 16). Two previous papers did examine how the photodegradation of PAHs differed between pure and natural waters, but their conclusions appear contradictory (2, 3). Zepp and Schlotzhauer (2) found the rate of pyrene loss was no different in natural waters compared to pure water, whereas Mill et al. (3) found the loss of benz[a]anthracene and benzo[a]pyrene in creek water was 30% and 50% slower than in pure water, respectively, and concluded that DOM was responsible for this inhibition. In this study, we have measured the kinetics and quantum yields for the photodegradation of a broad suite of PAHs in pure water, solutions of Suwannee River fulvic acid (SRFA), and natural waters using conditions similar to those expected in the environment-light in the solar range and low PAH concentrations (nM). Our goals for this study were 2-fold: (1) to determine how widely the quantum yields for photodegradation in pure water vary for PAHs exhibiting a broad range of structures and molecular properties, ultimately to probe what factors control the direct photodegradation of PAHs in aerated aqueous solutions and (2) to examine whether the photodegradation of PAHs is affected significantly by the presence of organic matter in natural waters. We find that the quantum yields for PAH photodegradation in pure water vary widely, from a low of 3.2 × 10-5 for fluoranthene to a high of 9 × 10-3 for acenaphthene. Rate constants were largely unchanged in natural waters. Thus, as a good approximation, quantum yields in pure water together with the absorption spectra of the PAHs and spectral irradiance can be employed to estimate the rates of loss in natural waters.
Experimental Section Materials. Acenaphthene (AC, 99%), anthracene (AN, 99%), benz[a]anthracene (BA, 99%), benzo[a]pyrene (BP, 97%), benzo[b]fluoranthene (BB, 98%), benzo[k]fluoranthene (BK, 98%), chrysene (CH, 98%), fluoranthene (FL, 99%), fluorene (FU, 99%), perylene (PE, 99.5%), phenanthrene (PH, 99.5%), pyrene (PY, 99%), and sodium hydroxide (99.999%) were obtained from Aldrich and used without further purification. Methanol (MeOH, HPLC grade) and acetonitrile (ACN, HPLC grade) were obtained from Fisher. SRFA was obtained from the International Humic Substance Society (lot # 1S101F). Synthetic air (UHP/zero grade) was obtained from Air Products. Water was obtained from a Millipore Milli-Q system. PAH stock solutions (0.5-10 µM) were prepared in ACN and stored in amber borosilicate vials at room temperature. Solutions of SRFA were prepared by diluting a 100 mg/L stock solution with pure water followed by addition of NaOH to adjust the pH to 6.0 and were then stored in the dark at 4 °C. Surface waters (Table 1) from the mouth of the Chesapeake Bay and Atlantic Ocean were collected and filtered through a Gelman 0.2 µm fluted filter on board the R/V Cape Henlopen during October 1999. Surface water from the mouth of the Choptank River was collected unfiltered in June 2001. All natural water samples were stored in the dark at 4 °C and filtered through a Nalgene 0.2 µm nylon syringe filter before use. Apparatus. Absorption spectra were measured with a Hewlett-Packard 8452A diode array spectrophotometer. Steady-state fluorescence was acquired using an SLM Aminco 10.1021/es025603k CCC: $22.00
2002 American Chemical Society Published on Web 09/18/2002
Series 2 luminescence spectrometer (AB2). The C-18 reverse phase HPLC system with fluorescence detection has been previously described (20). The irradiation system consisted of a 300 W Xe arc lamp housed in an ILC Technology R400-2 and powered by an ILC Technology PS300-1. Before being directed onto a 1.0 cm quartz cuvette, light from the lamp was first passed through 18 cm of water to remove infrared radiation and then directed through a long pass filter, either a Schott WG305, WG320, WG335, WG355, KV370, LF380, or WG405, where the number indicates the wavelength of 50% transmission. In experiments using BP, a 0.5 neutral density filter was placed after the filter to lower the irradiance. Irradiance was measured with a LICOR 1800 UW spectroradiometer (300-800 nm). Total power was checked daily with an IL-1700 radiometer to confirm that total lamp output was consistent among experiments. When using the WG305 filter, the irradiance at the surface of the cuvette was four to six times that of the surface solar irradiance at 400 nm for a midsummer day at 40 °N latitude (21) and was more enriched in the UV-B. Experimental Procedures. SRFA and natural water samples were warmed to room temperature and bubbled with air to ensure air equilibration. Ten to thirty microliters of PAH stock solution were added to 4.0 mL of water sample using a Hamilton gastight syringe. PAH concentrations were measured using either HPLC or steady-state fluorescence. Early PAH experiments employed HPLC using either ACN/ water or MeOH/water mobile phases with fluorometric detection (see Supporting Information Table 1). PAH concentrations were determined from a calibration curve of fluorescence peak area versus concentration of PAH injected onto the HPLC. Because steady-state fluorescence measurements provided results that were indistinguishable from those acquired with the HPLC (see Supporting Information), steady-state fluorescence was used primarily to measure concentrations of PAHs owing to its speed and reproducibility. For each PAH, fluorescence excitation and emission wavelengths were chosen to minimize CDOM background fluorescence (see Supporting Information Table 1). Fluorescence was detected at two emission wavelengths (16 nm band-pass) at the selected excitation wavelength (4 nm bandpass) and were used to acquire two independent measurements of the rate constant, which were averaged. The excitation shutter was opened, and the fluorescence signal at the each emission wavelength was acquired for 3.0 s at a data collection rate of 5 Hz. The resulting 15 fluorescence data points at each wavelength were averaged. After correcting for background fluorescence (eq 1), PAH concentration ([PAH]) was determined (eq 2) from the slope of a calibration curve (m) developed for each PAH in each water sample.
FPAH ) Ft - Fb
(1)
FPAH m
(2)
[PAH] )
Here FPAH is the fluorescence resulting from the PAH, and Ft is the total fluorescence of the solution representing both PAH fluorescence and background (Fb). Background fluorescence was due to both CDOM fluorescence and in some cases Raman scattering. Background fluorescence of CDOM solutions, which decreases with irradiation time (22), was measured by irradiating CDOM solutions containing no PAH using the same procedure. In all cases, calibration curves were linear with PAH concentration, although PAH fluorescence was reduced slightly ( 1). Photodegradation rate constants of many PAHs were unchanged both for solutions of SRFA or natural waters (Table 3). While the rate constant ratios of AC, BA, BP, BK, and PE were less than one and AN was greater than one, these changes were small (most < 40%) and variable between SRFA and natural waters. The origin of the possible weak inhibition of PAH photodegradation in natural waters is unclear. Although light attenuation by CDOM in our experiments can be excluded, binding of PAHs to DOM could be important (12-14). However, only for large PAHs such as PE, BP, BK, and BB, with binding constants (KOC) around 106 (13), would we anticipate significant interactions of PAHs with DOM. This binding could reduce the photodegradation rate constants by decreasing the lifetimes of the excited singlet or triplet states. Thus if this mechanism were operative, we would expect that larger PAHs would show greater rate constant reductions in natural waters; however such behavior was not observed (Table 3). Further, except for PE, only trivial reductions in fluorescence were observed for most PAHs in SRFA solutions or natural waters indicating little interaction (12-14). Finally, varying the concentration of SRFA over 3 orders of magnitude did not change the BA or BP photodegradation rate constants in a way that is consistent with this mechanism of inhibition (Table 4). Of all the PAHs examined, only the degradation of AN appeared to be sensitized in natural waters. However even in this case, the mechanism is unclear. Photochemically generated hydroxyl radical (7, 10, 30), singlet dioxygen (31), and other reactive intermediates (6) are capable of reacting with PAHs (1), but due to other sinks, the steady-state concentrations of these species in most natural waters are far too low to compete with direct photolysis (2, 6). PAHs bound to CDOM could potentially react with photochemically produced reactive oxygen species due to their proximity to this source, but this mechanism would likely apply to most VOL. 36, NO. 20, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
4367
TABLE 3. Ratios of PAH Photodegradation Rate Constants and Fluorescence in Natural Waters (k′ and F′, Respectively) Compared to Pure Water (k and F, Respectively)a 5 mg/L SRFA
Choptank River
Chesapeake Bay
Atlantic
PAH
k′/k
F′/F
k′/k
F′/F
k′/k
F′/F
k′/k
F′/F
AC BP BA AN PH FU CH PY PE BK BB FL
0.59 ( 0.06 1.1 ( 0.1 0.83 ( 0.09 1.7 ( 0.1
0.78 ( 0.02 0.79 ( 0.02 0.84 ( 0.04 0.81 ( 0.03
0.84 ( 0.04 0.94 ( 0.04 0.93 ( 0.02 0.89 ( 0.05
0.63 ( 0.05 0.75 ( 0.08 0.59 ( 0.04 2.3 ( 0.2
0.90 ( 0.03 0.98 ( 0.05 0.96 ( 0.03 0.90 ( 0.04
0.9 ( 0.3 0.91 ( 0.06 0.35 ( 0.03
0.86 ( 0.03 0.96 ( 0.03 0.93 ( 0.02 0.89 ( 0.03 1.0 ( 0.1 0.89 ( 0.03 0.8 ( 0.3 0.98 ( 0.08 0.63 ( 0.01
0.57 ( 0.05 0.70 ( 0.08 0.56 ( 0.07 1.3 ( 0.2
0.7 ( 0.4 0.90 ( 0.08 1.2 ( 0.3
0.57 ( 0.05 0.60 ( 0.06 0.59 ( 0.07 1.2 ( 0.1 0.8 ( 0.2 0.9 ( 0.1 1.3 ( 0.7 0.9 ( 0.1 0.7 ( 0.2
1.5 ( 0.7 0.77 ( 0.09 0.5 ( 0.1
0.8 ( 0.3 1.00 ( 0.08 0.67 ( 0.06
1.9 ( 0.6
0.9 ( 0.2
1.6 ( 0.9 0.77 ( 0.09 0.7 ( 0.2 0.4 ( 0.1 1.0 ( 0.3
0.8 ( 0.2 1.03 ( 0.07 0.94 ( 0.03 0.89 ( 0.05 0.88 ( 0.08
a
Rate constants were obtained by measuring PAH loss using polychromatic irradiation while employing the WG305 filter (see Experimental Section). Reported error represents the standard deviation from the mean of three or more experiments.
TABLE 4. Rate Constant Ratios (k′/k) of BA and BP Photodegradation in SRFA Solutions (k′) and Pure Water (k)a PAH
5000 µg/L SRFA
500 µg/L SRFA
50 µg/L SRFA
5 µg/L SRFA
BA BP
0.71 ( 0.06 1.02 ( 0.08
0.58 ( 0.07 0.73 ( 0.06
0.61 ( 0.05 0.78 ( 0.06
0.60 ( 0.03 0.76 ( 0.07
a Rate constants were obtained by measuring PAH loss using polychromatic irradiation while employing the WG305 filter (see Experimental Section). Reported error represents the standard deviation from the mean of three or more experiments.
of the compounds examined, not just one select PAH. Further, there is little evidence for binding under these conditions (Table 3), and thus this contribution is not expected to be significant. We are thus hesitant to conclude that the differences in rate constants observed between pure and natural waters, while being outside our analytical uncertainty, are real and not a result of small systematic errors. Other than its role in light attenuation, we propose that (C)DOM does not significantly affect PAH photochemical loss rates in natural waters. Quantum yields in pure water, along with the overlap between PAH absorption and solar irradiance, can be used to estimate the rates of PAH photodegradation in aquatic environments. Because the direct photolysis of PAHs is dependent on both the quantum yield and the spectral overlap of PAH absorption and solar irradiance (eq 5), the first-order rate constants predicted for their decay in an optically thin section of surface water does not follow the same trend as the quantum yields (kp in Table 2). Owing to relatively high quantum yields and good spectral overlaps, BP, PE, and PY decay relatively rapidly, whereas AC, which also exhibits a high quantum yield, decays slowly due to its low absorption over the solar wavelength range. Despite good spectral overlaps, BB and BK are predicted to decay slowly due to their low quantum yields. These results are inconsistent with a recent theoretical analysis by Chen et al. (29), who suggested that PAHs exhibiting smaller values of the HOMO-LUMO gap, (Elumo - Ehomo), will decay more rapidly, due primarily to their better spectral overlap. Moreover, they predict that PE should disappear 10 times faster than PY, while we calculate that PY and PE will have similar photochemical loss rates in the environment (Table 2). Our results indicate that a better understanding of the basic mechanism(s) of the photodegradation is needed to achieve more reliable predictive capabilities. 4368
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 36, NO. 20, 2002
Acknowledgments This work was supported by grants to N.V.B. from the Office of Naval Research (N00014-95-10201 and N00014-99-10034) and the United States Environmental Protection Agency through the Science to Achieve Results (STAR) program. We thank Rachel Fasnacht for help with manuscript preparation. We thank Mike Mallonee for collection of the Choptank River water.
Supporting Information Available Experimental details showing PAH photodegradation protocols as well as a comparison of steady-state fluorescence and HPLC detection methods. This material is available free of charge via the Internet at http://pubs.acs.org.
Literature Cited (1) Neff, J. M. Polycyclic Aromatic Hydrocarbons in the Aquatic Environment; Applied Science Publishers: London, 1979. (2) Zepp, R. G.; Schlotzhauer, P. F. In Polynuclear Aromatic Hydrocarbons; Jones, P. W., Leber, P., Eds.; Ann Arbor Science Publishers: Ann Arbor, MI, 1979; pp 141-158. (3) Mill, T.; Mabey, W. R.; Lan, B. Y.; Baraze, A. Chemosphere 1981, 10, 1281-1290. (4) Zepp, R. G.; Baughman, G. L.; Schlotzhauer, P. F. Chemosphere 1981, 10, 109-117. (5) Zepp, R. G.; Baughman, G. L.; Schlotzhauer, P. F. Chemosphere 1981, 10, 119-126. (6) Blough, N. V.; Zepp, R. G. In Active Oxygen in Chemistry; Foote, C. S., Valentine, J. S., Greenberg, A., Liebman, J. F., Eds.; Chapman and Hall: New York, 1995; pp 280-333. (7) Vaughan, P. P.; Blough, N. V. Environ. Sci. Technol. 1998, 32, 2947-2953. (8) Thomas-Smith, T. E.; Blough, N. V. Environ. Sci. Technol. 2001, 35, 2721-2726. (9) Blough, N. V.; Del Vecchio, R. In Biogeochemistry of Marine Dissolved Organic Matter; Hansell, D. A., Carlson, C. A., Eds.; Academic Press: 2002; pp 509-546. (10) Zafiriou, O. C.; True, M. B. Mar. Chem. 1979, 8, 33-42. (11) Zafiriou, O. C.; True, M. B. Mar. Chem. 1979, 8, 9-32. (12) Gauthier, T. D.; Shane, E. C.; Guerin, W. F.; Seitz, W. R.; Grant, C. L. Environ. Sci. Technol. 1986, 20, 1162-1166. (13) Backhus, D. A.; Gschwend, P. M. Environ. Sci. Technol. 1990, 24, 1214-1223. (14) Schlautman, M. A.; Morgan, J. J. Environ. Sci. Technol. 1993, 27, 961-969. (15) Sigman, M. E.; Zingg, S. P.; Pagni, R. M.; Burns, J. H. Tetrahedron Lett. 1991, 32, 5737-5740. (16) Sigman, M. E.; Schuler, P. F.; Ghosh, M. M.; Dabestani, R. T. Environ. Sci. Technol. 1998, 32, 3980-3985. (17) Beltran, F. J.; Ovejero, G.; Garcia-Araya, J. F.; Rivas, J. Ind. Eng. Chem. Res. 1995, 34, 1607-1615. (18) Lehto, K.; Vuorimaa, E.; Lemmetyinen, H. J. Photochem. Photobiol. A 2000, 136, 53-60. (19) Miller, J. S.; Olejnik, D. Water Res. 2001, 35, 233-243.
(20) Kieber, D. J.; Blough, N. V. Anal. Chem. 1990, 62, 2275-2283. (21) Zepp, R. G.; Cline, D. M. Environ. Sci. Technol. 1977, 11, 359366. (22) Del Vecchio, R.; Blough, N. V. Mar. Chem. 2002, 78, 231-253. (23) Warneck, P. Chemistry of the Natural Atmosphere; Academic Press: San Diego, CA, 1988; Vol. 41. (24) Harvey, R. G. Polycyclic Aromatic Hydrocarbons; Wiley-VCH: New York, 1997. (25) Kavarnos, G. J.; Turro, N. J. Chem. Rev. 1986, 86, 401-449. (26) Vauthey, E.; Haselbach, E.; Suppan, P. Helv. Chim. Acta 1987, 70, 347-353. (27) Paalme, L.; Uibopuu, H.; Gubergrits, M.; Jacquignon, P. Oxid. Commun. 1983, 4, 27-34. (28) Chen, J. W.; Kong, L. R.; Zhu, C. M.; Huang, Q. G.; Wang, L. S. Chemosphere 1996, 33, 1143-1150. (29) Chen, J.; Peijnenburg, W. J. G. M.; Quan, X.; Chen, S.; Martens, D.; Schramm, K.-W.; Kettrup, A. Environ. Pollut. 2001, 114, 137143. (30) Mopper, K.; Xianliang, Z. Science 1990, 250, 661-664. (31) Haag, W. R.; Hoigne, J. Environ. Sci. Technol. 1986, 20, 341348.
(32) Dabestani, R.; Ivanov, I. N. Photochem. Photobiol. 1999, 70, 10-34. (33) Bogan, B. W.; Lamar, R. T. Appl. Environ. Microbiol. 1995, 61, 2631-2635. (34) Bordwell, F. G.; Bausch, M. J. J. Am. Chem. Soc. 1986, 108, 24732474. (35) Uibopuu, H.; Vodzinskii, Y. U.; Tikhova, N. Y.; Kirso, U.; Jacquignon, P. C. J. Gen. Chem. USSR (Eng. Transl.) 1987, 12331236. (36) Karcher, W. Spectral Atlas of Polycyclic Aromatic Compounds Volume 2; Kluwer Academic Publishers: Dordrecht, 1988; Vol. 2. (37) Hamai, S. J. Am. Chem. Soc. 1989, 111, 3954-3957.
Received for review February 25, 2002. Revised manuscript received August 6, 2002. Accepted August 22, 2002. ES025603K
VOL. 36, NO. 20, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
4369
