Comparative study of electrochemical degradation ...

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Water Research 39 (2005) 2527–2534 www.elsevier.com/locate/watres

Comparative study of electrochemical degradation and ozonation of nonylphenol Jaeshin Kima, Gregory V. Korshina,, Alexander B. Velichenkob a Department of Civil and Environmental Engineering, University of Washington, Box 352700, Seattle, WA 98115-2700, USA Department of Physical Chemistry, Ukrainian State University of Chemical Technology, Gagarin Ave., 8 Dnepropetrovsk 49005, Ukraine

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Received 21 December 2004; received in revised form 6 April 2005; accepted 12 April 2005 Available online 24 June 2005

Abstract Treatment of solutions of nonylphenol (NP), Triton X-100 (TrX) and phenol in a flow-through undivided EC reactor equipped with a Co2+-promoted PbO2 anode and a stainless steel cathode was accompanied by consistent changes of absorbance, fluorescence and mass spectra of the effluents, and formation of aldehydes ranging from formaldehyde to decyl aldehyde. Deconvolution of the absorbance spectra of EC-treated NP, TrX and phenol and examination of their fluorescence indicated that the compounds are rapidly degraded in the reactor. For NP, the degradation of the target proceeded via the generation (at current densities o25 mA/cm2) of benzoquinone intermediates that yielded peaks with m=z ratios 223, 227, 235, and 241 D in the mass spectra. Their breakdown at current densities 410 mA/cm2 was accompanied by the release of aldehydes that were predominated by acetaldehyde and formaldehyde. The total yield of aldehydes increased with the current density, but their speciation showed little sensitivity to it. Deconvolution of the absorbance spectra of NP solutions subjected to ozonation, and analysis for reaction by-products formed in these conditions showed the reaction pathway in the latter case was likely to be similar to that observed for the EC treatment. r 2005 Elsevier Ltd. All rights reserved. Keywords: Absorbance spectroscopy; Alkylphenol; Degradation; Electrochemical; Mass spectroscopy; Nonylphenol; Ozonation

1. Introduction Alkylphenol ethoxylates (for instance, nonylphenol ethoxylates NPnEOs, where n is the number of ethoxy units, typically 8–12) are the main components of the nonionic group of surfactants. Varying levels of NPnEOs and similar alkylphenols have been found in indoor air and dust, sewage sludge and sediments, wastewater, landfill leachates, surface waters and treated drinking water; these compounds have been shown to cause endocrine disruption (e.g., Rudel et al., 2003; Corresponding author. Tel.: 206 543 2394; fax: 206 685 9185.

E-mail address: [email protected] (G.V. Korshin).

Wintgens et al., 2003; Petrovic et al., 2003; Montgomery-Brown and Reinhard, 2003; Berryman et al., 2004; Wenzel et al., 2004). Biodegradation and/or hydrolysis cause one or more ethoxylate (EO) groups in NPnEOs to cleave yielding intermediate products with a lesser number of EO units and, finally, nonylphenol (NP) (Sato et al., 2003; Franska et al., 2003). The endocrine disrupting activity of the NPnEOs degradation products appears to increase as the molecular weights of the breakdown products decrease. NP is relatively hydrophobic (log Kow 4.8–5.3) and tends to accumulate in sediments or sludge (Montgomery-Brown and Reinhard 2003); however, the water column continues to be affected due to the slow leaching from solid phases and also by effluents containing alkylphenol ethoxylates

0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2005.04.070

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J. Kim et al. / Water Research 39 (2005) 2527–2534

and their breakdown products. Because of that, it is reasonable to treat NPnEOs and/or NP-containing effluents prior to disposal. Advanced oxidation processes, AOP (e.g., ozonation, photolytic degradation) have been used to remove these compounds (Sherrard et al., 1994; Sherrard et al., 1996; Brand et al., 1998; Mizuno et al., 2002; Ike et al., 2002; Wintgens et al., 2003). Identified products of NPnEOs or NP degradation formed in AOP treatment include NPnEO with a smaller number of EO groups, carboxylic acids predominated by acetate, oxalate and formate, and aldehydes predominated by formaldehyde, acetaldehyde, glyoxal, and methylglyoxal (Sherrard et al., 1996; Mizuno et al., 2002). Electrochemical (EC) oxidation represents another option for treatment of NPnEO and allied species. Principal advantages of EC processes are the ease of operations, a wide range of treatment conditions and elimination of the need to generate, dispense and store treatment reagents. Even more important is the capability of EC treatment, at least in principle, to induce a very deep oxidation that can result in a virtually complete mineralization of the target (EC incineration, ECI) (Comninellis, 1994; Polcaro and Palmas, 1997; Houk et al., 1998; Johnson et al., 1999; Tahar and Savall, 1999; Feng and Li, 2003). The development of advanced anode materials has made the option of ECdriven removal of alkylphenol ethoxylates increasingly more viable. Specifically, boron-doped diamond (BDD) and transition metal dioxide-based (e.g., dimensionally stable anodes DSA, PbO2, SnO2) electrodes can produce upon anodic polarization large quantities of hydroxyl radicals and, in the case of PbO2, ozone. EC oxidation with these materials have been shown to rapidly degrade diverse compounds, including phenols and chlorophenols, pesticides, and endocrine disruptors such as bisphenol A (Comninellis, 1994; Boscoletto et al., 1994; Houk et al., 1998; Amadelli et al., 1999; Johnson et al., 1999; Boye et al., 2002; Da Silva et al., 2001; Velichenko et al., 2002; Montilla et al., 2002, Feng and Li, 2003; Borras et al., 2003, Zhi et al., 2003). The principal pathway EC degradation of phenols has been shown to include the formation of a benzoquinone intermediate followed by its breakdown accompanied by the release of oxidation products such as maleic and oxalic acids. In principle, alkylphenol ethoxylates and NP can be similarly degraded by EC oxidation, but EC reactions of these species have been studied very scarcely (Schumann et al., 1998; Tanaka et al., 1999; You et al., 2002; Kuramitz et al., 2002). EC oxidation of alkylphenol ethoxylates with conventional (e.g., carbon-based) electrode materials have been shown to result in a partial oxidation followed by their polymerization on the surface. There is little doubt that these species can be degraded further, but EC oxidation of NP and similar

compounds with transition metal dioxide-based anodes that are efficient in ECI, has not been adequately explored. The goal of this communication is to examine the mechanisms of breakdown of NP and related species by, and compare speciation of relevant products formed upon, EC treatment of these species using a PbO2 anode and, on the other hand, a more conventional treatment process exemplified in this study by ozonation.

2. Experimental High-purity water (18.2 MO-cm resistivity, residual dissolved organic carbon o0.1 mg/L) obtained with a Milli-Q Plus system was used to prepare all solutions. NP and Triton X-100 (TrX) were purchased from Aldrich. For EC experiments, these compounds were dissolved in the presence of a background electrolyte (0.01 M Na2SO4). Ozonation was carried out using similar solutions in the absence of background salts. NP solutions were filtered through a 0.45 mm filter to remove particulates. The concentration of NP was determined by means for measurements of concentrations of dissolved organic carbon (DOC) and UV absorbance at 275 nm. The molar extinction coefficient for NP was at 275 nm was determined to be equal 2837 M 1 cm 1. EC treatment was carried out using an undivided flow-through EC micro-cell (ElectroCell AB, Sweden) similar to that described in more detail by Johnson et al. (2000). Cobalt-promoted PbO2 (Co–PbO2) and stainless steel were used as the anode and cathode, respectively. The synthesis and properties of Co–PbO2 were described by Velichenko et al. (2002). The surface area of the electrodes was 10 cm2 and the inter-electrode distance was 0.5 cm. The range of current densities was from 0 to 40 mA/cm2. The flow rate was set at 25 mL/min, which corresponded at a 12 s hydraulic residence time in the cell. Ozone was produced using pure oxygen gas and a C2P-3 corona discharge ozone generator (PCI Ozone Corp). Ozone stock solution was prepared by sparging ozone into high-purity water held in a pre-cleaned glass container. Ozonation of NP was carried out by mixing a desired volume of ozone stock solution with solutions of NP. After one hour of ozonation time at room temperature, analyses for relevant water quality parameters (UV absorbance, fluorescence, concentrations of aldehydes) were carried out. Concentrations of DOC were measured using a Model 1010 OI Analytical carbon analyzer. Absorbance spectra were measured with a dual-beam Perkin-Elmer Lambda-18 spectrophotometer using 5 cm quartz cells. Fluorescence spectra were measured with a PerkinElmer LS-50B fluorometer. Analyses for aldehydes were carried out in accord with Method 6252 ‘‘Disinfection

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By-Products: Aldehydes’’ (Standard Methods, 1995). Aldehydes were derivatized to corresponding oximes by o-(2,3,4,5,6-pentafluorobenzyl)-hydroxylamine (PFBHA). The oximes were extracted with hexane and analyzed using a Perkin-Elmer AutoSystem gas chromatograph equipped with an RTX-1MS analytic column and an electron capture detector. The instrument was calibrated using five standards with varying concentrations of formaldehyde, acetaldehyde, propanaldehyde, glyoxal, and pyruvaldehyde. A Bruker Esquire ion trap mass spectrometer with electrospray ionization (ESI) was employed to examine molecular weights of products of EC oxidation of NP. Mass spectra were obtained in the negative mode with a probe tip potential of 3 kV, a skimmer voltage of 15 V, and a capillary exit voltage of 75 V. The temperature of ionization chamber was maintained at 200 1C. Sample of 15 mL was injected using a syringe pump at the rate of 10 mL/min. The absorbance of spectra of EC-treated and ozonated solutions of phenol, NP and TrX were processed using numerical deconvolution. Evolving factor analysis (EFA) was performed to determine the number of spectroscopically distinct species, and multivariate curve resolution (MCR) was used to determine the spectra of the species discerned by EFA and quantify the dependence of their contributions vs. current density or ozone dose. PLC Toolboxs and MatLabs software packages were employed for deconvolution procedures.

3. Results EC treatment of solutions of NP, TrX, and phenol was accompanied by consistent changes of their absorbance and fluorescence spectra, and formation of reaction by-products exemplified by aldehydes. The absorbance spectra of phenol, NP and TrX prior to EC treatment or ozonation were similar. For all three compounds, two distinct bands were present for wavelength 4200 nm (Fig. 1). One of them was located in the wavelength range 210–230 nm. That band had a maximum at 211, 220, and 223 nm for phenol, NP, and TrX, respectively. Another less intense but very distinct band characteristic for many phenolic compounds (Sadtler, 1979) was located in the range of wavelengths 240–290 nm. Its maximum was found at 270, 275, and 275 nm for phenol, NP, and TrX, respectively. Increase of the current passing through the EC cell caused the intensity of the bands to decrease monotonically but other features appeared in addition to them. For phenol, a separate band located in the range 230–260 nm was prominent at current densities 5–25 mA/cm2 (Fig. 1A). At higher currents, that band disappeared, along with the characteristic features of the absorbance spectrum of phenol per se. The remaining spectrum was featureless, with absorbance rapidly decreasing at l4200 nm. Similar featureless absorbance spectra were observed for NP and TrX at current densities 425 mA/cm2 (Fig. 1B and C). In the

0.05 Current density (mA/cm2)

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Fig. 1. Changes of UV absorbance spectra of phenol (A), nonyl phenol (B), and Triton X (C) at varying current densities. The concentrations of phenol, nonyl phenol, and Triton X 3.0, 2.8, and 3.0 mM, respectively. Background electrolyte 0.01 M Na2SO4.


Fluorescence intensity (a.u.)

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Fig. 2. Influence of current density of the intensity of fluorescence of 2.8 mM nonyl phenol. Excitation wavelength 250 nm.

Normalized absorbance

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Fig. 3. Comparison of deconvoluted spectra of the intermediate species (A) and final products formed as a result of EC treatment of phenol, nonyl phenol, and Triton X.

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intermediate range of current densities (10–25 mA/cm2), broad ‘‘humps’’ were observed in the wavelength range 230–260 nm for NP and TrX but the band noted for EC treated solutions of phenol could not be discerned. The fluorescence emission spectra of NP and TrX exhibited a less complex pattern of behavior. The emission spectrum of NP comprised a single band located in the range of wavelengths 290–360 nm; its maximum was at 308 nm (Fig. 2). (Because the fluorescence spectra of NP and TrX were almost identical, only the data for NP will be discussed.) EC treatment caused the emission intensity to decrease rapidly and virtually no fluorescence was observed for current densities 430 mA/cm2. Analysis of the emission spectra showed that, in contrast with the absorbance spectra of NP, no new features that could be assigned to fluorescing species other than NP were present in the emission spectra of EC treated solutions. Numerical deconvolution of the absorbance spectra of EC treated solution of phenol, NP, and TrX shown in Fig. 1 indicated that only three spectroscopically distinct species were present in all cases. These included the initial species, an intermediate species (denoted as I) predominant in the intermediate range of current densities, and an operationally defined final product (FP) associated with the largely featureless spectra observed at current densities 430 mA/cm2. The normalized absorbance spectra of the intermediate species and FP for phenol, NP, and TrX obtained by deconvolution are shown in Fig. 3A and B, respectively. The behavior of these species as a function of current density is shown in Fig. 4 for NP (the concentration of NP and estimated contributions of the intermediate and final product were normalized to the initial concentration of NP). It is to be recognized that EFA per se does not yield the absolute concentrations of reactants because their molar absorbances need to be determined by independent experiments. However, EFA provides a precise evaluation of the number of spectroscopically distinct components and their contributions to the overall absorbance. Correspondingly, EFA data provide

Relative concentration

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0.00

(mA/cm2)

Fig. 4. Influence of current density on the concentration of nonyl phenol (NP), intermediate species (I) and final product (FP) and total yield of aldehydes.

unambiguous evidence concerning the behavior of the significant system components as a function of current density or ozone concentration. The total yield of all aldehyde species normalized by the molarity of NP initially present in the solution is also shown in Fig. 4. It can be observed that the onset of the release of aldehydes took place at current densities 5–15 mA/cm2. This range of current densities coincides with the breakdown of the intermediate species. The total yield of all quantified aldehydes reached or exceeded a 100% level for current densities 430 mA/cm2. The concentrations of acetaldehyde and formaldehyde found in EC-treated solutions of NP were comparable, while the concentrations of glyoxal, propanaldehyde, and pyruvaldehyde were almost an order of magnitude

ARTICLE IN PRESS J. Kim et al. / Water Research 39 (2005) 2527–2534 0.50

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EC Ozone

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Acetaldehyde Formaldehyde Glyoxal Propanaldehyde Pyruvaldehyde

Current density (mA/cm2)

Fig. 5. Influence of current density on the molar yield of individual aldehydes. EC treatment of 2.8 mM nonyl phenol.

Fig. 7. Speciation of aldehydes formed in EC-treated and ozonated solutions of nonyl phenol.

released aldehydes, which was almost identical with that determined for EC-treated solutions of NP (Fig. 7). Relative concentration

1.2 1.0

4. Discussion

0.8 0.6 0.4

NP I FP Aldehydes

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4 6 Ozone dose (mol O3/mol NP)

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Fig. 6. Influence of ozone doze of the concentration of nonyl phenol (NP), intermediate species (I) and final product (FP) and total yield of aldehydes.

less than those of the two main aldehyde species (Fig. 5). In all cases, the concentrations of aldehydes increased rapidly at current densities 410 mA/cm2. Despite the sensitivity of the absolute concentrations of the aldehydes to the increase of current density, it did not significantly affect their speciation. Results obtained for ozonated solutions of NP and TrX were largely similar to those for EC treatment. The molar O3/NP ratio was varied in these experiments from 0 to ca. 10.4 mol O3 per mole NP. Deconvolution of the absorbance spectra of ozonated solutions of NP showed that, in addition to NP per se, only two spectroscopically distinct species were present. Their absorbance spectra were similar to those showed in Fig. 3 for the intermediate and final products formed in solutions of NP subjected to EC oxidation. The maximum molar yield of all aldehydes found in ozonated solutions of NP was close 100%, similarly to the results obtained for ECtreated solutions. In contrast with EC treatment of NP, for which the release of aldehydes took place for current densities exceeding a certain threshold value (10–15 mA/ cm2), the yield of aldehydes for ozonated NP increased proportionally to the ozone dose (Fig. 6). The concentration of ozone did not affect the speciation of the

Prior literature provides unambiguous evidence that EC treatment of aqueous solutions of phenol is accompanied by its oxidation to benzoquinone (BQ) followed by the breakdown of BQ to form various reaction products, notably maleic acid (Polcaro and Palmas, 1997; Houk et al., 1998; Johnson et al., 1999; Tahar and Savall, 1999; Boye et al., 2002; Feng and Li, 2003). Analysis of the spectroscopic data obtained for EC treated solutions of NP and TrX indicates that the mechanisms of EC oxidations of these compounds are essentially the same. Indeed, the spectrum of the intermediate species formed upon EC treatment of phenol and obtained via deconvolution (Fig. 3A) has a feature (a maximum at 243 nm) characteristic for BQ (Sadtler, 1979). For NP and TrX, the absorbance spectra of the respective intermediates are similar to that of BQ, although the maximum at 243 nm is less pronounced and/or shifted to longer wavelengths. It is possible that the features present in the spectra of the intermediates formed in EC-treated solutions of NP and TrX are less specific due to the superposition of absorbance bands of several products that are structurally similar to BQ (for instance, 4-nonyl-1,2- and 4nonyl-1,3-benzoquinone, and 4-nonyl-4-hydroxycyclohexa-2,5-dienone, as was observed by Schumann et al. (1998) and Nilson et al., 1973). On the other hand, the absorbance spectra of the operationally defined final product of EC oxidation (Fig. 3B) are not specific enough to ascribe them to any particular compound. They are likely to be associated with multiple species, notably carboxylic and keto-carboxylic compounds (maleic, ketoglutaric, fumaric acids) that are found in significant concentrations in EC-treated solutions of phenols (Johnson et al., 1999). Mass-spectra of EC-treated solutions of NP confirmed the trends established based on the data of



absorbance measurements. It was observed that the intensity of the peak corresponding to unaltered NP molecules (m=z 219 D) rapidly decreased as the current density increased. The relative intensity of a molecular fragment with a 235 D m=z ratio had a maximum at the current density 10 mA/cm2, while counts for fragments with m=z ratios 223, 227, and 241 D were highest at a 15 mA/cm2 current density. At current densities 415 mA/cm2, peaks that correspond to molecular fragments with lower molecular weights, notably with a 155 D m=z ratio become m more intense. The peak with a 235 D m=z ratio appears to correspond to a dihydroxy nonylphenol intermediate (e.g., 1,2-dihydroxy-4-nonylphenol) that was formed as a result of reactions between NP and OHd radicals generated at the surface of the PbO2 anode (Houk et al., 1998; Schumann et al., 1998; Boye et al., 2002; Feng and Li, 2003; Da Silva et al., 2001). However, 4-nonylbenzoquinone species with a 233 D m=z ratio, whose formation was anticipated to follow the formation of 4nonyl-dixydroxy phenol, was not detected. It is possible that it was unstable in conditions of ESI and tended to break down to form molecular fragments such as 2nonyl maleic acid (m=z ratio 241 D) and 2-nonyl maleic anhydride (m=z ratio 223). The peak with a 227 D m=z ratio was apparently associated with 2-octyl maleic acid, which was formed either as a result of loss of a CH2 unit from 2-nonyl maleic acid (or species preceding it) caused by an attack by OH radicals (Brand et al., 1998), or via the breakdown of octylphenol initially present as an impurity in NP. The conspicuous feature of the behavior of the species associated with m=z ratios 223, 227, 235, and 241 was that their abundance was correlated (R2 0.76) with the yield of the intermediate detected by deconvolutions of the absorbance spectra and prevalent in the range of current densities 5–30 mA/cm2, while the intensities of these peaks were not correlated with the formation of FP (Fig. 4). On the other hand, the intensity of the peak with a 155 m=z ratio associated with decyl aldehyde was strongly correlated with the yield of FP and exhibited a rapid increase for current densities 415 mA/cm2 (Fig. 8). The behavior of the peak with a 181 m=z ratio was very similar but the identity of the species that correspond to this peak remains uncertain. The above observations and the result that a virtually 100% yield of low molecular weight aldehydes was observed for EC-treated solutions of NP confirm that EC treatment leads to a rapid breakdown of the aromatic ring in the molecules of the target species. The data also indicate that microscopic mechanisms of NP and TrX degradation by EC oxidation and ozonation are largely similar. For instance, both the speciation of low molecular weight aldehydes formed in EC-treated and ozonated solutions of NP (Fig. 7) and

1.20 1.00

m/z 155 m/z 181

0.80 0.60 0.40 0.20 0.00

no current 10 mA/cm2 15 mA/cm2 20 mA/cm2 30 mA/cm2 40 mA/cm2

Fig. 8. Relative intensities of peaks with m=z ratios 155 and 181 D as a function of current density.

trends seem in the absorbance spectra of the relevant systems are virtually identical. Measurements of the DOC concentrations in EC treated and ozonated solutions of NP and TrX showed that, despite the rapid breakdown of the parent molecules, the concentration of organic carbon remained virtually unchanged. Correspondingly, EC incineration of NP and TrX was not observed in the experimental conditions utilized in this study. It is possible that further increase of current density and/or charge per volume of the treated solution in EC reactors with PbO2 anodes or other materials (notably, BDD) capable of producing high yields of OHd radical are likely to be accompanied by deeper mineralization of NP, TrX, and other alkylphenols, but this option requires that the anodic material be extremely resistant to corrosion. However, given the result of this study that the EC treatment of NP and TrX is accompanied by a complete fragmentation of the aromatic ring and a very high yield of the relevant oxidation products, it may be concluded that EC treatment can be carried out to treat effluents containing alkylphenols.

5. Conclusions Treatment of solutions of NP, TrX, and phenol in a flow-through undivided EC reactor equipped with a Co2+-promoted PbO2 anode and a stainless steel cathode was accompanied by consistent changes of absorbance, fluorescence, and ESI mass spectra of the effluents, and formation of aldehydes ranging from formaldehyde to decyl aldehyde. Deconvolution of the absorbance spectra of EC-treated NP, TrX, and phenol and examination of their fluorescence spectra indicated that the compounds are rapidly degraded in the reactor due to the attack by OHd radicals generated at the surface of PbO2. In the case of NP, the degradation of the target proceeded via the generation of benzoquinone


intermediates that yielded peaks with m=z ratios 223, 227, 235, and 241 D in the ESI mass spectra. These species were prevalent at current densities o25 mA/cm2. Their breakdown at current densities 410 mA/cm2 was accompanied by the release of aldehydes (and other products such as carboxylic acids). The total yield of aldehydes that were predominated by acetaldehyde and formaldehyde increased with the current density, but their speciation showed little sensitivity to it. Deconvolution of the absorbance spectra of NP solutions subjected to ozonation, and analysis for reaction byproducts formed in these conditions showed the reaction pathway in the latter case was largely similar to that observed for the EC treatment. The data indicate that EC treatment leads to a rapid breakdown of the aromatic ring in the molecules of the target species, but a complete EC incineration of NP and TrX can be achieved for conditions (e.g., treatment time, current density) exceeding those utilized in this study.

Acknowledgments This study was partially supported by Awwa Research Foundation (project # 2728). The views represented in this publication do not necessarily represent those of the funding agency. The authors also would like to thank graduate student Julien Jarrige (Universite´ de Poitiers, France) for his participation in the experiments.

References Amadelli, R., Armelao, L., Velichenko, A.B., Nikolenko, N.V., Girenko, D.V., Kovalyov, S.V., Danilov, F.I., 1999. Oxygen and ozone evolution at fluoride modified lead dioxide electrodes. Electrochimica Acta 45 (4/5), 713–720. Berryman, D., Houde, F., DeBlois, C., O’Shea, M., 2004. Nonylphenolic compounds in drinking and surface waters downstream from treated textile and pulp and paper effluents: a survey and preliminary assessment of their potential effects on public health and aquatic life. Chemosphere 56 (3), 247–255. Borras, C., Laredo, T., Scharifker, B.R., 2003. Competitive electrochemical oxidation of p-chlorophenol and p-nitrophenol on Bi-doped PbO2. Electrochimica Acta 48 (19), 2775–2780. Boscoletto, A.B., Gottardi, F., Milan, L., Pannocchia, P., Tartari, V., Tavan, M., 1994. Electrochemical treatment of bisphenol-A containing wastewaters. Journal of Applied Electrochemistry 24, 1052–1058. Boye, B., Dieng, M.M., Brillas, E., 2002. Degradation of herbicide 4-chlorophenoxyacetic acid by advanced electrochemical oxidation methods. Environmental Science Technology 36 (15), 3030–3035. Brand, N., Mailhot, G., Bolte, M., 1998. Degradation photoinduced by Fe(III): method of alkylphenol ethoxylates

2533

removal in water. Environmental Science Technology 32 (18), 2715–2720. Comninellis, C., 1994. Electrocatalysis in the electrochemical conversion/combustion of organic pollutants for waste water treatment. Electrochimica Acta 39 (11/12), 1857–1862. Da Silva, L.M., De Faria, L.A., Boodts, J.F.C., 2001. Green processes for environmental application. Electrochemical ozone production. Pure and Applied Chemistry 73 (12), 1871–1884. Feng, Y.J., Li, X.Y., 2003. Electro-catalytic oxidation of phenol on several metal-oxide electrodes in aqueous solution. Water Research 37 (10), 2399–2407. Franska, M., Franski, R., Szymanski, A., Lukaszewski, Z., 2003. A central fission pathway in alkylphenol ethoxylate biodegradation. Water Research 37 (5), 1005–1014. Houk, L.L., Johnson, S.K., Feng, J., Houk, R.S., Johnson, D.C., 1998. Electrochemical incineration of benzoquinone in aqueous media using a quaternary metal oxide electrode in the absence of a soluble supporting electrolyte. Journal of Applied Electrochemistry 28, 1167–1177. Ike, M., Asano, M., Belkada, F.D., Tsunoi, S., Tanaka, M., Fujita, M., 2002. Degradation of biotransformation products of nonylphenol ethoxylates by ozonation and UV/ TiO2 treatment. Water Science and Engineering 46 (11/12), 127–132. Johnson, D.C., Feng, J., Houk, L.L., 2000. Direct electrochemical degradation of organic wastes in aqueous media. Electrochimica Acta 46 (2/3), 323–330. Johnson, S.K., Houk, L.L., Feng, J., Houk, R.S., Johnson, D.C., 1999. Electrochemical incineration of 4-chlorophenol and the identification of products and intermediates by mass spectrometry. Environmental Science Technology 33 (15), 2638–2644. Kuramitz, H., Saitoh, J., Hattori, T., Tanaka, S., 2002. Electrochemical removal of p-nonylphenol from dilute solutions using a carbon fiber electrode. Water Research 36 (13), 3323–3329. Mizuno, T., Yamada, H., Tsuno, H., 2002. Characteristics of oxidation by-products formation during ozonation and ozone/hydrogen peroxide process in the aqueous solution of nonylhenol ethoxylates. Advances in Asian Environmental Engineering 2 (2), 33–42. Montgomery-Brown, J., Reinhard, M., 2003. Occurrence and behavior of alkylphenol polyethoxylates in the environment. Environmental Engineering Science 20 (5), 471–486. Montilla, F., Michaud, P.A., Morallon, E., Vazquez, J.L., Comninellis, C., 2002. Electrochemical oxidation of benzoic acid at boron-doped diamond electrodes. Electrochimica Acta 47 (21), 3509–3513. Nilson, A., Ronlan, A., Parker, V.D., 1973. Anodic oxidation of phenolic compounds. Part III. Anodic hydroxylation of phenols. A simple general synthesis of 4-alkyl-4-hydroxycyclohexa-2,5-dienones from 4-alkylphenols. Journal of Chemecal Society Perkin I, 2337–2345. Petrovic, M., Diaz, A., Ventura, F., Barcelo, D., 2003. Occurrence and removal of estrogenic short-chain ethoxy nonylphenolic compounds and their halogenated derivatives during drinking water treatment. Environmental Science Technology 37 (19), 4442–4448.



Polcaro, A.M., Palmas, S., 1997. Electrochemical oxidation of chlorophenols. Industrial Engineering and Chemical Research 36 (5), 1791–1798. Rudel, R.A., Camann, D.E., Spengler, J.D., Korn, L.R., Brody, J.G., 2003. Phthalates, pesticides, polybrominated diphenyl ethers, and other endocrine-disrupting compounds in indoor air and dust. Environmental Science Technology 37 (20), 4543–4553. Sadtler, 1979. In: Simons, W.W. (Ed.), Handbook of Ultraviolet Spectra. Sadtler Research Laboratories, Philadelphia, PA. Sato, H., Shibata, A., Wang, Y., Yoshikawa, H., Tamura, H., 2003. Characterization of biodegradation intermediates of nonionic surfactants by MALDI-TOF 2. Oxidative biodegradation profiles of uniform octylphenol polyethoxylate in 18 O labeled water. Biomacromolecules 4 (1), 46–51. Schumann, U., Gluschke, M., Grundler, P., Mikolajczyk, R., Mikolajczyk, L., 1998. Reaction pathways in electrochemical oxidation of non-ionic surfactants—comparison with ozonolysis. Tenside Sufactants Detergents 35 (5), 379–386. Sherrard, K.B., Marriott, P.J., McCormick, M.J., 1994. Electrospray mass spectrometric analysis and photocatalytic degradation of polyethoxylate surfactants used in wool scouring. Analytical Chemistry 66, 3394–3399. Sherrard, K.B., Marriott, P.J., Amiet, R.G., McCormick, M.J., Colton, R., Millington, K., 1996. Spectroscopic analysis of heterogeneous photocatalysis products of nonylphenol- and primary alcohol ethoxylate nonionic surfactants. Chemosphere 33 (10), 1921–1940. Standard Methods for Examination of Water and Wastewater, 1995. In: Eaton, A.D., Clesceri, L.S., Greenberg, A.E. (Eds.), American Public Health Association, Washington DC.

Tahar, N.B., Savall, A., 1999. A comparison of different lead dioxide coated electrodes for the electrochemical destruction of phenol. Journal of New Material for Electrochemical Systems 2 (1), 19–26. Tanaka, S., Nakata, Y., Kuramitz, H., Kawasaki, M., 1999. Electrochemical decomposition of bisphenol A and nonylphenol using a Pt/Ti electrode. Chemical Letters 28 (9), 943–944. Velichenko, A.B., Amadelli, R., Baranova, E.A., Girenko, D.V., Danilov, F.L., 2002. Electrodeposition of Co-poped lead dioxide and its physicochemical properties. Journal of Electroanalytical Chemistry 527 (1/2), 56–64. Wenzel, A., Boehmer, W., Mueller, J., Ruedel, H., SchroeterKermani, C., 2004. Retrospective monitoring of alkylphenols and alkylphenol monoethoxylates in aquatic biota from 1985 to 2001: results from the German environmental specimen bank. Environmental Science Technology 38 (6), 1654–1661. Wintgens, T., Gallenkemper, M., Melin, T., 2003. Occurrence and removal of endocrine disruptors in landfill leachate treatment plants. Water Science Technology 48 (3), 127–134. You, T., Niwa, O., Tomita, M., Ichino, T., Hirono, S., 2002. Electrochemical oxidation of alkylphenols on ECR-sputtered carbon film electrodes with flat sub-nanometer surfaces. Journal of Electrochemical Society 149 (12), E479–E484. Zhi, J.F., Wang, H.B., Nakashima, T., Rao, T.N., Fujishima, A., 2003. Electrochemical incineration of organic pollutants on boron-doped diamond electrode. Evidence for direct electrochemical oxidation pathway. Journal of Physical Chemistry B 107 (48), 13389–13395.