Environ. Sci. Technol. 1905, 19, 1072-1076
Ishihara, Y., presented a t the Dry Limestone Injection Process Symposium, Gilbertsville, KY, June 1970. Borgwardt, R. H.; Harvey, R. D. Environ. Sei. Technol. 1972,6, 350-360. Bhatia, S. K.; Perlmutter, D. D. AIChE J. 1981,27,226-234. Borgwardt, R. H.; Gillis, G. R.; Roach, N. F., presented at the Joint EPA/EPRI Symposium on Stationary Source Combustion NO, Control, Dallas, TX, Nov 1982. Coutant, R. W.; Simon, R.; Campbell, B.; Barrett, R. E. Oct 1971, EPA APTD-0802. McClellan, G. H., presented a t the Dry Limestone Injection Process Symposium, Gilbertsville, KY, June 1970. Harvey, R. D. July 1971, EPA APTD-0920. Seeker, W. R.; Samuelsen, G. S.; Heap, M. P.; Trolinger, J. D. 18th Symp. (Int.) Combust. [Proc.] 1981, 18th, 1213-1226. Hamor, R. J.; Smith, I. W. Fuel 1970, 50, 394-404. Altenkirch, R. A.; Peck, R. E.; Chen, S. L. Powder Technol. 1978,20, 189-196. Beer, J. M.; Chigier, N. A. “Combustion Aerodynamics”; Applied Science Publishers, Ltd.: London, 1972; p 18.
(17) Boynton, R. S. “Chemistry and Technology of Limestone”, 2nd ed.; Wiley: New York, 1980. (18) Borgwardt, R. H. AZChE J . 1985, 31, 103-111. (19) Glasson, D. R. J . Appl. Chem. 1967, 17, 91-96. (20) Beruto, D.; Barco, L.; Searcy, A. W.; Spinolo, G. J. Am. Chem. Soc. 1980,63, 439-443. (21) Beruto, D.; Barco, L.; Belleri, G.; Searcy, A. W. J . Am. Chem. SOC.1981,64, 74-80. (22) Levin, E. M.; Robbins, C. R.; McMurdie, H. F. “Phase Diagrams for Ceramists”; Reser, M. K., Ed.; American Ceramic Society: Columbus, OH, 1964; p 102. Received for review November 2, 1984. Revised manuscript received June 13,1985. Accepted June 27,1985. This research was supported by the US.Environmental Protection Agency under Contract 68-02-3633. Although the research described in this article has been funded by the Environmental Protection Agency, it has not been subjected to Agency review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred.
Interactions between Polycyclic Aromatic Hydrocarbons and Dissolved Humic Material: Binding and Dissociation John F. McCarthy” and Braullo D. Jlmenez Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, Tennessee 3783 1
Binding of polycyclic aromatic hydrocarbons (PAH’s) to dissolved humic material (DHM) was examined by using equilibrium dialysis and fluorescence techniques. There was a direct relationship between the hydrophobicity of the PAH and the affinity for binding to DHM. The binding affinity Pa for benzo[a]pyrene (BaP), benzanthracene, and anthracene decreased slightly as the concentration of DHM increased. The binding of BaP to DHM was completely reversible and the extent of reversibility was unrelated to the sorption time. The rate of binding of BaP to DHM, measured by the quenching of BaP fluorescence, was very rapid and was completed within 5-10 min. The results suggest that the presence of DHM, or other sorptive components of the dissolved organic pool, may affect binding to sediment or suspended particles and thus alter the fate and transport of organic contaminants in aquatic systems.
Very little is known, however, about the interaction between contaminants and DHM or about the influence of this interaction on the binding of contaminants to particles. In this study, the association of a series of PAH’s with DHM and the reversibility of that association are examined by using equilibrium dialysis and fluorescence techniques. There is continuing discussion in the literature as to whether association of organic contaminants with sediment involves a process of adsorption or partitioning (14,15);we attempt to avoid that controversy by referring to “binding” or “association”, without suggesting any specific mechanism. We refer to the humic material used in this study as “dissolved” since it is resistant to centrifugation and passes through a 0.3-ym filter. This is a functional definition and is not intended to address the physicochemical distinction between dissolved and colloidal material.
Introduction The partitioning of hydrophobic organic contaminants within the environment is of fundamental importance in determining their fate and transport. A number of studies have examined the adsorption of organic contaminants to sediments or to suspended particulates and have shown that the affinity for association of a contaminant with a sediment is well correlated both with the hydrophobicity of the contaminant (expressed as the octanol-water partition coefficient, KOw) and with the organic content of the sediment (1,2). Another potentially important, but less obvious, sorbent for organic contaminants is naturally occurring dissolved organic matter (DOM) such as dissolved humic material (DHM). DHM has been shown to form stable complexes with several trace organic contaminants (3-5). More recently, the binding of contaminants such as polycyclic aromatic hydrocarbons (PAH’s), insecticides, and herbicides to DHM has been measured quantitatively (6-9). Binding of PAH’s to DHM has been shown to greatly reduce the availability of the PAH’s for uptake and bioaccumulation by aquatic organisms (20-13).
Materials and Methods Humic acid was obtained from Aldrich Chemical Co. (Milwaukee, WI). Before it was used in any experiment, the humic acid was dissolved in water, centrifuged at lOOOOg for 30 min, and then filtered through precombusted glass-fiber filters (type A-E, nominal retention of 0.3 ym; Gelman Sciences, Inc., Ann Arbor, MI) to remove any particulate material. The silica content of the DHM solution was measured by atomic absorption (Perkin-Elmer Model 603 equipped with a Model 2200 graphite furnace) and by ICP (Instruments SA) to determine if substantial amounts of clay microparticles were present. The organic carbon content of the DHM stock solutions was measured with an 01 carbon analyzer (01Corp., College Station, TX). DHM concentrations are expressed as grams of organic carbon per liter. Water used in these experiments either was dechlorinated tap water that was passed through a bed of activated charcoal to remove any DOM or was deionized distilled water free of DOM; both types of water were filtered through precombusted Gelman type A-E glass-fiber filters to remove particulates.
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Radiolabeled (14C) PAH’s [naphthalene (NPH), anthracene (ANTH), benzanthracene (BA), 3-methylcholanthrene (MC), and benzo[a]pyrene (BaP)] were obtained from Pathfinder Laboratories, Inc. (St. Louis, MO) and California Bionuclear (Sun Valley, CA). [ 14C]PAH’~ were repurified by high-performance liquid chromatography (10) before there use in experiments since the presence of radiolabeled polar impurities can produce underestimates of the association coefficient. NonradioactivePAH’s were obtained from Aldrich (gold label) or from Mallinckrodt Chemical Co. Radioactive and nonradioactive PAH’s were dissolved in a carrier (acetone, methanol, or dioxane), and a small volume of the carrier stock solution (500 nm) to minimize photodegradation of the PAH’s. Dialysis tubing (Spectra/Por 6; molecular weight cutoff of 1000) was washed in distilled water, 1M Na2C03,1 M NaHC03, and distilled water to remove the sodium benzoate preservative. Some of the tubing was packed in sodium azide; this tubing was washed in distilled water. Equilibrium dialysis experiments to measure the binding of PAH’s to DHM were performed by placing 4 or 8 mL of a DHM solution in dialysis tubing and clamping the ends of the tubing with dialysis tubing clamps (Fisher Scientific). The dialysis bag was then placed in a 100-mL glags jar containing an aqueous solution of a PAH (80 mL). Sodium azide (0.02%) was added to prevent bacterial transformation of the PAH. The jar was sealed with a Teflon-lined cap and shaken in the dark at 23 “C for at least 4 days unless otherwise noted. Control experiments demonstrated that equilibration was complete within 24-48 h. A sample of the solutions inside and outside the dialysis bag was analyzed for 14C radioactivity by using ACS scintillation cocktail (Amersham)and a CD460 liquid scintillation counter (Packard). Data are expressed as moles of PAH on the basis of the specific activity of the PAH stock solutions. The radioactivity outside the dialysis tubing represents the PAH that is dissolved in the water (free), and the radioactivity inside the tubing represents the sum of dissolved PAH and PAH bound to the DHM. The concentration of PAH bound to DHM (mol of PAH/g of C) was calculated by subtracting the PAH concentrations inside and outside the dialysis bag (mol of PAH/mL) and dividing by the concentration of DHM within the bag (g of C/mL). The dissociation of PAH from DHM was also measured by equilibrium dialysis. After equilibration of the PAH and DHM as described above, the dialysis bag was rinsed in distilled water and transferred to a jar containing 80 mL of uncontaminated water with sodium azide. The jar was covered and shaken in the dark at 23 “C for at least 4 days unless otherwise noted (dissociation equilibrium was completed within 24-48 h). The time course of binding of PAH to DHM was analyzed by measuring the quenching of the fluorescence of BaP on a SPF-500 spectroflurometer (American Instrument Co., Silver Spring, MD). BaP was excited at 380 nm, and emission was measured at 405 nm. A small volume of DHM solution (10-100 ILL)was added to a cuvette containing a solution of BaP and rapidly mixed. Fluorescence data were corrected for the contribution of DHM fluorescence to the observed fluorescence and for the inner filter effect resulting from the absorption of excitation and emission light by the DHM (16, 17). Both of these corrections were minimal since relatively low
-
BINDING
-e-DISSOCIATION
6 0
en5
1
v
I
1‘
-i
1 I
0
0
0.5
I
1
I
1.5 2.0 2.5 FREE BaP (pmol/mL)
1.0
I 3.0
3.5
Figure 1. Isotherms describing the association (circles and solid line) and dissociation (squares and dashed line) of BaP and DHM (2.5 mg of C/L). Association continued for 7 days, and dissociation lasted 4 days. The slopes, (Pa = 1.8 X lo6,r 2 = 0.97;P, = 2.16 X lo6,r 2 = 0.98)were not significantly different @ > 0.05).
concentrations of DHM were used in these experiments. The quenching of BaP fluorescence is quantified as the ratio of F,/F, where Fo is the fluorescence of BaP when no DHM is present and F is the observed fluorescence when DHM is present. Quenching results in a decrease in the value of F relative to Fo and is measured as an increase in the ratio Fo/F. Results were analyzed by linear regression and analysis of variance using SAS (Statistical Analysis System, Carey, NC). Results and Discussion Binding Affinities of PAH’s. Organic carbon constituted 50.2% of the DHM used in these experiments, and the silica content of the DHM was 2.6 mg of Si/g of C. The association of PAH’s to DHM was measured by equilibrium dialysis. Isotherms were linear up to the solubility limit of the PAH. A typical association curve is illustrated in Figure 1. The association coefficient, Pa, is calculated as the ratio of the concentration of free vs. DHM-bound PAH after equilibration or as the slope of the association curve estimated by a linear least-squares fitting procedure. It should be noted that sorption onto the glassware and onto the dialysis membrane does not affect the results of the equilibrium dialysis experiments. When the system is at equilibrium, the free and DHM-bound PAH’s are in equilibrium with each other and with PAH bound to the glass or to the membrane. Binding to the glass and membrane decreases the total amount of PAH in solution (free or bound to DHM) but does not affect the final equilibrium between these components of interest (6). The possibility that a low molecular weight component of the DHM passed through the dialysis bag and bound PAH outside the bag was considered. The DHM solution was predialyzed for 4 days against several changes of distilled water before the bag was transferred to a jar with radiolabeled PAH. It is assumed that any portion of the DHM that could pass through the bag would have been removed by the predialysis step. There was no significant Environ. Sci. Technol., Vol. 19,No. 1 1 , 1985
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Table I. Effect of Association Time on Reversibility of Binding of BaP to DHM (2.5 mg of C/L)O dissociation time, days
Pa x 106
2 7 7
2 2 4
2.02 f 0.30 (4) 1.95 & 0.12 (4) 1.94 f 0.36 (12)
2.59 f 0.55 (4)
2.36 & 1.20 (4) 2.22 f 0.59 (12)
-0.24 -0.41 -0.45
2.03 f 0.43b
2.42 f 0.66b
-0.41 f 0.6gb
Pd
x
lo6
association time, days
(pa
- p d ) x lo6
"Paand Pd are calculated from the ratio of bound and free BaP after association or dissociation equilibration, respectively, and are expressed as the mean f standard deviation. The number of replicates is indicated in parentheses. See text for results of statistical analyses. *Mean f SD.
106
t
1 t 1 t
0 B
to5
Y
104
/
f4TH
log K O W = 0.48+(0.97xlog
r 2 ~0.96
PO Figure 2. Relationship between the KO, of the PAH's and their affinity for binding to DHM, expressed as the association coefficient, Pa.
difference ( p > 0.1) between the results of association experiments with or without predialysis of DHM. The affinity for binding to the DHM was related to the hydrophobicity of the PAH, expressed as the KO,(Figure 2). In fact, the Pa is approximately equal to the K,,. The P,'s for binding of BaP, BA, and ANTH to DHM reported here agree within a factor of 2 of those measured by a reverse-phase technique using the same DHM from Aldrich (8). These P i s are approximately an order of magnitude lower than those reported for the binding of NPH and ANTH to natural marine colloids, which was mebured by using a hollow fiber technique (9). This discrepancy may reflect variability in the binding affinities of different sources of DOM. The P,'s for the binding of PAH's to Aldrich DHM agree well with the partition coefficients describing the binding of the PAH's to sediment, when those coefficients are corrected for the organic content of the sediment [i.e., expressed as the K, or K,, (1,2)]. This result is consistent with the hypothesis that the binding of nonionic organic contaminants such as PAH to the sediment is controlled primarily by interaction of the contaminant with the organic coating on the sediment particle. Because most sediments have a low organic carbon content relative to DHM (100% organic matter when expressed on a gram of C per liter basis), DHM can have a much higher capacity for binding PAH's in the water column than will the same mass of suspended sediment particles. For example, equivalent amounts of PAH would be bound by a solution containing 10 mg of C/L of DHM, or 200 mg/L suspended sediment particles with an organic content of 5 % Reversibility of Binding. The dissociation of BaP bound to DHM was examined quantitatively by using equilibrium dialysis. The equilibrium dialysis method would be invalid for measuring dissociation of the BaP-
.
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DHM complex if substantial amounts of BaP bound to the dialysis membrane were released into the water during the dissociation equilibration or if substantial amounts of freed BaP were removed from the system due to sorption to the glassware. Although there was undoubtedly reequilibration among all these components, there was no net change in the mass of BaP associated with the equilibration between the free BaP and the BaP bound to DHM. The total mass of BaP transferred when the dialysis bag was introduced into uncontaminated water at the beginning of the dissociation (free and DHM-bound BaP, exclusive of any radiolabeled material associated with the dialysis membrane) was within 98% (+7%, n = 13) of the total mass of free and DHM-bound BaP in the dialysis system after dissociation equilibrium (exclusive of any radiolabeled material associated with the dialysis membrane or glassware). A partition coefficient for dissociation, Pd, is defined as the ratio of the concentration of PAH bound to the DHM vs. that dissolved in the water at dissociation equilibrium or as the slope of the dissociation isotherm (Figure 1). The binding of BaP to DHM is fully reversible, since the slopes of the association and dissociation curves (Paand Pd, respectively) were not significantly different (p > 0.05). The length of time that BaP and DHM associated did not affect the reversibility of the binding (Table I). Pa was not affected by the association time ( p > 0.20), and Pd was not affected by either the association or dissociation times ( p > 0.66 or 0.92, respectively). Evaluation of the differences between Pa and Pd (Pa- Pd reflects the reversibility of binding under the different combinations of time) showed that there was no significant effect due to association or dissociation time (p > 0.85). However, analysis of Pa - Pd did reveal that there was a tendency for Pd to be greater than Pa at all combinations of time (0.02 < p < 0.05). On the basis of the lack of significant differences in the slopes of the isotherms in Figure 1and the marginal differences between Pa and Pd in Table I, it seems reasonable to conclude that the binding of BaP to DHM is fully reversible. Effect of DHM Concentration on Binding Affinity. The effect of the DHM concentration on the binding affinity, Pa,was examined for a series of PAH's (Figure 3). Pa decreased slightly at increasing concentrations of DHM for BaP, BA, and ANTH. The Pa for NPH increased, but at DHM concentrations that were much higher than those used with the other PAH's. Although the change in Pa at increasing concentrations of DHM is statistically significant (p > 0.1) for the four PAH's, the magnitude of the change is quantitatively small; Padid not change by more than a factor of 2 over the range of DHM concentrations examined. Similar results have been reported for BaP and ANTH (8). Time Course of Binding. The time course of the quenching of BaP fluorescence due to the addition of DHM is shown in Figure 4. The quenching of fluorescence by DHM followed the classic Stern-Volmer relationship with the ratio F o / F being linearly related related to the
41
I
I
I
I
BENZO[o] PYRENE
r
1
I
I
I
I
I
I
1
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3
2.5
2.0 -
ygx
I BENZANTHRACENE
I
(6)
1.5‘1
a“
-
1.0 -
a
:
- #
*
--
.
LL
I
I
I
0
I
-I
0.7 mg C/L 0
&O6
0.3 mg C/L 0
f-
0
20
40
60
TIME (min)
Flgure 4. Time course of quenching of BaP fluorescence following the addition of DHM at time zero. DHM concentrations are indicated. FoIF is the ratio of fluorescence In the absence of DHM to that In the presence of DHM.
concentration of quencher (DHM). Since the quantum yield of the BaP-DHM complex is not known, no attempt was made to directly quantify the extent of binding by fluorescence measurements. The association of BaP with DHM was very rapid. Quenching is complete within 5-10 min after the addition of DHM (Figure 4). Karickhoff (18) has proposed that there are two kinetically distinct components associated with the binding of contaminants to sediment, a “fast” and a “slow”. The latter component is postulated to reflect migration of the contaminant to less accessible sites within the sediment matrix. The data in Figure 4 suggest that binding of BaP to DHM exhibits only a fast kinetic component (or, arguably, a very fast and fast component). The apparent speed with which BaP binds to DHM would be consistent with the smaller molecular distances involved in migrating into the interior of a DHM molecule, rather than to a sediment particle. The possibility that further binding to DHM may occur over long periods of time cannot be excluded on the basis of these short-term fluorescence data. However, dissociation of BaP from DHM is equally reversible regardless of the association time (Table I). This behavior
contrasts with that for desorption of contaminants from sediment. In the latter system, recovery of sorbed pollutants depends on the incubation time of the pollutant in the sediment system. Longer incubation times are postulated to result in migration of an increasingly large fraction of the pollutant into the less readily desorbed slow compartment within the sediment particles (18). Comparison of the results in the two sorbent systems suggests that binding to DHM does not involve a slow component comparable to that postulated in sediment systems. Environmental Implications. The results of the present study demonstrate that DHM can reversibly bind organic contaminants with an affinity comparable to that of the organic coating of sediment particles [KO,(1,2 ) ] . Furthermore, binding to DHM appears to be more rapid than comparable binding to sediment (1,2). It should be noted that different sources of DHM in natural waters can have different affinities for binding organic Contaminants (6-9) and that the quantitative relationships developed in this study cannot be applied universally. Nevertheless, our results suggest that the presence of DHM or other sorptive components of the dissolved organic carbon pool can affect the environmental transport and fate of organic contaminants by competing with sediment or suspended particles for the binding of contaminants. Hasset and Anderson (19) have shown that binding of cholesterol and a PCB by particles from river water or treated sewage was less efficient when the aqueous phase consisted of concentrated DOM from sewage or river compared to unconcentrated samples or to distilled water. Voice et al. (20) have postulated that the decrease in binding affinity of organic contaminants with sediment at increasing solids concentrations is due to release of solute-binding material, in the form of DOM or microparticulates, to the liquid phase. Thus, it would appear that DHM (and possibly other components of DOM) have the potential to affect the partitioning of contaminants to the sediment. The presence of DHM might be expected to decrease the amount of hydrophobic contaminant that will bind to suspended particles or be sequestered in the sediment and increase the amount of contaminant that will remain stabilized within the water column. In lotic systems, contaminants could be transported much further downstream from their source since settling of particles will have less effect in limiting their transport. The increased amount of contaminant in the water column (nonparticulate) could be a source of biological concern since, in Environ. Sci. Technol., Vol. 19, No. 11, 1985
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general, contaminant bound to sediment or particles is much less available for uptake by aquatic organisms than is contaminant dissolved in the water (10,21,22). Such concerns appear to be unnecessary, however, since contaminants bound to DHM are largely unavailable for uptake by aquatic organisms (10-13). Our studies on the interactions of DOM with organic contaminants have demonstrated that DOM is an important factor to be considered to understand processes controlling the transport, fate, and biological effect of hydrophobic contaminants in aquatic systems.
Landrum, P. F.; Nihart, S.R.; Eadie, B. J.; Gardner, W.
S.Enuiron. Sci. Technol. 1984, 18, 187-192. Wijayaratne,R.; Means, J. C. Mar. Environ. Res. 1984,11, 77-89.
McCarthy,J. F. Arch. Environ. Contam. Toxicol. 1983,12, 559-568.
Leversee, G. J.; Landrum, P. F.; Giesy, J. P.; Fannin, T. Can. J. Fish. Aquat. Sci. 1983, 40 (Suppl. 2), 63-69. McCarthy, J. F.; Jimenez, B. D. Enuiron. Toxicol. Chem. 1985, 4, 511-521.
McCarthy, J. F.; Jimenez, B. D. Aquat. Toxicol., in press. Chiou, C. T.; Porter, P. E.; Schmeddling,D. W. Environ. Sci. Technol. 1983, 17, 227-31.
Acknowledgments We thank D. K. Whitmore and M. C. Black for technical assistance, J. J. Beauchamp for help with the statistical analyses, and G. R. Southworth for valuable discussions and suggestions. Registry No. NPH, 91-20-3; ANTH, 120-12-7; BA, 56-55-3; MC, 56-49-5; BaP, 50-32-8. Literature Cited Karickhoff, S. W.; Brown, D. S.; Scott, T. A. Water Res. 1979,13, 241-248.
Means, J. C.; Wood, S.G.; Hassett, J. J.; Banwart, W. L. Environ. Sci. Technol. 1980, 14, 1524-1528. Warshaw, R. L.; Burcar, P. J.; Goldberg, M. C. Enuiron. Sci. Technol. 1969, 3, 271-273. Boehm, P. D.; Quinn, J. G. Geochim. Cosmochim. Acta 1973,37, 2459-2477. Poirrier, N. A,; Bordelor, B. R.; Laseter, J. L. Environ. Sci. Technol. 1972,6, 1033-1035. Carter, C. W.; Suffett, I. H. Environ. Sci. Technol. 1982, 16, 735-740. Means, J. C.; Wijayaratne, R. Science (Washington,D.C.) 1982, 215, 968-970.
Mingelgrin,U.; Gerstl, Z. J. Environ. Qual. 1983,12, 1-11. Parker, C. A. “Photoluminescenceof Solutions”;Elsevier: New York, 1968; pp 220-232. Lloyd, J. B. “Standards in Fluorescence”;Miller, J. N., Ed.; Chapman Hall: New York, 1981; pp 33-35. Karickhoff, S.W. “Contaminantsand Sediments”;Baker, P. A., Ed.; Ann Arbor Science: Ann Arbor, 1980; Vol. 2, pp 193-205.
Hassett, J. P.; Anderson, M. A. Water Res. 1982, 16, 681-686.
Voice, T. C.; Rice, C. P.; Weber, W. J. Environ. Sci. Technol. 1983, 17, 513-518.
Landrum, P. F.; Scavia, D. Can. J. Fish. Aquat. Sci. 1983, 40, 298-305. Muir, D. C. G.; Townsend, B. E.; Lockhart, W. L. Enuiron. Sci. Technol. 1983, 2, 269-281. Received for review November 5,1984. Accepted June 10,1985. This research was sponsored by the Office of Health and Environmental Research, Ecological Research Division, U.S. Department of Energy, under Contract DE-AC05-840R21400 with Martin Marietta Energy Systems, Inc. Publication No. 2534, Environmental Sciences Division, Oak Ridge National Laboratory.
Seasonal Variations in Weathering and Toxicity of Crude Oil on Seawater under Arctic Conditions Lelv K. Sydnes,” Tor H. Hemmlngsen, Smlvi Skare, and Slssel H. Hansen
Department of Chemistry, University of Tromser, N-900 1 Tromsca, Norway Inger-Brltt Falk-Petersen and Sunnlva Lmnnlngt
Institute of Biology and Geology, University of Tromser, N-9001 Tromsca, Norway Kjetlll 0stgaard
Institute of Marine Biochemistry, University of Trondheim, N-7034 Trondheim NTH, Norway The present study shows that oil in Arctic marine environment is modified by several processes. From October to February the oil composition is mainly affected by evaporation. During the rest of the year the composition is considerably influenced by photooxidation and subsequent dissolution of polar oxidation products in the water phase. These products are toxic and may represent a hazard to marine organisms in the Arctic spring and summer. Introduction Oil discharged into the marine environment is affected by a number of processes that change its chemical and
physical characteristics (1-9). One of the processes is photochemical oxidation, which is capable of transforming a variety of oil components into oxygenated derivatives (5, 10-13). Many of these products are fairly water soluble and show significant toxicity to algae (14,15),bacteria (16, 17),marine invertebrates, and fish (18, 19). Most of the studies performed to assess the importance of photochemical weathering of oil have been carried out under laboratory conditions with various light sources (2, 10, 11, 20-26). Although some experiments have been performed with simulated sunlight, it is of importance that the conditions prevailing in nature are studied directly. This is particularly the in the Arctic area where the influx of solar radiation shows great seasonal changes in total energy as well as wavelength distribution. The purpose of this study was to observe how the natural Y
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