Photochemical degradation of nonylphenol and nonylphenol polyethoxylates in natural waters

Chemosphere,Vol. 28, No. 7, pp. 1361-1368, 1994

Pergamon 0045-6535(94)E0066-3

Photochemical

Copyright © 1994 Elsevier Science Ltd Printed in Great Britain. All rights reserved 0045-6535/94 $7.00+0.00

Degradation of Nonylphenol and Polyethoxylates in Natural Waters

Nonylphenol

Marijan Ahel 1, Frank E. Scully2, Jr., Jiirg Hoign6 and Walter Giger Swiss Federal Institute for Environmental Science and Technology (EAWAG), 8600 Dtibendorf, Switzerland Present address:

1Center for Marine Research Zagreb, Rudjer Bo.~kovid Institute, 41000 Zagreb, Croatia 2Department of Chemistry and Biochemistry, Old Dominion University, Norfolk, Virginia 23529-0126, USA (Received in Germany 24 November 1993; accepted 4 Jantlary 1994)

ABSTRACT The rates of photochemical transformation of nonylphenol (NP) and nonylphenol polyethoxylates (NPnEO) in natural waters were assessed by exposing their solutions in filtered lake water (DOC=4 rag/L) to sunlight. The first-order rate constant of sunlight photolysis (kp) for NP was estimated at 0.09 m2/(kWh). This corresponds to a half-life of 10-15 hours under continous clear sky, noon, summer sunlight in the surface layer of natural waters. The photolysis rate in the deeper layers is strongly attenuated, being approximately 1.5 times slower at depths of 20-25 cm than at the surface. The photochemical oxidation of NPnEO was shown to be significantly slower than that of NP. Additional laboratory experiments using a merry-go-round reactor (MGRR) have shown that the photochemical degradation of both NP and NPnEO was due mainly to sensitized photolysis whilst direct photolysis was comparatively slow. Moreover, experiments with D20 revealed that singlet oxygen was not an important photooxidant of NP at pH values usually found in natural waters.

INTRODUCTION Nonylphenol is a pollutant of high environmental concern (Giger et al., 1984). It was well known that its acute toxicity to aquatic organisms is rather high (MeLeese et al., 1981) but recent Studies have shown that sublethal toxic effects were detected at concentrations as low as 6 ttg/L (Naylor et al., 1992). In addition, NP was shown to significantly accumulate in various aquatic organisms (Ekelund et al., 1990; Ahel et al., 1993). The main source of NP in the aquatic environment is the application of nonionic surfactants of the nonylphenol polyethoxylate type (NPnEO) of which more than 360 thousand tons per year are used in Western Europe, USA and Japan (Richtler and Knaut, 1988; similar production rates are estimated for 1992).

1361

1362 During conventional mechanical-biological sewage treatment approximately 30-35 % of NPnEO are converted to NP and short-chain NPnEO (Ahel et al., 1994a). Consequently, relatively high concentrations of NP were found in Swiss freshwaters that receive inputs from secondary effluents (Ahel, 1987; Ahel et al., 1994b). In spite of a strong reduction of NPnEO usage in Switzerland and Germany NPnEO continue to be extensively used in many other countries. Therefore, studying processes that could contribute to the removal of NP from natural waters is of great interest for the proper assessment of the environmental acceptability of NPnEO surfactants. Photochemical transformation of phenolic compounds in natural waters has recently been intensively studied (Hwang et al., 1986; Scully and Hoign6, 1987; Faust and Hoign6, 1987; Hoign6, 1990). In natural surface water reaction of chlorophenols with singlet oxygen was shown to be important only if they are dissociated at given pH conditions (Scully and Hoign6, 1987; Tratnyek and Hoign6, 1991). In a comprehensive study by Faust and Hoign6 (1987) on sensitized photooxidation of alkylphenols the organic peroxy radical-type was postulated as a more efficient transient oxidant. Halflives of different alkylphenols in the top meter of a freshwater were estimated to vary in the range of 0.9-72 days. This report aims at estimating the photochemical behaviour of nonylphenol and nonylphenol polyethoxylates in natural waters.

EXPERIMENTAL Aqueous solutions of 4-nonylphenol (NP; 0.39-1 Izmol/L), octylphenol (OP; 0.48 limol/L) and nonylphenol monoethoxylate (NP1EO; 0.33-1 I~mol/L) were prepared from their saturated solutions obtained by a generator column technique (Abel and Giger, 1993). NP was a mixture of differently branched isomers (Giger et al., 1984), while 95 % of OP was 4-(1,1,3,3-tetramethyl)butyl phenol. Both chemicals were purchased from Fluka, Switzerland. NP1EO was preparatively isolated from the commercial mixture Imbetin N/7A (Dr. Kolb, Hedingen, Switzerland). Sunlight phototransformation of NP and NPnEO was performed in 50 mL quartz tubes which were suspended in a shallow flat-bottomed container filled with tap water or in Chriesbach creek at a depth of 20-25 cm. The solutions of NP and NPnEO were prepared in filtered (0.45 ~tm) lake water (Greifensee, DOC=4 mg/L; pH=8.4). The spectra of the lake water in which solutions were prepared and waters of Chriesbach creek showed no difference in the UV and visible ranges. During the experiment the creek water was clear and the temperature, varied between 14.5-17 °C, depending on the time of day. The temperature in the shallow vessel was adjusted (addition of ice) to be similar to that of the creek (17+3 °C), The total sunlight irradiation was determined by integrating the values which were recorded in time intervals of 10 min. The experiments were performed twice: on l l t h September, 1985 (from 1120 to 1720 h; total duration 351 min) and on 18th September, 1985 (from 1105 to 1605 h; total duration 288 min). The average sun irradiation intensities during the experiments were 0.705 kW/m2 and 0.760 kW/m 2 respectively. The first experiment was performed at the original pH value of lake water (8.4), while in the second it was adjusted to 9.4 by the addition of NaOH. Laboratory experiments were performed in a merry-go-round reactor (MGRR) which was supplied with various light sources (medium pressure mercury lamp, tungsten-

1363 halogen lamp). The function and construction of the apparatus were described in more detail elsewhere (Haag and Hoign6, 1986). Quantitative determinations of the analytes were performed by normal-phase HPLC after a simple extraction of the water samples with n-hexane (Ahel and Giger, 1985; Ahel, 1987). The photolysis rate constants (kp) were calculated by presuming first-order kinetics (Scully and Hoign6, 1987) as described by the following equation: kp=-ln(C/Co)/t where CO and C are the initial concentration of analyte and its concentration at the time t, respectively. Consequently, the photolysis half-life was calculated from: t1/2=O.693/kp RESULTS AND DISCUSSION

Sunlight Photolysis The results of the NP photolysis by sunlight in lake water (DOC=4 mg/L) are given in Figure 1. As can be seen, the photolysis was much faster (kp=0.09 h -1) in tubes suspended in a water-filled flat-bottomed container than in the tubes suspended in a creek (kp=0.06 and 0.05 h-l). 100 - ~

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Figure 1. Sunlight photolysis of nonylphenol (NP) and nonylphenol monoethoxylate (NP1EO) in lake water: (A) 11/09/1985, (B) 18/09/1985; open symbols: sample tubes suspended in a fiat-bottomed container filled with tap water; closed symbols: sample tubes suspended in a creek at a depth of 20-25 cm. This difference could be attributed to light attenuation in the creek, as proved by actinometry with p-nitroanisol (NA). Radiation intensity in the flat-bottomed container (kpNA=0.12 h -1) was estimated to be about three times higher than that in the creek at a depth of 20-25 cm (kpNA=0.04 h'l).

1364 The photolysis rate of OP was very similar to that determined for NP (kp of 0.10 and 0.05 h -1 in the flat-bottomed container and creek respectively). Although sunlight photolysis rates of both NP and OP were found to be much slower than previously reported for some other alkylphenols (Faust and Hoign*, 1987), the results suggest that a significant portion (30 %) of these compounds could be photoehemically degraded in the surface layer of natural waters within one day. By contrast, during the same experiments NP1EO concentration remained essentially unchanged. It should be mentioned that in the second experiment (18th September) the pH value of the exposed water sample was increased to 9.4 in order to enhance the photolysis rate but during the experiment the pH decreased to 8.7. Such a change of the pH value did not have any significant effect on the photolysis rate as compared to the original pH of the lake water (pH 8.4). The half-life of the photochemical degradation was estimated from the irradiation dose needed for 50 % degradation. The estimated tl/2 values range from 15-20 hours and correspond to a continous sunlight intensity of 0.700 kW/m2 which is a typical value for late summer at noon in Diibendorf. In June and July (sunlight intensity of aproximately 1000 kW/m2; Haag and Hoign6, 1986) the half-life values are expected to be considerably lower (approximately 10-15 hours). Consequently, it is expected that in clear and shallow waters photochemical degradation could play a role in the elimination of NP. By comparison, photochemical degradation of NPnEO seems to be insignificant. Direct versus Sensitized Photolysis in Merry-Go-Round Reactor (MGRR)

Direct photolysis of NP (1.13 ~tmol/L) and NP1EO (1 ~tmol/L) in distilled water using a MGRR is presented in Fig. 2. The mercury lamp (700 W) and Solidex glass filter (~.>280 nm) employed in the apparatus provided approximately 10 times more light intensity than sunlight during a sunny summer day (Haag and Hoign6, 1986). Rate constants for the direct photolysis (kp) for NP and NP1EO were 0.20 h -1 and 0.06 h -1 respectively. Most of the direct photolysis is due to absorption of UV-B light below 300 nm (Zepp and Cline, 1977). The absorption spectra of NP and NPnEO (~max=277 nm) exibit sharp decreases in their absorption between 280 nm (wavelength cutoff of applied Solidex filter) and 295 nm (wavelength cutoff of terrestrial light). It is well known that photochemical degradation of many phenolic compounds can be much faster in the presence of other dissolved substances (Zepp et al., 1981; Faust and Hoign6, 1987). Photochemical degradation of NP and NPIEO was investigated in filtered lake water (Greifensee) containipg 4 mg/L of DOC. Other conditions were identical to those of the direct photolysis. The results given in Fig. 2 show that the photolysis in the presence of natural organic matter is much faster for both NP (kp=0.92 h -l) and NP1EO (kp=0.52 h -1) than the direct photolysis. Obviously, the degradation was due mainly to sensitized photolysis. In a preliminary experiment the photolysis of higher NPnEO oligomers (nEO=3-18) was also examined. It was observed that they are photolyzed even slower than NP1EO. No significant difference in the photolysis rate was observed between individual oligomers in the range NP6EO-NP17EO (kp=0.026 h-l).

1365

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Figure 2. Photolysis of nonylphenol (NP) and nonylphenol monoethoxylate (NPIEO) in distilled water and in filtered lake water using a merry-go-round reactor (Hg-lamp 700 W; Solidex glass filter). Singlet Oxygen Mechanism It has been shown that the photo-oxidation mechanism via singlet oxygen ( 1 0 2 ) can play an important role in the photolysis of phenolate (Scully and Hoign6, 1987). Therefore it was appropriate to test whether this pathway is of relevance for the photodegradation of nonylphenolic compounds. Alkylphenols are weak acids which can be present in aqueous solutions either in the undissoeiated or deprotonated species. In the analogy with some other phenols (Scully and Hoign¢, 1987; Tratnyek and Hoign¢, 1991) it can be presumed that deprotonated NP reacts about an order of magnitude faster than the undissociated compound. The fraction of the dissociated species depends on the dissociation constant (i.e. on its pKa) and on the pH of the aqueous solution. Therefore the dependence of the photolysis rate of NP on pH of the solution was examined in the range from 7.0 to 10.7 (Fig. 3). For these experiments a tungsten-halogen lamp (60 V) was used instead of the mercury lamp. A solution of potassium dichromate was used as a filter solution ( k > 5 3 0 nm). Rose bengal dye (5 mg/L) was added as a sensitizer to NP solutions (1 $tmol/L) in a phosphate buffer to produce singlet oxygen. 'The photolysis of NP in the pH range between 7.0 and 9.0 is virtually constant (kp=0.45 and 0.50 h-l), while the photolysis at pH 10.7 is only slightly faster (kp=0.92 h-I). This small increase in the apparent

1366 reaction rate constant observed in the pH region of significant dissociation of NP (Johnson, 1973) shows that singlet oxygen cannot control the sensitized photo-oxidation reaction. At pH values typical of natural waters (pH 6-9) much less than 10 % of NP is dissociated but the rate constants are only one order of magnitude larger than those of the undissociated compound. 100

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Figure 3. Photolysis of nonylphenol in H20 and D20 containing 5 mg/L rose bengal dye at different pH values using a merry-go-round reactor (tungsten-halogen lamp, 60V; see Scully and Hoignd, 1987). Since the half-life of singlet oxygen in heavy water (D20) is much longer than in ordinary water (H20) (Merkel et al., 1972), the comparison of sensitized photolysis rates for these two solvents can be used as a further test for the hypothesis of a singlet oxygen mechanism. Fig. 3 gives the comparison of NP photolysis in H20 and D20 containing 10 % H20 using rose begal as a selective sensitizer to produce singlet oxygen. The results indicate that the photolysis via singlet oxygen is important for alkylphenols only at higher pH values (pH>10). At pH 9.0, the ratio of photolysis rates in H20 (kp=0.36 h -1) and D20 (kp=0.68 h -l) is only 1.8, as compared with the ratio of 4.6 at pH 10.7. Consequently, the singlet oxygen is not expected to be a significant transient photooxidant of NP in natural waters as it has already been shown for a number of other alkylphenols (Faust and Hoign~, 1987). In natural waters some types of photooxidants which are derived from the triplet excited natural organic matter may act as the rate controlling photooxidants of phenolic compounds (Canonica and Hoign6, submitted to Environ. Sci. Technol.)

1367 ACKNOWT~DGF_aMEN~ This project was supported in part by the Swiss National Science Foundation (Nationales Forschungsprogramm 7D, research project on "Organic Pollutants in Sewage Sludge"). Financial support of the National Research Council Fund for Science of the Republic of Croatia is acknowledged. We thank B. Faust and V. Sturzenegger for helpful discussions.

Ahel M. (1987) Biogeochemical Behaviour of Alkylphenol Polyethoxylates in the Aquatic Environment. Ph.D. Thesis, University of Zagreb, Croatia, pp 200, Ahel M. and Giger W. (1985) Determination of alkylphenols and alkylphenol mono- and diethoxylates in environmental samples by high-performance liquid chromatography. Anal. Chem. 57, 1577-1583. Ahel M. and Giger W. (1993) Aqueous solubility of alkylphenols and alkylphenol polyethoxylates. C h e m o s p h e r e 2 6 , 1461-1470. Ahel M., Giger W. and Koch M. (1994a) Behaviour of alkylphenol polyethoxylate surfactants in the aquatic environment I. Occurrence and transformation in sewage treatment. Wat. Res. 28, in press. Ahel, M., Giger W. and Schaffner C. (1994b) Behaviour of alkylphenol polyethoxylate surfactants in the aquatic environment II. Occurrence and transformation in rivers. Wat. Res. 28, in press. Ahel M., McEvoy J. and Giger W. (1993) Bioaccumulation of the lipophilic metabolites of nonionic surfactants in freshwater organisms, Environ. Pollut. 79, 243-248. Ekelund R., Bergman A., Granmo A. and Berggren M. (1990) Bioaccumulation of 4nonylphenol in marine animals-a reevaluation. Environ. Pollut. 64, 107-120. Faust B.C. and Hoign6 J. (1987) Sensitized photooxidation of phenols by fulvic acid and in natural waters. Environ. Sci. Technol. 21, 957-964. Haag W.R. and Hoign6 J. (1986) Singlet oxygen in surface waters. 3. Photochemical formation and steady-state concentrations in various types of waters. Environ. Sci. Technol. 20, 341-348. Hoign6 J. (1990) Formulation and calibration of environmental reaction kinetics; Oxidation by aqueous photooxidants as an example. In Aquatic Chemical Kinetics, Stumm W. (Ed.), J. Wiley, New York, pp 43-70. Hwang H-M., Hodson R. E. and Lee R.F. (1986) Degradation of phenol and chlorophenols by sunlight and microbes in.estuarine water. Environ. Sci. Technol. 20, 1002-1007. Giger W., Brunner P.H. and Schaffner C. (1984) 4-Nonylphenol in sewage sludge: accumulation of toxic metabolites from nonionic surfactants. Science 225, 623-625. Johnson C.D. (1973) The Hammett Equation, Cambridge University Press, 1973, p 30. McLeese D.W., Zitko V., Sergeant D.B., Burridge L. and Metcalfe C.D. (1981) Lethality and accumulation of alkylphenols in aquatic fauna. Chemosphere 10, 723-730. Merkel P.B., Nilsson R. and Kearns D.R. (1972) Deuterium effects on singlet oxygen lifetimes in solutions. A new test of singlet oxygen reactions. J. Am. Chem. Soc. 94, 1030-1031. Naylor C.G., Mieure J.P., Morici I. and Romano R.R. (1992) Alkylphenol ethoxylates in the environment. In Proceedings of the 3rd CESIO International Surfactants Congress, Section E, F & LCA Seminar, London, 1-5 June 1992, pp 111-124. Richtler H.J. and Knaut J. (1988) World prospect for surfactants. In Proceedings of the Second World Surfactant Conference, May 24-27, Paris, 3-58.

1368 Scully F.E. Jr. and Hoign6 J. (1987) Rate constants for reactions of singlet oxygen with phenols and other compounds in water. Chemosphere 16, 681-694. Tratnyek P.G. and Hoign6 J. (1991) Oxidation of substituted phenols in the environment: A QSAR analysis of rate constants for reaction with singlet oxygen. Environ. Sci. Technol. 25, 1596-1604. Zepp R.G. and Cline D.M. (1977) Rates of direct photolysis in aquatic environment. Environ. Sci. Technol. 11, 359-366. Zepp R.G., Baughman G.L. and Schlotzhauer P.L. (1981) Comparison of photochemical behaviour of various humic substances in water: I. Sunlight induced reactions of aquatic pollutants photosensitized by humic substances. Chemosphere 10, 109-117.