- Email: [email protected]
Degradation of the herbicide 2,4-dichlorophenoxyacetic acid by ozonation catalyzed with Fe2+ and UVA light
Applied Catalysis B: Environmental 46 (2003) 381–391
Degradation of the herbicide 2,4-dichlorophenoxyacetic acid by ozonation catalyzed with Fe2+ and UVA light Enric Brillas∗ , Juan Carlos Calpe, Pere-Llu´ıs Cabot Laboratori de Ciència i Tecnologia Electroqu´ımica de Materials, Departament de Qu´ımica F´ısica, Facultat de Qu´ımica, Universitat de Barcelona, Mart´ı i Franquès 1-11, 08028 Barcelona, Spain Received 8 March 2003; received in revised form 2 June 2003; accepted 16 June 2003
Abstract Solutions of the herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) with concentrations up to near saturation at pH 3.0 and at 25 ◦ C have been treated with ozone and ozonation catalyzed with Fe2+ and/or UVA light. Direct ozonation yields a slow depollution, while all contaminants are completely removed under UVA irradiation. The highest oxidizing power is achieved when Fe2+ and UVA light are combined, since greater amounts of oxidizing hydroxyl radical are generated and Fe3+ complexes are photodecomposed. The initial mineralization rate is enhanced when herbicide concentration increases and more hydroxyl radicals are produced by the catalyzed ozonation processes. The herbicide decay always follows a pseudo first-order reaction. Reverse-phase chromatography allows the detection and quantification of aromatic intermediates such as 2,4-dichlorophenol, 4,6-dichlororesorcinol and chlorohydroquinone. In all treatments, fast dechlorination reactions take place leading to chloride ion accumulation in the medium. The evolution of generated carboxylic acids such as glycolic, glyoxylic, maleic, fumaric and oxalic has been followed by ion-exclusion chromatography. Only oxalic acid remains stable in the O3 system, being quickly mineralized to CO2 by hydroxyl radicals formed in the O3 /UVA one. A high stability of oxalic acid in the O3 /Fe2+ system has also been found, since it yields Fe3+ -oxalato complexes. These species are photodecarboxylated under UVA irradiation in the O3 /Fe2+ /UVA system. A possible reaction pathway for 2,4-D mineralization involving all intermediates detected is proposed. © 2003 Elsevier B.V. All rights reserved. Keywords: 2,4-Dichlorophenoxyacetic acid; Ozone; Fe2+ catalyst; UVA light; Water treatment; TOC removal; Oxidation products
1. Introduction Ozone is a chemical agent widely used for the mineralization (i.e. conversion into CO2 and inorganic ions) of pesticides and related biorecalcitrant organic contaminants in waters [1–3]. However, the high energy cost for its generation limits many practical applications of direct ozonation (O3 system) to ∗ Corresponding author. Tel.: +34-93-4021223; fax: +34-93-4021231. E-mail address: [email protected] (E. Brillas).
water treatment. Alternative approaches have then been taken to try to improve its oxidizing power, also reducing its economic cost. In this way, more efficient pollutant removals have been reported by combining ozonation with UV light [1,4–9], homogeneous catalysts such as H2 O2 [1,10,11] and Fe2+ [7–9,12,13], and photocatalysis with TiO2 [7,14–21]. It is well known that when the O3 system is used to degrade acid waters, the organic matter is directly oxidized by molecular ozone, which can attack selectively pollutants, especially aromatic compounds,
0926-3373/$ – see front matter © 2003 Elsevier B.V. All rights reserved. doi:10.1016/S0926-3373(03)00266-2
382
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
following second-order reactions with small rate constants [1–3,10,22,23]. In contrast, the different catalyzed ozonation systems have greater degradative ability because of the production of a stronger, non-selective oxidizing agent such as hydroxyl radical (OH• ). This species can react more rapidly with organics yielding dehydrogenated or hydroxylated derivatives. Under irradiation of the solution with UV light at λ > 300 nm (O3 /UVA system), the main source of OH• is the direct reaction of ozone with H2 O2 , previously formed from its initial photolysis [1]: O3 + H2 O + hν → H2 O2 + O2
(1)
O3 + H2 O2 → OH• + HO2 • + O2
(2)
In this process, weaker oxidizing hydroperoxyl radical (HO2 • ) is also generated. When Fe2+ is present in the acid solution (O3 /Fe2+ system), it catalyzes ozone decomposition to give FeO2+ as intermediate by reaction (3) [12]. This species can further evolve to OH• by reaction (4) or can oxidize Fe2+ to Fe3+ at slower rate by reaction (5), then limiting the generation of hydroxyl radical for high Fe2+ concentration [7]. Fe2+ + O3 → FeO2+ + O2
(3)
FeO2+ + H2 O → Fe3+ + OH• + OH−
(4)
FeO2+ + Fe2+ + 2H+ → 2Fe3+ + H2 O
(5)
A faster oxidation of some organic pollutants by Fe2+ catalyzed ozonation under UVA irradiation (O3 /Fe2+ /UVA system) has been recently reported [7–9]. The acceleration of the mineralization process can be accounted for several parallel ways involving: (i) photolysis of complexes of Fe3+ with generated carboxylic acids, e.g. oxalic acid [24]; (ii) photoreduction of Fe3+ to give Fe2+ and an additional OH• via reaction (6) [25]; and (iii) additional oxidation of Fe2+ with H2 O2 generated from reaction (1) to yield more OH• by the classical Fenton reaction (7) [26,27]. Fe3+ + H2 O + hν → Fe2+ + OH• + H+
(6)
Fe2+ + H2 O2 → Fe3+ + OH• + OH−
(7)
Fe2+ regeneration from reaction (6) propagates the catalytic reaction (3) and larger amounts of oxidizing
OH• can be formed by reaction (4). In addition, the resulting Fe3+ from Fenton reaction (7) can be photoreduced again by reaction (6). 2,4-Dichlorophenoxyacetic acid (2,4-D) is a herbicide with potential toxicity for humans and animals that is used wordwide on a large scale for the selective control of weeds in gardens and farming. It is poorly biodegradable and has been detected as major pollutant in ground and surface waters [28,29]. For this reason, there is great interest in the study of oxidation methods with ability enough to mineralize 2,4-D in aqueous medium to try to avoid its dangerous accumulation in the aquatic environment. Recently, the degradation of this herbicide has been explored by different chemical, photochemical and photocatalytic systems such as O3 [2–4,7,22,23], O3 /Fe2+ [7], O3 /H2 O2 [11], H2 O2 /Fe2+ (Fenton’s reagent) [26,27], H2 O2 /UV [11], H2 O2 /Fe3+ /UV [26–29], TiO2 /UV [14–17,21], O3 /UVA [4,7], O3 /TiO2 /UVA [7] and O3 /Fe2+ /UVA [7], as well as by advanced electrochemical oxidation treatments [29–32]. Although an efficient depollution of 2,4-D solutions has been reported for different catalyzed ozonation processes, less is known about the degradative paths involved in them. This knowledge is necessary to understand the behavior of such treatments, reason for which more exhaustive research is required considering valuable information on the evolution of this herbicide and by-products formed. In this paper, we present a detailed study on the degradative behavior of 2,4-D in acidic medium by ozonation processes catalyzed with Fe2+ and/or UVA light. Comparative treatments for herbicide concentrations up to near saturation (ca. 620 ppm at 25 ◦ C) have been carried out at pH 3.0 using the O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. For the two last methods, 1 mM Fe2+ has been chosen as catalyst concentration, since higher content of this ion in solution accelerates the competitive reaction (5) limiting the destruction of organics [7]. The same solutions have also been treated by the O3 system to confirm the higher efficiency of the catalyzed ozonation methods. In each process, the decay kinetics of the herbicide and the evolution of its aromatic products and generated carboxylic acids have been followed by chromatographic techniques. A general reaction sequence for the mineralization
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
of the herbicide involving all the detected species is proposed. 2. Experimental 2.1. Chemicals 2,4-D, 2,4-dichlorophenol, 4,6-dichlororesorcinol, chlorohydroquinone, glycolic acid, glyoxylic acid, maleic acid, fumaric acid and oxalic acid were reagent grade from Merck, Fluka and Panreac. All the acidic solutions tested were prepared with high-purity water obtained from a Millipore Milli-Q system, with conductivity <6 × 10−8 S cm−1 at 25 ◦ C. The initial pH of each solution was adjusted with analytical grade sulfuric acid supplied by Merck. Analytical grade heptahydrated ferrous sulfate used as catalyst was purchased from Fluka. Organic solvents and the other chemicals employed were either HPLC or analytical grade from Merck, Fluka and Aldrich. 2.2. Instruments Ozone was generated with an Erwin Sander 300.5 ozonizer fed with pure O2 at 1.1 bar and at a flow rate of 60 l h−1 . The solution pH was measured with a Crison 2000 pH-meter. Total organic carbon (TOC) was determined on a Shimadzu 5050 TOC analyzer. Reverse-phase chromatography was performed with a Waters 600 HPLC liquid chromatograph, fitted with a Spherisorb ODS2 5 m column (150 mm × 4.6 mm (i.d.)) at room temperature, along with a Waters 996 photodiode array detector selected at 286 nm, controlled through a Millennium-32® program. The same HPLC chromatograph fitted with a Bio-Rad Aminex HPX 87H column (300 mm × 7.8 mm (i.d.)) at 35 ◦ C, in conjunction with the photodiode array detector selected at 210 nm, were used for ion-exclusion chromatography. Gas chromatography-mass spectrometry (GC-MS) was carried out with a Hewlett-Packard system consisting of a HP 5890 gas chromatograph fitted with a HP-1 0.25 m column (120 m ×0.25 mm (i.d.)) and a HP 5989A mass spectrometer operating in El mode at 300 ◦ C. Chloride ion concentration was determined by potentiometric titration with AgNO3 using a Metrohm Titrino 702 SM automatic titrator, with a detection limit of 5 ppm for Cl− .
383
2.3. Ozonation system All degradation experiments were conducted in an open, thermostated cylindrical Pyrex cell containing 100 ml of herbicide solution under stirring with a magnetic bar. The temperature was maintained at 25.0 ± 0.1 ◦ C. The O2 + O3 mixture generated by the ozonizer was supplied to the solution through a stainless steel diffuser. Its flow rate was regulated to 20 l h−1 yielding a constant production of 1.3 g O3 h−1 , as determined by standard iodometry. A 6 W Philips fluorescent black light blue tube was used to irradiate the solution with UVA light. This tube emitted in the wavelength range 300–420 nm, with λmax = 360 nm, and was placed at the top of the open cell at 4 cm over the sample. The flux of incident photons by unit reactor volume was 8.3×10−7 einstein l−1 min−1 , as detected with an uranyl actinometer active to photons in the range 250–500 nm. Solutions with a 2,4-D concentration between 608 ppm (close to saturation) and 5 ppm were acidified with 0.25 M H2 SO4 to pH 3.0 and further mineralized by the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. For the two last methods, 1 mM FeSO4 was added to each sample as catalyst before treatment. 2.4. Analysis procedures Before analysis of samples extracted from degraded solutions, they were filtered with 0.45 m PTFE filters supplied by Whatman. The mineralization of herbicide solutions was monitored for their TOC decay, using the standard non-purgeable organic carbon (NPOC) method. The evolution of 2,4-D and its aromatics products was followed by reverse-phase chromatography, with a 50:45:5 (v/v/v) methanol/phosphate buffer (pH 2.5)/pentanol mixture as eluent at 1.0 ml min−1 . For the quantification of generated carboxylic acids by ion-exclusion chromatography, 4 mM H2 SO4 was employed as eluent at 0.6 ml min−1 . In both techniques, 20 l aliquots were injected into the HPLC chromatograph. To identify the stable intermediates formed during the mineralization of 2,4-D, the organic components of several solutions degraded by the different systems for short times were extracted with 100 ml of dichloromethane and after evaporation of the solvent, the remaining solids were dissolved in 5 ml of methanol to be further analyzed by GC-MS.
384
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
3. Results and discussion 3.1. Comparative TOC removal All initial 2,4-D solutions were adjusted to pH 3.0 to be further treated by the different ozonation methods at 25 ◦ C during 2 h as maximum. A gradual depollution with raising degradation time was always found. Reproducible values for TOC removal were obtained under all experimental conditions tested. In these trials, a slow acidification was observed, reaching the final solutions a pH value between 2.5 and 2.9. As an example, Fig. 1 shows the comparative TOC abatement for 100 ml of a 230 ppm herbicide solution (with 100 ppm of initial TOC) using the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. From these data, one can determine an initial mineralization rate of 104 ppm TOC h−1 (Fig. 1a), 116 ppm TOC h−1 (Fig. 1c), 456 ppm TOC h−1 (Fig. 1b) and 612 ppm TOC h−1 (Fig. 1d), respectively. These results indicate that the oxidizing ability of the O3 system at the beginning of the process is accelerated in the presence of Fe2+ , as expected if the herbicide and its oxidation products are more quickly destroyed by OH• produced from reaction (4) than by ozone. The fact that a much higher initial TOC decay is found for the O3 /Fe2+ system than for the O3 /UVA one, evidences more efficient generation of OH• from reaction (4) than from reaction (2). The highest initial
Fig. 1. TOC removal with time for the mineralization of 100 ml of a 230 ppm 2,4-D solution of pH 3.0 under the following conditions: (a) O3 (䊊); (b) O3 + 1 mM Fe2+ (䊐); (c) O3 + UVA (䉭); (d) O3 + 1 mM Fe2+ + UVA (䉫). Ozone flow rate 1.3 g h−1 . Temperature 25 ◦ C.
mineralization rate for the O3 /Fe2/ UVA treatment can then be related to the production of more oxidizing OH• by the simultaneous action of reactions (2), (4), (6) and (7). The relative oxidizing ability of the above processes varies when degradation time is prolonged. As can be seen in Fig. 1a, direct ozonation gives rise to a progressive, but slow, depollution up to attain 59% of TOC removal at 2 h. Under UVA irradiation, all organics are much more rapidly mineralized by reaction with OH• formed from reaction (2) and the solution TOC is reduced by 95% after 2 h of the O3 /UVA treatment (see Fig. 1c). In contrast, a very different behavior is obtained in the presence of catalytic Fe2+ . While at the first oxidation stages of the O3 /Fe2+ system a fast degradation rate of contaminants takes place, at times longer than 30 min the mineralization process becomes very hard (see Fig. 1b) because of the formation of very stable products, probably complexes of Fe3+ with generated carboxylic acids [24]. At the end of this treatment, only a 66% of TOC is removed, a value slightly lower than 60% obtained at 30 min. When the solution is also illuminated with UVA light using the O3 /Fe2+ /UVA system, the process is much more efficient, giving a 92% of TOC removal at 1 h (see Fig. 1d). This behavior can be accounted for by the fast photodecomposition of Fe3+ complexes formed [24] and by the enhancement of the generation rate of oxidizing OH• by reactions (2), (4), (6) and (7). Similar relative degradative trends to those described above were found for solutions containing a 2,4-D concentration between 608 and 58 ppm. Complete mineralization (more than 95% of depollution) was only achieved under UVA illumination, usually after 2 h of the O3 /UVA treatment and less than 1.5 h for the O3 /Fe2+ /UVA system. The quick and total degradation obtained in the latter conditions is shown in Fig. 2. Table 1 summarizes the percentage of TOC removal determined for the solutions tested after 1 h of each treatment, along with the corresponding initial mineralization rate. A similar depollution can be observed for all herbicide solutions from each method used, since at 1 h their TOC is reduced by 36–44, 63–78, 51–64 and 85–95% for the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems, respectively. Each method is then able to mineralize an extent of organics practically independent of the initial 2,4-D concentration, indicating that at long degradation time
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
385
the action of Fe2+ . For example, for 608 ppm of 2,4-D this parameter increases rapidly from 315 ppm TOC h−1 for the O3 /UVA system to 1050 ppm TOC h−1 for the O3 /Fe2+ one, but it only raises to 1211 ppm TOC h−1 for the O3 /Fe2+ /UVA method. This confirms that reaction (4) is the main source of oxidizing OH• in the Fe2+ catalyzed ozonation processes. 3.2. 2,4-D decay and reaction kinetics
Fig. 2. TOC time-course for the mineralization of 100 ml of different 2,4-D solutions of pH 3.0 by the O3 /Fe2+ /UVA system. Herbicide concentration: 608 ppm (䊊); 519 ppm (䊐); 486 ppm (䉭); 460 ppm (䉫); 345 ppm (䊉); 230 ppm (䊏); 115 ppm (䉱); 58 ppm (䉬). Ozone flow rate 1.3 g h−1 . Temperature 25 ◦ C.
its oxidizing power remains quite constant under the experimental conditions considered, being limited by the reactive ability of its oxidants (O3 and/or OH• ). However, an inspection of Table 1 shows a fast fall of initial mineralization rate with decreasing herbicide concentration. This tendency is not surprising and can be simply related to the existence of smaller amounts of degradable organics in the medium. Comparison of the data collected in Table 1 for the different catalyzed ozonations allows the conclusion that the initial mineralization rate for all solutions is mainly enhanced by
The kinetics for the disappearance of 2,4-D by reaction with the oxidizing species involved in each treatment was studied using a 230 ppm herbicide solution. The decay of this compound was followed by reverse-phase chromatography, where it displayed a well-defined peak with a retention time (tr ) of 3.65 min. The evolution of its concentration is depicted in Fig. 3. As can be seen, 2,4-D is rapidly destroyed for all procedures, disappearing from the medium in a time ranging between 8 and 14 min. The fast herbicide removal by the O3 system is due to its direct attack with molecular ozone, having a second-order rate constant (k2 ) of 298 M−1 s−1 [3]. Its quicker destruction from the different catalyzed ozonations can then be ascribed to its parallel reaction with the stronger oxidizing OH• produced in all them, because this compound is not photolyzed by UVA light. Herbicide concentration decays reported in Fig. 3 were analyzed from kinetic equations related to different reaction orders. Good linear plots, with
Table 1 Percentage of TOC removal obtained after 1 h of treatment of 100 ml of different 2,4-D solutions at pH 3.0 and at 25 ◦ C by O3 and catalyzed ozonation processes [2,4-D]0 (ppm)a
Method
608 519 486 460 345 230 115 58
41 36 38 37 36 44 42 42
O3 (192) (169) (179) (159) (116) (104) (58) (31)
O3 /UVA
O3 /Fe2+
O3 /Fe2+ /UVA
64 63 68 77 75 76 78 76
52 54 61 63 63 62 64 51
93 92 93 93 95 92 86 85
(315) (219) (206) (205) (176) (116) (63) (27)
(1050) (877) (895) (893) (753) (456) (156) (84)
(1211) (1167) (1075) (1062) (828) (612) (188) (74)
The corresponding initial mineralization rate (ppm TOC h−1 ) is given in parenthesis. a Initial 2,4-D concentration.
Fig. 3. 2,4-D concentration decay for a 230 ppm herbicide solution under the same experimental conditions as in Fig. 1. Applied system: (a) O3 (䊊); (b) O3 /Fe2+ (䊐); (c) O3 /UVA (䉭); (d) O3 /Fe2+ /UVA (䉫). The panel inset presents the corresponding kinetic analysis assuming a pseudo first-order reaction for 2,4-D.
386
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
regression coefficients >0.997, were only obtained when they were fitted to a pseudo first-order reaction. This kinetic analysis is presented in the inset of Fig. 3, giving a pseudo first-order rate constant (k1 ) of 6.3 × 10−3 , 7.3 × 10−3 , 9.4 × 10−3 and 1.17 × 10−2 s−1 for the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems, respectively. This behavior suggests the presence of a steady concentration of oxidizing species (O3 and/or OH• ) in all systems during their initial degradation times, so that the corresponding pseudo first-order rate constant can be expressed, in general, as follows: k1 = k2 [O3 ] + k [OH• ] k
(8)
denotes the second-order rate constant for where the reaction of 2,4-D with hydroxyl radical. Thus, taking k1 = 6.3 × 10−3 s−1 , k2 = 298 M s−1 [3] and k = 0 for the O3 system, one can determine from Eq. (8) that injected ozone acts with a steady concentration of 2.1 × 10−5 M. The increase in k1 for the catalyzed ozonations depends on the relative proportion of OH• produced in them, since k > k2 . The higher k1 value for the O3 /Fe2+ system (9.4 × 10−3 s−1 ) than for the O3 /UVA one (7.3×10−3 s−1 ) can be explained again by the much more efficient OH• generation in the presence of Fe2+ from reaction (4) than under UVA illumination from reaction (2). When both catalysts Fe2+ and UVA light are combined, the additional OH• enhancement by reactions (6) and (7) can justify the highest k1 value of 1.17 × 10−2 s−1 found for the O3 /Fe2+ /UVA treatment. To gain a better insight into the competitive reactions of organics with oxidizing species taking place in the O3 /Fe2+ /UVA system, the influence of herbicide concentration upon its decay kinetics was studied. As can be seen in Fig. 4, 2,4-D is more slowly destroyed when its starting concentration increases. For example, the time needed for its complete removal is about 18 min for 608 ppm, but only close to 8 min for 230 ppm. This behavior is also reflected by the concomitant drop of its pseudo first-order rate constant. From the corresponding kinetic analysis depicted in the inset of Fig. 4, decreasing k1 values of 1.17×10−2 , 9.1 × 10−3 , 7.0 × 10−3 , 6.6 × 10−3 , 6.0 × 10−3 and 5.8 × 10−3 s−1 (with regression coefficients >0.997) are obtained for increasing 2,4-D concentrations of 230, 345, 460, 486, 519 and 608 ppm, respectively. This trend can be explained by a progressive consump-
Fig. 4. Destruction of 2,4-D concentration with time for same solutions of Fig. 2 degraded by the O3 /Fe2+ /UVA system. Initial herbicide concentration: 608 ppm (䊊); 519 ppm (䊐); 486 ppm (䉭); 460 ppm (䉫); 345 ppm (䊉); 230 ppm (䊏). The corresponding kinetics analysis considering a pseudo first-order reaction for the herbicide is shown in the inset.
tion of more injected ozone when reacts directly with higher amounts of herbicide, and also of some oxidation products produced, so that a smaller proportion of this oxidant is transformed into OH• from reactions (2) and (4) leading to a gradual tall in k1 , as predicted from Eq. (8). These findings allow us to consider that in the O3 /Fe2 /UVA system, the degradative path via OH• is favored at low pollutant concentrations, whereas reactions with ozone become more important as greater amounts of contaminants are contained in the solution. 3.3. Identification and evolution of oxidation products Several 230 ppm 2,4-D solutions were treated by the different systems for 10 min to detect the stable aromatic intermediates formed by GC-MS. For all procedures, the gas chromatogram of collected organics displayed three peaks which were identified by mass spectrometry and ascribed to the primary product 2,4-dichlorophenol [m/e = 162 (100, M+ ), 164 (69 (M + 2)+ ), 166 (10 (M + 4)+ )] and its derivatives 4,6-dichlororesorcinol (m/e = 178 (100, M+ ), 180 (66 (M + 2)+ ), 182 (10 (M + 4)+ )] and clorohydroquinone [m/e = 144 (100, M+ ), 146 (25 (M + 2)+ )]. Reverse-phase chromatograms of the same degraded solutions also exhibited peaks related to chlorohydroquinone at tr = 1.69 min, 4,6-dichlororesorcinol
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
387
Fig. 5. Evolution of the concentration of: (a) 4-chlorophenol; (b) 4,6-dichlororesorcinol; (c) chlorohydroquinone; (d) chloride ion, during the treatment of a 230 ppm 2,4-D solution of pH 3.0 at 25 ◦ C by the system: O3 (䊊); O3 /Fe2+ (䊐); O3 /UVA (䉭); O3 /Fe2+ /UVA (䉫).
at tr = 2.34 min and 2,4-dichlorophenol at tr = 5.75 min. These peaks were unequivocally identified by comparing their retention times and UV-Vis spectra with those of pure compounds, measured on the photodiode array detector. Similar oxidation products have been detected during the degradation of 2,4-D by photocatalysis combined with TiO2 [14–16] and advanced electrochemical methods [30–32]. The evolution of the above aromatics intermediates during the treatments of the 230 ppm 2,4-D solution was followed by determining their concentrations as a function of degradation time with calibration via standard compounds. Fig. 5a shows that the primary product 2,4-dichlorophenol is rapidly formed and destroyed in all cases, being completely removed at 25–30 min. Surprisingly, a greater maximum concentration of this product is accumulated between 4 and 6 min as the applied treatment pro-
duces more oxidizing OH• , varying from 2.1 ppm for direct ozonation to about 5 ppm for the O3 /Fe2+ /UVA system. This suggests that direct reaction of 2,4-D with O3 and/or OH• can also give other undetected products, although more OH• generation favors the formation of 2,4-dichlorophenol. Similar trends can be observed in Fig. 5b and c for its derivatives 4,6-dichlororesorcinol and chlorohydroquinone, respectively, although both species are accumulated in less extent and more rapidly removed from the medium, usually in 15–20 min. The little effect of UVA light on the evolution of all aromatic intermediates discards their direct photodecomposition. Mineralization of 2,4-D leads to the loss of chlorine atoms of its aromatic ring. The overall rate of dechlorination reactions involved in such process was followed by determining the concentration of Cl− accumulated in the above-degraded solutions. The
388
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
evolution of this ion is presented in Fig. 5d. As can be seen, 74 ppm of Cl− corresponding to 100% of initial chlorine are found in the medium after ca. 20 min of the O3 /Fe2+ /UVA treatment, i.e. when all the initial herbicide (see Fig. 3) and its detected chloroaromatic derivatives (see Fig. 5a–c) have been removed. For the other methods with smaller oxidizing power, a slower Cl− accumulation takes place, although more than 70 ppm of this ion is released after 30 mim (see Fig. 5d). These findings are indicative of the existence of fast dechlorination reactions at the beginning of the degradative process of 2,4-D with formation of stable chloride ion in all procedures tested. Ion-exclusion chromatograms of the same treated solutions exhibited peaks associated with generated carboxylic acids such as oxalic at tr = 6.62 min, maleic at tr = 7.96 min, glyoxylic at tr = 9.42 min, glycolic at tr = 12.3 min and fumaric at tr = 15.6 min. Good defined peaks for these products were obtained
from 20 min of degradation, since at lower times a large peak due to the strong adsorption at least of the initial compound inside the column prevented their quantification. Note that glycolic acid is expected to be formed during the breaking of the C(1)–O bond of 2,4-D. Glycolic acid is oxidized to oxalic acid via glyoxylic acid, while maleic and fumaric acids coming from the breaking of aromatic rings of intermediates are also converted into oxalic acid [32]. The time-course of glycolic acid concentration is depicted in Fig. 6a. A great accumulation of this compound close to 55 ppm can be observed at 20 min of direct ozonation, indicating its complete release from initial herbicide. At longer times it decays rapidly up to reach about 5 ppm from 80 min. Fig. 6a also shows that glycolic acid is more rapidly destroyed by the O3 /UVA and O3 /Fe2+ systems, disappearing at 60 min. Similar trends are found for glyoxylic (see Fig. 6b) and maleic (see Fig. 6c) acids. Both products, as well as
Fig. 6. Time-course of the concentration of generated carboxylic acids: (a) glycolic; (b) glyoxylic; (c) maleic; (d) oxalic, detected during the degradation of a 230 ppm 2,4-D solution of pH 3.0 at 25 ◦ C by the system: O3 (䊊); O3 /Fe2+ (䊐); O3 /UVA (䉭); O3 /Fe2+ /UVA (䉫).
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
fumaric acid, undergo a slow decay by direct ozonation, remaining in the medium even at 2 h, but disappear after 70–90 min of the O3 /UVA treatment. It can then be inferred that all these carboxylic acids are slowly destroyed by direct attack with molecular ozone, progressively increasing their degradation rate as more OH• is produced in the medium by the catalyzed ozonation processes. For this reason, they are always undetected when using the O3 /Fe2+ /UVA system, where the highest OH• concentration is generated from reactions (2), (4), (6) and (7).
389
However, Fig. 6d shows a very different behavior for oxalic acid, since it is removed by the O3 /UVA and O3 /Fe2+ /UVA systems, being accumulated in the O3 and O3 /Fe2+ treatments up to a steady concentration of 62 and 106 ppm, respectively. These results suggest an enhancement of the degradative path to oxalic acid when more OH• is produced. Note that Zuo and Hoignè [24] reported that this acid forms stable complexes with Fe3+ , which can be efficiently photodecarboxylated by UV light. When the O3 /Fe2+ and O3 /Fe2+ /UVA methods are applied to the 230 ppm
Fig. 7. General reaction pathway proposed for 2,4-D mineralization by O3 and ozonation catalyzed with Fe2+ and/or UVA light at pH 3.0.
390
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
2,4-D solution, Fe3+ is generated from reactions (4) and (5) and hence, a large proportion of Fe3+ -oxalate complexes is expected in the corresponding degraded solutions. The formation of such complexes limits the oxidizing ability the O3 /Fe2+ system. In fact, the steady concentration of 106 ppm of oxalic acid for this method corresponds to 29 ppm of TOC, a value very close to 34 ppm of TOC attained by the 230 ppm herbicide solution at 2 h (see Fig. 1b). The final solution of the O3 /Fe2+ treatment is then expected to be mainly composed of stable Fe3+ -oxalato complexes. The efficient destruction of such species under the action of UVA light can account for the fast and complete mineralization of 2,4-D by the O3 /Fe2+ /UVA system, thus explaining the highest oxidizing ability of this procedure. In contrast, the steady oxalic acid concentration of 62 ppm obtained for direct ozonation only corresponds to 17 ppm of TOC, i.e. a 41% of the overall TOC obtained after 2 h of degradation by this procedure (see Fig. 1a). Consequently, a larger proportion of other hardly oxidable products are present in such final solution. These products, as well as oxalic acid, are decomposed by the O3 /UVA system, yielding the overall mineralization of the herbicide. 3.4. Reaction pathways for 2,4-D degradation Fig. 7 shows a general reaction pathway for 2,4-D mineralization at pH 3.0 taking into account all intermediates detected. The main oxidizing agents are O3 for direct ozonation and O3 and OH• for catalyzed ozonations. The proposed sequence involves the degradative path to oxalic acid as the ultimate generated carboxylic acid, a way which is more important as higher amounts of OH• are produced by the catalyzed systems in the order O3 /UVA < O3 /Fe2+ < O3 /Fe2+ /UVA. The process is initiated by the attack of O3 and/or OH• on the C(1)–O bond of 2,4-D, causing the breaking of its lateral chain to give 2,4-dichlorophenol and glycolic acid. This reaction occurs in larger extent in the O3 /Fe2+ /UVA system when more amounts of OH• can react with the initial reactant. 2,4-Dichlorophenol can then be hydroxylated either on C(5)-position yielding 4,6-dichlororesorcinol or on C(4)-position leading to chlorohydroquinone with loss of a chlorine atom. This species is also a strong oxidizing agent, being reduced to chloride ion. Further degra-
dation with dechlorination of chlorobenzoquinone and 4,6-dichlororesorcinol by reaction with O3 and/or OH• gives maleic and fumaric acids, which are oxidized to oxalic acid. This acid can also be formed by a parallel path involving hydroxylation of glyoxylic acid, previously formed from dehydrogenation of the initially generated glycolic acid. The transformation of all these carboxylic acids into oxalic acid is accelerated as more OH• is generated in the medium by catalyzed ozonations. Oxalic acid remains stable in the O3 system, being mineralized to CO2 by the action of OH• produced from reaction (2) in the O3 /UVA system. In the presence of Fe2+ , stable Fe3+ -oxalato complexes are formed, which can be rapidly photodecomposed to CO2 with loss of Fe2+ [24] under UVA irradiation in the O3 /Fe2+ /UVA system.
4. Conclusions It has been demonstrated that the initial mineralization rate of acidic 2,4-D solutions in the O3 system is accelerated in the presence of Fe2+ as catalyst due to the production of oxidizing OH• by reaction (4), although their degradation is inhibited by the formation of complexes of Fe3+ with intermediates. Complete mineralization can be achieved by the O3 /UVA system, and more rapidly using the O3 /Fe2+ /UVA one, since it produces more OH• from reactions (2), (4), (6) and (7) and can photodecompose Fe3+ complexes by the action of UVA light. The initial mineralization rate is always enhanced by increasing herbicide concentration. The pseudo first-order rate constant for 2,4-D decay increases as higher OH• concentration is generated by the catalyzed ozonation process, whereas in the O3 /Fe2+ /UVA system it drops with raising herbicide concentration due to a smaller OH• production. 2,4-Dichlorophenol, 4,6-dichlororesorcinol and chlorohydroquinone have been quantified as stable aromatic intermediates by reverse-phase chromatography. The initial chlorine is released to the medium as chloride ion. Ion-exclusion chromatography reveal the generation of carboxylic acids such as glycolic, glyoxylic, maleic, fumaric and oxalic, which are slowly destroyed by direct ozonation and efficiently removed by catalyzed ozonations, except oxalic acid. This acid is stable in the O3 system, being quickly mineralized to CO2 by the O3 /UVA one. For the
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
O3 /Fe2+ system, it yields stable Fe3+ -oxalato complexes, which are photodecarboxylated to CO2 under UVA irradiation in the O3 /Fe2+ /UVA one. Acknowledgements The authors thank financial support of MCYT (Ministerio de Ciencia y Tecnolog´ıa, Spain) under project BQU2001-3712. References [1] G.R. Peyton, in: N.M. Ram, R.F. Christan, K.P. Canto (Eds.), Significance and Treatment of Volatile Organic Compounds in Water Supplies, Lewis Pu, 1990, pp. 313–362. [2] S. Chiron, A. Fernandez-Alba, A. Rodriguez, E. Garcia-Calvo, Water Res. 34 (1999) 366. [3] H. Jian-Ying, M. Takeshi, M. Yasumoto, A. Takako, Water Res. 34 (2000) 2215. [4] J. Prado, J. Arantegui, E. Chamorro, S. Esplugas, Ozone: Sci. Technol. 16 (1994) 235. [5] R. Andreozzi, R. Marotta, J. Hazard Mater. 69 (1999) 303. [6] S.W. Kuo, Chemosphere 39 (1999) 1853. [7] E. Piera, J.C. Calpe, E. Brillas, X. Domènech, J. Peral, Appl. Catal. B: Environ. 27 (2000) 169. [8] R. Sauleda, E. Brillas, Appl. Catal. B: Environ. 29 (2001) 135. [9] L. Sánchez, X. Domènech, J. Casado, J. Peral, Chemosphere 50 (2003) 1085. [10] C. Adams, P.A. Scanlan, N.D. Secrist, Environ. Sci. Technol. 28 (1994) 1812.
391
[11] G.F. Ijpelaar, M. Groenendijk, R. Hopman, J.C. Kruithof, Water Sci. Technol., Water Supply 2 (2002) 129. [12] T. Løgager, J. Holeman, K. Sehested, T. Pedersen, Inorg. Chem. 31 (1992) 3523. [13] D.L. Sedlack, J. Hoigné, M.M. David, R.N. Colvile, E. Seyffer, K. Acker, W. Wiepercht, L. Wolfgang, J.A. Lind, S. Fuzzi, Atmos. Environ. 31 (1997) 2515. [14] M. Trillas, J. Peral, X. Doménech, Appl. Catal. B: Environ. 5 (1995) 377. [15] M. Trillas, J. Peral, X. Doménech, J. Chem. Technol. Biotechnol. 67 (1996) 237. [16] L. Sánchez, J. Peral, X. Domènech, Electrochim. Acta 41 (1996) 1981. [17] A.D. Modestov, O. Lev, J. Photochem. Photobiol. A: Chem. 112 (1998) 261. [18] L. Sánchez, J. Peral, X. Domènech, Appl. Catal. B: Environ. 19 (1998) 59. [19] L. Rivas, I.R. Bellodono, F. Ascari, Chemosphere 37 (1998) 1033. [20] S. Hay, T.N. Obee, J. Adv. Oxid. Technol. 4 (1999) 209. [21] S.T. Muller, Z. Sun, M.P.K. Gireesh, K. Itoh, M. Murabayashi, J. Adv. Oxid. Technol. 4 (1999) 179. [22] D.C.C. Yao, W.R. Haag, Water Res. 25 (1991) 761. [23] W. Chu, M.H. Ching, Water Res. 37 (2003) 39. [24] Y. Zuo, J. Hoignè, Environ. Sci. Technol. 26 (1992) 1014. [25] B.C. Faust, J. Hoignè, Atmos. Environ. 24 (1990) 79. [26] J.J. Pignatello, Environ. Sci. Technol. 26 (1992) 944. [27] Y. Sun, J.J. Pignatello, Environ. Sci. Technol. 27 (1993) 304. [28] D.G. Hoover, G.E. Borgonovi, S.H. Jones, M. Alexander, Appl. Environ. Microb. 51 (1986) 226. [29] J.J. Aaron, M.A. Oturan, Turk. J. Chem. 25 (2001) 509. [30] M.A. Oturan, J.J. Aaron, N. Oturan, J. Pinson, Pesticide Sci. 55 (1999) 558. [31] M.A. Oturan, J. Appl. Electrochem. 30 (2000) 475. [32] E. Brillas, J.C. Calpe, J. Casado, Water Res. 34 (2000) 2253.
Degradation of the herbicide 2,4-dichlorophenoxyacetic acid by ozonation catalyzed with Fe2+ and UVA light Enric Brillas∗ , Juan Carlos Calpe, Pere-Llu´ıs Cabot Laboratori de Ciència i Tecnologia Electroqu´ımica de Materials, Departament de Qu´ımica F´ısica, Facultat de Qu´ımica, Universitat de Barcelona, Mart´ı i Franquès 1-11, 08028 Barcelona, Spain Received 8 March 2003; received in revised form 2 June 2003; accepted 16 June 2003
Abstract Solutions of the herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) with concentrations up to near saturation at pH 3.0 and at 25 ◦ C have been treated with ozone and ozonation catalyzed with Fe2+ and/or UVA light. Direct ozonation yields a slow depollution, while all contaminants are completely removed under UVA irradiation. The highest oxidizing power is achieved when Fe2+ and UVA light are combined, since greater amounts of oxidizing hydroxyl radical are generated and Fe3+ complexes are photodecomposed. The initial mineralization rate is enhanced when herbicide concentration increases and more hydroxyl radicals are produced by the catalyzed ozonation processes. The herbicide decay always follows a pseudo first-order reaction. Reverse-phase chromatography allows the detection and quantification of aromatic intermediates such as 2,4-dichlorophenol, 4,6-dichlororesorcinol and chlorohydroquinone. In all treatments, fast dechlorination reactions take place leading to chloride ion accumulation in the medium. The evolution of generated carboxylic acids such as glycolic, glyoxylic, maleic, fumaric and oxalic has been followed by ion-exclusion chromatography. Only oxalic acid remains stable in the O3 system, being quickly mineralized to CO2 by hydroxyl radicals formed in the O3 /UVA one. A high stability of oxalic acid in the O3 /Fe2+ system has also been found, since it yields Fe3+ -oxalato complexes. These species are photodecarboxylated under UVA irradiation in the O3 /Fe2+ /UVA system. A possible reaction pathway for 2,4-D mineralization involving all intermediates detected is proposed. © 2003 Elsevier B.V. All rights reserved. Keywords: 2,4-Dichlorophenoxyacetic acid; Ozone; Fe2+ catalyst; UVA light; Water treatment; TOC removal; Oxidation products
1. Introduction Ozone is a chemical agent widely used for the mineralization (i.e. conversion into CO2 and inorganic ions) of pesticides and related biorecalcitrant organic contaminants in waters [1–3]. However, the high energy cost for its generation limits many practical applications of direct ozonation (O3 system) to ∗ Corresponding author. Tel.: +34-93-4021223; fax: +34-93-4021231. E-mail address: [email protected] (E. Brillas).
water treatment. Alternative approaches have then been taken to try to improve its oxidizing power, also reducing its economic cost. In this way, more efficient pollutant removals have been reported by combining ozonation with UV light [1,4–9], homogeneous catalysts such as H2 O2 [1,10,11] and Fe2+ [7–9,12,13], and photocatalysis with TiO2 [7,14–21]. It is well known that when the O3 system is used to degrade acid waters, the organic matter is directly oxidized by molecular ozone, which can attack selectively pollutants, especially aromatic compounds,
0926-3373/$ – see front matter © 2003 Elsevier B.V. All rights reserved. doi:10.1016/S0926-3373(03)00266-2
382
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
following second-order reactions with small rate constants [1–3,10,22,23]. In contrast, the different catalyzed ozonation systems have greater degradative ability because of the production of a stronger, non-selective oxidizing agent such as hydroxyl radical (OH• ). This species can react more rapidly with organics yielding dehydrogenated or hydroxylated derivatives. Under irradiation of the solution with UV light at λ > 300 nm (O3 /UVA system), the main source of OH• is the direct reaction of ozone with H2 O2 , previously formed from its initial photolysis [1]: O3 + H2 O + hν → H2 O2 + O2
(1)
O3 + H2 O2 → OH• + HO2 • + O2
(2)
In this process, weaker oxidizing hydroperoxyl radical (HO2 • ) is also generated. When Fe2+ is present in the acid solution (O3 /Fe2+ system), it catalyzes ozone decomposition to give FeO2+ as intermediate by reaction (3) [12]. This species can further evolve to OH• by reaction (4) or can oxidize Fe2+ to Fe3+ at slower rate by reaction (5), then limiting the generation of hydroxyl radical for high Fe2+ concentration [7]. Fe2+ + O3 → FeO2+ + O2
(3)
FeO2+ + H2 O → Fe3+ + OH• + OH−
(4)
FeO2+ + Fe2+ + 2H+ → 2Fe3+ + H2 O
(5)
A faster oxidation of some organic pollutants by Fe2+ catalyzed ozonation under UVA irradiation (O3 /Fe2+ /UVA system) has been recently reported [7–9]. The acceleration of the mineralization process can be accounted for several parallel ways involving: (i) photolysis of complexes of Fe3+ with generated carboxylic acids, e.g. oxalic acid [24]; (ii) photoreduction of Fe3+ to give Fe2+ and an additional OH• via reaction (6) [25]; and (iii) additional oxidation of Fe2+ with H2 O2 generated from reaction (1) to yield more OH• by the classical Fenton reaction (7) [26,27]. Fe3+ + H2 O + hν → Fe2+ + OH• + H+
(6)
Fe2+ + H2 O2 → Fe3+ + OH• + OH−
(7)
Fe2+ regeneration from reaction (6) propagates the catalytic reaction (3) and larger amounts of oxidizing
OH• can be formed by reaction (4). In addition, the resulting Fe3+ from Fenton reaction (7) can be photoreduced again by reaction (6). 2,4-Dichlorophenoxyacetic acid (2,4-D) is a herbicide with potential toxicity for humans and animals that is used wordwide on a large scale for the selective control of weeds in gardens and farming. It is poorly biodegradable and has been detected as major pollutant in ground and surface waters [28,29]. For this reason, there is great interest in the study of oxidation methods with ability enough to mineralize 2,4-D in aqueous medium to try to avoid its dangerous accumulation in the aquatic environment. Recently, the degradation of this herbicide has been explored by different chemical, photochemical and photocatalytic systems such as O3 [2–4,7,22,23], O3 /Fe2+ [7], O3 /H2 O2 [11], H2 O2 /Fe2+ (Fenton’s reagent) [26,27], H2 O2 /UV [11], H2 O2 /Fe3+ /UV [26–29], TiO2 /UV [14–17,21], O3 /UVA [4,7], O3 /TiO2 /UVA [7] and O3 /Fe2+ /UVA [7], as well as by advanced electrochemical oxidation treatments [29–32]. Although an efficient depollution of 2,4-D solutions has been reported for different catalyzed ozonation processes, less is known about the degradative paths involved in them. This knowledge is necessary to understand the behavior of such treatments, reason for which more exhaustive research is required considering valuable information on the evolution of this herbicide and by-products formed. In this paper, we present a detailed study on the degradative behavior of 2,4-D in acidic medium by ozonation processes catalyzed with Fe2+ and/or UVA light. Comparative treatments for herbicide concentrations up to near saturation (ca. 620 ppm at 25 ◦ C) have been carried out at pH 3.0 using the O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. For the two last methods, 1 mM Fe2+ has been chosen as catalyst concentration, since higher content of this ion in solution accelerates the competitive reaction (5) limiting the destruction of organics [7]. The same solutions have also been treated by the O3 system to confirm the higher efficiency of the catalyzed ozonation methods. In each process, the decay kinetics of the herbicide and the evolution of its aromatic products and generated carboxylic acids have been followed by chromatographic techniques. A general reaction sequence for the mineralization
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
of the herbicide involving all the detected species is proposed. 2. Experimental 2.1. Chemicals 2,4-D, 2,4-dichlorophenol, 4,6-dichlororesorcinol, chlorohydroquinone, glycolic acid, glyoxylic acid, maleic acid, fumaric acid and oxalic acid were reagent grade from Merck, Fluka and Panreac. All the acidic solutions tested were prepared with high-purity water obtained from a Millipore Milli-Q system, with conductivity <6 × 10−8 S cm−1 at 25 ◦ C. The initial pH of each solution was adjusted with analytical grade sulfuric acid supplied by Merck. Analytical grade heptahydrated ferrous sulfate used as catalyst was purchased from Fluka. Organic solvents and the other chemicals employed were either HPLC or analytical grade from Merck, Fluka and Aldrich. 2.2. Instruments Ozone was generated with an Erwin Sander 300.5 ozonizer fed with pure O2 at 1.1 bar and at a flow rate of 60 l h−1 . The solution pH was measured with a Crison 2000 pH-meter. Total organic carbon (TOC) was determined on a Shimadzu 5050 TOC analyzer. Reverse-phase chromatography was performed with a Waters 600 HPLC liquid chromatograph, fitted with a Spherisorb ODS2 5 m column (150 mm × 4.6 mm (i.d.)) at room temperature, along with a Waters 996 photodiode array detector selected at 286 nm, controlled through a Millennium-32® program. The same HPLC chromatograph fitted with a Bio-Rad Aminex HPX 87H column (300 mm × 7.8 mm (i.d.)) at 35 ◦ C, in conjunction with the photodiode array detector selected at 210 nm, were used for ion-exclusion chromatography. Gas chromatography-mass spectrometry (GC-MS) was carried out with a Hewlett-Packard system consisting of a HP 5890 gas chromatograph fitted with a HP-1 0.25 m column (120 m ×0.25 mm (i.d.)) and a HP 5989A mass spectrometer operating in El mode at 300 ◦ C. Chloride ion concentration was determined by potentiometric titration with AgNO3 using a Metrohm Titrino 702 SM automatic titrator, with a detection limit of 5 ppm for Cl− .
383
2.3. Ozonation system All degradation experiments were conducted in an open, thermostated cylindrical Pyrex cell containing 100 ml of herbicide solution under stirring with a magnetic bar. The temperature was maintained at 25.0 ± 0.1 ◦ C. The O2 + O3 mixture generated by the ozonizer was supplied to the solution through a stainless steel diffuser. Its flow rate was regulated to 20 l h−1 yielding a constant production of 1.3 g O3 h−1 , as determined by standard iodometry. A 6 W Philips fluorescent black light blue tube was used to irradiate the solution with UVA light. This tube emitted in the wavelength range 300–420 nm, with λmax = 360 nm, and was placed at the top of the open cell at 4 cm over the sample. The flux of incident photons by unit reactor volume was 8.3×10−7 einstein l−1 min−1 , as detected with an uranyl actinometer active to photons in the range 250–500 nm. Solutions with a 2,4-D concentration between 608 ppm (close to saturation) and 5 ppm were acidified with 0.25 M H2 SO4 to pH 3.0 and further mineralized by the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. For the two last methods, 1 mM FeSO4 was added to each sample as catalyst before treatment. 2.4. Analysis procedures Before analysis of samples extracted from degraded solutions, they were filtered with 0.45 m PTFE filters supplied by Whatman. The mineralization of herbicide solutions was monitored for their TOC decay, using the standard non-purgeable organic carbon (NPOC) method. The evolution of 2,4-D and its aromatics products was followed by reverse-phase chromatography, with a 50:45:5 (v/v/v) methanol/phosphate buffer (pH 2.5)/pentanol mixture as eluent at 1.0 ml min−1 . For the quantification of generated carboxylic acids by ion-exclusion chromatography, 4 mM H2 SO4 was employed as eluent at 0.6 ml min−1 . In both techniques, 20 l aliquots were injected into the HPLC chromatograph. To identify the stable intermediates formed during the mineralization of 2,4-D, the organic components of several solutions degraded by the different systems for short times were extracted with 100 ml of dichloromethane and after evaporation of the solvent, the remaining solids were dissolved in 5 ml of methanol to be further analyzed by GC-MS.
384
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
3. Results and discussion 3.1. Comparative TOC removal All initial 2,4-D solutions were adjusted to pH 3.0 to be further treated by the different ozonation methods at 25 ◦ C during 2 h as maximum. A gradual depollution with raising degradation time was always found. Reproducible values for TOC removal were obtained under all experimental conditions tested. In these trials, a slow acidification was observed, reaching the final solutions a pH value between 2.5 and 2.9. As an example, Fig. 1 shows the comparative TOC abatement for 100 ml of a 230 ppm herbicide solution (with 100 ppm of initial TOC) using the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems. From these data, one can determine an initial mineralization rate of 104 ppm TOC h−1 (Fig. 1a), 116 ppm TOC h−1 (Fig. 1c), 456 ppm TOC h−1 (Fig. 1b) and 612 ppm TOC h−1 (Fig. 1d), respectively. These results indicate that the oxidizing ability of the O3 system at the beginning of the process is accelerated in the presence of Fe2+ , as expected if the herbicide and its oxidation products are more quickly destroyed by OH• produced from reaction (4) than by ozone. The fact that a much higher initial TOC decay is found for the O3 /Fe2+ system than for the O3 /UVA one, evidences more efficient generation of OH• from reaction (4) than from reaction (2). The highest initial
Fig. 1. TOC removal with time for the mineralization of 100 ml of a 230 ppm 2,4-D solution of pH 3.0 under the following conditions: (a) O3 (䊊); (b) O3 + 1 mM Fe2+ (䊐); (c) O3 + UVA (䉭); (d) O3 + 1 mM Fe2+ + UVA (䉫). Ozone flow rate 1.3 g h−1 . Temperature 25 ◦ C.
mineralization rate for the O3 /Fe2/ UVA treatment can then be related to the production of more oxidizing OH• by the simultaneous action of reactions (2), (4), (6) and (7). The relative oxidizing ability of the above processes varies when degradation time is prolonged. As can be seen in Fig. 1a, direct ozonation gives rise to a progressive, but slow, depollution up to attain 59% of TOC removal at 2 h. Under UVA irradiation, all organics are much more rapidly mineralized by reaction with OH• formed from reaction (2) and the solution TOC is reduced by 95% after 2 h of the O3 /UVA treatment (see Fig. 1c). In contrast, a very different behavior is obtained in the presence of catalytic Fe2+ . While at the first oxidation stages of the O3 /Fe2+ system a fast degradation rate of contaminants takes place, at times longer than 30 min the mineralization process becomes very hard (see Fig. 1b) because of the formation of very stable products, probably complexes of Fe3+ with generated carboxylic acids [24]. At the end of this treatment, only a 66% of TOC is removed, a value slightly lower than 60% obtained at 30 min. When the solution is also illuminated with UVA light using the O3 /Fe2+ /UVA system, the process is much more efficient, giving a 92% of TOC removal at 1 h (see Fig. 1d). This behavior can be accounted for by the fast photodecomposition of Fe3+ complexes formed [24] and by the enhancement of the generation rate of oxidizing OH• by reactions (2), (4), (6) and (7). Similar relative degradative trends to those described above were found for solutions containing a 2,4-D concentration between 608 and 58 ppm. Complete mineralization (more than 95% of depollution) was only achieved under UVA illumination, usually after 2 h of the O3 /UVA treatment and less than 1.5 h for the O3 /Fe2+ /UVA system. The quick and total degradation obtained in the latter conditions is shown in Fig. 2. Table 1 summarizes the percentage of TOC removal determined for the solutions tested after 1 h of each treatment, along with the corresponding initial mineralization rate. A similar depollution can be observed for all herbicide solutions from each method used, since at 1 h their TOC is reduced by 36–44, 63–78, 51–64 and 85–95% for the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems, respectively. Each method is then able to mineralize an extent of organics practically independent of the initial 2,4-D concentration, indicating that at long degradation time
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
385
the action of Fe2+ . For example, for 608 ppm of 2,4-D this parameter increases rapidly from 315 ppm TOC h−1 for the O3 /UVA system to 1050 ppm TOC h−1 for the O3 /Fe2+ one, but it only raises to 1211 ppm TOC h−1 for the O3 /Fe2+ /UVA method. This confirms that reaction (4) is the main source of oxidizing OH• in the Fe2+ catalyzed ozonation processes. 3.2. 2,4-D decay and reaction kinetics
Fig. 2. TOC time-course for the mineralization of 100 ml of different 2,4-D solutions of pH 3.0 by the O3 /Fe2+ /UVA system. Herbicide concentration: 608 ppm (䊊); 519 ppm (䊐); 486 ppm (䉭); 460 ppm (䉫); 345 ppm (䊉); 230 ppm (䊏); 115 ppm (䉱); 58 ppm (䉬). Ozone flow rate 1.3 g h−1 . Temperature 25 ◦ C.
its oxidizing power remains quite constant under the experimental conditions considered, being limited by the reactive ability of its oxidants (O3 and/or OH• ). However, an inspection of Table 1 shows a fast fall of initial mineralization rate with decreasing herbicide concentration. This tendency is not surprising and can be simply related to the existence of smaller amounts of degradable organics in the medium. Comparison of the data collected in Table 1 for the different catalyzed ozonations allows the conclusion that the initial mineralization rate for all solutions is mainly enhanced by
The kinetics for the disappearance of 2,4-D by reaction with the oxidizing species involved in each treatment was studied using a 230 ppm herbicide solution. The decay of this compound was followed by reverse-phase chromatography, where it displayed a well-defined peak with a retention time (tr ) of 3.65 min. The evolution of its concentration is depicted in Fig. 3. As can be seen, 2,4-D is rapidly destroyed for all procedures, disappearing from the medium in a time ranging between 8 and 14 min. The fast herbicide removal by the O3 system is due to its direct attack with molecular ozone, having a second-order rate constant (k2 ) of 298 M−1 s−1 [3]. Its quicker destruction from the different catalyzed ozonations can then be ascribed to its parallel reaction with the stronger oxidizing OH• produced in all them, because this compound is not photolyzed by UVA light. Herbicide concentration decays reported in Fig. 3 were analyzed from kinetic equations related to different reaction orders. Good linear plots, with
Table 1 Percentage of TOC removal obtained after 1 h of treatment of 100 ml of different 2,4-D solutions at pH 3.0 and at 25 ◦ C by O3 and catalyzed ozonation processes [2,4-D]0 (ppm)a
Method
608 519 486 460 345 230 115 58
41 36 38 37 36 44 42 42
O3 (192) (169) (179) (159) (116) (104) (58) (31)
O3 /UVA
O3 /Fe2+
O3 /Fe2+ /UVA
64 63 68 77 75 76 78 76
52 54 61 63 63 62 64 51
93 92 93 93 95 92 86 85
(315) (219) (206) (205) (176) (116) (63) (27)
(1050) (877) (895) (893) (753) (456) (156) (84)
(1211) (1167) (1075) (1062) (828) (612) (188) (74)
The corresponding initial mineralization rate (ppm TOC h−1 ) is given in parenthesis. a Initial 2,4-D concentration.
Fig. 3. 2,4-D concentration decay for a 230 ppm herbicide solution under the same experimental conditions as in Fig. 1. Applied system: (a) O3 (䊊); (b) O3 /Fe2+ (䊐); (c) O3 /UVA (䉭); (d) O3 /Fe2+ /UVA (䉫). The panel inset presents the corresponding kinetic analysis assuming a pseudo first-order reaction for 2,4-D.
386
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
regression coefficients >0.997, were only obtained when they were fitted to a pseudo first-order reaction. This kinetic analysis is presented in the inset of Fig. 3, giving a pseudo first-order rate constant (k1 ) of 6.3 × 10−3 , 7.3 × 10−3 , 9.4 × 10−3 and 1.17 × 10−2 s−1 for the O3 , O3 /UVA, O3 /Fe2+ and O3 /Fe2+ /UVA systems, respectively. This behavior suggests the presence of a steady concentration of oxidizing species (O3 and/or OH• ) in all systems during their initial degradation times, so that the corresponding pseudo first-order rate constant can be expressed, in general, as follows: k1 = k2 [O3 ] + k [OH• ] k
(8)
denotes the second-order rate constant for where the reaction of 2,4-D with hydroxyl radical. Thus, taking k1 = 6.3 × 10−3 s−1 , k2 = 298 M s−1 [3] and k = 0 for the O3 system, one can determine from Eq. (8) that injected ozone acts with a steady concentration of 2.1 × 10−5 M. The increase in k1 for the catalyzed ozonations depends on the relative proportion of OH• produced in them, since k > k2 . The higher k1 value for the O3 /Fe2+ system (9.4 × 10−3 s−1 ) than for the O3 /UVA one (7.3×10−3 s−1 ) can be explained again by the much more efficient OH• generation in the presence of Fe2+ from reaction (4) than under UVA illumination from reaction (2). When both catalysts Fe2+ and UVA light are combined, the additional OH• enhancement by reactions (6) and (7) can justify the highest k1 value of 1.17 × 10−2 s−1 found for the O3 /Fe2+ /UVA treatment. To gain a better insight into the competitive reactions of organics with oxidizing species taking place in the O3 /Fe2+ /UVA system, the influence of herbicide concentration upon its decay kinetics was studied. As can be seen in Fig. 4, 2,4-D is more slowly destroyed when its starting concentration increases. For example, the time needed for its complete removal is about 18 min for 608 ppm, but only close to 8 min for 230 ppm. This behavior is also reflected by the concomitant drop of its pseudo first-order rate constant. From the corresponding kinetic analysis depicted in the inset of Fig. 4, decreasing k1 values of 1.17×10−2 , 9.1 × 10−3 , 7.0 × 10−3 , 6.6 × 10−3 , 6.0 × 10−3 and 5.8 × 10−3 s−1 (with regression coefficients >0.997) are obtained for increasing 2,4-D concentrations of 230, 345, 460, 486, 519 and 608 ppm, respectively. This trend can be explained by a progressive consump-
Fig. 4. Destruction of 2,4-D concentration with time for same solutions of Fig. 2 degraded by the O3 /Fe2+ /UVA system. Initial herbicide concentration: 608 ppm (䊊); 519 ppm (䊐); 486 ppm (䉭); 460 ppm (䉫); 345 ppm (䊉); 230 ppm (䊏). The corresponding kinetics analysis considering a pseudo first-order reaction for the herbicide is shown in the inset.
tion of more injected ozone when reacts directly with higher amounts of herbicide, and also of some oxidation products produced, so that a smaller proportion of this oxidant is transformed into OH• from reactions (2) and (4) leading to a gradual tall in k1 , as predicted from Eq. (8). These findings allow us to consider that in the O3 /Fe2 /UVA system, the degradative path via OH• is favored at low pollutant concentrations, whereas reactions with ozone become more important as greater amounts of contaminants are contained in the solution. 3.3. Identification and evolution of oxidation products Several 230 ppm 2,4-D solutions were treated by the different systems for 10 min to detect the stable aromatic intermediates formed by GC-MS. For all procedures, the gas chromatogram of collected organics displayed three peaks which were identified by mass spectrometry and ascribed to the primary product 2,4-dichlorophenol [m/e = 162 (100, M+ ), 164 (69 (M + 2)+ ), 166 (10 (M + 4)+ )] and its derivatives 4,6-dichlororesorcinol (m/e = 178 (100, M+ ), 180 (66 (M + 2)+ ), 182 (10 (M + 4)+ )] and clorohydroquinone [m/e = 144 (100, M+ ), 146 (25 (M + 2)+ )]. Reverse-phase chromatograms of the same degraded solutions also exhibited peaks related to chlorohydroquinone at tr = 1.69 min, 4,6-dichlororesorcinol
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
387
Fig. 5. Evolution of the concentration of: (a) 4-chlorophenol; (b) 4,6-dichlororesorcinol; (c) chlorohydroquinone; (d) chloride ion, during the treatment of a 230 ppm 2,4-D solution of pH 3.0 at 25 ◦ C by the system: O3 (䊊); O3 /Fe2+ (䊐); O3 /UVA (䉭); O3 /Fe2+ /UVA (䉫).
at tr = 2.34 min and 2,4-dichlorophenol at tr = 5.75 min. These peaks were unequivocally identified by comparing their retention times and UV-Vis spectra with those of pure compounds, measured on the photodiode array detector. Similar oxidation products have been detected during the degradation of 2,4-D by photocatalysis combined with TiO2 [14–16] and advanced electrochemical methods [30–32]. The evolution of the above aromatics intermediates during the treatments of the 230 ppm 2,4-D solution was followed by determining their concentrations as a function of degradation time with calibration via standard compounds. Fig. 5a shows that the primary product 2,4-dichlorophenol is rapidly formed and destroyed in all cases, being completely removed at 25–30 min. Surprisingly, a greater maximum concentration of this product is accumulated between 4 and 6 min as the applied treatment pro-
duces more oxidizing OH• , varying from 2.1 ppm for direct ozonation to about 5 ppm for the O3 /Fe2+ /UVA system. This suggests that direct reaction of 2,4-D with O3 and/or OH• can also give other undetected products, although more OH• generation favors the formation of 2,4-dichlorophenol. Similar trends can be observed in Fig. 5b and c for its derivatives 4,6-dichlororesorcinol and chlorohydroquinone, respectively, although both species are accumulated in less extent and more rapidly removed from the medium, usually in 15–20 min. The little effect of UVA light on the evolution of all aromatic intermediates discards their direct photodecomposition. Mineralization of 2,4-D leads to the loss of chlorine atoms of its aromatic ring. The overall rate of dechlorination reactions involved in such process was followed by determining the concentration of Cl− accumulated in the above-degraded solutions. The
388
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
evolution of this ion is presented in Fig. 5d. As can be seen, 74 ppm of Cl− corresponding to 100% of initial chlorine are found in the medium after ca. 20 min of the O3 /Fe2+ /UVA treatment, i.e. when all the initial herbicide (see Fig. 3) and its detected chloroaromatic derivatives (see Fig. 5a–c) have been removed. For the other methods with smaller oxidizing power, a slower Cl− accumulation takes place, although more than 70 ppm of this ion is released after 30 mim (see Fig. 5d). These findings are indicative of the existence of fast dechlorination reactions at the beginning of the degradative process of 2,4-D with formation of stable chloride ion in all procedures tested. Ion-exclusion chromatograms of the same treated solutions exhibited peaks associated with generated carboxylic acids such as oxalic at tr = 6.62 min, maleic at tr = 7.96 min, glyoxylic at tr = 9.42 min, glycolic at tr = 12.3 min and fumaric at tr = 15.6 min. Good defined peaks for these products were obtained
from 20 min of degradation, since at lower times a large peak due to the strong adsorption at least of the initial compound inside the column prevented their quantification. Note that glycolic acid is expected to be formed during the breaking of the C(1)–O bond of 2,4-D. Glycolic acid is oxidized to oxalic acid via glyoxylic acid, while maleic and fumaric acids coming from the breaking of aromatic rings of intermediates are also converted into oxalic acid [32]. The time-course of glycolic acid concentration is depicted in Fig. 6a. A great accumulation of this compound close to 55 ppm can be observed at 20 min of direct ozonation, indicating its complete release from initial herbicide. At longer times it decays rapidly up to reach about 5 ppm from 80 min. Fig. 6a also shows that glycolic acid is more rapidly destroyed by the O3 /UVA and O3 /Fe2+ systems, disappearing at 60 min. Similar trends are found for glyoxylic (see Fig. 6b) and maleic (see Fig. 6c) acids. Both products, as well as
Fig. 6. Time-course of the concentration of generated carboxylic acids: (a) glycolic; (b) glyoxylic; (c) maleic; (d) oxalic, detected during the degradation of a 230 ppm 2,4-D solution of pH 3.0 at 25 ◦ C by the system: O3 (䊊); O3 /Fe2+ (䊐); O3 /UVA (䉭); O3 /Fe2+ /UVA (䉫).
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
fumaric acid, undergo a slow decay by direct ozonation, remaining in the medium even at 2 h, but disappear after 70–90 min of the O3 /UVA treatment. It can then be inferred that all these carboxylic acids are slowly destroyed by direct attack with molecular ozone, progressively increasing their degradation rate as more OH• is produced in the medium by the catalyzed ozonation processes. For this reason, they are always undetected when using the O3 /Fe2+ /UVA system, where the highest OH• concentration is generated from reactions (2), (4), (6) and (7).
389
However, Fig. 6d shows a very different behavior for oxalic acid, since it is removed by the O3 /UVA and O3 /Fe2+ /UVA systems, being accumulated in the O3 and O3 /Fe2+ treatments up to a steady concentration of 62 and 106 ppm, respectively. These results suggest an enhancement of the degradative path to oxalic acid when more OH• is produced. Note that Zuo and Hoignè [24] reported that this acid forms stable complexes with Fe3+ , which can be efficiently photodecarboxylated by UV light. When the O3 /Fe2+ and O3 /Fe2+ /UVA methods are applied to the 230 ppm
Fig. 7. General reaction pathway proposed for 2,4-D mineralization by O3 and ozonation catalyzed with Fe2+ and/or UVA light at pH 3.0.
390
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
2,4-D solution, Fe3+ is generated from reactions (4) and (5) and hence, a large proportion of Fe3+ -oxalate complexes is expected in the corresponding degraded solutions. The formation of such complexes limits the oxidizing ability the O3 /Fe2+ system. In fact, the steady concentration of 106 ppm of oxalic acid for this method corresponds to 29 ppm of TOC, a value very close to 34 ppm of TOC attained by the 230 ppm herbicide solution at 2 h (see Fig. 1b). The final solution of the O3 /Fe2+ treatment is then expected to be mainly composed of stable Fe3+ -oxalato complexes. The efficient destruction of such species under the action of UVA light can account for the fast and complete mineralization of 2,4-D by the O3 /Fe2+ /UVA system, thus explaining the highest oxidizing ability of this procedure. In contrast, the steady oxalic acid concentration of 62 ppm obtained for direct ozonation only corresponds to 17 ppm of TOC, i.e. a 41% of the overall TOC obtained after 2 h of degradation by this procedure (see Fig. 1a). Consequently, a larger proportion of other hardly oxidable products are present in such final solution. These products, as well as oxalic acid, are decomposed by the O3 /UVA system, yielding the overall mineralization of the herbicide. 3.4. Reaction pathways for 2,4-D degradation Fig. 7 shows a general reaction pathway for 2,4-D mineralization at pH 3.0 taking into account all intermediates detected. The main oxidizing agents are O3 for direct ozonation and O3 and OH• for catalyzed ozonations. The proposed sequence involves the degradative path to oxalic acid as the ultimate generated carboxylic acid, a way which is more important as higher amounts of OH• are produced by the catalyzed systems in the order O3 /UVA < O3 /Fe2+ < O3 /Fe2+ /UVA. The process is initiated by the attack of O3 and/or OH• on the C(1)–O bond of 2,4-D, causing the breaking of its lateral chain to give 2,4-dichlorophenol and glycolic acid. This reaction occurs in larger extent in the O3 /Fe2+ /UVA system when more amounts of OH• can react with the initial reactant. 2,4-Dichlorophenol can then be hydroxylated either on C(5)-position yielding 4,6-dichlororesorcinol or on C(4)-position leading to chlorohydroquinone with loss of a chlorine atom. This species is also a strong oxidizing agent, being reduced to chloride ion. Further degra-
dation with dechlorination of chlorobenzoquinone and 4,6-dichlororesorcinol by reaction with O3 and/or OH• gives maleic and fumaric acids, which are oxidized to oxalic acid. This acid can also be formed by a parallel path involving hydroxylation of glyoxylic acid, previously formed from dehydrogenation of the initially generated glycolic acid. The transformation of all these carboxylic acids into oxalic acid is accelerated as more OH• is generated in the medium by catalyzed ozonations. Oxalic acid remains stable in the O3 system, being mineralized to CO2 by the action of OH• produced from reaction (2) in the O3 /UVA system. In the presence of Fe2+ , stable Fe3+ -oxalato complexes are formed, which can be rapidly photodecomposed to CO2 with loss of Fe2+ [24] under UVA irradiation in the O3 /Fe2+ /UVA system.
4. Conclusions It has been demonstrated that the initial mineralization rate of acidic 2,4-D solutions in the O3 system is accelerated in the presence of Fe2+ as catalyst due to the production of oxidizing OH• by reaction (4), although their degradation is inhibited by the formation of complexes of Fe3+ with intermediates. Complete mineralization can be achieved by the O3 /UVA system, and more rapidly using the O3 /Fe2+ /UVA one, since it produces more OH• from reactions (2), (4), (6) and (7) and can photodecompose Fe3+ complexes by the action of UVA light. The initial mineralization rate is always enhanced by increasing herbicide concentration. The pseudo first-order rate constant for 2,4-D decay increases as higher OH• concentration is generated by the catalyzed ozonation process, whereas in the O3 /Fe2+ /UVA system it drops with raising herbicide concentration due to a smaller OH• production. 2,4-Dichlorophenol, 4,6-dichlororesorcinol and chlorohydroquinone have been quantified as stable aromatic intermediates by reverse-phase chromatography. The initial chlorine is released to the medium as chloride ion. Ion-exclusion chromatography reveal the generation of carboxylic acids such as glycolic, glyoxylic, maleic, fumaric and oxalic, which are slowly destroyed by direct ozonation and efficiently removed by catalyzed ozonations, except oxalic acid. This acid is stable in the O3 system, being quickly mineralized to CO2 by the O3 /UVA one. For the
E. Brillas et al. / Applied Catalysis B: Environmental 46 (2003) 381–391
O3 /Fe2+ system, it yields stable Fe3+ -oxalato complexes, which are photodecarboxylated to CO2 under UVA irradiation in the O3 /Fe2+ /UVA one. Acknowledgements The authors thank financial support of MCYT (Ministerio de Ciencia y Tecnolog´ıa, Spain) under project BQU2001-3712. References [1] G.R. Peyton, in: N.M. Ram, R.F. Christan, K.P. Canto (Eds.), Significance and Treatment of Volatile Organic Compounds in Water Supplies, Lewis Pu, 1990, pp. 313–362. [2] S. Chiron, A. Fernandez-Alba, A. Rodriguez, E. Garcia-Calvo, Water Res. 34 (1999) 366. [3] H. Jian-Ying, M. Takeshi, M. Yasumoto, A. Takako, Water Res. 34 (2000) 2215. [4] J. Prado, J. Arantegui, E. Chamorro, S. Esplugas, Ozone: Sci. Technol. 16 (1994) 235. [5] R. Andreozzi, R. Marotta, J. Hazard Mater. 69 (1999) 303. [6] S.W. Kuo, Chemosphere 39 (1999) 1853. [7] E. Piera, J.C. Calpe, E. Brillas, X. Domènech, J. Peral, Appl. Catal. B: Environ. 27 (2000) 169. [8] R. Sauleda, E. Brillas, Appl. Catal. B: Environ. 29 (2001) 135. [9] L. Sánchez, X. Domènech, J. Casado, J. Peral, Chemosphere 50 (2003) 1085. [10] C. Adams, P.A. Scanlan, N.D. Secrist, Environ. Sci. Technol. 28 (1994) 1812.
391
[11] G.F. Ijpelaar, M. Groenendijk, R. Hopman, J.C. Kruithof, Water Sci. Technol., Water Supply 2 (2002) 129. [12] T. Løgager, J. Holeman, K. Sehested, T. Pedersen, Inorg. Chem. 31 (1992) 3523. [13] D.L. Sedlack, J. Hoigné, M.M. David, R.N. Colvile, E. Seyffer, K. Acker, W. Wiepercht, L. Wolfgang, J.A. Lind, S. Fuzzi, Atmos. Environ. 31 (1997) 2515. [14] M. Trillas, J. Peral, X. Doménech, Appl. Catal. B: Environ. 5 (1995) 377. [15] M. Trillas, J. Peral, X. Doménech, J. Chem. Technol. Biotechnol. 67 (1996) 237. [16] L. Sánchez, J. Peral, X. Domènech, Electrochim. Acta 41 (1996) 1981. [17] A.D. Modestov, O. Lev, J. Photochem. Photobiol. A: Chem. 112 (1998) 261. [18] L. Sánchez, J. Peral, X. Domènech, Appl. Catal. B: Environ. 19 (1998) 59. [19] L. Rivas, I.R. Bellodono, F. Ascari, Chemosphere 37 (1998) 1033. [20] S. Hay, T.N. Obee, J. Adv. Oxid. Technol. 4 (1999) 209. [21] S.T. Muller, Z. Sun, M.P.K. Gireesh, K. Itoh, M. Murabayashi, J. Adv. Oxid. Technol. 4 (1999) 179. [22] D.C.C. Yao, W.R. Haag, Water Res. 25 (1991) 761. [23] W. Chu, M.H. Ching, Water Res. 37 (2003) 39. [24] Y. Zuo, J. Hoignè, Environ. Sci. Technol. 26 (1992) 1014. [25] B.C. Faust, J. Hoignè, Atmos. Environ. 24 (1990) 79. [26] J.J. Pignatello, Environ. Sci. Technol. 26 (1992) 944. [27] Y. Sun, J.J. Pignatello, Environ. Sci. Technol. 27 (1993) 304. [28] D.G. Hoover, G.E. Borgonovi, S.H. Jones, M. Alexander, Appl. Environ. Microb. 51 (1986) 226. [29] J.J. Aaron, M.A. Oturan, Turk. J. Chem. 25 (2001) 509. [30] M.A. Oturan, J.J. Aaron, N. Oturan, J. Pinson, Pesticide Sci. 55 (1999) 558. [31] M.A. Oturan, J. Appl. Electrochem. 30 (2000) 475. [32] E. Brillas, J.C. Calpe, J. Casado, Water Res. 34 (2000) 2253.