A New Sono-electrochemical Method For Enhanced Detoxification

Ultrasonics Sonochemistry 11 (2004) 365–372 www.elsevier.com/locate/ultsonch

A new sono-electrochemical method for enhanced detoxification of hydrophilic chloroorganic pollutants in water Yakov Yasman a, Valery Bulatov a, Vladimir V. Gridin a, Sabina Agur a, Noah Galil b, Robert Armon b, Israel Schechter a,* a

b

Department of Chemistry, The Grand Water Research Institute, Technion-Israel Institute of Technology, Technion City, Haifa 32000, Israel Department of Civil Engineering, The Grand Water Research Institute, Technion-Israel Institute of Technology, Technion City, Haifa 32000, Israel Received 24 April 2003; accepted 12 October 2003 Available online 24 January 2004

Abstract A new method for detoxification of hydrophilic chloroorganic pollutants in effluent water was developed, using a combination of ultrasound waves, electrochemistry and Fenton’s reagent. The advantages of the method are exemplified using two target compounds: the common herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) and its derivative 2,4-dichlorophenol (2,4-DCP). The high degradation power of this process is due to the large production of oxidizing hydroxyl radicals and high mass transfer due to sonication. Application of this sono-electrochemical Fenton process (SEF) treatment (at 20 kHz) with quite a small current density, accomplished almost 50% oxidation of 2,4-D solution (300 ppm, 1.2 mM) in just 60 s. Similar treatments ran for 600 s resulted in practically full degradation of the herbicide; sizable oxidation of 2,4-DCP also occurs. The main intermediate compounds produced in the SEF process were identified. Their kinetic profile was measured and a chemical reaction scheme was suggested. The efficiency of the SEF process is tentatively much higher than the reference degradation methods and the time required for full degradation is considerably shorter. The SEF process maintains high performance up to concentrations which are higher than reference methods. The optimum concentration of Fe2þ ions required for this process was found to be of about 2 mM, which is lower than that in reference techniques. These findings indicate that SEF process may be an effective method for detoxification of environmental water. Ó 2003 Elsevier B.V. All rights reserved. Keywords: Sonoelectrochemistry; Fenton reagent; Ultrasound wave

1. Introduction Halogenated organic pollutants are toxic materials that are often present in industrial effluents and sometimes in drinking water. Many of these materials are very stable and resist traditional biodegradation treatments. Several different approaches are known for the decomposition of these compounds. These include chemical oxidation, advanced oxidation processes, photochemical and photocatalytic reactions, sonochemical and sono-electrochemical processes. These degradation mechanisms were suited to several specific

*

Corresponding author. Fax: +972-4-8292579. E-mail address: [email protected] (I. Schechter).

1350-4177/$ - see front matter Ó 2003 Elsevier B.V. All rights reserved. doi:10.1016/j.ultsonch.2003.10.004

compounds, since their efficiency is compound-dependent. Therefore, we shall restrict our discussion to our target material, which is 2,4-dichlorophenoxyacetic acid (2,4-D). This herbicide is used in controlling broadleaf weed in cereal crops, sugarcane, turf, and pastures. After 50 years of use, 2,4-D is the most widely used herbicide worldwide [1]. Although its mild toxicity, (oral LD50 ranges from 370 to 666 mg/kg in small animals), it can be converted into highly toxic chloroorganic products [2]. Therefore, a maximum concentration level of 70 ppb is recommended in drinking water by the World Health Organization [4]. 2,4-D is a poorly biodegradable pollutant [5]. In fresh water it can be mineralized (i.e. converted to CO2 and chloride ions) by various microbiological schemes only at concentration levels that do

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not exceed 10 ppb. Therefore, alternative methods must be considered for this compound in groundwater. Many oxidation schemes were applied in attempts to mineralize 2,4-D in aqueous solutions [3–11]. Some of the processes were based on the chemical [3,4] and photochemical [4–11] degradation, mediated by hydroxyl radicals. Significant accumulation of chloride ions was reported on the time scale of 4 h; [10]. 2,4Dichlorophenol (2,4-DCP), chlorohydroquinone and hydroquinone were detected as the intermediates of the processes involved. Complete degradation of a 0.1 mM herbicide solution at pH approximately 3 was reported using a chemical degradation approach [4]. Similar results were achieved by using the Fenton reagent, where considerable degradation of 2,4-D was detected after 10 min of the oxidative process. Note, however, that these results were achieved in pure solutions of relatively low initial herbicide concentrations. At high concentrations and in the presence of various salts, the classical Fenton reaction is not so effective anymore. As the pollutant concentration approaches the solubility limit, the inorganic salt effect results in formation of suspension, which is actually the case for many effluents. Therefore, an alternative degradation method is required for effluent water at rather high pollutant concentrations. Effective electrochemical methods were also developed for the treatment of wastewater polluted by toxic and stable organic compounds [12–17]. Direct electrochemical techniques [12] for water purification involve anodic or cathodic decomposition; however, these are commonly associated with anodic oxidation of water (release of O2 ). Therefore, only low current yields can be achieved this way. Additional problems, which are related to the low miscibility of most organics in water and to the hindered mass transfer to the anode, are responsible for rather low space-time yields. Considerably better results may be achieved using indirect electrochemical oxidation, where the pollutants are oxidized in the bulk by a mediator in a high oxidation state. For example, the application of the electroFenton process to degradation of 2.4-D resulted in 67% TOC removal, while only 16% were achieved by direct anodic oxidation (100 mA for 4 h) [17]. In addition to the above chemical, photochemical and electrochemical approaches, sono-chemical techniques have also been developed. Since these are closely related to our target technique, they are more elaborated in the following. Sono-chemical reactions involving chloroorganic solutes in aqueous solutions have received intensive attention since 1950 due to the work of Weissler et al., Parke and Taylor [18,19]. Studies of sono-chemical modifications of carbon tetrachloride and chloroform in aqueous environment [20] then followed. Since 1990, there has been increasing interest in the ultrasound-mediated

degradation of various organic contaminants present both in natural and industrial water resources. Ultrasound was applied to decomposition of such chlororganic compounds as: chloroform [21], carbon tetrachloride [22], chlorofluorocarbons [23], chlorophenols [24–27], chlorinated aromatic compounds, biphenyls and pesticides [28]. A complete sonolytic degradation of 2,4-D 0.2 mM, performed at high frequency ultrasound radiation of 640 kHz and at a pH of 2.2, has occurred within 21 min [29]. Under alkaline conditions (pH ¼ 11.8), a longer degradation time was required (38 min) [29]. The chemical aspects of continuous and pulsed ultrasound treatment of various aqueous solutions are believed to be associated with acoustic cavitations. There are three regions of importance in respect of the aqueous sono-chemical processes. The first region is the interior of the collapsing cavitation bubbles in which extreme thermodynamic conditions, due to high local temperature and pressure, rapidly set in [30–33]. In this region, a fast pyrolysis of volatile solutes takes place; water molecules also undergo thermal decomposition to produce H atoms and  OH radicals [34]. The second region is the interfacial boundary between the gaseous and the liquid phases where the temperature is lower than inside the bubbles, yet still high enough to cause thermal decomposition of organic solutes. It is believed that the reactive radicals formed from water decomposition in gas bubbles are localized in this region. The third region is the bulk of solution (usually at ambient temperature) where various reactions of organic solutes with either  OH radicals or H atoms, which escape from the bubbles’ interface, may occur [35]. Sono-electrochemistry was also successful in wastewater detoxification [15]. However, as far as we know, the combination of ultrasonic waves, electrical field and Fenton’s process, has never been examined for water treatment. This approach seems promising, due to the power of combining its individual components which may enhance the overall performance. In this study we addressed the sono-electrochemical Fenton (SEF) process and evaluated its effect upon degradation of chlorinated aromatic compounds in environmental water. We studied the intermediate compounds formed in this process, as well as the various factors affecting its efficiency. Kinetic investigation provided insight to the chemical reactions that take place and to the fate of the various intermediates.

2. Experimental procedures 2.1. Materials and reagents Both target compounds: 2,4-dichlorophenoxyacetic acid (Aldrich, 98%) and 2,4-dichlorophenol (Fluka,

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95%) were tested by HPLC and used as supplied with no further purification. Spectroscopy purity dichlorometane (Carlo Erba, 99.8%) was used after testing by GLC. Analytical grade reagents of sodium sulfate (Agan, Israel), ferrous sulfate (Mallinckrodt, USA) and hydrogen peroxide (Carlo Erba, 30% ) were used. 2.2. Sono-electrochemical facility The schematic diagram of our sono-electro-oxidation facility is shown in Fig. 1. The reactor was a glass cylinder vessel (internal diameter 25 mm, effective sample volume of 10 ml). The solution temperature was maintained at 25 ± 1 °C by circulating water in a doublejacket cooling array. Sonication was achieved at low frequencies (20 kHz) using an ultrasonic generator Sonic @ Materials Model 2020, fitted with a horn that emits ultrasound via a titanium alloy tip (13 mm in diameter) dipped in the studied solutions from the top of the reactor. The average output electric power of the generator was 75 W. Both cathode and anode were made of nickel foil (0.125 mm thin) in the form of cylindrical segments of 11 mm radius and 20 mm height. They were placed around a horn (11 mm radius). The support electrolyte was Na2 SO4 (0.5 g/l). Sono-electro-oxidation was carried out in the galvanostatic mode at current intensities not exceeding 100 mA. Similar to recent reports [17], the SEF scheme used here appeared to be insensitive to the current density, within the experimentally tested range of 10–100 mA/cm2 . The dissolution of the electrodes in the acidic media, under the combined application of electrolysis and sonication, was checked. The maximum concentration on Ni ions was less than 1 mg/l, which was negligible compared to the electrolyte concentration.

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All experiments were carried out under initial concentration of 0.25–1.5 mM 2,4-D and 0.35–1.5 mM 2,4DCP and pH 3. In order to address the effect of variable concentration of Fenton’s reagent upon the decomposition of 2,4-D and 2,4-DCP, the experiments were carried out by using Fe2þ in the range from 0.5 to 50 mM, while keeping H2 O2 unchanged at a level of 30 mM. Stock solutions (250 ml each) at various reagent concentrations were prepared, and samples were inserted into the reaction cell and left for given durations. They were taken out of the process at the specified time and analyzed. In order to ensure that the solutions left in the reactor are not affected by the analytical sampling, a fresh sample was used for each time-duration point in our diagrams. Classical Fenton experiments were also carried out for comparison, under the following conditions: 2,4-D concentration of 1,2 mM, Fe2þ ––3.0 mM and H2 O2 –– 3.0 mM. All experiments were performed under intensive stirring. 2.3. Analytical cross-referencing Variations of 2,4-D and 2,4-DCP concentration occurring under either sono-Fenton’s or sonoelectroFenton’s processes were readily monitored using HP 8453 UV–Visible photodiode array Hewlett Packard spectrophotometer. In order to identify stable intermediates formed during degradation of 2,4-D and 2,4-DCP, the organic components present in small samples (5 ml) of the tested solution were further extracted using 5 ml of dichlorometane. The solvent was allowed to evaporate and the resulting analyte was then tested by GC and GC-MS. Hewlett-Packard gas chromatograph (model 5890 with FID detector) and Finnigan TSQ-70B mass-spectrometer were used. The chloride ion concentration in the sampled solutions was determined by potentiometric titration using AgNO3 .

3. Results and discussion

Fig. 1. Schematic diagram of sono-electro-oxidation facility: (1) potentiostat, (2) electrodes, (3) 20 kHz transducer, (4) water jacketed reactor, (5) frequency generator, (6) water criostat.

In this study we intend to evaluate the performance of sono-electrochemical Fenton process for decomposition of 2,4-D and 2,4-DCP in water. As a reference, we should consider traditional oxidation processes. We examined several such processes for degradation of chloroorganic pollutants in aqueous solutions. First, we present the performance of the classical Fenton oxidation process, which takes place in the presence of Fe2þ ions. (Hereafter we shell use the term ‘‘degradation’’ for decomposition of the toxic compounds and their derivatives, such as 2,4-D and 2,4-DCP, chlorinated phenols and quinones. Full degradation means full removal of these compounds from the reaction mixture; we assume

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3.0

1.2 mM 2.4-D

Initial concentration Fenton (60s)

2.5 Absorbance

of the electrochemical oxidation of hydrophilic phenols [12] and the homolytical reactions of phenoxyacetic acid [3] (reaction (3) and (4)):

2.0 SF (60s)

1.5 1.0

SEF (60s)

0.5 0.0 250

275 300 Wavelength / nm

325

Fig. 2. UV absorbance spectra of as-prepared 1.2 mM of 2,4-D; and following 60 s of Fenton, SF and SEF processes. [Fe2þ ] ¼ 3.0 mM, [H2 O2 ] ¼ 3.0 mM.

that the remaining chloroorganic non-aromatic compounds offer no significant problem for standard methods of effluent water purification.) The results for classical Fenton oxidation of 2,4-D are shown in Fig. 2, where the initial spectrum and that obtained after application of Fenton process for 7 h, are provided. Clearly, sizeable changes are observed only after a rather long time. The changes are mainly attributed to the Fenton oxidation, since the half lifetime of 2,4-D in ambient water is much longer (6–170 days, depending on the environmental conditions it is kept under) [36]. In contrast to the above results, dramatic increase in the degradation rate for both 2,4-D and 2,4-DCP occurs with either sono-Fenton (SF) or sono-electrochemical Fenton (SEF) schemes. Typical results obtained after just 60 s of application of SF and SEF to aqueous solutions of 2,4-D are shown in Fig. 2. A considerably much faster decomposition of the starting compound is observed in both processes. The faster degradation of 2,4-D and 2,4-DCP within SF and SEF schemes (reactions (1) and (2)) is believed to be due to the high efficiency for the production of  OH radicals as well as to the ultrasonic cleaning of electrode’s active surfaces during these processes.

ArOH ! ArO þ Hþ þ e


ð3Þ

ArOCH2 COOH ! ArOCH2 COO þ Hþ þ e


ð4Þ

We believe that the oxidative and reductive electrochemical processes and the reactions involving HO are significantly more effective under the SEF process. In all cases, degradation of halogenated organic compounds takes place via several intermediate molecules. These side-products may be further decomposed until degradation is accomplished, or may remain in the solution. Regarding the decomposition of 2,4-D and 2,4DCP, recent reports indicate that the intermediates might be even more harmful than the parent compounds themselves [38]. In order to understand the SF and the SEF processes and the variations in their efficiencies, and in order to evaluate the usefulness of these processes, we need to identify and analyze the intermediates formed. Then we can carry out a kinetic investigation of the parent molecules and intermediate compounds. 3.1. Study of the intermediates The intermediates produced in the SEF process were detectable soon after the initiation of the oxidation. The highest concentrations of such compounds were generally observed within the first 30–180 s. This is exemplified for 2,4-D in Fig. 3, where UV spectra were acquired at various times after the initiation of the SEF process on 2,4-D contaminated water. The presence of intermediate products is evident from the modifications of the original UV absorption band at 250–270 nm. The absorption at these wavelengths increases in correlation with the rapid decrease of the 2,4-D peaks. GC/MS identification patterns of the intermediates produced in 3.0 2.5

ð1Þ Ultrasonic reaction : H2 O ! H þ HO

ð2Þ

In this context, note that among very reactive species suitable for SE oxidation treatments (such as O2
, OH , HO2 , ROO ) the hydroxyl radical is certainly the most reactive [37]. It is a very strong one-electron oxidizing agent, which seldom reacts as an electron transfer reagent; it is very reactive in hydrogen atom abstraction and in electrophilic addition processes [37]. Probably phenoxy––ArO and phenoxyacetic ArOCH2 COO radicals are formed in the initial stages

Absorbance

Fenton reaction : H2 O2 þ Fe2þ ! OH
þ HO þ Fe3þ

0 sec

SEF process 2.4-D, 2 mM

2.0

30sec

1.5

60sec 120sec

1.0

180sec

0.5

250

275

300

325

Wavelength / nm Fig. 3. UV absorbance spectra taken during SEF treatment of 2 mM of 2,4-D after various times. [Fe2þ ] ¼ 3.0 mM, [H2 O2 ] ¼ 3.0 mM.

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Table 1 GC/MS identification of intermediates obtained in SEF process of 2,4-D (M + 1)þ

(M + 1)þ

Structure

Structure

þ

221

162 (M )

OCH2 COOH CI

OH CI

CI

CI

2.4-D

2.4-DCP þ

191

178 (M )

OCHO

OH CI

CI

CI

CI Formate

Quinone

219

351

O

O CI

O

O

OCH2CH2O

O Cl

O

2,4-D OCH2COOH Cl

Cl

Cl

OCHO Cl

Cl

2,4-DCP

Cl

SEF decomposition of 2,4-D, which may be responsible for these intermediate compounds, is shown in Fig. 4. Worth noting is a close similarity of the intermediate compounds obtained with SEF treatments to those reported for the UV-photolysis tests [26]. The latter, however, are by far slower processes (days) than the former ones (minutes). Moreover, a dimerization stage, which is quite common in UV-photolysis [34], seems absent in both SF and SEF processes. This could be attributed to instability of dimers within a doublecharge layer situated either nearby the electrodes (as for SEF processes) or at the cavitation-bubble interfaces (as for both SF and SEF). 3.2. Kinetic investigation

OH

Cl

CI

Trace: diphenoxyethane

both SF and SEF degradation of 2,4-D were studied. The corresponding molecules are summarized in Table 1. A suggested chemical scheme for the first stages of

Cl

CI

CI

Lactone

Cl

O CH2 C H2 O CI

CI

Cl

OH

OH

Cl

OH Cl

Cl Fig. 4. Schematic presentation of the proposed first chemical stages in SEF decomposition of 2,4-D.

Kinetic study revealed the rate constants associated with the studied decomposition. The corresponding rate constant for the decomposition of 2,4-D is 0.01 s
1 ; it is about 10 times smaller for 2,4-DCP. This finding is somewhat surprising, since analysis of the intermediate products obtained during SEF decomposition of 2,4-D shows that 2,4-DCP is formed. Therefore, one might have anticipated SEF decomposition of such intermediates to be the rate limiting process for the decomposition of 2,4-D herbicide. According to our observations, however, this was not the case. This finding could be attributed to the partial contribution of 2,4-DCP to the

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100

(a)

0.8 0.6 0.4 [2,4-DCP] 0.2 [Formate] [2,4-D]

0.0 0

10

20

30

40

50

60

70

80

1.6 1.2

0.4

0.8

0.2

0.4

Cl- Concentration / mM

(b)

0.6

[2,4-D] 0.0

0.0 100

200

300 400 Time / s.

40 20 SF 2,4-DCP 0.2

0.4

0.6

0.8

1.0

SF 2,4-D 1.2

1.4

Initial concentration / mM

[Chloride-ions]

0

SEF 2,4-DCP

0.0

2.0

1.0 0.8

60

0

Time / s.

2,4-D Concentration / mM

SEF 2,4-D

80 Degradation / %

Concentration / mM

1.0

500

600

Fig. 5. (a) Kinetic monitoring of the main pollutant (2,4-D) and the major intermediate compounds produced in SEF process (2,4-DCP and 2,4-dichlorophenolformate). (b) Build-up of the chloride-ions and the pollutant decomposition during application of SEF process to 2,4D. [Fe2þ ] ¼ 3.0 mM, [H2 O2 ] ¼ 3.0 mM.

overall byproducts, or to differences in the reaction schemes in the decomposition of these compounds, induced by the presence of the other intermediates. This point requires further investigation. Typical decomposition kinetics obtained for 0.9 mM aqueous solution of 2,4-D as well as a fast formation of the intermediate products are exemplified in Fig. 5a for SEF. Quite naturally, degradation of 2,4-D is associated with formation of chloride ions, which soon becomes the main byproduct. This is exemplified in Fig. 5b, where after 10 min of oxidation, close to 99% of chloride ions, which corresponds to a practical absence of 2,4-D, was detected. Only trace amounts of smaller chlorinated components were still present. In other words, a practically full degradation of 2,4-D down to CO2 and chloride-ions seems feasible by the SEF scheme. 3.3. Efficiency of SE and SEF processes The efficiency of either SE or SEF processes was found to decrease as a function of increasing concentration of the parent compound. The degradation percentages obtained within the first 60 s of these treatments are shown as a function of the initial con-

Fig. 6. The initial concentration dependent efficiency of material decomposition: degradation of 2,4-D (including the 2,4-DCP byproduct) and 2,4-DCP in SEF and in SF processes. For both treatments the data correspond to 60 s long application. [Fe2þ ] ¼ 3.0 mM, [H2 O2 ] ¼ 3.0 mM.

centration. Results are presented for decomposition of 2,4-D and for decomposition of 2,4-DCP (Fig. 6). Note that the decomposition of 2,4-D forms some 2,4-DCP as intermediate, therefore, the figures presented here regarding this material refer to the decomposition of the parent compound and its intermediate. Observe that in respect of decomposition of 2,4-D and 2,4-DCP the SEF process begins to sizably depend on the initial concentration from about 0.8 mM for the former and around 0.4 mM for the latter compound. This finding is consistent with the previously noted kinetic data suggesting that despite of 2,4-DCP being an intermediate stage for the 2,4-D decomposition process, this does not affect the decay rates obtained for the latter herbicide. In the SE process, the decrease in decomposition efficiency starts much sooner for both materials studied. This difference may be associated with the enhanced activity of the oxidative  OH radicals inside the doublecharged layers nearby the electrodes (as in SEF) in addition to the already present sono-chemical processes at the bubbles’ interfaces (as in both SF and SEF processes). Overall, the SEF scheme appears to be more efficient either in decomposing the parent compounds or their respective intermediates. 3.4. Influence of Fe2þ concentration For many practical purposes the amount of Fe2þ used in either SF or SEF processes might be limited by various environmental regulations [39]. Sensitivity of decomposition efficiency for both target compounds to the amount of Fe2þ used is presented in Fig. 7 for the SEF process. The efficiency of degradation is peaked around 2 mM for both parent compounds. These results suggest that within the SEF scheme the Fenton processes appear to be quite a powerful source of oxidative

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Fig. 7. Dependence of the efficiency of the SEF process upon [Feþ ] concentration. Decomposition of 1.2 mM 2,4-D and of 0.5 mM 2,4DCP was performed. The concentration of H2 O2 (3.0 mM) was held unchanged.

radicals generating  OH radicals from H2 O2 in the presence of Fe-ions. The details of the results shown in this Figure are not fully understood and require further investigation. However, it should be noted that typical homogeneously catalyzed reactions of this kind in aqueous solutions need up to 30–60 mM of Fe-ions to be present [3,25]. This could be a considerable advantage of the SEF process over homogeneously catalyzed degradation.

4. Conclusions Feasibility of the new SEF method as an effective electrochemical oxidation process for destruction of chloroorganic compounds was demonstrated. The chemical intermediates involved in this process were identified and their kinetic profile was studied. It was found out that the actual efficiency of the SEF process is currently much higher than all other reference degradation processes. Thus, the time required for full degradation was considerably shorter for the SEF scheme. Moreover, when comparing the efficiency drop with increasing the initial concentrations, it seems that the SEF process maintains its high performance up to considerably higher concentrations. The optimum concentration of Fe2þ ions required for this process was found to be of about 2 mM, which is considerably lower than in the reference techniques (in some cases by more than an order of magnitude). This low concentration allows for environmental application of the SEF process. Pilot conditions for performing SEF induced degradation of target compounds appear to be both readily engineered and of low energy consumption (the ultra-

371

sound power used in SEF reactor is 100 W/l and the typical time duration for full degradation of either 2mM 2,4-D or its toxic metabolite––2,4-DCP is no longer than 600 s). It seems that further enhancement of the degradation efficiency may be achieved with higher frequency ultrasound sources (500 kHz) as well as in flowing electrochemical reactor. Prospective reagents for Fenton processes appear to be Fe and Cu ions. Investigation of SEF processes in regards of other contaminants, such as polycyclic aromatic compounds is planed. In this study we have proven the feasibility of the SEF process and its advantages over other techniques, however, we have not resolved the details of its mechanism. We believe that it is related to the oxidative and reductive electrochemical processes and the reactions involving HO , which are of enhanced efficiency under the SEF process. In this respect, note also the similarity of the intermediate compounds obtained with SEF to those known for UV-photolysis. This may be an indication of the mechanism, however, the SEF is a much faster process. It seems that further investigation of the mechanisms involved in SEF is needed. Acknowledgements This research was supported by the Grand Water Research Institute (Technion) and by the James Frank Program in Laser––Matter Interaction. Y.Y., V.B. and V.V.G. are grateful for financial support by the Israel Ministry of Absorption provided to new immigrant scientists. References [1] Industry Task Force II on 2,4-D research Data Web Site, 1999. Available from . [2] D. Kaiumova, F. Kaiumov, G. Opelz, C. Susal, Chemosphere 43 (2001) 801. [3] R. Brown, S. Jamison, U. Pandit, J. Pinkus, G. White, H. Braendlin, J. Org. Chem. 29 (1964) 146. [4] J. Pignatello, Environ. Sci. Technol. 26 (1992) 944. [5] Y. Sun, J.J. Pignatello, Agric. Food Chem. 41 (1993) 308. [6] Y. Sun, J.J. Pignatello, Agric. Food Chem. 41 (1993) 1139. [7] J. Matos, J. Laine, J. Hermann, J. Catal. 200 (2001) 1010. [8] W. Chu, Chemosphere 44 (2001) 935. [9] A. Modestov, V. Glezer, I. Marjasin, O. Lev, J. Phys. Chem. B 101 (1997) 4623. [10] M. Trillas, J. Peral, X. Domenech, Appl. Catal. B: Environ. 5 (1995) 377. [11] E. Piera, J. Calpe, E. Brillas, X. Domenech, J. Peral, Appl. Catal. B: Environ. 27 (2000) 169. [12] N. Bunce, S. Merica, J. Lipkowski, Chemosphere 35 (1997) 2719. [13] U. Galla, P. Kritzer, J. Bringmann, H. Schmieder, Chem. Eng. Technol. 3 (2000) 230. [14] Q. Wang, A. Lemley, Environ. Sci. Technol. 35 (2001) 4509. [15] F. Trabelsi, H. A€ıt-Lyazidi, B. Ratsimba, A. Wilhelm, H. Delmas, P.-L. Fabre, J. Berlan, Chem. Eng. Sci. 51 (1996) 1857. [16] M. Oturan, J. Appl. Electrochem. 30 (2000) 475.

372

Y. Yasman et al. / Ultrasonics Sonochemistry 11 (2004) 365–372

[17] E. Brillas, J. Calpe, J. Casado, Water Res. 34 (2000) 2253. [18] A. Weissler, H. Cooper, S. Snyder, J. Am. Chem. Soc. 72 (1950) 1769. [19] A.V.M. Parke, D. Taylor, J. Chem. Soc. (1956) 4442. [20] B.H. Jennigs, S.N. Townsend, J. Phys. Chem. 65 (1961) 1574. [21] K. Inazu, Y. Nagata, Y. Maeda, Chem. Lett. (1993) 57. [22] I. Hua, M. Hoffmann, Environ. Sci. Technol. 30 (1996) 864. [23] K. Hirai, Y. Nagata, Y. Maeda, Ultrason. Sonochem. 3 (1996) S205. [24] C. Petrier, Y. Jiang, M. Lamy, Environ. Sci. Technol. 32 (1998) 1316. [25] B. Lin, Y. Ma, J. Environ. Eng. 2 (2000) 130. [26] K. Teo, Y. Xu, C. Yang, Ultrason. Sonochem. 8 (2001) 241. [27] Y. Ku, K.-Y. Chen, K.-C. Lee, Water Res. 31 (1997) 929. [28] H. Okuno, B. Yim, Y. Mizukoshi, Y. Nagata, Y. Maeda, Ultrason. Sonochem. 7 (2000) 261.

[29] J. Peller, O. Wiest, P. Kamat, J. Phys. Chem. A 105 (2001) 3176. [30] T.J. Mason, J.P. Lorimer, Applied Sonochemistry, Wiley VCH, 2002. [31] E. Hart, C. Fischer, A. Henglein, Radiat. Phys. Chem. 36 (1990) 511. [32] Mi
sik, N. Miyoshi, P. Riesz, J. Phys. Chem. 99 (1995) 3605. [33] E.B. Flint, K.S. Suslick, Science 253 (1991) 1357. [34] K. Mokino, M. Mossoba, P. Riesz, J. Phys. Chem. 87 (1983) 1369. [35] A. Henglein, Ultrasonics 25 (1987) 6. [36] V. Chang, J. Liu, S. Yuan, Sci. Total Environ. 215 (1998) 1. [37] M. Oturan, J.J. Pinson, J. Phys. Chem. 99 (1995) 13948, and references cited there. [38] A. Hirvonen, M. Trapido, J. Hentunen, J. Tarhanen, Chemospere 41 (2000) 1211. [39] EPA 40CFR Ch.1 (7-1-00 edition) Part 143. National Secondary Drinking Water Regulations.