Aqueous pesticide degradation by hydrogenperoxide/ultraviolet irradiation and Fenton-typeadvanced oxidation processes: a review
Keisuke Ikehata and Mohamed Gamal El-Din
Abstract: Pesticide pollution of surface water and groundwater has been recognized as a major problem in many countriesbecause of the persistence of pollutants in aquatic environments and the consequent potential adverse health effects.Various hydrogen peroxide-based advanced oxidation processes, such as hydrogen peroxide/ultraviolet irradiation, Fenton,photo-Fenton, and electro-Fenton processes are likely key technologies for degrading and detoxifying these pollutants inwater and wastewater. In this paper, the hydrogen peroxide-based advanced oxidation treatment of eight major groups ofpesticides, namely aniline-based compounds, carbamates, chlorophenoxy compounds, organochlorines, organophosphates,pyridine and pyrimidine derivatives, triazines, and substituted ureas, as well as that of several miscellaneous pesticides, isreviewed. The degree of pesticide degradation, reaction kinetics, identity and characteristics of degradation by-products andintermediates, and possible degradation pathways are covered and discussed.
Résumé : La pollution de l’eau de surface et de l’eau souterraine par des pesticides a été reconnue comme étant unproblème majeur dans plusieurs pays en raison de la persistance des polluants dans les environnements aquatiqueset des effets néfastes potentiels conséquents sur la santé. Divers procédés d’oxydation avancée basés sur le peroxyded’hydrogène, tels que les procédés au peroxyde d’hydrogène/irradiation aux ultraviolets, Fenton, photo-Fenton et électro-Fenton, sont probablement des technologies clés pour dégrader et détoxifier ces polluants dans l’eau et les eaux usées. Leprésent article examine le traitement d’oxydation avancée basé sur le peroxyde d’hydrogène de huit groupes principaux depesticides, notamment les composés à base d’aniline, les carbamates, les chlorophénoxydes, les composés organochlorés,les composés organophosphorés, les dérivés de la pyridine et de la pyrimidine, les triazines et les urées de substitution,ainsi que ceux de plusieurs différents pesticides. Le degré de dégradation des pesticides, la réaction cinétique, l’identité etles caractéristiques des sous-produits et des produits intermédiaires de la dégradation, ainsi que les voies de dégradationpossibles sont traités et discutés.
A large number of pesticide active ingredients have beenregistered and marketed for pest control purposes around theworld. They are divided into several major types depending
on their usage. The categories include herbicides, insecticides,fungicides, rodenticides, nematicides, microbiocides, and plantand insect growth regulators. Pesticide residues are widespreadin streams and shallow groundwater, and their occurrence fol-lows the patterns of geographic and seasonal use of pesticides
Received 13 December 2004. Revision accepted 14 November 2005. Published on the NRC Research Press Web site at http://jees.nrc.ca/ on10 March 2006.
K. Ikehata and M. Gamal El-Din.1 Department of Civil and Environmental Engineering, 3-093 Markin/CNRL Natural Resources EngineeringFacility, University of Alberta, Edmonton, AB T6G 2W2, Canada.
Written discussion of this article is welcomed and will be received by the Editor until 31 July 2006.
in the area (Kolpin et al. 1998; Gilliom et al. 1999). For ex-ample, high concentrations of herbicides are often found in theareas with extensive agricultural activity, whereas many insec-ticides are found in urban streams. Possible causes of pesticidecontamination in drinking water sources include agriculturaland urban runoffs, direct application of pesticides to controlaquatic insects and vegetation, domestic usage, leaching frompesticide wastes, and industrial-scale pest control operations.Because pesticides are so difficult to remove from aquatic envi-ronments, and because of the potential health risks they present,pesticide pollution in surface water and groundwater has beenrecognized for many years as an important issue in a numberof countries.
Various treatment processes have been investigated to re-duce pesticide concentrations in water, and to minimize thepotential health risks associated with exposure to the chem-icals through consumption of contaminated waters (Al Mo-mani et al. 2004). Two types of contaminated aqueous mediashould be considered here: wastewaters from pesticide manu-facturing plants, agricultural fields, and equipment rinsing op-erations (rinse water or rinsate), as well as surface water andgroundwater. Whereas wastewaters often contain very high lev-els (milligram per litre or more) of pesticides, surface water andgroundwater usually contain only trace amounts of pesticides(microgram per litre or less), but these often occur as a morecomplex mixture (Felsot 1996; Kolpin et al. 1998). Thus, ap-plicable treatment options should be different for wastewaterand surface–groundwater treatments. For wastewater treatment,physical treatments such as lined evaporative beds and activatedcarbon adsorption, chemical treatments such as photolysis, hy-drolysis and chemical oxidation, and biological treatments suchas activated sludge, biobeds, and constructed wetlands havebeen evaluated (Felsot 1996; Felsot et al. 2003). Felsot et al.(2003) suggested that the combination of physical or chemicalmethods with biological treatment was likely a feasible optionfor the detoxification of pesticide wastewater.
On the other hand, biodegradation is not usually involvedin the removal of trace pesticides during water treatment, ex-cept in the case of inoculated granular activated carbon filtration(Feakin et al. 1995; van der Hoek et al. 1999). Physical unit pro-cesses such as nanofiltration and reverse osmosis (Agbekodo etal. 1996; Berg et al. 1997; Van der Bruggen et al. 1998; Bous-sahel et al. 2000), slow sand filtration (Lambert and Graham1995), and activated carbon adsorption (Baldauf 1993; Gicquelet al. 1997; Thacker et al. 1997), as well as chemical oxidationsuch as ozonation and advanced oxidation processes (Reynoldset al. 1989; Camel and Bermond 1998), have proven usefulin the removal of pesticide residues during water treatment.However, neither physical nor chemical approaches alone canachieve the complete removal of pesticides. Physical processesmerely transfer pollutants to another phase, and it is necessaryto destroy rejected or adsorbed pesticides. In addition, activatedcarbon adsorption is not applicable to polar substances (Baldauf1993). Chemical oxidation can lead to incomplete destructionof pesticide molecules and thus, the formation of undesirable
by-products. Therefore, a combination of physical and chemi-cal unit processes is required and actually employed to ensurethe removal of pesticide residues and by-products from drinkingwater.
Chemical oxidation is apparently a key technology for solv-ing the pesticide removal problems in both water and waste-water treatments. Various chemical oxidants have been eval-uated for such purpose, including chlorine, chlorine dioxide,potassium permanganate, ozone (O3), and hydrogen peroxide(H2O2). However, these chemical oxidants, with the exceptionof ozone, which has a relatively high oxidation potential (2.07V;relative to the hydrogen electrode), are not effective enough todegrade highly refractory synthetic organic chemicals like pes-ticides. On the other hand, combinations of chemical oxidants(such as O3 and H2O2), iron salts, semiconductors (such as tita-nium dioxide, TiO2), and (or) ultraviolet-visible light (UV-Vis)irradiation yield hydroxyl radicals, which are very powerful ox-idants with an even higher oxidation potential (2.8 V) than thatof molecular ozone (Andreozzi et al. 1999). Such processes arecollectively known as advanced oxidation processes (AOPs),and have gained considerable popularity in recent years in thefields of water and wastewater treatment, including pesticidewaste treatment (Rice 1997; Peñuela and Barceló 1998; Felsotet al. 2003; Ikehata and Gamal El-Din 2004).
In this review, a critique of the recent literature published inthe past 15 years concerning the H2O2-based AOPs, includingH2O2/UV and various Fenton-type processes for aqueous pes-ticide degradation, is presented to gain up-to-date informationon such topics as the degree of degradation, reaction kinetics,identity and characteristics of oxidation by-products, and pos-sible degradation pathways. The pesticides reviewed here aregrouped according to their unique chemical structures as listedin Table 1. Brief descriptions of the general characteristics ofpesticides are also provided in the following sections.
Basic characteristics of pesticides
Pesticide typesPesticides can be classified into several different types de-
pending on the particular pest organisms targeted. Categoriesinclude herbicides, insecticides, fungicides, rodenticides, ne-maticides, and microbiocides (antimicrobials). In addition tothese six major groups, a number of subtypes are also definedby some regulatory agencies. The chemical substances that ac-tually have pesticidal potentials are called pesticide active in-gredients (AIs). More than 1000 AIs are listed in the USEPA’s1998 Rainbow Report (USEPA 1998).
Commercial pesticide formulations (products) contain vari-ous additives besides AIs, and often have many different tradenames. In addition to these trade names, AIs themselves alsohave several different common names, as well as a chemicalname determined by the International Union of Pure and Ap-plied Chemistry (IUPAC). In this review, however, the commonnames of pesticides referred to in the publications reviewed aregenerally used; other names are mentioned occasionally in the
text. For the sake of simplicity, terms such as “pesticide”, “her-bicide”, and “insecticide” are used in this review to account forthe corresponding AIs.
Pesticides can also be classified according to their uniquechemical structures (or functional groups) that exert pesticidaleffects on target organisms. These include anilides, carbamates,chlorophenoxy carboxylic acids, organochlorines, organophos-phates, substituted ureas, triazines, and others (see Table 1).As one can expect, substances with similar chemical structuresnormally have similar pesticidal effects. For example, most ofthe triazine compounds, such as atrazine, can be used to selec-tively kill weeds by inhibiting photosynthesis and are generallyregarded as herbicides (US EPA 2002), although, as always,there are some exceptions.
Environmental and public health significance ofpesticides
Pesticides in streams are a potential concern for human healthif they affect a drinking water source or occur in recreationaluse areas, such as those frequented for bathing and swimming.Pesticides also present a potential threat for aquatic life andecosystems in all streams. Although the primary issue relatedto groundwater is drinking water quality, groundwater may alsofunction as a source of pesticides for surface water. Variousadverse health effects are known to result from many pesti-cides. Ecological effects of trace pesticide residues, and thefate of these residues in the environment, are also importantissues for both the scientific community and the general public.Readers can refer to several online resources, such as thoseof the USEPA (http://www.epa.gov/pesticides/), the ATSDR(http://www.atsdr.cdc.gov/toxfaq.html), and the US NationalToxicology Program (http://ntp-server.niehs.nih.gov/), for theupdates on potential health and environmental effects, as well
as information on some physicochemical properties of partic-ular substances. The limits and guideline values for pesticidesin drinking water and some environmental waters issued bythe World Health Organization (WHO), Australia, the UnitedStates, New Zealand, Japan, Canada, the European Union, andTaiwan have recently been reviewed and published (Hamiltonet al. 2003).
Hydrogen peroxide, Fenton, and advancedoxidation processes
Hydrogen peroxide (H2O2) is a strong oxidant, having oxi-dation potentials of 1.80 and 0.87 V at pH 0 and 14, respectively(Neyens and Baeyens 2003).Although hydrogen peroxide is notstrong enough to oxidize most pesticides by itself, combiningit with other chemical and physical agents facilitates the for-mation of hydroxyl radicals that have a much higher oxidativepotential (2.8 V) than the parent oxidant (Neyens and Baeyens2003). The chemical and physical agents include ferrous ion(Fe2+), ozone, and UV irradiation as shown below:
The first and second combinations are referred to as theFenton process and O3/H2O2 (not covered in this review), re-spectively. The third reaction is usually referred asH2O2/UV (or UV/H2O2) because the wavelength at which H2O2molecules absorb radiation falls within the ultraviolet region.The use of hydrogen peroxide as an oxidant has a number ofadvantages over other chemical treatments such as chlorina-tion and ozonation: its commercial availability, thermal stabil-ity and storage on-site, infinite solubility in water, no mass-
transfer problems associated with gases, minimal capital in-vestment, and no formation of disinfection by-products suchas halogenated hydrocarbons and bromate ion (Legrini et al.1993; Symons and Zheng 1997). There are a few varietiesof Fenton process covered in this review including Fenton-like, photo-Fenton, electro-Fenton, photoelectro-Fenton,peroxi-coagulation, and chelating agent assisted Fenton-typeprocesses (Table 2). The fundamental aspects of these variousFenton and H2O2/UV processes, as well as general hydroxylradical reactions are briefly described below. Readers can referto literature for the pesticide degradation by AOPs and relatedtechnologies other than the ones reviewed here including ozone-based processes (Ikehata and Gamal El-Din 2005a, 2005b) anddirect photolysis and various photochemical processes such asTiO2/hν (Burrows et al. 2002).
General hydroxyl radical reactionsHydroxyl radical is a short lived, extremely potent oxidizing
agent that can oxidize organic compounds (RH) through hy-drogen abstraction (eq. [4]), electrophilic addition (eq. [5]), orelectron-transfer reaction (eq. [6]) (Legrini et al. 1993).
[4] RH + ·OH → H2O + R·
[5]
R
R
R
R
R
R
R
ROH+ ·OH
[6] ·OH + RX → OH− + RX·+Of these hydroxyl radical reactions, hydrogen abstraction is a
major mechanism in most cases. The generated organic radicals
(R·) can be oxidized further by hydrogen peroxide or molecularoxygen. As a result, hydroxyl radicals or peroxyl radicals willbe generated, respectively.
[7] R· + H2O2 → ROH + ·OH
[8] R· + O2 → ROO·The sequence of above reactions may lead eventually to min-
eralization of organic compounds into carbon dioxide, water,and inorganic ions. At the same time, hydroxyl radicals cancouple each other to form hydrogen peroxide (eq. [9]) or reactwith hydrogen peroxide to form water and hydroperoxyl rad-icals (eq. [10]). Dimerization of two organic radicals can alsooccur (eq. [11]). The values for the rate constants in eqs. [9]and [10] are taken from Duesterberg et al. (2005).
[9] 2 ·OH → H2O2 (k = 5.2 × 109 M−1· s−1)
[10] ·OH+H2O2 → H2O+HO2 · (k = 3.3×107 M−1· s−1)
[11] 2R· → R-R
Hydroperoxyl radicals (HO2·) generated in eq. [10] havemuch less oxidizing potential and likely do not contribute tothe oxidative degradation of organic substances (Legrini et al.1993). Thus, hydrogen peroxide can act as a hydroxyl radicalscavenger as well as an initiator (Neyens and Baeyens 2003).Carbonate and bicarbonate ions as well as natural organic matterare also known hydroxyl radical scavengers (Glaze et al. 1995;Crittenden et al. 1999). These reactions slow down the degra-dation of organic substances and lower the process efficiencyin advanced oxidation processes.
Classic Fenton and Fenton-like processesFenton process was first demonstrated by Fenton (1894) uti-
lizing a mixture of hydrogen peroxide and ferrous salts (i.e.,Fenton’s reagent) to oxidize tartaric acid to dihydroxy maleicacid. Haber and Weiss (1934) later suggested the formationof hydroxyl radicals upon the iron-catalyzed decomposition ofhydrogen peroxide. In addition to the ferrous ion (Fe2+), a num-ber of other metal ions, including Cu+, Ti3+, Cr+, Co2+, werefound to have the oxidative features of Fenton’s reagent in thepresence of hydrogen peroxide (Goldstein et al. 1993).
The generation and subsequent reaction of hydroxyl radicalsin Fenton process involves complex chain reactions (Neyensand Baeyens 2003). The first step is the chain initiation and ratelimiting (eq. [1], k ∼= 70 M−1·s−1). This reaction requires aproton to initiate; thus the Fenton process generally needs tobe employed under acidic conditions. Radical chain reactionspropagate further through the reactions shown in eqs. [4] to [8].Hydroxyl radicals may also react with ferrous ion to terminatethe chain reactions (eq. [12]). The values for the rate constantsshown in eqs. [12] to [16] are taken from Neyens and Baeyens(2003).
[12] ·OH+Fe2+ → OH−+Fe3+ (k = 3.2×108 M−1· s−1)
The ferric ion may react with hydrogen peroxide in the fol-lowing manner:
[13] Fe3+ + H2O2 → Fe-OOH2+ + H+
(k = 0.001–0.01 M−1· s−1)
The iron (II) hydroperoxyl ion (Fe-OOH2+) decomposes intohydroperoxyl radical and ferrous ion.
[14] Fe-OOH2+ → HO2· + Fe2+
The reaction of hydrogen peroxide with ferric ion (eqs. [13]and [14]) is often referred to as a “Fenton-like” reaction. Hy-droperoxyl radicals may react with ferric or ferrous ions.
[15] Fe2+ + HO2· → Fe3+ + HO2−
(k = 1.3 × 106 M−1· s−1 at pH 3)
[16] Fe3+ + HO2· → Fe2+ + O2 + H+
(k = 1.2 × 106 M−1· s−1 at pH 3)
In addition, organic radicals R· generated in eq. [4] can beoxidized by ferric ion or reduced by ferrous ion.
[17] R· + Fe3+ → R+ + Fe2+
[18] R· + Fe2+ → R− + Fe3+
The reaction in eq. [17] regenerates ferrous ion and concludesthe catalytic cycle of Fenton process. Meanwhile, the ferrousions generated during Fenton process described above may be
hydrated, and subsequently react with hydroxide ions to formferric hydroxo complexes that precipitate between pH 3 and 7.
[19] [Fe(H2O)6]3+ + H2O → [Fe(H2O)5OH]2+ + H3O+
[20] [Fe(H2O)5OH]2+ + H2O
→ [Fe(H2O)4(OH)2]+ + H3O+
[21] 2[Fe(H2O)5OH]2+ → [Fe2(H2O)8(OH)2]4+ + 2H2O
These reactions account for the coagulation capacity of Fen-ton’s reagent (Neyens and Baeyens 2003). Suspended solids canbe captured and precipitated along with the ferric complexes.This feature may be useful for the treatment of wastewaters hav-ing high suspended solids content. At the same time, the ferricions become unavailable for further reactions. The requirementof pH control is one of the drawbacks of classic Fenton andFenton-like processes, as well as the generation and disposal ofsludge at the end of treatment.
Photo-Fenton processThe rate of organic pollutant degradation following the Fen-
ton process is strongly accelerated by irradiating UV-Vis rays(Pignatello 1992; Oppenländer 2003). This modified Fentonprocess is called photo-assisted Fenton (or photo Fenton) pro-cess and involves photolysis of the hydroxyl complex of ferricion [Fe(OH)2+] into a hydroxyl radical and a ferrous ion.
[22] Fe(OH)2+ + hν(λ < 400 nm) → Fe2+ + ·OH
The hydroxyl complex has a maximum UV absorbance atabout 300 nm, and the quantum yields (�) of the above pho-tochemical reaction are 0.14 at 313 nm and 0.017 at 360 nm(Faust and Hoigné 1990). The regeneration of ferrous ion andadditional generation of hydroxyl radical during the photolysisfacilitate the degradation of organic compounds. In addition totypical UV lamps, UV-A and visible light sources such as poly-chromatic Hg lamps, black light bulbs, fluorescent light bulbs,and sunlight can be used as a light source for photo-Fentonprocess (Sun and Pignatello 1993c; Malato et al. 2003b). Inaddition to the hydroxyl radical reactions, direct photolysis oforganic compounds and hydroxyl radical generation by hydro-gen peroxide decomposition may need to be accounted for pol-lutant degradation pathways when a light source emitting shortwavelength UV-radiation (below 300 nm) is used (Wadley andWalte 2002). Acidification of reaction medium before the treat-ment and sludge disposal are still required in the photo-Fentonprocess.
Chelating ligand-assisted Fenton/photo-Fenton processAddition of chelating agents, such as oxalate and citrate, to
the photo-Fenton system is known to broaden the range of wave-length applicable to photolysis of ferric complexes to ferrousions (Zepp et al. 1992). Chelating ligands facilitate the dissolu-tion of ferric ions at pH ranging from 3 to 8. The Fe3+-chelate
photodissociates by a ligand-to-metal charge transfer excitation(Sun and Pignatello 1992):
[23] L-Fe3+ + hν → Fe2+ + L·where L represents an organic ligand. The photo-reduced ironcan react with hydrogen peroxide in the Fenton reaction (eq. [1]).Unlike the simple photo-Fenton process (eq. [22]), generationof hydroxyl radicals from the Fe3+-chelate complexes is notlikely to occur because the charge transfer to iron from an or-ganic ligand is favored over a hydroxide ligand (Sun and Pig-natello 1993a). It was also reported that some chelating organiclegands, such as gallic acid and picolinic acid, could enhancehydroxyl radical generation in Fenton and Fenton-like reactionswith or without UV irradiation (Sun and Pignatello 1992). Inaddition to hydroxyl radicals, formation of ferryl complexessuch as Fe4+ = O and [L·+]Fe4+ = O was suggested as activeoxidants involved in this system (Sun and Pignatello 1992). Or-ganic ligands are not very stable and can also be degraded bythe hydroxyl radicals as the treatment proceeds (Sun and Pig-natello 1992). Thus, the ligand does not act as a catalyst in thistype of Fenton process. Despite the significant enhancement inorganic pollutant degradation of the Fenton/photo-Fenton pro-cess by the addition of such a chelating ligand it introduces anadditional component into the wastewater, although the ligandcan be destroyed by the hydroxyl radicals.
Electro-Fenton, photoelectron-Fenton, and peroxi-coagulation processes
Various direct and indirect electrochemical treatment systemshave been evaluated for the destruction of toxic organic pollu-tants in aqueous medium. Anodic oxidation is an example ofdirect methods, in which degradation of organic compounds oc-curs by reaction with adsorbed hydroxyl radicals formed at thesurface of a high-oxygen overvoltage anode made of platinum,PbO2, IrO2, or boron-doped diamond (Johnson et al. 1999; Gan-dini et al. 2000; Brillas et al. 2003a).
[24] H2O → ·OHads + H+ + e−
On the other hand, electro-Fenton processes that involve elec-trogeneration of hydrogen peroxide and (or) ferrous ion areconsidered indirect electrochemical treatment methods. Hydro-gen peroxide can be electrochemically generated by the two-electron reduction of dioxygen in acidic medium on graphite,reticulated vitreous carbon, mercury pool, carbon-felt, andoxygen-diffusion cathodes (Oturan and Pinson 1995; Oturanet al. 1999; Brillas et al. 2003a).
[25] O2(g) + 2H+ + 2e− → H2O2
Controlled and continuous generation of hydrogen peroxideis one of the advantages of this type of electro-Fenton process.Ferrous salt, such as ferrous sulfate, is usually added before theelectrolysis to an undivided electrolytic cell.
Alternatively, ferrous ion can be generated by the two-electronoxidation of sacrificial metal iron electrodes (Roe and Lemley1997).
[26] Fe(s) → Fe2+ + 2e−
Simultaneously, electrolysis of water takes place at the cath-ode.
[27] 2H2O + 2e− → H2 + 2OH−
In this case, hydrogen peroxide needs to be delivered sep-arately to the system. Controlled delivery of ferrous ion is anadvantage of this type of electro-Fenton process. However, thisprocess has a disadvantage due to the generation of hydroxideion that causes precipitation of iron hydroxides. To overcomethis problem, two half cells can be separated by a salt bridgeor an anion exchange membrane (Saltmiras and Lemley 2002;Wang and Lemley 2002a). In this modified system, Fenton reac-tions only occur in the anodic half cell in which the solution pHis maintained low. Thus, this type of electro-Fenton process isalso called anodic Fenton process (Saltmiras and Lemley 2002;Lemley et al. 2004).
A combination of anodic and cathodic Fenton processes yieldsthe so-called peroxi-coagulation process (Brillas et al. 2003c;Durán Moreno et al. 2004). Hydrogen peroxide and ferrousion are simultaneously and continuously generated in situ viaeqs. [25] and [26]. Ferric ion and hydroxyl radical can be formedas a result of classical Fenton reaction and the excess of ferricion precipitates as hydrated Fe(OH)3 solids. Pollutants can beremoved by typical hydroxyl radical reactions as well as by thecoagulation with the iron hydroxides in the peroxi-coagulationprocess.
It is known that addition of UV irradiation to the electro-Fenton and peroxi-coagulation processes significantly improvesthe degradation efficiency of organic pollutants (Brillas et al.2003c, 2003d). These processes are called photoelectro-Fentonand photoperoxi-coagulation processes, respectively. Com-plexes of ferric ion and organic acids, which are resistant todegradation in the dark electro-Fenton processes, can be photo-decomposed and the organic acids can be mineralized in sucha system (Brillas et al. 2003c, 2003d).
H2O2/UV processThe application of the H2O2/UV process to waste treatment
was first proposed in the late 1970s, and has been investigatedfor the degradation of various aqueous organic pollutants fordecades (Legrini et al. 1993; Glaze et al. 1995; Stefan et al.1996; Cater et al. 2000). In the H2O2/UV process a hydro-gen peroxide molecule is cleaved into two hydroxyl radicals(eq. [3]) by UV photolysis at 253.7 nm with a quantum yieldof 0.98 (Legrini et al. 1993). The rate of photolysis of aqueoushydrogen peroxide is pH dependent and increases when morealkaline conditions are used. It has been suggested that peroxideanion (HO2
−) formation might be responsible for the enhancedgeneration of hydroxyl radicals under such conditions (pKa of
H2O2 = 11.6), since peroxide anion has a higher molar absorp-tion coefficient at 253.7 nm (240 M−1·cm−1) than hydrogenperoxide itself (18.6 M−1·cm−1). At the same time, a dismuta-tion reaction of hydrogen peroxide also occurs at alkaline pHsas shown below:
[28] H2O2 + HO2− → H2O + O2 + ·OH
The H2O2/UV process particularly suits groundwater anddrinking water treatment because unlike ozonation and ozone-based AOPs it does not produce bromate ion, which is a sus-pected carcinogen and disinfection by-product of ozone-basedtreatments from bromide in water (Von Gunten and Hoigné1994; Symons and Zheng 1997). Besides the degradation ofaqueous organic pollutants, it is also effective in inactivatingpathogenic microorganisms in water (Gardner and Shama 1998;Koivunen and Heinonen-Tanski 2005). However, there are alsoseveral drawbacks encountered with the H2O2/UV process. Themajor disadvantage is the limited rate of hydroxyl radical for-mation due to the rather small molar absorptivity of hydrogenperoxide at 254 nm. The photolysis of hydrogen peroxide canbe also inhibited in the presence of dissolved and suspendedsolids by reducing available photons for the photolysis throughlight scattering. Direct UV photolysis of organic pollutants alsoneeds to be taken into consideration depending on the chemicalstructure of the compounds.
The primary objective of aqueous pesticide treatment byAOPs, such as Fenton and H2O2/UV processes, is the detoxifi-cation of harmful pesticides or their transformation into formsmore amenable to biodegradation. The latter is only applicableto the treatment of wastewater for pesticides, not to the watertreatment. The hydroxyl radicals generated by these AOPs reactwith pesticide molecules in water with varying degrees of affin-ity depending on the chemical and physical properties of themolecules. Pesticide molecules are transformed into primaryoxidation products, which often undergo either spontaneoustransformation (e.g., protonation, dimerization, isomerization,and tautmerization) or further oxidation by another hydroxylradical. If UV-Vis irradiation and (or) another source of en-ergy are involved, multiple reaction mechanisms need to beconsidered. These reactions can occur competitively, depend-ing on reaction conditions such as pH and the presence of otherinorganic and organic species. Formulated pesticide productsalso contain ingredients other than AIs, so-called “inert ingre-dients” such as solvents, surfactants, carriers, and intensifiers.As a result, the treated pesticide solution tends to contain vari-ous by-products originating from the parent pesticide and otherconstituents in the solution.
It is very difficult to determine the degree of pesticide degra-dation by Fenton and H2O2/UVAOPs primarily due to the com-
plexity of the reactions mentioned above. Although often usedin the literature, complete disappearance (conversion) of pesti-cides as monitored by high performance liquid chromatography(HPLC), for example, is often not enough to detoxify the pesti-cide solution as potentially more toxic degradation by-productsmay be present in the treated solution. Alternately, the degree ofpesticide degradation and mineralization can be measured bythe reduction in organic content, such as total organic carbon(TOC), dissolved organic carbon (DOC), and chemical oxygendemand (COD), in the solution, or by the evolution of radioac-tive carbon dioxide (14CO2) from 14C labeled pesticide com-pounds. Although a high degree of mineralization is a morelikely sign of detoxification, it still cannot be counted uponas a guarantee, especially when other organic constituents arepresent. This is often the case in natural waters and real pesticidewastewaters. Therefore, more direct analyses, such as toxicityassays and biodegradability tests, are required to ensure thequality of treated water and wastewater.
Kinetic parameters such as reaction rate constants and acti-vation energy (Ea) are useful to model the treatment processand predict the behavior of chemical species during treatment.These parameters can also be used to compare the reactivity ofpesticides toward ozonation orAOPs, if they are determined un-der comparable reaction conditions. The kinetic rate constantscommonly reported for pesticide degradation by Fenton-type orH2O2/UV processes are overall second order rate constants forhydroxyl radical reactions and pseudo first order rate constantsfor pesticide conversion. It should be noted that the latter is onlyuseful for internal comparison and is covered here accordingly.
Most of the studies reviewed here were conducted at roomtemperature (20 ◦C), unless otherwise noted in the text. A tabu-lated summary is also presented in Appendix A, although onlyessential reaction conditions, such as initial pH and initial pesti-cide concentration, are shown due to space limitations. Readersare encouraged to refer to the original works cited for additionaldetails.
Aniline-based compoundsThe aniline-based herbicides reviewed here include alachlor,
butachlor, metazachlor, metolachlor, propachlor, and trifluralin.These compounds are chlorinated acetoanilides, with the ex-ception of trifluralin, which is a dinitroaniline derivative. Thechemical structures and formula weights of these compoundsare shown in Fig. 1.
AlachlorAlachlor is one of the most heavily used chlorinated ace-
toanilide herbicides. Various Fenton processes were examinedfor alachlor degradation in aqueous solution.Arnold et al. (1996)reported nearly complete conversion of 26.4 mg·L−1 of alachlorby Fenton process (295 mg·L−1 Fe2+, 170 mg·L−1 H2O2,pH 2.5, 25 ◦C) in a pesticide rinse water containing four otherpesticides, including atrazine, cyanazine, metolachlor, andEPTC.
Fig. 1. Chemical structure of aniline-based pesticides (formulaweight is shown in brackets).
N OCl
O
NN
NCl
O
NO
O
Cl
N
N
O
ON
O
O
F FF
N
Cl
O
O
NCl
O
alachlor (269.77)
metazachlor (277.75)
metolachlor (283.80) trifluralin (335.28)
butachlor (311.85)
propachlor (211.69)
Pratap and Lemley (1994) evaluated classical Fenton andelectrochemical peroxide (electro Fenton) treatments for thedegradation of several pesticides including alachlor. This elec-trochemical Fenton process involved the generation of ferrousion by electrolysis with a constant current of 0.3 A in thepresence of hydrogen peroxide. After 5 to 20 min of elec-trolysis, the electrodes were removed (t = 0), and Fenton-type reactions started. Although nearly complete conversion of30 mg·L−1 of alachlor was achieved by electro Fenton treat-ment with 200 mg·L−1 generation of iron after 3.5 h, the clas-sical Fenton process with less iron (50 mg·L−1) was found tobe more efficient (7.5 g·L−1 of H2O2). Alachlor conversionwas much faster in the latter process, with complete conversionachieved after 30 min.
Huston and Pignatello (1999) reported complete conversionof 54 mg·L−1 of alachlor (72% removal as TOC) by photoFenton treatment at pH 2.8 within 120 min (λ = 300–400 nm,1.2 × 1919 photons·L−1·s−1, 2.8 mg·L−1 of Fe3+, 340 mg·L−1
of H2O2). They also observed that the degradation of alachlorby the photo Fenton process was slower when a commercialformulation of this compound was treated. Formulation typealso had an impact on the degradation of this pesticide.
Quantitative dechlorination and the formation of nitrate wereobserved after the photo Fenton treatment (Huston and Pig-natello 1999). Pratap and Lemley (1994) qualitatively char-acterized the organic by-products of alachlor by Fenton andelectrochemical Fenton treatments, and suggested the hydrox-ylation of alachlor during the treatments. No further attemptwas made to identify these by-products.
Recently, Farré et al. (2005) evaluated catalytic ozonationprocesses including the photo-Fenton combined with ozona-tion (PhFO) and TiO2 photocatalysis combined with ozona-tion (PhCO) as well as conventional O3/UV AOP for degra-dation of several EU Priority Hazardous Substances includingalachlor. The PhFO process was apparently superior to the othertwo AOPs, and nearly 70% of initial TOC was removed from50 mg·L−1 of alachlor solution in about 1 h of treatment at
pH 3 and 25 ◦C (1.6 g·h−1 of applied ozone, 6 W black light,0.48 g·L−1 of H2O2, 2 mg·L−1 of FeSO4·7H2O). However,the Microtox toxicity of the pesticide solution increased afterthe treatment for 3 h, suggesting the formation of toxic by-product such as 2,6-diethylaniline (Farré et al. 2005). Similarresults were reported using a solar-driven photo-Fenton pro-cess with 2 to 55 mg·L−1 of FeSO4·7H2O and 0.48 g·L−1 ofH2O2 (Hincapié et al. 2005). Their observation suggested thatdechlorination would be a key step to detoxify this chlorinatedacetoanilide herbicide by AOPs.
Butachlor and propachlorBenitez et al. (2004a) investigated the kinetics of several
acetoanilide herbicides, including butachlor, propachlor, andmetolachlor, in UV direct photolysis, H2O2/UV, and O3/UVprocesses using a low pressure mercury vapor lamp that emit-ted monochromatic radiation at 254 nm (2.03 × 10−6 Eins·s−1
at 3.5 cm). The quantum yields for the photolysis of butachlor,propachlor, and metolachlor were determined as 0.78, 0.127,and 0.56 mol·Eins−1, respectively. A kinetic model was alsodeveloped to predict the conversion of these acetoanilide herbi-cides by H2O2/UV (and O3/UV) AOP in natural waters. Degra-dation by-products or intermediates were not identified.
MetazachlorHessler et al. (1993) investigated the degradation of
10.3 mg·L−1 of metazachlor in aqueous solution with UV ir-radiation (λ = 254 nm, 6.3 × 10−7 Eins·L−1·s−1) in both thepresence and absence of hydrogen peroxide. They observedstrong increases in the quantum yields of metazachlor conver-sion at pH 3 and pH 7 in the presence of 6.8–68 mg·L−1 ofH2O2, as compared with direct photolysis, whereas increas-ing H2O2 concentration to more than 68 mg·L−1 lowered thequantum yields. This is most likely due to the inner filter ef-fect of H2O2 molecules in water that hinders the photon pen-etration (Hessler et al. 1993). No dark reaction of H2O2 withmetazachlor was observed. No attempt was made to identify thedegradation products of metazachlor by the H2O2/UV AOP.
MetolachlorMetolachlor is another heavily used chlorinated acetoanilide
herbicide. Benitez et al. (2004a) investigated UV direct pho-tolysis, H2O2/UV, and O3/UV processes for the degradationof this herbicide (see Butachlor, propachlor). Complete con-version by Fenton process of 59 mg·L−1 of metolachlor in apesticide rinse water was reported (Arnold et al. 1996). Nearlycomplete mineralization of 28.4 mg·L−1 of metolachlor was re-ported using a photo Fenton system with 340 mg·L−1 of H2O2,55.8 mg·L−1 of Fe3+, and UV-Vis irradiation (λ = 300–400 nm,1×1018 photons·L−1·s−1) (Pignatello and Sun 1995). After 6 hof reaction, the herbicide, including its aromatic ring, was com-pletely mineralized to chloride, ammonia, nitrate, and CO2. Theanalysis of organic intermediates during oxidation indicatedthat the metolachlor degradation occurred by the non-selectiveattack of hydroxyl radicals on the molecules. A comparativestudy performed on the photo Fenton treatment of metolachlor
Fig. 2. Metolachlor degradation products of the photo Fentonprocess (Pignatello and Sun 1995) and electro Fenton (inset)(Pratap and Lemley 1998).
N
O
Cl
O
O
NCl
O
OHNH
Cl
O
HNO
NO
O
Cl
HO
NO
O
Cl
HO
HNOH
and other pesticides indicated that this herbicide was relativelyeasy to degrade by this process (Huston and Pignatello 1999).
Pratap and Lemley (1994, 1998) investigated the electro Fen-ton process for the removal of 45–170 mg·L−1 of metolachlorin water. Although their initial attempt at utilizing electroly-sis to produce ferrous ion for the Fenton reaction resulted ina less efficient conversion of this herbicide than that achievedby classical Fenton treatment (Pratap and Lemley 1994), stepaddition of hydrogen peroxide as well as near-UV illumination(λ = 330–400 nm, 201 mg·L−1 of iron, 7.5 g·L−1 of H2O2;photoelectro Fenton) accelerated herbicide conversion (Pratapand Lemley 1998).
Several degradation by-products and intermediates of meto-lachlor produced by Fenton-type processes were reported. For-mation of chloroacetate, oxalate, formate, serin, and severalaromatic products (Fig. 2) was noticed, and their concentra-tions were monitored during the photo Fenton treatment ofmetolachlor (Pignatello and Sun 1995). One possible reactionproduct of the electro Fenton treatment of metolachlor was2-[(2-ethyl-6-methylphenyl)-amino]-1-propanol (Fig. 2; inset),although this compound was a minor product according to theHPLC profile obtained (Pratap and Lemley 1998).
TrifluralinTrifluralin is a fluorinated dinitroaniline herbicide. Sun and
Pignatello (1993a) reported a very slow conversion of trifluralinby Fe3+-chelate/H2O2 (dark) treatment. Three types of chelat-ing agents were evaluated, including picolinic acid, gallic acid,and rhodizonic acid (Fig. 3). These chemicals form complexeswith ferric ion and promote the generation of hydroxyl radi-cals, even in the absence of UV-Vis irradiation. The addition ofoxalate, another chelating agent often used in the photo Fentonprocess, did not show any effect under the dark condition. Onlyup to 40% conversion of 2.4 mg·L−1 of trifluralin was achievedafter 6 h of treatment with 55 mg·L−1 of Fe3+, 340 mg·L−1 ofH2O2, and 1 mM chelator at pH 6 and 25 ◦C.
Fig. 3. Chelating agents evaluated for the degradation oftrifluralin by Fe3+-chelate/H2O2 process.
NOH
O
picolinic acid
OH
HO OH
OHO
gallic acid
OH
OH
O
O
O
O
rhodizonic acid
Saltmiras and Lemley (2001) reported the use of anodic Fen-ton treatment to degrade a commercial formulation of this dini-troaniline herbicide.Anodic Fenton treatment is a modified ver-sion of the electrochemical peroxidation (or electro Fenton, seealso Metolachlor) process. This improved system utilizes twoseparated half-cells connected with a salt bridge. Fenton reac-tions occur only in the anode half-cell, in which ferrous ionsare produced by a sacrificial iron anode (Saltmiras and Lemley2001). Up to 70–85% conversion of 5–33.5 mg·L−1 of trifluralinwas achieved with the anodic Fenton process with continuousdelivery of ferrous ions (20.8 mg·L−1·min−1) and H2O2 with1:1 or 1:10 (for 5 mg·L−1 of trifluralin) molar ratio for 1 h.Significant volatilization of herbicide (37–54%) was also ob-served. Saltmiras and Lemley (2001) also observed a far morerapid removal of the inert ingredient that enhanced solubilityof trifluralin in water. They suggested that the rapid degrada-tion of this inert ingredient might accelerate the volatilizationof trifluralin during the treatment. No degradation by-productsof this dinitroaniline herbicide were identified in any of thesestudies.
Summary of aniline-based pesticidesThe chlorinated acetoanilides reviewed here are readily
degradable by Fenton-type AOPs, whereas trifluralin, a flu-orinated dinitroaniline, appears to be more resistant. This islikely due to the fluorination and high hydrophobicity of thelatter compound. Major mineralization of alachlor and meto-lachlor has been demonstrated, and several inorganic and or-ganic degradation by-products have been identified for thesecompounds. Increasing toxicity has been observed duringalachlor degradation by photo-Fenton and photo-Fenton/O3 pro-cesses. No study has been reported on the biodegradability im-provement of these pesticides in H2O2-based AOPs.
CarbamatesThe carbamate pesticides reviewed here include aldicarb,
asulam, bendiocarb, carbaryl (NAC), carbofuran, dioxacarb,EPTC, fenobucarb (BPMC), formetanate, methomyl, oxamyl,promecarb, propamocarb, and propoxur (Baygon). The chem-ical structures and formula weights of these compounds areshown in Fig. 4. Among the compounds listed above, EPTCis a thiocarbamate pesticide. Ethylene thiourea is also coveredhere because it is a potentially carcinogenic impurity and a ma-jor degradation product of dithiocarbamate pesticides such asmancozeb. These compounds are reviewed in the latter part ofthis section.
Fig. 4. Chemical structure of carbamate pesticides (formulaweight is shown in brackets).
HN
O
ON
S
H2N
S
O
OHN
O
O
O
O
ONH
O
O
HN
OO
O
O
NHO O
O NH
O
O
HN
O
N
O
O
NH
N
S NO
O
NH
O
HN
O
ON
S N
O
O
HN
N NH
O
O
O
O
NH
O
S
O
NNHHN
S
aldicarb (190.26) asulam (230.24)
bendiocarb (223.23)
carbaryl (201.22)
carbofuran (221.26)dioxacarb (223.23)
fenobucarb (207.27)
formetanate (221.26)
methomyl (192.23)
promecarb (207.27)
oxamyl (219.26) propamocarb (188.27)
propoxur (209.24) EPTC (189.32)
ethylene thiourea(102.15)
AldicarbThe rate constant for the reaction of aldicarb with hydroxyl
radical generated by Fenton’s process was determined as 8.1 ×109 M−1·s−1 using the competition kinetics method (Haag andYao 1992). Huston and Pignatello (1999) reported the completeconversion of 38 mg·L−1 of aldicarb and 62% reduction in TOCusing the photo Fenton process within 120 min (see Alachlor, ananiline-based pesticide, for the reaction conditions). They alsoobserved the formation of nitrate and sulfate after the photoFenton treatment.
AsulamCatastini et al. (2002) evaluated catalytic photodegradation
of asulam with photoreduction of iron (III) aquacomplexes([Fe(OH)(H2O)5]+2) to ferrous ions and hydroxyl radicals inthe presence of molecular oxygen at pH 3.0–3.4. Ferrous ionsare oxidized back to ferric ions through various pathways suchas photooxidation and oxidation by hydrogen peroxide gener-ated within the system, where another hydroxyl radical forms.Complete conversion and nearly complete TOC reduction of23 mg·L−1 of asulam was achieved with 16.7 mg·L−1 of Fe3+by irradiation at 365 nm and by solar irradiation, within 17 h
Fig. 5. Proposed degradation pathway of bendiocarb bymembrane anodic Fenton treatment (Wang and Lemley 2003b).
O
O
ONH
O
O
O
OH
O
O
O
OH
bendiocarb
and 28–30 h, respectively (Catastini et al. 2002). Degradationby-products or intermediates were not identified.
BendiocarbAaron and Oturan (2001) evaluated electro Fenton, H2O2/UV,
photo Fenton, and direct photolysis for the 112–188 mg·L−1
of bendiocarb degradation. Unlike the system used by Pratapand Lemley (1994) (see Alachlor, an aniline-based compound),H2O2 was generated electrochemically from the dissolved oxy-gen in Aaron and Oturan’s system, and ferric chloride (FeCl3)was introduced to supply iron (55.8 mg·L−1 as Fe3+). The con-version of this insecticide was apparently much faster in theH2O2/UV and photo Fenton processes (λ = 254 nm,68 mg·L−1 of H2O2, 55.8 mg·L−1 of Fe3+) than in the otherprocesses.Aaron and Oturan also proposed a degradation mech-anism of bendiocarb; however, the identification of intermedi-ates or by-products was not actually performed.
Wang and Lemley (2003b) studied the competitive degrada-tion by membrane anodic Fenton treatment (continuous Fe2+generation and H2O2 addition with a 1:10 molar ratio, at 25 ◦C)of six carbamate insecticides, including bendiocarb, carbaryl,carbofuran, dioxacarb, fenobucarb, and promecarb. The mem-brane anodic Fenton treatment is a variety of anodic Fenton pro-cess (see also Trifluralin, an aniline-based compound) that usesan ion exchange membrane to separate the cathode and anodehalf-cells (Wang and Lemley 2002a). The rate constant for ben-diocarb with hydroxyl radicals was determined as8.9 × 109 M−1·s−1 using a competitive degradation method.The order of reactivity of the six carbamates tested was as fol-lows: dioxacarb and carbaryl > fenobucarb > promecarb >
bendiocarb > carbofuran. The degradation products were iden-tified by gas chromatography – mass spectrometry (GC–MS)analysis, and the degradation pathway was proposed (Fig. 5).The toxicity of a carbamate mixture to an earthworm (Eiseniafoetida) was also reduced after the membrane anodic Fentontreatment (Wang and Lemley 2003b).
Carbaryl (NAC)Sun and Pignatello (1993a) demonstrated the complete con-
version within 10 min of 20 mg·L−1 of carbaryl with Fe3+-chelate/H2O2 (dark) processes using either picolinic, gallic orrhodizonic acid as a chelating agent (see Fig. 3 and Trifluralin,an aniline-based compound, for the reaction conditions). In thepresence of the chelating agent, the conversion of carbaryl bythe Fenton type reactions was highly enhanced.
Carbaryl degradation was also studied using the membraneanodic Fenton treatment (Wang and Lemley 2002a, 2003b).
Fig. 6. Proposed degradation pathways of carbaryl by membraneanodic Fenton treatment (Wang and Lemley 2002a).
O
HN
O
O
HO
O
OH O
O
O
NH
Ocarbaryl
Complete conversion of 20.1 mg·L−1 of carbaryl was achievedwithin 4 min of the electrolytic Fenton treatment (see alsoBendiocarb). As was the case with bendiocarb, the rate con-stant and activation energy for the reaction of carbaryl and hy-droxyl radicals were determined using the membrane anodicFenton process. A kinetic model to predict the degree of car-baryl conversion by the process was presented (Wang and Lem-ley 2002a). The durability of the ion exchange membrane forrepeated use (100 times) was also demonstrated. The degra-dation products and possible degradation pathway were alsodetermined (Fig. 6).
CarbofuranNearly complete conversion of 22–100 mg·L−1 of
carbofuran by H2O2/UV AOP was reported (Scheunert et al.1993; Benitez et al. 1995b, 2002). Benitez et al. (1995b) alsoshowed that elevated temperature (up to 40 ◦C) improved thecarbofuran conversion by the H2O2/UV process. The tempera-ture dependent rate constants and quantum yields for the conver-sion of carbofuran by the H2O2/UV process were determined(Benitez et al. 1995b). Their kinetic analysis revealed the ma-jor contribution of hydroxyl radical reactions to the carbofuranconversion during the H2O2/UV treatment over the direct pho-tolysis (Benitez et al. 1995b, 2002).
Various Fenton-type processes were also evaluated forcarbofuran degradation. Benitez et al. (2002) reported an im-provement in performance over the classical Fenton processwith the photo Fenton process using polychromatic UV irra-diation on carbofuran conversion. Benitez et al. (2002) alsodetermined the rate constant for the hydroxyl radical reactionas 4.0 × 109 M−1·s−1 at pH 3. Huston and Pignatello (1999)reported complete conversion of 53 mg·L−1 of carbofuran andmore than 90% TOC reduction in the solution by photo Fentontreatment within 120 min (see Alachlor, an aniline-based com-pound, for the reaction conditions). Oxalate was detected as anorganic ionic species after the treatment (Huston and Pignatello1999). The membrane anodic Fenton process (see also Ben-diocarb) was also employed for the degradation of carbofuran(Wang and Lemley 2003a, 2003b). Nearly 80% COD reductionwas achieved by the treatment of 44 mg·L−1 of carbofuran solu-tion for 14 min (Wang and Lemley 2003a). An improvement inbiodegradability and significant earthworm toxicity reductionwas also observed after the membrane anodic Fenton treatment
Fig. 7. Proposed degradation pathway of carbofuran by membraneanodic Fenton treatment (Wang and Lemley 2003a).
O
O
O
H
O
O
O
NH
O
OH
O
OH
HO O
OH
O
carbofuran
Fig. 8. Proposed degradation pathways of dioxacarb, fenobucarb,and promecarb by membrane anodic Fenton treatment (Wang andLemley 2003b).
O O
O NH
O
O O
OHOH
OO
OH
OH
HO
O
HN
O
O
H
O
O
HN
O
O
H
O
O
HO
O
O
H
O
O
O
HN
OOH
HO
OH
dioxacarb
fenobucarb
promecarb
HO HO
O
(Wang and Lemley 2003a, 2003b). Degradation products weredetermined, and possible degradation pathways were also pro-posed as shown in Fig. 7 (Wang and Lemley 2003a).
Dioxacarb, fenobucarb (BPMC), and promecarb
Wang and Lemley (2003b) reported the application of mem-brane anodic Fenton treatment to the degradation of dioxacarb,fenobucarb, and promecarb, in addition to the three carbamateinsecticides already reviewed above. Rate constants for thehydroxyl radical reaction and activation energies as well asdegradation products were determined, and possible degrada-tion pathways were proposed for each compound (Fig. 8).
FormetanateA series of investigations have been carried out on the degra-
dation of formetanate by photo-catalytic reactions driven byUV radiation, as well as by solar radiation using pilot-scalecompound parabolic collectors (Blanco et al. 1999; Fallmannet al. 1999a, 1999b; Malato et al. 2002a, 2003b). Two typesof photocatalytic degradation were evaluated including photoFenton and TiO2/hν processes. A formetanate formulation inaqueous solution (initial TOC = 100 mg·L−1) was success-fully degraded and mineralized up to 90% by a laboratory-scale photo Fenton process (λ = 300–450 nm, 14 mg·L−1 ofFe2+, 680 mg·L−1 of H2O2, pH 2.8, 20–55 ◦C) (Fallmann etal. 1999b). Complete conversion of 50 mg·L−1 of formetanateand effective TOC reduction (>90%) was also demonstratedin pilot-scale solar reactors even at smaller iron concentrations(2.8 mg·L−1 of Fe2+) (Malato et al. 2002a). A synthetic waste-water containing 10 pesticides including formetanate (initialTOC = 100 mg·L−1) was also successfully treated by the photoFenton process on both laboratory- (Fallmann et al. 1999b) andpilot-scales (Fallmann et al. 1999a, 1999b). Although still ef-fective, the TiO2/hν process (200 mg·L−1 of TiO2, slurry) wasconsistently less efficient than the photo Fenton process (Blancoet al. 1999; Fallmann et al. 1999b). The kinetics of formetanatedegradation by the solar-photo Fenton process were also studied(Malato et al. 2002a, 2003b). For further details regarding thesolar reactor developments reviewed here, refer to the overviewarticles by Malato et al. (2002b) and Malato andAgüera (2004).
Methomyl (Lannate)Malato et al. (2002a, 2003b) evaluated the solar-driven photo
Fenton and TiO2/hν processes for methomyl degradation anddetoxification in water. As was the case with formetanate,the photo Fenton process was more efficient in degrading50 mg·L−1 of methomyl (monitored as TOC) than was theTiO2/hν process, although both processes were capable of min-eralizing more than 90% of this carbamate insecticide (Malatoet al. 2002a). For the pesticide solution treated by both pro-cesses, several toxicity bioassays were also performed withDaphnia magna, Selenastrum capricornotum, and Vibrio fis-cheri (Fernández-Alba et al. 2002; Malato et al. 2003b). Inthe case of photo Fenton treatment, the toxicity to microalgaeS. capricornotum initially increased and then later decreased.This implies the formation of a potentially more toxic degrada-tion intermediate(s) of methomyl (not identified). The results ofMalato et al. (2003b) also indicated that the TOC reduction didnot correlate with toxicity reduction, and emphasized the impor-tance of continuous monitoring of toxicity during wastewatertreatment, in addition to TOC reduction and parent compoundconversion. No degradation by-products and (or) intermediates,except for ammonium and sulfate ions, were determined in anyof these studies.
Oxamyl (Vydate) and propamocarbThe solar driven photo Fenton and TiO2 photocatalytic pro-
cesses described above (see Formetanate) were also appliedto the degradation of oxamyl and propamocarb (Blanco et al.
Fig. 9. Primary degradation products of ethylene thiourea byFenton treatment (Saltmiras and Lemley 2000).
NHHN
O
ethylene urea
NHN
SO3H
2-imidazolin-2-yl sulfonic acid
1999; Fallmann et al. 1999a, 1999b). These pesticides weretreated in the form of formulated products. Propamocarb, as itwas monitored as TOC, was one of the hardest pesticides todegrade by the photo Fenton process (Fallmann et al. 1999a).Oxamyl was also relatively recalcitrant (Fallmann et al. 1999a).
The rate constant for the reaction of oxamyl with hydroxylradicals generated by classical Fenton’s process was determinedas 2.0 × 109 M−1·s−1 (Haag and Yao 1992). Degradation by-products were not identified in any of these studies.
Propoxur (Baygon)A Fenton-type Fe3+-chelate/H2O2 (dark) treatment was eval-
uated for propoxur (Baygon) degradation (Sun and Pignatello1993a).All three chelating agents, including picolinic acid, gal-lic acid, and rhodizonic acid (Fig. 3; also see Trifluralin, ananiline-based compound, for the reaction conditions), catalyti-cally enhanced propoxur conversion. Gallic acid was the mosteffective, with complete conversion of 21 mg·L−1 of propoxurbeing achieved in less than 2 min. No degradation by-productsof propoxur were determined.
EPTCArnold et al. (1996) demonstrated nearly complete conver-
sion of 30 mg·L−1 of ETPC by the Fenton process (295 mg·L−1
of Fe2+, 170 mg·L−1 of H2O2, pH 2.5, 25 ◦C) in a pesticiderinse water containing four other pesticides, including atrazine,cyanazine, metolachlor, and alachlor. No degradation productswere determined.
Ethylene thioureaSaltmiras and Lemley (2000) investigated the degradation
of 20.4 mg·L−1 of ethylene thiourea using three types of Fen-ton process, namely anodic Fenton (described in Trifluralin, ananiline-based compound), electro Fenton, and classical Fenton.Although all of these Fenton processes were found effective inthe conversion of ethylene thiourea, the primary degradationproducts, including ethylene urea and 2-imidazolin-2-yl sul-fonic acid (Fig. 9), were more persistent than the parent com-pound. Among the processes evaluated, anodic Fenton treat-ment was found to be the most effective method of breakingdown these degradation products, leaving them completely de-graded within 10 min of treatment.
Summary of carbamatesIt appears that all of the carbamate pesticides are fairly re-
active toward Fenton-type processes, although H2O2/UV AOP,having only been evaluated for bendiocarb and carbofuran, has
Fig. 10. Chemical structure of chlorophenoxy compounds(formula weight is shown in brackets).
ClO
HO
OO
Cl
ClO
HO
Cl
Cl
OO
HO
Cl
Cl
Cl
OO
HOO
HO
O Cl
ClOO
HO
4-chlorophenoxyaceticacid (186.59)
2,4-D (221.04)2,4-DP (235.07)
2,4,5-T (255.48)MCPP (214.65)MCPA (200.62)
not yet been thoroughly evaluated with regards to the degrada-tion of this class of pesticide. Considerable mineralization hasbeen demonstrated in many cases, especially using the photoFenton process. Possible degradation pathways have been pro-posed for bendiocarb, carbaryl, carbofuran, dioxacarb, fenobu-carb, and promecarb using the membrane anodic Fenton pro-cess.Varieties of inorganic and organic degradation by-productsand (or) intermediates have also been determined for other car-bamates. Toxicity reduction in various aqueous carbamates hasbeen demonstrated using either the membrane anodic Fentonor the solar-driven photo Fenton process.
Chlorophenoxy compoundsThe chlorophenoxy herbicides reviewed here include
4-chlorophenoxyacetic acid (4-CPA), 2,4-D, 2,4-DP (dichlor-prop), 2,4,5-T, MCPA, and MCPP (mecprop). The chemicalstructures and formula weights of these compounds are shownin Fig. 10.
4-Chlorophenoxyacetic acid (4-CPA)Degradation of 4-chlorophenoxyacetic acid (4-CPA) was
studied using various electrochemical processes. Boye et al.(2002) evaluated anodic oxidation (with graphite cathodewithout H2O2 generation; 100 mA), anodic oxidation with elec-trogenerated H2O2, electro Fenton (anodic oxidation with elec-trogenerated H2O2 plus 55.8 mg·L−1 of ferrous ion),and photoelectro Fenton (electro Fenton with UV irradiation;λ = 360 nm) for the degradation of this herbicide (up to387 mg·L−1) at pH 3 at 35 ◦C. It should be noted that hy-drogen peroxide was generated electrochemically by bubblingoxygen gas at the cathode in an undivided cell having a plat-inum anode in this electro Fenton system (Boye et al. 2002).This introduces a significant difference from the system de-veloped by Lemley and coworkers (e.g., Wang and Lemley2003b) described in previous sections, where ferrous ions weregenerated through the electrolysis of an iron electrode in twoseparated half-cells (Pratap and Lemley 1994, 1998; Saltmirasand Lemley 2000). The photoelectro Fenton process was appar-ently superior to other electrochemical processes in degrading4-CPA. Nearly complete mineralization (monitored as TOC)was achieved in approximately 3 h. Degradation intermediates
Fig. 11. Complete mineralization of 4-chlorophenoxyacetic acid(4-CPA) by photoelectro Fenton process (Boye et al. 2002).
4-CPA
Cl
O
OHO
OHO
OH
Cl
OH
OH
O
H
OH
OH
Cl
O
OH
H
O
OHHO
O
OH
HO
O
O
H
HO
O
H
O
O
OH
OH
CO2
OO
O
O
HO
OH
Fe3+
-oxalatecomplexes
of 4-CPA were determined by GC–MS, and possible degra-dation pathways were proposed (Fig. 11). The mineralizationof carboxylic acid intermediates such as oxalate, which wereknown to persist in other electrochemical treatments, was alsodemonstrated using the photoelectro Fenton treatment (Boyeet al. 2002) as well as the improved versions of electro-Fentonand anodic oxidation processes using a boron-doped diamondelectrode (Brillas et al. 2004).
Brillas et al. (2003c) reported the improved versions of elec-tro Fenton and photoelectro Fenton processes for 4-CPA degra-dation, called peroxi-coagulation and photoperoxi-coagulationtreatments, respectively. The photoperoxi-coagulation processutilizes a sacrificial iron anode, instead of the platinum onefound in the former configurations, to supply ferrous ions. Thisis similar to the one developed by Lemley and coworkers; how-ever, in this case, both Fe2+ and H2O2 are produced elec-trochemically. As the name suggests, the peroxi-coagulationtreatment employs precipitation of Fe(OH)3, in addition to theFenton-type degradation, for the removal of organic pollutants.The contribution of each mechanism to 4-CPA removal var-ied depending on the reaction conditions. The TOC reduction(not mineralization) in 40–388 mg·L−1 of 4-CPA solution atpH 3 was greatly improved over the electro and photoelectroFenton processes, and was slightly faster with the photoperoxi-coagulation than with the peroxi-coagulation, due to the addi-tional effect of photochemical reactions in the former process.The same degradation intermediates appearing in Fig. 11 weredetected during the photoperoxi-coagulation treatments of 4-CPA. More than 85% dechlorination was observed (Brillas et al.2003b). Brillas et al. (2003b) also reported that four chlorophe-noxy herbicides, including 4-CPA, 2,4-D, 2,4,5-T, and MCPA,were degraded at comparable rates by the peroxi-coagulationprocess.
Fig. 12. Some degradation products of 2,4-D by H2O2/UV AOP(Scheuer et al. 1995; Alfano et al. 2001).
OH
Cl
Cl
OH
OH
Cl
HOOH
O
HO OH
O O
HO
OH
OHO
HOOH
O
O
2,4-dichlorophenol chlorohydroquinone
glycolic acid malonic acid
maleic acid
fumaric acid
of 2,4-D by H2O2/UV AOP was reported (Scheuer et al. 1995;Alfano et al. 2001; Kwan and Chu 2003), and the conversion wasgreatly enhanced from the direct UV photolysis alone (Alfanoet al. 2001; Kwan and Chu 2003). Scheuer et al. (1995) reportedthe formation of various anionic degradation products during2,4-D degradation by H2O2/UV AOP (λ = 185–254 nm, 50–200 mg·L−1 of H2O2, 41 mg·L−1 of 2,4-D), including organicacids such as acetate, glycolate, formate, malonate, maleate,oxalate, and fumarate (Fig. 12), as well as inorganic anions in-cluding chloride and nitrate.Almost quantitative dechlorinationwas observed. Detection of malonate, malenate, and fumarateimplies the destruction of the aromatic ring. Degradation ofthese organic acids was also accomplished by prolonged UVirradiation (Scheuer et al. 1995). Alfano et al. (2001) devel-oped a kinetic model for 2,4-D degradation by H2O2/UV AOP(λ = 254 nm, 0–221 mg·L−1 of H2O2, 30–90 mg·L−1 of2,4-D, 25 ◦C), which also accounted for the formation anddegradation of two aromatic intermediates, including2,4-dichlorophenol and chlorohydroquinone (Fig. 12). Benitezet al. (2004b) also studied the kinetics of 2,4-D degradationby direct photolysis, H2O2/UV, and O3/H2O2 processes. Thequantum yield and second order rate constant at 20 ◦C weredetermined as 8.1 × 10−3 mol·Eins−1 for direct photolysis and5.1 × 109 M−1·s−1 for hydroxyl radical reactions of 2,4-D,respectively.
Varieties of Fenton-type processes have been evaluated for2,4-D degradation. Slightly faster and higher conversion of2,4-D was achieved by the classic Fenton reaction than byUV photolysis (Pignatello 1992; Kwan and Chu 2003). 2,4-Dichlorophenol was detected as a degradation intermediate ofthe Fenton treatment. Pignatello (1992) demonstrated that al-though the Fenton-type Fe3+/H2O2 process (dark, oxygenated;55.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2, 22.1 mg·L−1 of2,4-D, pH 2.7) was slower than the classic Fenton, a higher de-gree of mineralization was achieved with the former process.Nearly quantitative (90–100%) dechlorination was observed inthe oxygenated 2,4-D solution treated by the Fe3+/H2O2 pro-cess (Pignatello 1992). Carbon dioxide evolution was also ob-served from the 14C labeled aromatic ring (up to 69% with100 mM [H2O2]0) and the labeled carboxy group (37% with10 mM [H2O2]0) of 2,4-D. The presence of inorganic anionssuch as sulfate and chloride inhibited the degradation of2,4-D by the Fe3+/H2O2 process, probably through complexa-
Fig. 13. Aromatic intermediates of 2,4-D by Fe3+-chelate/H2O2
treatment with or without UV (Sun and Pignatello 1993b).
Cl
O
O
H
Cl
Cl
O
Cl
Cl
Cl
O
Cl O
O
2,4-dichlorophenolformate
2,4-dichloro-1-(chloromethoxy) benzene
6,8-dichloro-2H-1,4-benzodioxan-3-one
tion and radical scavenging (Pignatello 1992). Chu et al. (2004b)recently developed a kinetic model to describe the degradationof 2,4-D by Fenton process based on a two-stage pattern of thepesticide decay.
Sun and Pignatello (1992, 1993a) reported that the conver-sion of 2,4-D by Fe3+/H2O2 (dark) was accelerated by the ad-dition of some types of chelating agents such as picolinic acid,gallic acid, and rhodizonic acids (Fig. 3) at pH 6, although com-plete mineralization was still not possible due to the formationof stable ferric complexes. Oxalic acid and citric acid, whichwere active under photo irradiation, had no effect on the conver-sion under dark conditions (Sun and Pignatello 1993a; Kwanand Chu 2003). In addition to 2,4-dichlorophenol and organicacids, several aromatic intermediates of 2,4-D treated by Fe3+-chelate/H2O2, with or without UV irradiation, were determinedas shown in Fig. 13 (Sun and Pignatello 1993b).
The photo-assisted Fenton process was studied in the absence(Chu et al. 2004a) and in presence of chelating agents such asoxalate, citrate, picolinate, gallate, and rhodizonate (Pignatello1992; Sun and Pignatello 1993a; Kwan and Chu 2003, 2004a,2004b, 2004c). As expected, the addition of a chelating agentand UV/Vis irradiation (λ = 290–700 nm) enhanced the degra-dation of 2,4-D by Fe2+ or Fe3+/H2O2 in water (Pignatello1992; Kwan and Chu 2004b, 2004c). Both conversion and min-eralization of 2,4-D was further accelerated and enhanced inthe presence of a chelating agent (Sun and Pignatello 1993a;Kwan and Chu 2003), although these complexes were less sta-ble under UV (λ = 300–400 nm) radiation (Sun and Pignatello1993a). Sun and Pignatello (1993c) also demonstrated the con-tribution of molecular oxygen to the mineralization of 2,4-D bythe photo Fe3+-chelate/H2O2 process. As was the case with thedark Fenton-like reactions, negative effects of phosphate andsulfate on the photo Fenton reactions were reported (Pignatello1992; Lee et al. 2003), although these may be overcome bythe addition of chelating agents that form stronger complexeswith ferric ions (Lee et al. 2003). Kwan and Chu (2004a) iden-tified several degradation intermediates of 2,4-D degradationby the oxalate-mediated photo Fenton process and proposeda degradation pathway similar to the one proposed for 4-CPAdegradation by Boye et al. (2002) using a photoelectro Fen-ton process (Fig. 11). According to the pathway, dechlorinationand hydroxylation occurs at the aromatic rings of 2,4-D fol-lowed by removal of the glycolic acid group and subsequentring opening. Paterlini and Nogueira (2005) recently investi-gated the optimization of 2,4-D degradation and mineralization
Fig. 14. Degradation of chlorophenoxy herbicides by electroFenton process (Oturan et al. 1999).
n(Cl)
O
R
O
OH
n(Cl)
OH
R
OH
O
OH
n(Cl)
OH ·OHn(Cl) (OH)m
·OH
·OH
+
aliphatic alcohols,esters,carboxylic acids
by oxalate-mediated photo Fenton process using the responsesurface methodology. Up to 93% of TOC from 0.1 mM 2,4-Dwas mineralized under the optimized reaction condition involv-ing 0.6 mM ferrioxalate, 8 mM H2O2, and 10 min of irradiationwith a 15 W black light lamp.
Three types of electrochemical methods were employed togenerate ferrous ions, hydrogen peroxide, or both, for Fentondegradation of 2,4-D. Oturan et al. (1999, 2001) investigated thedegradation of several chlorophenoxy herbicides by an electro-chemical Fenton process generating H2O2 and reducing Fe3+with continuous bubbling of O2 and electrolysis (60 mA). Fer-rous ammonium sulfate, (NH4)2·Fe(SO4)2·6H2O, was suppliedas a source of ferric ions (55.8 mg·L−1 of Fe3+). More than75% TOC reduction (220 mg·L−1 of 2,4-D) was achieved bythe electro Fenton treatment after 6 h (Oturan 2000). Varioushydroxylated aromatic intermediates of 2,4-D degradation weredetected, and possible degradation pathways were proposed asshown in Fig. 14 (Oturan et al. 1999; Oturan 2000). This degra-dation pathway generally agrees with the one proposed for 4-CPA degradation by photoelectro Fenton treatment (Fig. 11).Detailed reaction mechanisms of the hydroxylation and aro-matic ring opening reactions were also proposed (Oturan 2000).Brillas et al. (2004) recently reported the improved performanceof electro-Fenton process on 2,4-D degradation using a boron-doped diamond electrode. Oxalate generated upon the 2,4-Ddegradation could be degraded and mineralized in their im-proved system.
Wang and Lemely (2001) investigated the effects of temper-ature, H2O2:Fe2+ ratio, and initial herbicide concentration on2,4-D degradation by anodic Fenton process in which ferrousions were supplied by electrolysis of an iron anode (see Triflu-ralin, an aniline-based compound for the process description).Complete conversion of 11–88 mg·L−1 of 2,4-D was demon-strated with the anodic Fenton process within 6 min, and pos-itive effects of increasing temperature (up to 34 ◦C) were ob-served. The kinetic model for the 2,4-D degradation was devel-oped, and the activation energy was calculated as 26.1 kJ·mol−1
(Wang and Lemley 2001).Brillas et al. (2003b) also reported the 2,4-D degradation
by another variety of electrochemical Fenton treatment calledperoxi-coagulation, wherein both H2O2 and ferrous ions weresupplied electrochemically (see 4-Chlorophenoxyacetic acid).As was the case with 4-CPA, peroxi-coagulation was more effi-cient in removing dissolved organic carbon (DOC) derived from
Fig. 15. A degradation intermediate (2,4,5-trichlorophenol) of2,4,5-T by Fenton-type processes (Pignatello 1992; Oturan et al.1999).
Cl
Cl
Cl
HO
2,4-D than was the electro Fenton process (without continuousFe2+ generation).
2,4-DP (dichlorprop)The chlorophenoxy herbicide 2,4-DP is different from
2,4-D by one methyl carbon on the alkyl side chain (Fig. 10);therefore, its reactivity to ozonation and advanced oxidationis very similar to that of the latter compound. Degradationintermediates detected during the electro Fenton treatment of2,4-D and 2,4-DP were essentially the same, including2,4-dichlorophenol, 1,2-dihydroxy-4,6-dichlorobenzene, andaliphatic compounds (Oturan et al. 1999). No further study wasreported on the degradation of this herbicide with H2O2/UV orother Fenton-type AOPs.
2,4,5-TThe herbicide 2,4,5-T is 2,4,5-trichlorophenoxyacetic acid,
and is different from 2,4-D by one extra chlorine substitution onthe aromatic ring (Fig. 15). Several Fenton-type processes wereevaluated for the degradation of 2,4,5-T. Sun and Pignatello(1993a) reported the complete conversion and nearly 80% min-eralization (of the 14C labeled aromatic ring) of 25.5 mg·L−1 of2,4,5-T by the Fe3+-chelate/H2O2 treatment (see also 2,4-D).A by-product, 2,4,5-trichlorophenol (Fig. 15) was detected byGC–MS, although it was subsequently converted to other prod-ucts. Nearly quantitative dechlorination was observed in the2,4,5-T solution treated by the Fe3+/H2O2 (dark) process (Pig-natello 1992). Complete mineralization of 14C labeled aromaticring of 2,4,5-T was achieved by the photo-assisted Fe3+/H2O2process (Pignatello 1992).
Several varieties of electro Fenton-type processes were alsoevaluated. Oturan et al. (1999) reported the formation of 2,4,5-trichlorophenol and its hydroxylated products after electro Fen-ton treatment of aqueous 2,4,5-T. Boye et al. (2003b) com-pared the performance of various electrochemical treatmentprocesses, including anodic oxidation, anodic oxidation in thepresence of electro generated H2O2, electro Fenton, and pho-toelectro Fenton for 2,4,5-T degradation and TOC reduction(also see 4-Chlorophenoxyacetic acid). Of these processes, pho-toelectro Fenton treatment was the quickest and most efficient.Complete conversion of 200 mg·L−1 of 2,4,5-T and nearly com-plete mineralization (as TOC) were achieved by the photoelec-tro Fenton treatment. Various degradation intermediates weredetermined (Fig. 16), and their evolution and subsequent degra-dation was monitored during the treatment.Although all of thesearomatic compounds, and some of the organic acids, could bedegraded very quickly by either electro or photoelectro Fentonprocesses, oxalic acid was more persistent and could not be
Fig. 16. Degradation pathway of 2,4,5-T by photoelectro Fentonprocess (Boye et al. 2003b).
Cl
Cl
Cl
OH
OH
Cl
Cl
OH
Cl
Cl
Cl
O
O OH
HO
OHO
Cl
Cl
HO
OH
HO
OH
O
O
OH
O
O
OH
OH
H O
O OH
O
H OHO
O
OH
HO
CO2
O
O
HO
OH
O
OH
HO
O
2,4,5-T
Fe3+-oxalatecomplexes
degraded by the aforementioned process alone, requiring pho-tochemical reactions (Boye et al. 2003b) or a novel boron-dopedelectrode (Brillas et al. 2004) to complete the mineralization.
In addition, peroxi-coagulation treatment was evaluated forthe degradation and removal of 2,4,5-T in water (Boye et al.2003b; Brillas et al. 2003b). As was the case with the otherchlorophenoxy herbicides reviewed above, continuous genera-tion of H2O2 and Fe2+ by electrolysis in this system enhancedthe removal of aqueous 2,4,5-T (270 mg·L−1) through Fe(OH)3precipitation. Ultraviolet irradiation (λ = 300–400 nm;photoperoxi-coagulation) further improved the TOC reductionand led to the degradation of ferrioxalate complex (Boye et al.2003b). A degradation pathway essentially the same as the oneproposed for the photoelectro Fenton process (Fig. 16) was alsopresented for the photo-peroxi-coagulation treatment of 2,4,5-T(Boye et al. 2003b).
MCPAMCPA (4-chloro-2-methylphenoxyacetic acid) is structurally
different from 4-chlorophenoxyacetic acid by one methyl groupon the aromatic ring. Benitez et al. (2004b) studied the kinet-ics of MCPA degradation by direct photolysis, H2O2/UV, andO3/H2O2 processes. The quantum yield and second order rateconstant at 20 ◦C were determined as 0.15 mol·Eins−1 in the pHrange from 5 to 9 for direct photolysis and 5.1 × 109 M−1·s−1
for hydroxyl radical reactions of MCPA, respectively. Based ontheir kinetic study, a significant contribution of direct photoly-sis in the MCPA degradation during the H2O2/UV process wassuggested (Benitez et al. 2004b).
As with other chlorophenoxy herbicides, such as4-chlorophenoxyacetic acid, 2,4-D, and 2,4,5-T, a series ofstudies were reported on MCPA degradation by photo/electro-chemical Fenton-type processes (Boye et al. 2003a; Brillas et al.2003b, 2003d, 2004), and comparable results were obtained forthis compound as well. The primary degradation intermediate
Fig. 17. 4-Chloro-o-cresol, the primary intermediate of MCPAdegradation by electro Fenton-type processes (Boye et al. 2003a;Brillas et al. 2003d).
OH
Cl
identified was 4-chloro-o-cresol (Fig. 17). A reaction scheme,very similar to that of the 4-chlorophenoxyacetic acid shownin Fig. 11, was also proposed for the degradation of MPCA byphotoelectro Fenton and photo peroxi-coagulation (Boye et al.2003a; Brillas et al. 2003d).
MCPP (mecoprop)Complete conversion of MCPP by the electro Fenton process
(65–70 mA, 112 mg·L−1 of Fe2+, electrogeneration of H2O2)was reported (Oturan et al. 1999). The major degradation in-termediate, 4-chloro-o-cresol (see Fig. 17), and its degradationproducts were also detected.
Summary of chlorophenoxy compoundsMuch effort has been invested in the development and im-
provement of degradation procedures for chlorophenoxy herbi-cides, such as 2,4-D, using various Fenton-type and H2O2/UVAOPs. Complete conversion and a major degree of mineral-ization, using photo Fenton and electrochemical Fenton pro-cesses, have been reported for all of the chlorophenoxy her-bicides reviewed here. The degradation intermediates and (or)by-products of chlorophenoxy herbicides have been well deter-mined, and degradation pathways, which are analogous to eachother, have also been proposed in most cases. However, no studyhas been reported on the evolution of toxicity during treatment,although fairly toxic chlorophenols have often been detectedas the major degradation intermediates. Also, biodegradabilityis likely improved by the AOPs, but this is not clearly demon-strated in the studies reviewed here.
Organochlorine compoundsThe organochlorine pesticides reviewed here include chloro-
thalonil, chlordane, dalapon, DDT, dicamba, endrin, endosul-fan, hexachlorocyclopentadiene, lindane, methoxychlor(DMDT), pentachlorophenol, and toxaphene. The chemicalstructures and formula weights of these compounds are shownin Fig. 18.
Chlorothalonil
Park et al. (2002) evaluated the degradation of chlorothalonilby dark and photo Fenton processes. They found that ferricnitrate was more effective than chloride and sulfate salts inconverting this organochlorine fungicide by the Fenton-likeprocess (Fe3+/H2O2), and that the conversion of chlorothalonilwas enhanced by UV irradiation (wavelength, intensity un-known; 55.8–111.6 mg·L−1 of Fe3+, 3.4 g·L−1 of H2O2,2 mg·L−1 of chlorothalonil). More dechlorination (30–61%)
Fig. 18. Chemical structure of organochlorines (formula weight isshown in brackets).
Cl
Cl
Cl
O O
N
Cl
NCl
Cl
ClCl
Cl Cl
Cl
Cl
Cl
ClCl
OH
Cl
Cl
Cl
Cl
Cl
O OH
O
Cl
Cl
Cl
Cl
HO
O
ClCl Cl
ClCl
DDT (354.49)
Cl
ClCl
Cl
ClCl
O
O
S
O
Cl
Cl
Cl
Cl
ClCl O
Cl
Cl
Cl
ClCl
ClCl Cl
ClCl
Cl Cl
n(Cl)
methoxychlor (345.65)
chlorothalonil (265.91)
chlordane (409.78)
pentachlorophenol(266.34)
dicamba (221.04)
dalapon (142.97)
endrin (380.91)
endosulfan (406.92)
lindane (290.83)hexachlorocyclopentadiene
(272.77)
toxaphene
Fig. 19. Degradation products and degradation pathway ofchlorothalonil by Fenton-like Fe3+/H2O2 process (Park et al.2002). Further dechlorination of trichloro-m-phthalodinitrile alsooccurred (not shown).
N
Cl
NCl
Cl
Cl
N
N
Cl3
HO
Cl3
N
O
OHCl3
N
O
OH
HO
Cl3
N
Cl3
O
OH
NH2O
chlorothalonil
was achieved with photo Fenton treatment than with dark treat-ment (13–35%). Several degradation products of chlorothalonilwere identified, and a possible degradation pathway was pro-posed (Fig. 19).
ChlordaneA low reactivity of chlordane toward hydroxyl radicals gen-
erated by the photo Fenton process was reported (Haag andYao 1992). The reaction rate constant was smaller than thosereported for heptachlor and chlordene, which are the impurities
Fig. 20. Chemical structures of heptachlor and chlordene.
Cl Cl
Cl
Cl
Cl
ClCl
Cl
Cl
Cl
Cl
ClCl
heptachlor chlordene
of chlordane insecticide bearing another orefin group in theirmolecular structures (Fig. 20). No reaction by-products weredetermined. It should be noted that the use of chlordane is alsobanned in many countries.
DalaponA slow reaction of the anionic form of dalapon (pKa = 2.06)
with hydroxyl radicals generated by the photo Fenton processat pH 3.4 was reported (Haag and Yao 1992). No further studyhas been reported on the degradation of this organochlorineherbicide with Fenton-type or H2O2/UV AOPs.
DDTDDT is a well-known organochlorine insecticide that per-
sists in the environment, the use of which is banned in manycountries. Barbusinski and Filipek (2001) reported the Fentontreatment of industrial wastewater containing 47 µg·L−1 ofDDT and other pesticides including lindane, DMDT (methoxy-chlor), fenitrothion, and chlorfenvinphos. Nearly complete con-version of DDT and other pesticides was achieved by the Fen-ton treatment with 2.74 g·L−1 of Fe2+ and 5 g·L−1 of H2O2 atpH 3.0–3.5. Toxicity of wastewater to Vibrio fischeri was alsoreduced completely after the treatment. Degradation productsof DDT were not identified in this study.
DicambaVarious Fenton-type processes were evaluated for the degra-
dation of dicamba in water. Huston and Pignatello (1999) re-ported complete conversion of 48 mg·L−1 of dicamba and 90%TOC reduction by the photo Fenton process at pH 2.8 and 25 ◦Cwithin 120 min (λ = 300–400 nm, 1.2×1919 photons·L−1·s−1,2.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2). Nearly quantitativedechlorination was observed. Organic acids including Formateand oxalate were also detected as degradation by-products ofdicamba associated with use of the photo Fenton process.
As was the case with chlorophenoxy herbicides such as2,4-D, several electrochemical Fenton-type treatments wereevaluated for dicamba degradation, including electro Fenton,photoelectro Fenton, and peroxi-coagulation (Brillas et al.2003a, 2003b). Among them, photoelectro Fenton and peroxi-coagulation processes (115–230 mg·L−1 of dicamba, 55 mg·L−1
of Fe2+ for photoelectro Fenton, 100–450 mA, 4–6 h, pH 3,35 ◦C) were very effective and resulted in nearly completeTOC reduction, although the removal mechanisms of the twoprocesses were different. Whereas mineralization of this her-bicide occurred with photoelectro Fenton treatment (Brillas etal. 2003a), most of the TOC were coagulated and removed asFe(OH)3 precipitates by peroxi-coagulation treatment (Brillas
et al. 2003b). They also noted that the removal of this herbicide(as TOC) was much faster with the peroxi-coagulation treat-ment than with the photoelectro Fenton treatment (Brillas et al.2003b). Organic acids including oxalate, maleate, and formatewere detected and subsequently removed during the photoelec-tro Fenton treatment of dicamba (Brillas et al. 2003a).
EndrinEndrin is an organochlorine insecticide and (or) rodenti-
cide that is highly toxic and persists in the environment. Thus,its use is prohibited in many countries. The rate constant forthe reaction of endrin with hydroxyl radicals generated by ei-ther the Fenton or photo Fenton process, was determined as7.5 × 108 M−1·s−1 at pH 2.8–3.4 (Haag and Yao 1992). Nostudy was published identifying the degradation products ofendrin by Fenton or H2O2/UV treatment.
Endosulfan (α- and β-isomers)Endosulfan is another highly toxic and persistent organochlo-
rine insecticide. There are α- and β-isomers of this compoundthat are sometimes referred to as endosulfan-I and endosulfan-II, respectively. More than 80% TOC reduction (maximumTOC = 100 mg·L−1) in the pesticide solutions containing en-dosulfan α-β was demonstrated using photo Fenton (λ = 300–450 nm, 14 mg·L−1 of Fe2+, 680 mg·L−1 of H2O2, pH 2.8),although degradation of endosulfan itself was not clearly shown(Blanco et al. 1999; Fallmann et al. 1999a, 1999b). Degradationproducts were not determined in these studies.
HexachlorocyclopentadieneThe rate constant for the reaction of hexachlorocyclopenta-
diene with hydroxyl radicals generated by the Fenton processwas reported as 2.3 × 109 M−1·s−1 (Haag and Yao 1992). Nofurther study was reported on the degree of reaction or on thedegradation by-products of this organochlorine insecticide.
Lindane (γ -hexachlorocyclohexane)Lindane is one of the six isomers of hexachlorocyclohex-
ane (HCH) and is used as an insecticide or fungicide. Thisorganochlorine insecticide has also been used as a lotion orshampoo to control lice. Haag and Yao (1992) noted the slowreaction of lindane in water with hydroxyl radicals generated byFenton and photo Fenton processes. More recently, Barbusin-ski and Filipek (2001) demonstrated nearly complete conver-sion by Fenton treatment of lindane (54–62 µg·L−1) and twoother isomers of HCH in a wastewater also containing severalother pesticides (see also DDT). No degradation by-products oflindane were identified.
Methoxychlor (DMDT)Barbusinski and Filipek (2001) reported almost complete
conversion of 92 µg·L−1 of methoxychlor (DMDT) by the clas-sical Fenton process (2.74 g·L−1 of Fe2+ and 5 g·L−1 of H2O2,pH 3.0–3.5) in a pesticide wastewater also containing othercompounds such as DDT, lindane, fenitrothion, and chlorfen-vinphos. In contrast, Huston and Pignatello (1999) reported that
this organochlorine insecticide (2.2 mg·L−1) was somewhat re-sistant to degradation by the photo Fenton process (see Dicambafor the reaction conditions). No attempt has been made to iden-tify degradation by-products in these studies.
Pentachlorophenol (PCP)Trapido et al. (1997) reported fast conversion of 10.7 mg·L−1
of PCP by H2O2/UV AOP (1.36 g·L−1 of H2O2, λ = 254 nm,0.8 W·L−1) under both acidic (pH 2.5) and basic conditions(pH 9.5). Although the conversion was faster in acidic media,dechlorination under such conditions (28%) occurred consid-erably less often than in basic media (67%). Hirvonen et al.(2000) detected two isomers of tetrachlorobenzendiols. Theyalso reported the formation of a refractory dimeric product dur-ing the H2O2/UV treatment of PCP. They suggested that thisdimer was possibly derived from a tetrachlorophenoxy radicaland PCP.
Various Fenton-type processes were evaluated for the treat-ment of aqueous PCP. Lee and Carberry (1992) evaluated theFenton process with a view to enhancing the biodegradabilityof pentachlorophenol (PCP), an organochlorine fungicide andwood preservative highly resistant to microbial degradation. Avery high concentration of PCP (266 mg·L−1) was treated bythe Fenton process with 34 mg·L−1 of H2O2 and 50 mg·L−1 ofFe2+ for 4 h. Although the reduction in PCP concentration wasvery minor (reduced to 258 mg·L−1), biodegradation of the PCPsolution was substantially enhanced with Fenton pretreatment.
Engwall et al. (1999) reported the successful detoxification ofsynthetic wastewater saturated with two types of wood preser-vatives, including PCP and creosote, and containing a variety oforganic compounds such as polyaromatic hydrocarbons, pheno-lics, and heterocyclics. More than 80% reduction in TOC (ini-tially 46.6 mg·L−1) was achieved by the photo Fenton treatment(55.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2, λ = 300–400 nm,1.4 × 10−3 mol/L photons·min−1, pH 2.75, 3 h). Nearly quan-titative dechlorination of 8 mg·L−1 of PCP was also observed.The acute toxicity to fathead minnows (Pimephales promelas)and Daphnia pulex was reduced or eliminated after the photoFenton treatment.
Fukushima and Tatsumi (2001) studied the degradation ofPCP by the photo Fenton process (5.6 mg·L−1 of Fe3+,34 mg·L−1 of H2O2, λ > 370 nm, pH 5.0) in the presenceof humic acid. The degradation (or removal) of 50 µg·L−1 ofPCP was apparently enhanced in the presence of 50 mg·L−1 ofhumic acid. Fukushima and Tatsumi suggested that humic acidacted as a radical coupling partner for the phenoxy radical gen-erated by the Fenton process (Fig. 21). Addition of humic acidalso suppressed the formation of octachlorodibenzo-p-dioxinduring the treatment. In addition, tetrachloro-hydroquinone andtetrachlorocatechol were also detected as degradation products.
Oturan et al. (2001) applied an electro Fenton process, inwhich H2O2 was generated in an electrochemical reaction ofoxygen, to the degradation of aqueous PCP. Complete conver-sion of 8–26.6 mg·L−1 of PCP and more than 80% TOC reduc-tion were achieved by the electro Fenton treatment (−0.5 V vs.
Fig. 21. Proposed degradation pathway of pentachlorophenol(PCP) by the photo Fenton process in the presence and in theabsence of humic acid (Fukushima and Tatsumi 2001).
O
Cl
Cl
Cl
Cl
Cl
OHOH
Cl
Cl
Cl
Cl
Cl
O
ClCl
Cl
Cl Cl
O
Cl
Cl
Cl
Cl
PCP
OH
O
Cl
Cl
OH
Cl
Cl
O
O
Cl
Cl
Cl
Cl
Cl
Cl
Cl
Cl
O
Cl
Cl
Cl
Cl
HO
H+
OH OHOH
Cl
Cl
OH
Cl
Cl
HOOH
Cl
Cl
Cl
Cl
HO
HO
n(Cl)
HO
OH
OHO
O
tetrachloro-catechol
tetrachloro-hydroquinone
octachlorodibenzo-p-dioxin
withouthumic acid
phenolic groupin humic acid
further degradationcoupling products
withhumic acid
SCE, electrical charge: 1500 C) at pH 3. In addition, quantitativedechlorination was also observed. Recently, the solar-drivenphoto-Fenton (Hincapié et al. 2005) and photo-Fenton/O3 pro-cesses (Farré et al. 2005) were also successfully evaluated forthe mineralization and detoxification of 50 mg·L−1 of PCP(see Alachlor, an aniline-based pesticide, for the reaction con-ditions).
In addition to the classical Fenton, photo Fenton, and elec-tro Fenton processes, a number of Fenton-like systems werereported to be effective in degrading aqueous PCP, includingheme/H2O2 (Chen et al. 1999), porphyrin-derivative/ascorbicacid (Fukushima et al. 2002), and a microbial-driven Fentonprocess utilizing an Fe(III) reducing facultative anaerobe She-wanella putrefaciens strain 200 (McKinzi and Dichristina 1999).However, these systems are somewhat departed from the con-ventional Fenton process, and thus are not covered in this re-view.
ToxapheneToxaphene is a mixture of polychlorinated camphene (see
Fig. 18) and is highly resistant to biodegradation. The use ofthis organochlorine insecticide is banned in some countries, ofwhich the United States is one. As was the case with chlor-dane, a range of rate constants were reported for the reaction
of toxaphene and hydroxyl radicals generated by the Fentonprocess (Haag and Yao 1992). This implies that the reactivitytoward the hydroxyl radical reaction is substantially differentamong the components in the toxaphene mixture. No furtherstudy was published on the degradation of this compound inwater.
Summary of organochlorinesThe organochlorine pesticides reviewed here have varying
reactivity toward AOPs, and contradictory results have beenreported in some cases. Heavily chlorinated compounds suchas lindane, hexachlorocyclopentadiene, and endrin react veryslowly with hydroxyl radicals generated by the Fenton process.With the exceptions of PCP and dicamba, it would seem that lessattention has been paid to the advanced oxidation of organochlo-rines as compared with other major pesticide classes such aschlorophenoxy compounds, organophosphates, and triazines.This is probably due to the fact that some of these organochlo-rines are banned in many countries because of their high toxicityand high environmental impacts.
Organophosphate compoundsThe organophosphate pesticides reviewed here include
acephate, azinphos-methyl, chlorfenvinphos, chlorpyrifos, di-azinon, dichlorvos, disulfoton, edifenphos, EPN, fenitrothion,glyphosate, malathion, methamidofos (methamidophos),methyl-parathion, parathion, and phorate. The chemical struc-tures and formula weights of these compounds are shown inFig. 22.
Acephate, chlorpyrifos, edifenphosYu (2002) reported the treatment of concentrated solutions
(1 g·L−1) of several organophosphate insecticides includingacephate, edifenphos, chlorpyrifos, and methamidophos with aFenton process at pH 2.8 (concentration of Fe2+ and H2O2 notshown). After the Fenton treatment, COD in the insecticide so-lution was reduced by more than 95%. No degradation productswere determined.
Azinphos-methylHuston and Pignatello (1999) demonstrated the complete
conversion (within 30 min) and 56% removal as TOC (within2 h) of 25 mg·L−1 of azinphos-methyl by photo Fentontreatment (λ = 300–400 nm, 1.2 × 1919 photons·L−1·s−1,2.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2). Inorganic ions in-cluding nitrate, sulfate, and phosphate, as well as an organic acid(formate), were detected after the photo Fenton treatment. Phos-phate was found to slightly inhibit the degradation of azinophos-methyl.
Chlorfenvinphos and fenitrothionChlorfenvinfos degradation was recently evaluated using the
photo-Fenton/O3 (Farré et al. 2005) and solar-driven photo-Fenton processes (Hincapié et al. 2005). Nearly 80% TOC re-duction in 50 mg·L−1 of chlorfenvinfos solution was achieved
Fig. 22. Chemical structure of organophosphates (formula weightis shown in brackets).
O
NH
P OS
O NN
N
O
SP O
O
S Cl
Cl
ClO
P OO
O
N
Cl
Cl
Cl
OP
S
OO
N
N
OP
S
OO
Cl
Cl
OP
OO
O
SS
P OO
S
S
PO
O
S
N
O
O
O P
S
O
N
O
O
OP
S
OO
HOP
O
NH
OH
O
OH
O
O O
O
S
P
OO
S
O
PH2N O
S
N
O
O
OP
S
OON
O
O
OP
S
OO
S SP
S
OO
acephate (183.16)azinphos-methyl
(317.32) chlorfenvinfos (359.57)
chlorpyrifos (350.58)
diazinon (304.34)
dichlorvos (220.98)
disulfoton (274.39)
edifenphos (310.37)EPN (323.30)
fenitrothion (277.23)
glyphosate (169.07)
malathion (330.35)
methamidofos(141.12)
methyl-parathion(263.20)
parathion (291.26) phorate (260.36)
in both processes (see Alachlor, an aniline-based pesticide, forthe reaction conditions).
Derbalah et al. (2004) investigated the degradation and min-eralization of 0.5 mg·L−1 of fenitrothion in three photo-catalyticprocesses including photo-Fenton, Fe3+/UV, and H2O2/UV pro-cesses using a 300 W Xenon lamp with a glass filter (λ <
300 nm). Among the processes evaluated, photo-Fenton pro-cess was found the most efficient, and up to 93% mineralization(as DOC) of 0.5 mg·L−1 of fenitrothion was achieved within10 h of the photo-Fenton treatment with 50 mg·L−1 of ferricchloride and 0.7 mg·L−1 of H2O2.
Barbusinski and Filipek (2001) reported the complete conver-sion of 30–313 µg·L−1 of chlorfenvinphos and 44–377 µg·L−1
of fenitrothion by the Fenton process (2.74 g·L−1 of Fe2+,5 g·L−1 of H2O2, pH 3.0–3.5) in industrial wastewater also con-taining other pesticides including DDT, lindane, and methoxy-chlor. The toxicity of the wastewater to Vibrio fischeri was alsoreduced after treatment (Barbusinski and Filipek 2001). Sim-ilar pesticide wastewater was also treated by H2O2/UV AOP(Kowalska et al. 2004). Nearly all pesticides (0.87 mg·L−1 of
Fig. 23. Chemical structure of diazoxon (Wang and Lemley2002b).
N
N
OP
O
OO
chlorfenvinphos, 0.25 mg·L−1 of fenitrothion, 65 µg·L−1 of2,4-D, and other trace pesticides) were destroyed after 40 minof irradiation to the wastewater with a 150 W medium-pressuremercury vapor lamp in the presence of 112 mg·L−1 of H2O2.
DiazinonVarious Fenton-type AOPs, as well as H2O2/UV, were evalu-
ated for the degradation of diazinon. Doong and Chang (1998)reported the complete conversion of 10 mg·L−1 of diazinonwithin 3 h of irradiation (λ = 253–578 nm, 230 µW·cm−1
at a distance of 50 cm) by H2O2/UV, photo Fenton, andFe0/H2O2/UV processes at pH 7.0 with 10 mM phosphate bufferat 25 ◦C. The concentrations of chemicals were as follows:20 mg·L−1 of H2O2, 2.8 mg·L−1 of Fe2+, and (or) 1 g·L−1 ofFe0. Although the dark Fenton process was also tested undercomparable conditions, it was found to be much less effective.
Wang and Lemley (2002b) reported aqueous diazinon degra-dation by the anodic Fenton process wherein ferrous ion wasgenerated electrochemically from a sacrificial iron anode. Thediazinon (30 mg·L−1) was completely converted within 5 minby anodic Fenton treatment in water at 25 ◦C (continuous de-livery of H2O2, molar ratio of H2O2:Fe2+ = 10.1). Diazoxon(Fig. 23) was detected as an intermediate of diazinon oxida-tion after 0.5 min of treatment and rapidly disappeared within5 min. No other toxic intermediates were detected. A kineticmodel for the degradation of diazinon by the anodic Fenton pro-cess was presented, and the activation energy was calculated as12.6 kJ·mol−1. Wang and Lemley also reported that the con-version of this organophosphate insecticide was slower whentreated as a formulated insecticide than when treated as a purechemical. A reduction in COD in the diazinon solution was alsoobserved.
DichlorvosLu et al. (1997, 1999) reported the rapid degradation of
25–100 mg·L−1 of dichlorvos by Fenton process (14 mg·L−1
of Fe2+, 170 mg·L−1 of H2O2, pH 3). Nearly quantitativedechlorination was observed during the treatment (Lu et al.1997). They also found that the inhibition of the Fenton reac-tion by phosphate anions was likely due to the formation ofthe phosphate-Fe3+ complex (Lu et al. 1997). The relation-ship between the pseudo first order rate constants for dichlor-vos conversion and the concentrations of H2O2 and Fe2+ wasalso established (Lu et al. 1999). In addition to the Fenton pro-cess, complete conversion of 29.5 mg·L−1 of dichlorvos byH2O2/UVAOP (λ = 240–300 nm, 45.4–54.5 mg·L−1 of H2O2,pH 7) was also reported (Nitoi et al. 2001). No degradation by-products were determined in these studies.
DisulfotonHuston and Pignatello (1999) reported the degradation of
disulfoton by photo Fenton process (see Azinphos-methyl forthe reaction conditions). Although this organophosphate in-secticide was rapidly converted during the photo Fenton treat-ment (complete conversion of 15.5 mg·L−1 of disulfoton within30 min), only minor TOC reduction was achieved, even af-ter prolonged treatment (16% after 120 min). A quantitativeamount of phosphate and 84% of sulfate were recovered after120 min of photo Fenton treatment. Substantial amounts of or-ganic acids, including formate and acetate, were also detected.
EPNAs was the case with diazinon, three types of UV-assisted
AOPs were evaluated for the degradation of EPN in water(Doong and Chang 1998). Direct photolysis removed nearly80% of 10 mg·L−1 of EPN within 6 h. Whereas EPN wassomewhat resistant to H2O2/UV treatment, photo Fenton-typeprocesses (Fe2+/H2O2/UV and Fe0/H2O2/UV) were effectivein oxidizing this organophosphate insecticide. The dark Fen-ton process was less effective. No degradation products weredetermined in this study.
GlyphosateThe rate constant for the reaction of glyphosate and hy-
droxyl radicals generated by photo Fenton was reported as 1.8×108 M−1·s−1 at pH 3.8 (Haag and Yao 1992). Huston andPignatello (1999) reported 35% reduction in TOC after a 2-h photo Fenton treatment of 34 mg·L−1 of glyphosate solution(see Azinphos-methyl for the reaction conditions). A substantialyield (62%) of phosphate anion was confirmed after the treat-ment, although no organic acids, such as formate or acetate,were detected. No other degradation products were determined.
MalathionDowling and Lemley (1995) reported the complete con-
version of 145 mg·L−1 of malathion by Fenton and copper-amended Fenton processes in water (10 mg·L−1 of Fe2+,7.5 g·L−1 of H2O2). The addition of 18.4 mg·L−1 of Cu2+ ac-celerated the conversion of this organophosphate insecticide bya factor of three. Several degradation products of malathion, in-cluding malaoxon, were identified, and a degradation pathwaywas proposed as shown in Fig. 24. These degradation productspersisted during the Fenton treatment and could not be degradedcompletely (Dowling and Lemley 1995).
The complete conversion of 10 mg·L−1 of malathion was re-ported using three types of UV-assisted AOPs includingH2O2/UV, photo Fenton, and Fe0/H2O2/UV processes (see Di-azinon for the reaction conditions) (Doong and Chang 1998).Huston and Pignatello (1999) also employed the photo Fen-ton process for the degradation of 68 mg·L−1 of malathion(see Azinphos-methyl for the reaction conditions). Althoughthe photo Fenton treatment was effective in converting thisorganophosphate insecticide in water at pH 2.8, no reductionin TOC was observed. Quantitative sulfate, 35% of quantitative
Fig. 24. Proposed degradation pathway of malathion by Fentonprocess (Dowling and Lemley 1995). Some side reactions areomitted for clarity.
O
O O
O
S
P
OO
S
O
O O
O
S
P
OO
O
O
O
SH
O
O
O
PO
O
OH
O
O
O
O
O
O
O
O
OH
O
O
O
O
OH
malathion malaoxon
phosphate, and substantial amounts of formate, oxalate, and ac-etate were detected by ion chromatography after the treatment.
Roe and Lemley (1997) evaluated electro/photoelectro Fen-ton processes for malathion degradation, wherein ferrous ionwas produced by the electrolysis of an iron electrode. In thissystem, the Fenton reaction initiated as the electrolysis started.Although ferrous ion was generated by the electrolysis only fora certain time (2.5–10 min, 50–200 mg·L−1), the Fenton re-action continued to occur afterwards. Step addition of H2O2(75 mg·L−1) enhanced the conversion of this insecticide. Inthe optimized system, 30 mg·L−1 of malathion was completelyconverted, either in the presence or in the absence of UV ir-radiation (λ = 370 nm). The mineralization of malathion (upto 55%) by the electro Fenton process was also confirmed bymeasuring the 14C labeled compound in the solution.
Methamidofos (methamidophos)Several Fenton-type processes and the H2O2/UV AOP were
evaluated for the degradation of methamidofos. More than 85%of 400 mg·L−1 of methamidofos was converted by the classicalFenton process (10 mg·L−1 of Fe2+, 7.5 g·L−1 of H2O2) in thepresence of 18.4 mg·L−1 of Cu2+ (Dowling and Lemley 1995).As was achieved with malathion, diazinon, and EPN, completeconversion of 10 mg·L−1 of methamidofos was demonstratedusing the H2O2/UV, photo Fenton, and Fe0/H2O2/UV processesat pH 7 (Doong and Chang 1998). Fallmann et al. (1999a) re-ported very slow TOC removal from a commercially formu-lated solution of this insecticide (initial TOC = 100 mg·L−1;λ = 300–450 nm, 14 mg·L−1 of Fe2+, 680 mg·L−1 of H2O2,pH 2.8, 20–55 ◦C) by the photo Fenton process. No degradationby-products were determined in any of these studies.
Methyl-parathion (parathion-methyl)Dowling and Lemley (1995) reported the complete conver-
sion of 50 mg·L−1 of methyl-parathion by the classical Fen-ton process (see Malathion for the reaction condition). Methyl-paraoxon and p-nitrophenol were detected as degradation in-termediates (Fig. 25), and were subsequently degraded to more
Fig. 25. Proposed degradation pathway of methyl-parathion byFenton process (Dowling and Lemley 1995).
N
O
O
OP
S
OO
N
O
O
OP
O
OO
N
OH
O O
methyl-parathion methyl-paraoxon
p-nitrophenol
more polar products
polar organic compounds by the Fenton treatment. As was thecase with malathion and methamidofos, the addition of copperaccelerated the degradation of methyl-parathion and its degra-dation intermediates (Dowling and Lemley 1995).
Pignatello and Sun (1995) reported the nearly complete min-eralization of 26.3 mg·L−1 of methyl-parathion by photo Fen-ton treatment within 2 h (λ = 300–400 nm, 1 × 1018 photons·L−1·s−1, 55.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2, pH 2.8).During the degradation process the following inorganic and or-ganic compounds were detected: sulfate, nitrate, phosphate, p-nitrophenol, dimethylphosphate, and oxalate. The generationof methyl-paraoxon was also noted; however, its concentrationappeared to be very small (<1%).
Roe and Lemley (1997) reported the complete conversionof 12 mg·L−1 of methyl-parathion and up to 38% mineral-ization, as monitored by 14C in the solution, using the sameelectro/photoelectro Fenton process previously described (seeMalathion). It was also noted that UV irradiation (λ = 370 nm)had no impact on the degradation of methyl-parathion in theirsystem.
ParathionParathion is a highly toxic organophosphate insecticide. Its
use was recently banned in some countries, including the UnitedStates. Chen et al. (1998) compared the performance of di-rect photolysis, H2O2/UV, TiO2/UV, and TiO2/H2O2/UV pro-cesses for the degradation of parathion in a buffered solution(λ = 253–578 nm, 20 mg·L−1 of H2O2, 1 g·L−1 of TiO2,pH 7 phosphate buffer, 25 ◦C). Nearly complete conversion of10 mg·L−1 of parathion was achieved in all the processes tested;however, 15 times higher UV radiation input was required forthe photolysis and TiO2/UV systems. Chen et al. (1998) notedthat the TiO2/H2O2/UV process was the most efficient of theseprocesses. No degradation products were determined.
PhorateComplete conversion of 10 mg·L−1 of phorate was reported
with the H2O2/UV, photo Fenton, Fe0/H2O2/UV processes de-scribed previously (see Diazinon) (Doong and Chang 1998).No degradation products were determined.
Fig. 26. Chemical structure of pyridine and pyrimidine derivatives(formula weight is shown in brackets).
N
Cl
N
NH
N N
O
ON
N
Br-
Br-
N
NHN
N
Cl NH2
Cl
Cl
HO
O
imidacloprid (255.66)
diquat (dibromide; 344.05)
pyrimethanil (199.25)
picloram (241.46)
Summary of organophosphatesAll of the organophosphates reviewed here are amenable to
various Fenton-type and H2O2/UV AOPs. Considerable de-grees of mineralization have also been demonstrated for some ofthe organophosphates using the photo Fenton or electro Fentonprocesses, even though these organophosphates are resistant tomineralization in some cases, such as the case of malathiondegradation by the photo Fenton process. Inhibitory effectsof phosphate ion have also been suggested as an impact onthe degradation of organophosphates in Fenton-type processes,possibly through the complex formation with ferric ion.Very lit-tle kinetic information is available for organophosphate degra-dation. Degradation by-products and intermediates have beendetermined for diazinon, malathion, and methyl-parathion, anda common type of intermediate is the oxons, such as diazoxon(Fig. 23). These oxons are often more toxic than the parentorganophosphates and more resistant to degradation (Ohashi etal. 1994). The Fenton processes reviewed here are often capableof degrading these oxons to organic acids. No study has beenreported on the biodegradability or toxicity of organophosphatesolutions treated with Fenton-type or H2O2/UV AOPs, exceptin the cases of chlorfenvinphos and fenitrothion (toxicity re-duction).
Pyridine and pyrimidine derivativesThe pyridine and pyrimidine derivatives (including bipyridy-
lyium and chloro-nicotinyl compounds) reviewed here includediquat, imidacloprid, picloram, and pyrimethanil. The chemicalstructures and formula weights of these compounds are shownin Fig. 26.
Diquat dibromideA rate constant of 8.0 × 108 M−1·s−1 at pH 3.1 was reported
for the reaction of the diquat and hydroxyl radicals generatedby the Fenton process (Haag and Yao 1992). No further studywas reported on the degradation of this bipyridylium herbicidewith the Fenton or H2O2/UV AOP.
ImidaclopridMalato et al. (2001, 2002a) reported complete conversion
and 95% mineralization (as TOC) of 50 mg·L−1 of imidacloprid
Fig. 27. Proposed degradation pathway of imidacloprid bysolar-assisted photo Fenton and TiO2/hν processes (Malato et al.2001).
N
Cl
N
NH
N N
O
O
NH
O
N
Cl
NH2
O
N
Cl
H
O N
Cl
N
NH
O
N
Cl
OH
O
imidacloprid
C9HxClNyOz
by solar-assisted photo Fenton and TiO2/hν processes in pilot-scale solar reactors. They noted that the photo Fenton process(2.8 mg·L−1 of Fe2+, 510 mg·L−1 of H2O2, pH 2.7) was moreefficient than the TiO2/hν process (200 mg·L−1 of TiO2) in de-grading this chloro-nicotinyl insecticide. Residual degradationproducts in the treated solution exerted no toxicity on Daphniamagna (Malato et al. 2001). An aqueous solution of a commer-cial insecticide containing imidacloprid, as well as a solutionof pesticide mixture containing 10 types of pesticides includ-ing imidacloprid, were also successfully treated with the samesolar reactor system (initially containing 100 mg·L−1 TOC inboth cases) (Fallmann et al. 1999a, 1999b).
Various inorganic and organic ions, including chloride, ni-trate, ammonium, oxalate, formate, and acetate, were detectedduring the degradation of imidacloprid by the two solar-assistedprocesses in water (Malato et al. 2001). In addition, five ex-tractable degradation intermediates of imidacloprid were identi-fied (plus one compound unidentified), and their evolution mon-itored by liquid chromatography – mass spectrometry (LC–MS)over the course of the photo Fenton and TiO2/hν treatments. Adegradation pathway was also proposed as shown in Fig. 27(Malato et al. 2001).
PicloramVarious Fenton-type processes were evaluated for the degra-
dation of picloram. The reaction of picloram with hydroxylradicals produced by the Fenton process was reported to bevery fast (Haag and Yao 1992). The reaction rate constant wasindependent of pH, which ranged from 2.1 to 3.7 (pKa of pi-cloram carboxylic group = 3.6). This suggests that the reactionoccurs through addition to the ring, or through H abstractionfrom the amine; neither of these processes should be affectedby the ionization of the carboxylic group (Haag andYao 1992).
Sun and Pignatello (1993a) demonstrated rapid and competeconversion of 24 mg·L−1 of picloram by the Fe3+-chelate/H2O2process within 2 h (55 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2,1 mM chelator, pH 6, 25 ◦C). The chelating agents tested in-cluded picolinic acid, gallic acid, and rhodizonic acid (Fig. 3),the last of which was the most effective in catalyzing the degra-dation of picloram. This pyridine herbicide was also amenableto photo Fenton treatment (λ = 300–400 nm); the completeconversion of 50 mg·L−1 of herbicide and 90% mineralization
as TOC were achieved within 30 min and 2 h, respectively (Hus-ton and Pignatello 1999). Of all the pesticides tested, includingalachlor, aldicarb, atrazine, and malathion, picloram was oneof the most readily degradable using the photo Fenton process.Quantitative dechlorination, as well as the formation of nitrate,formate, oxalate, and acetate was also observed after the 2-hphoto Fenton treatment (Huston and Pignatello 1999).
Pratap and Lemley (1994) compared the performance of theelectrochemical Fenton (electro Fenton) process to that of theclassical Fenton process for the degradation of 30 mg·L−1 ofpicloram. This electro Fenton process utilizes ferrous ion gen-erated by the electrolysis of an iron electrode at a constant cur-rent of 0.3A in a reaction medium containing NaCl (electrolyte)and 7.5 g·L−1 of H2O2 at room temperature (22–27 ◦C) (Pratapand Lemley 1994). Both electro and classical Fenton processeswere effective in converting picloram, although the latter wasmore efficient in terms of the iron required (200 mg·L−1 forelectro Fenton vs. 50 mg·L−1 for classical Fenton). Pratap andLemley also noted that this herbicide could not be removed byelectrochemical treatment in the absence of H2O2.
Degradation products and intermediates of picloram, otherthan inorganic and organic ions, were not identified in any ofthe studies reviewed above.
PyrimethanilThe photo Fenton treatment of an aqueous solution of a com-
mercial pesticide containing pyrimethanil was reported (Fall-mann et al. 1999a). More than 80% TOC reduction (initial TOC= 100 mg·L−1) was achieved with the photo Fenton process(λ = 300–450 nm, 14 mg·L−1 of Fe2+, 680 mg·L−1 of H2O2,pH 2.8, 20–55 ◦C), although the degradation of pyrimethanilwas not clearly demonstrated. An aqueous solution of a mixtureof commercial pesticides, one of which contained this pyrimi-dine fungicide, was also treated with the photo Fenton, as wellas with the TiO2/hν process, in pilot-scale solar reactors (Blancoet al. 1999; Fallmann et al. 1999a, 1999b). No degradation by-products were determined in any of these studies.
Summary of pyridine and pyrimidine derivativesHigh degrees of mineralization have been demonstrated for
the pyridine and pyrimidine derivatives reviewed here, with theexception of diquat during photo Fenton-type processes. Somedegradation intermediates and (or) by-products have been iden-tified in the case of imidacloprid photo Fenton treatment, and areduction in toxicity has also been demonstrated. For the otherpyridine and pyrimidine pesticides, organic degradation inter-mediates and their toxicities are largely unknown. No reportshave been published on the application of the H2O2/UV processto the degradation of this class of pesticides.
TriazinesThe triazine pesticides reviewed here include ametryne,
atrazine, cyanazine, cyanuric acid, and simazine. Metribuzin, atriazinone herbicide, is also covered in this section. The chem-ical structures and formula weights of these compounds areshown in Fig. 28.
Fig. 28. Chemical structure of triazine and triazinone pesticides(formula weight is shown in brackets).
N
N
N
Cl
NH
NH
N
N
N
NH
S NH
N
N N
HN
Cl
NH
N
N
N
N
OH
HO OH N
N
N
HN
Cl
HN
N
N
N
S
NH2
O
atrazine (215.69)ametryne (227.37)
cyanazine (240.69)
cyanuric acid (129.08)
simazine (201.66)metribuzin (214.28)
AmetryneAmetryne is a methylthiolated atrazine analogue. McMartin
et al. (2003) reported direct photolysis (pH 7–7.5) and photoFenton-type oxidation (pH 3) of 0.1–1 mg·L−1 of ametryne ingroundwater and soil samples (λ = 300–400 nm or 254 nm,15 W). In the case of photo Fenton-type oxidation, ferric chlo-ride was added to the solution to reach a concentration of0.33 mg·L−1 of Fe3+, but no hydrogen peroxide was added.Ametryne was converted faster by the photo Fenton-type systemthan by direct photolysis, although the former process showedbiphasic kinetics, where a substantial reaction rate drop oc-curred after approximately 15 min of treatment. Ametryne wasapparently more readily degradable than atrazine. McMartin etal. (2003) suggested that the addition of H2O2 would increasethe reaction rate by simultaneously re-oxidizing the Fe2+ andgenerating hydroxyl radicals. No degradation products wereidentified.
AtrazineVarious types of Fenton-type processes were evaluated for
the degradation of atrazine, an s-triazine herbicide banned inEurope due to concern over the environmental persistence ofits degradation products, but still heavily used in the UnitedStates. Prados et al. (1995) reported 29–52% conversion of3.5 mg·L−1 of atrazine by a Fenton process with 10 mg·L−1
of Fe2+ and 5 mg·L−1 of H2O2 at pH 5.5–5. Arnold et al.(1995a, 1995b) reported complete conversion of 29 mg·L−1 ofatrazine by a Fenton process, and the treatment was optimizedto reduce the formation of chlorinated degradation products. Upto 55% dechlorination of this triazine herbicide was observedafter the Fenton treatment, and the major degradation prod-ucts were chlorodiamino-s-triazine and its mono-acetoamide(Arnold et al. 1995a). Other degradation intermediates and (or)by-products were also determined (Arnold et al. 1995a, 1995b),and the degradation pathway was proposed as shown in Fig. 29.Arnold et al. (1996) also reported that selected microorganismsRhodococcus corallinus NRRL B-15444R and Pseudomonassp. NRRL B-12228, known to have dehalogenase activity andbroad substrate specificity for chlorinated atrazines, respec-tively, could mineralize these atrazine degradation products up
Fig. 29. Proposed degradation pathway of atrazine by Fentonprocess (Arnold et al. 1995a).
N
N
N
Cl
NH
NH
N
N
N
Cl
NH
NH
N
N
N
Cl
NH2NH
N
N
N
OH
NH
NH
N
N
N
Cl
NH2H2N
N
N
N
Cl
NH
H2N
O
N
N
N
Cl
NH
H2N
N
N
N
Cl
NH
NH
O
N
N
N
OH
NH2H2N
O
O
atrazine
deethylatrazine
chlorodiamino-s-triazine
deisopropylatrazine
ammeline
to 73% as 14CO2 from [2,4,6-14C] atrazine. Kinetic modelingof atrazine conversion by the Fenton process was recently re-ported (Chan and Chu 2003a, 2003b).
Huston and Pignatello (1999) reported the complete con-version of 49 mg·L−1 of atrazine and 46% TOC reductionby the photo Fenton process (λ = 300–400 nm, 1.2 × 1919
photons·L−1·s−1, 2.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2,pH 2.8, 25 ◦C). They also observed quantitative dechlorina-tion after 2 h of photo Fenton treatment, although no organicacids, such as formate, oxalate, or acetate, were detected. Thisimplies that no s-triazine ring opening occurred. Balmer andSulzberger (1999) demonstrated that the addition of oxalate(1.6–16.2 mg·L−1) enhanced the performance of the photo Fen-ton process for the conversion of atrazine. The largest effectof the addition of oxalate was observed at pH 4.6 (Balmer andSulzberger 1999).As with the case of ametryne reviewed above,another type of photo Fenton-like degradation of 0.1–1 mg·L−1
of atrazine was reported in a natural groundwater sample with-out the addition of H2O2 (McMartin et al. 2003). Atrazine wasapparently more resistant than ametryne to degradation by thephoto Fenton-like process. The rate constant for the reaction ofatrazine with hydroxyl radicals generated by the photo Fentonprocess was also reported (Haag and Yao 1992). Resistance ofatrazine to complete mineralization and detoxification was alsoreported in the solar-driven photo-Fenton (Hincapié et al. 2005)and photo-Fenton/O3 processes (Farré et al. 2005).
Sun and Pignatello (1993a) demonstrated that the addition of1 mM of a chelating agent, such as picolinic acid, gallic acid,or rhodizonic acid (Fig. 3), also enhanced the conversion of21.5 mg·L−1 of atrazine by the dark Fenton-like Fe3+/H2O2process. Among the chelating agents tested, gallic acid was themost effective. Although complete conversion of this triazineherbicide was achieved, destruction of its s-triazine ring was not
Fig. 30. Chemical structure of phenolic substances evaluated forFe3+-chelate/H2O2 (dark) process (Rivas et al. 2002).
OH
OH
OHO
OHOH
OO
p-hydroxybenzoicacid
tyrosol(4-hydroxyphenethyl
alcohol)
4-hydroxycinnamicacid
observed after the Fe3+-chelate/H2O2 treatment (Sun and Pig-natello 1993a). More recently, Rivas et al. (2002) reported thepositive effect of a phenolic compound p-hydroxybenzoic acid(1 mM, Fig. 30) on the conversion of 21.6 mg·L−1 of atrazineby the dark Fe3+/H2O2 process. The addition of other pheno-lic compounds, including tyrosol (4-hydroxyphenethyl alcohol)and 4-hydroxycinnamic acid, also enhanced the conversion ofthis triazine herbicide. Rivas et al. (2002) also presented a ki-netic model for the conversion of atrazine by the Fe3+/H2O2process in the presence of these phenolic compounds.
A series of studies were reported on the degradation of aque-ous atrazine by the electrochemical Fenton-type process (Pratapand Lemley 1994, 1998; Saltmiras and Lemley 2002). As de-scribed above (see Picloram, a pyridine derivative), the initialgeneration of the electro Fenton process was less efficient thanclassical Fenton treatment (Pratap and Lemley 1994). How-ever, efficiency was improved, and the complete conversionof 26–30 mg·L−1 of atrazine was made possible, when theprocess was combined with near UV-irradiation (photoelectroFenton; λ = 330–400 nm, 201 mg·L−1 of Fe2+, 7.5 g·L−1 ofH2O2) (Pratap and Lemley 1998). Several degradation productswere detected after the photoelectro Fenton treatment, and threewere identified as deethylatrazine, chlorodiamino- s-triazine,and ammeline (see Fig. 29 for their structures). Another va-riety of electrochemical Fenton process, anodic Fenton, wasalso evaluated (Lemley et al. 2004), and its effectiveness on thedegradation of atrazine and seven atrazine degradation inter-mediates was demonstrated (Saltmiras and Lemley 2002). Theanodic Fenton process employs two separated half-cells, andFenton reactions occur only in the anode half-cell where fer-rous ion was generated by a sacrificial iron electrode. By sepa-rating the half-cells, it is possible to prevent increases in pH dueto hydroxyl ion generation. Using the anodic Fenton process,29 mg·L−1 of atrazine was completely converted within 3 min,and ammeline and chlorodiamino- s-triazine were obtained asprimary degradation end-products after 10 min of treatment(Saltmiras and Lemley 2002). A degradation pathway, essen-tially the same as the one shown in Fig. 29, was also proposedfor the anodic Fenton treatment of atrazine (Saltmiras and Lem-ley 2002). Destruction of the s-triazine ring was not observedto a measurable extent during any of these electrochemical pro-cesses.
A few groups of researchers evaluated the H2O2/UV pro-cess for the degradation of atrazine (Beltrán et al. 1993, 1996;Hessler et al. 1993; Prados et al. 1995; Prado and Esplugas1999). Generally, these studies indicated that the conversion ofthis triazine herbicide was enhanced by the addition of H2O2 ascompared with direct photolysis alone. However, Beltrán et al.(1993) reported that a high concentration (>3.4 g·L−1) of H2O2inhibited the atrazine conversion ([atrazine]0 = 10 mg·L−1),a result likely due to the scavenging of hydroxyl radicals byH2O2. Significant dark reactions of 5.4–8.2 µg·L−1 of atrazinewere also observed in the presence of a high concentration(0.68 g·L−1) of H2O2 (Hessler et al. 1993). The photochem-ical conversion (λ =254 nm, 6.3 × 10−7 Eins·L−1·s−1) of5.8 mg·L−1 of atrazine was strongly enhanced at pH 3 and7 in the presence of 6–60 mg·L−1 of H2O2, but the effect wasless pronounced at pH 11 (Hessler et al. 1993). The presenceof natural radical scavengers, including humic substances andbicarbonate, inhibited atrazine conversion by the H2O2/UV pro-cess (Beltrán et al. 1993). The kinetic constant for the hydroxylradical reactions and quantum yields for the photochemical re-actions were reported as shown in Appendix A (Beltrán et al.1993; Hessler et al. 1993). Prado and Esplugas (1999) comparedseveral ozone-based AOPs, UV photolysis, and the H2O2/UVprocess for atrazine conversion. They demonstrated the fastestconversion of this herbicide in a given set of reaction condi-tions at neutral pH, although they and other researchers alsosuggested that the H2O2/UV process would not work in col-ored water (Prados et al. 1995; Prado and Esplugas 1999). Sev-eral degradation by-products of atrazine, including deethyla-trazine, deisopropylatrazine and their 2-hydroxyl derivatives,were identified after the H2O2/UV treatment (Hessler et al.1993).
CyanazineBenitez et al. (1995a) reported a kinetic study of cyanazine
conversion by the H2O2/UV process. The nearly complete con-version of 100 mg·L−1 of cyanazine was achieved by106 mg·L−1 of H2O2 and UV irradiation (λ = 254 nm) atpH 9 and 20 ◦C. Whereas the conversion of this herbicide bythe H2O2/UV process was insensitive to pH in the tested range(pH 5–9); however, it was accelerated by increasing the tem-perature up to 40 ◦C.
Pratap and Lemley (1994) reported the complete conversionof 30 mg·L−1 of cyanazine using the classical Fenton process(50 mg·L−1 of Fe2+, 7.5 g·L−1 of H2O2, pH 2.3–3.2, 22–27 ◦C). Complete conversion of 132 µg·L−1 of cyanazine wasalso reported using the Fenton process in pesticide rinse wa-ter also containing atrazine, EPTC, alachlor, and metolachlor(295 mg·L−1 of Fe2+, 170 mg·L−1 of H2O2, pH 2.5, 25 ◦C)(Arnold et al. 1996). These studies consistently indicated thatcyanazine appeared to be more resistant than atrazine to degra-dation by hydroxyl radicals. After the Fenton treatment, theformation of several degradation products was observed. Theseproducts were assumed to be dealkylated cyanazine derivatives,although they were not identified (Pratap and Lemley 1994).An
earlier generation of the electro Fenton process (see also Piclo-ram, a pyridine derivative) was also evaluated for cyanazinedegradation; however, it was less efficient than the classicalFenton process (Pratap and Lemley 1994). No further studywas reported on the degradation of cyanazine with Fenton-typeor H2O2/UV AOPs.
Cyanuric acidThe very high chemical stability of cyanuric acid, a micro-
biocide and end product of triazine herbicide degradation, isevident from the fact that it has no oxidizable side chains, andthree triazine carbons are fully oxidized. Cyanuric acid cannotbe degraded by conventional AOPs including O3/UV, Fenton,and H2O2/UV processes, or by direct photolysis (λ > 340 nm)and TiO2/hν (De Laat et al. 1994; Minero et al. 1997; Goutailleret al. 2001). It is known that this compound can be degraded byhydrolysis, pyrolysis, γ -radiolysis, laser flash photolysis, vac-uum UV irradiation (λ = 100–200 nm), or biodegradation withselected microorganisms (Minero et al. 1997; Manoj et al. 2002;Horikoshi et al. 2003). It is unlikely that it can be degraded byoxidation.
MetribuzinScherer et al. (2004) investigated the membrane anodic Fen-
ton process as a chemical pretreatment of metribuzin to im-prove biodegradability of the pesticide solution (see Trifluralin,a aniline-based compound for the process description). Metri-buzin (85.7 mg·L−1) was quickly oxidized by this type of elec-tro Fenton treatment in 12.5 min with a continuous deliveryof H2O2 and ferrous ion (molar ratio = 2.5:1, 2.6 ng·min−1 ofH2O2 + electrolysis of sacrificial iron electrode at 0.1 A). Threeoxidation by-products were identified by GC–MS analysis asshown in Fig. 31. Biodegradability monitored by BOD5/CODratio was improved from 0.03 to 0.35 as a result of the anodicFenton treatment. This improvement was apparently due to theremoval of amino and methylthio groups from the metribuzinmolecule (Fig. 31). A kinetic model was also developed to pre-dict the degradation of metribuzin by the anodic Fenton process(Wang et al. 2004). A weak interaction between metribuzin andferric ion was observed, which slowed the herbicide degrada-tion during the treatment.
SimazineSimazine is very similar to atrazine in terms of its chemical
structure. There is one notable modification; simazine featuresanother N-ethyl group on the s-triazine ring instead of the N-isopropyl group found in atrazine. Thus, the reactivity of this tri-azine herbicide toward chemical oxidation is somewhat similarto that of atrazine. Haag andYao (1992) reported a rate constantfor the reaction of simazine with hydroxyl radicals generatedby the photo Fenton process as 2.8 × 109 M−1·L−1 at pH 2.0.This value is very close to the one determined for atrazine,2.6 × 109 M−1·L−1 at pH 1.6. More recently, Huston and Pig-natello (1999) reported that the conversion of simazine by photoFenton was faster than that of atrazine. As discussed previously,
Fig. 31. Proposed degradation pathway of metribuzin by anodicFenton process (Scherer et al. 2004).
N
N
N
S
NH2
O
N
N
N
SO
N
N
N
O
NH2
O
N
N
N
OO
metribuzin deamino-metribuzin
diketo-metribuzindeamino-diketo-metribuzin
the s-triazine ring was not destroyed during the photo Fentontreatment (see also Atrazine).
Beltrán et al. (2000) reported a kinetic study of simazinedegradation by the H2O2/UV AOP, along with ozonation andozone-based AOPs. Complete conversion of 5 mg·L−1 ofsimazine was confirmed with the H2O2/UV AOP (λ = 254 nm,1.9 × 10−6 Eins·L−1·s−1, 136 mg·L−1 of H2O2, pH 7). Thekinetic analysis of Beltrán et al. revealed the major contribu-tion (68%) of the radical reaction pathway to the conversion ofsimazine by the H2O2/UV process over that of direct photolysis(32%).
No degradation products of simazine have been determinedin any of the studies reviewed here, although N-deethylation ofsimazine is known to occur during Fenton treatment (Lai et al.1995).
Summary of triazinesAll of the triazines (and metribuzin) reviewed here are fairly
reactive toward the hydroxyl radicals generated by Fenton-typeor H2O2/UVAOPs. Simazine and ametryne are apparently morereactive than atrazine, and cyanazine is more resistant. Nev-ertheless, the destruction of the s-triazine ring of these com-pounds, including cyanuric acid, has proven to be very difficultusing conventional oxidation processes, including Fenton andH2O2/UV AOPs. It should be noted that cyanuric acid is muchless toxic than the parent herbicides and can be degraded mi-crobiologically. However, incomplete degradation of triazineherbicides would lead to the formation of varieties of toxic by-products; thus, careful process monitoring and managementshould be implemented. Kinetic data for the degradation ofthese herbicides by AOPs are well documented.
Substituted urea compoundsThe substituted urea pesticides reviewed here include diuron,
fenuron, isoproturon, lufenuron, linuron, metoxuron, metobro-muron, and monolinuron. Most of them are phenylurea her-bicides, with the exception of lufenuron (benzoylurea insecti-cide). The chemical structures and formula weights of thesecompounds are shown in Fig. 32.
DiuronMalato et al. (2002a, 2003a, 2003b, 2003c) and Hincapié et
al. (2005) reported a series of investigations into the degradationof diuron by solar-driven photo Fenton (2.8 mg·L−1 of Fe2+,510 mg·L−1 of H2O2 maintained, pH 2.7–2.8) and TiO2/hν
(200 mg·L−1) processes in pilot-scale solar reactors (refer toan overview article by Malato et al. (2002b) for the details ofthe solar reactors). The complete conversion of 30 mg·L−1 ofdiuron and a substantial TOC reduction were achieved with bothprocesses, although the photo Fenton process consistently per-formed better than the TiO2/hν process (Malato et al. 2002a).It was found that approximately 10% of initial TOC was diffi-cult to degrade (Hincapié et al. 2005). Quantitative dechlorina-tion was achieved in both cases within a relatively early stageof treatment (Malato et al. 2003c). A number of degradationby-products and (or) intermediates were identified during thesolar-driven photo Fenton and TiO2/hν processes (Malato etal. 2003a, 2003c), and a degradation pathway was proposed inwhich competitive and interchangeable dechlorination, hydrox-ylation, and N-demethylation reactions are involved (Fig. 33).Various organic acids including oxalate, formate, and acetatewere also generated and subsequently degraded during the treat-ment. The photo-Fenton/O3 process (see Alachlor, an aniline-based pesticide, for the reaction conditions) was also evaluatedfor diuron degradation (Farré et al. 2005). It was found that di-uron mineralization was relatively slow as compared with otherpesticides including PCP, alachlor, chlorfenvinfos, and isopro-turon, and that about 50% of initial TOC was removed from50 mg·L−1 of diuron solution after 1.5 h of treatment.
The toxicity of the diuron solution was also monitored bybioassays with Daphnia magna and microalgae Selenastrumcapricornotum as well as by Microtox during the solar-drivenphotocatalytic treatment (Malato et al. 2003c; Hincapié et al.2005). Although the toxicity initially decreased quickly, un-known and very toxic degradation intermediate(s) formed. Con-sequently, the toxicity increased to nearly the original value, and
Fig. 33. Proposed degradation pathway of diuron by photo Fentonand TiO2/hν process (Malato et al. 2003c).
Cl
Cl
HN
N
O
Cl
OH
HN
N
O
OH
OH
HN
N
O
OH
OH
HN
N
O
OH
Cl
Cl
HN
N
O
H
O
Cl
Cl
HN
HN
O
Cl
Cl
HN
H2N
O
Cl
Cl
HN
N
O
OH
Cl
OH
HN
N
O
OH
ClHN
N
O
HO
OH
OH
diuron
dechlorination
N-demethylation
hydroxylation
CO2 +NO2-+ NH4
3++ Cl
-
then quickly dissipated. This result suggests that complete con-trol over treatment processes needs to be attained to assure thequality of treated water and wastewater (Malato et al. 2003c).
FenuronAcero et al. (2002) evaluated UV photolysis, H2O2/UV, Fen-
ton, and photo Fenton processes, as well as ozonation and sev-eral ozone-based AOPs for the degradation of fenuron (λ =185–436 nm, 1.76 × 10−5 Eins·s−1, 34–170 mg·L−1 of H2O2,2.8–5.6 mg·L−1 of Fe2+, pH 3 for Fenton and photo Fenton,pH 5 for UV/H2O2). Complete conversion of this phenylureaherbicide (16.4 mg·L−1) was achieved in all cases with the givenreaction conditions, with the exception of direct photolysis, bywhich the conversion occurred very slowly. Various kinetic pa-rameters to model the overall reaction of fenuron degradationwere determined, including quantum yield for direct photoly-sis, and rate constants for direct ozone reactions and hydroxylradical reactions (Acero et al. 2002). No attempt was made todetermine the degradation by-products and (or) intermediatesof fenuron generated during the AOPs.
Isoproturon and metobromuronParra et al. (2000) reported the degradation, and subsequent
biodegredation, of isoproturon and metobromuron in a num-ber of variations of Fenton-type and TiO2/hν processes withsimulated solar irradiation (λ = 300–800 nm, 80 mW·cm−2).The best result was obtained using the photo Fenton process(55.8 mg·L−1 of Fe3+, 850 mg·L−1 of H2O2), by which morethan 80% TOC reduction was achieved for both compounds(43.3 mg·L−1 of isoproturon or 241 mg·L−1 of metobromuron).The TiO2/H2O2/hν (1 g·L−1 of TiO2, 850 mg·L−1 of H2O2) and
the dark Fenton process were also effective, although TiO2/hν
without the addition of H2O2 was less effective in reducingTOC. Direct photolysis, Fe3+/UV, and H2O2/UV processes un-der comparable conditions were not very effective in reducingTOC (Parra et al. 2000).
The evolution of aliphatic and aromatic degradation interme-diates (not identified), TOC, the BOD5/COD ratio, and Micor-tox toxicity were also monitored over the course of the photoFenton treatment of these two phenyl urea herbicides (Parra etal. 2000). In the case of isoproturon, the herbicide solution wassuccessfully detoxified and the biodegradability was substan-tially improved after the treatment. By contrast, although themajority of TOC and toxicity was removed from the metobro-muron solution by the photo Fenton treatment the biodegrad-ability was virtually unchanged and recalcitrant by-products(presumably brominated organics) remained after the treatment.By coupling the photo Fenton treatment and biodegradation,95% mineralization was achieved in the case of isoproturon(Parra et al. 2000).
Isoproturon degradation was recently evaluated also in thephoto-Fenton/O3 (Farré et al. 2005) and solar-driven photo-Fenton processes (Hincapié et al. 2005). Nearly 70% and 90%TOC reduction in 50 mg·L−1 of isoproturon solution wasachieved in these processes, respectively (see Alachlor, ananiline-based pesticide, for the reaction conditions). No specificdegradation products and (or) intermediates were determinedexcept inorganic ions including bromide, nitrate, and ammoniain these studies.
Lufenuron
Fallmann et al. (1999a) reported the treatment of a commer-cial insecticide containing the benzoyl urea compound lufe-nuron using the photo Fenton process with simulated solar irra-diation (λ = 300–450 nm, 70 W·L−1) in water. More than 90%of TOC (initially 100 mg·L−1) was removed from the insecti-cide solution after the photo Fenton treatment with 14 mg·L−1
of Fe2+ and 680 mg·L−1 of H2O2 at pH 2.8 and 20–55 ◦C(temperature uncontrolled). The same group of authors also de-monstrated successful treatment of a mixture of 10 commercialpesticides, one of which contained lufenuron, in a similar man-ner in laboratory and pilot-scale photo (solar) reactors (Blancoet al. 1999; Fallmann et al. 1999a, 1999b). Degradation by-products were not determined in any of the studies describedabove.
Linuron and monolinuron
Barlas (2000) evaluated the degradation of 25 mg·L−1 oflinuron and 32–43 mg·L−1 of monolinuron by Fenton andH2O2/UV processes at pH 2–3. The former phenylurea her-bicide was less reactive than the latter during both types ofadvanced oxidation treatment, a result most likely due to thepresence of the extra chlorine atom on the aromatic ring (seeFig. 32). Degradation products were determined for neithercompound in any of the cases mentioned.
Fig. 34. Proposed degradation pathway of metoxuron by O3/UVand H2O2/UV (Mansour et al. 1992). Some intermediates areomitted for clarity.
OHN
N
O
OHN
N
O
O
H
OHN
HN
O
OHN
NH2
O
OHN
N
O
OHN
N
O
Cl
HO
Cl
Cl
Cl
Cl
OH
OHN
N
O
OH
HO
OHN
N
O
O
O
OHHN
N
O
Cl
OHHN
N
O
OH
metoxuron
ring opening,generation of polar
compounds
Metoxuron
Nearly complete conversion of 20 mg·L−1 of metoxuronwas achieved through the H2O2/UV process (λ > 290 nm,0.36 mg·L−1 of H2O2). Degradation by-products and (or) in-termediates of metoxuron were also identified, and a proposeddegradation pathway is shown in Fig. 34 (Mansour et al. 1992).
Summary of substituted urea compounds
Effective degradation of all of the substituted urea com-pounds reviewed here has been demonstrated using eitherFenton-type or H2O2/UV AOPs. Degradation by-products and(or) intermediates have been determined for diuron and metox-uron, although they are largely unknown for the other com-pounds. Toxicity reduction has been confirmed in the case ofphoto Fenton treatment of isoproturon and metobromuron. Itis interesting to note that, although the biodegradability of theisoproturon solution was improved after the photo Fenton treat-ment, it was unchanged in the case of metobromuron. This ob-servation emphasizes the importance of identifying and charac-terizing degradation by-products in terms of both toxicity andbiodegradability in order to validate the performance of AOPs.
Miscellaneous pesticides
Several miscellaneous pesticides shown in Fig. 35 are also re-viewed here, including acrinatrin, abamectin (avermectin), ben-tazone, captan, and carbetamide. Acrinatrin, captan, and carbe-tamide are a pyrethroid insecticide, a thiophthalimide fungicide,and an amide herbicide, respectively. Abamectin and bentazoneare an unclassified insecticide and herbicide, respectively.
Fig. 35. Chemical structure of miscellaneous pesticides (formulaweight is shown in brackets).
O
N
O
O
O
O
F
FF
F
F
F
HN
N
S
O
O
O
N
O
O
S ClCl
Cl
HN
O
O
NH
O
O
O
HOHO
O
O
OH
H
H
O
HO
HOO
OH H
HO HH
O
H
HH
acrinatrin (541.45)
bentazone (240.28) captan (300.59)
carbetamide (236.27)
avemectin B1a
(abamectin, 873.09)
Acrinatrin and abamectin (avermectin B1)Commercial formulations of acrinatrin and abamectin were
treated with photo Fenton or TiO2/hν processes individually oras a mixture of 10 pesticides in aqueous solution in laboratoryand pilot-scale photo (solar) reactors (Blanco et al. 1999; Fall-mann et al. 1999a, 1999b). More than 70% and 90% TOC (ini-tially 100 mg·L−1) was removed from the solutions of acrina-trin and abamectin, respectively, after photo Fenton treatmentwith 14 mg·L−1 of Fe2+ and 680 mg·L−1 H2O2 at pH 2.8 and20–55 ◦C (temperature uncontrolled) (Fallmann et al. 1999a).Degradation by-products of these pesticides were not identified.
BentazoneBeltrán-Heredia et al. (1996) reported the kinetic study
of bentazone conversion by the H2O2/UV AOP. The quan-tum yields of UV direct photolysis (λ = 239–366 nm, 4.52 ×10−5 Eins·s−1) of bentazone and the rate constants for the re-action of hydroxyl radicals with bentazone were determined atdifferent pH and temperature (10–40 ◦C) values. For example,the quantum yield and the rate constant at pH 7 and 20 ◦C were1.25 and 2.92 × 109 M−1·s−1, respectively. The temperature-dependent Arrhenius equation was also established for both ofthe kinetic parameters at pH 7.
CaptanHuston and Pignatello (1999) reported the complete con-
version of 0.88 mg·L−1 captan by the photo Fenton processwithin 10 min (λ = 300–400 nm, 1.2×1919 photons ·L−1 ·s−1,2.8 mg·L−1 of Fe3+, 340 mg·L−1 of H2O2, pH 2.8, 25 ◦C). Noattempt was made to determine degradation by-products.
Fig. 36. Proposed degradation pathway of carbetamide by O3/UVand H2O2/UV (Mansour et al. 1992).
HN
O
O
NH
O
OH
NH
O
N
OO
O
HN
O
O
NH2
O
OH
NH2
O
HN
O
O
NH
O
O
O
NH
O
H2N
(OH)n
carbetamide
n = 1, 2
CarbetamideMansour et al. (1992) investigated the various photochem-
ical degradation (λ = 290 nm) of carbetamide (amide herbi-cide) in water. A nearly complete conversion of 20 mg·L−1 ofcarbetamide was achieved with the H2O2/UV (0.17 mg·L−1
H2O2), TiO2/UV (150 mg·L−1 TiO2), and O3/UV processes(1.44 mg·L−1 applied ozone dose). The latter process appearedto be more effective than the former ones in the conversion ofthis herbicide. Mansour et al. (1992) also identified degradationby-products and (or) intermediates of metoxuron, and proposedthe degradation pathway shown in Fig. 36.
Summary of miscellaneous pesticidesIt is impossible to generalize or to compare the data for the
miscellaneous pesticide degradation processes reviewed herebecause of the structural diversity among these pesticides. Verylittle data are available for most of the pesticides. Degradationby-products have only been determined for the carbetamidedegradation by H2O2/UV AOP. No data is available on the tox-icity and biodegradability of the degradation by-products.
Concluding remarks
It has been clearly shown that the majority of the pesticidesreviewed in this paper, including aniline-based compounds,carbamates, chlorophenoxy compounds, organochlorines,organophosphates, pyridine and pyrimidine derivatives,triazines, substituted ureas, and some miscellaneous pesticides,are substantially reactive to, and readily degradable by, vari-ous Fenton-type AOPs. Remarkable performances of variousversions of photo-assisted Fenton and electrochemical Fentonprocesses have been demonstrated for aqueous pesticide degra-dation. In addition to the complete conversion of pesticides, amajor reduction (>50%) in TOC, as well as nearly completedechlorination (when a chlorinated compound was treated) wasachieved in most cases.
On the other hand, H2O2/UV AOP has yet to be as activelyinvestigated with respect to the degradation of aqueous pesti-cides as the Fenton-type processes have been. One apparentdisadvantage of the H2O2/UV process is its requirement of UV
irradiation at shorter wavelengths at which many organic com-pounds, as well as suspended solids, absorb photons and there-fore interfere with the activation of H2O2 to hydroxyl radicals.Consequently, this type ofAOP is likely unsuitable for the treat-ment of opaque pesticide wastewater and colored surface watercontaminated with pesticides. Instead, it is suitable for clearwater contaminated with pesticides.
In most of the studies reviewed here, pesticide concentra-tions fall into the 10–100 mg·L−1 level that represents pesticidewastewater rather than contaminated surface water and ground-water. Data regarding pesticide removal from contaminated sur-face water and groundwater with Fenton-type and H2O2/UVAOPs are scarce, whereas ozonation and ozone-based AOPs areoften employed for such purposes (Ikehata and Gamal El-Din2005a, 2005b). Because Fenton-type processes require morecareful pH control and sludge disposal, H2O2/UV AOP may bemore suitable for water treatment. Decomposition of residualH2O2 after treatment is required in both cases.
Although it has been demonstrated that nearly complete min-eralization of pesticides can be achieved using various Fenton-type or H2O2/UV AOPs, it may not be necessary to accom-plish such total decomposition in real situations because morecost effective biological treatment processes, such as activatedsludge and trickling filters, can be implemented after the AOPpretreatment of wastewater. Therefore, more studies may berecommended to evaluate Fenton-type and H2O2/UV AOPs aspretreatment processes to improve biodegradability of recalci-trant pesticide solutions.
Kinetic data on the degradation of certain pesticides are avail-able in the case of H2O2/UV and relatively simple classical andphoto Fenton processes. However, such data are limited forother Fenton-type processes, partly because of their complexsystems where pseudo first order kinetics cannot be applied.Several kinetic models have also been presented in some casesfor specific treatment systems, although further validation undermore realistic reaction conditions (i.e., in the presence of var-ious organic and inorganic compounds, mixture of pesticides)may be required before their practical uses can be evaluated.
Various degradation by-products, as well as intermediates,have been determined for many combinations of pesticidesand treatment processes. The degradation of major pesticides,such as metolachlor, 2,4-D, pentachlorophenol, malathion, andatrazine, has been studied intensively, and many publicationsare dealing with the identification of associated degradation by-products and (or) intermediates. However, such data are limitedfor many other pesticides, or, in some cases, no data is avail-able. This is likely due to the fact that the relative importance ofsuch pesticides is not as high as that of the major ones, or thatthe toxicity and (or) environmental impact of these compoundsare believed to be low. Nevertheless, improper treatment of pol-luted water and wastewater may lead to incomplete destructionof pesticides and the accidental release of potentially toxic pes-ticides and their degradation by-products. Thus, it is necessaryto have at least some data on the identity, properties, and fateof degradation by-products.
Another important aspect in the AOP application to contam-inated water and wastewater treatment in general is the processefficiency based on their energy consumption to achieve a cer-tain level of contaminant degradation or TOC removal (Legriniet al. 1993; Cater et al. 2000), although this is often overlooked.Such information is particularly useful for utilities to evaluatethe cost-effectiveness of the processes; therefore, it should beassessed and included in future studies.
The studies reviewed here have demonstrated that the acutetoxicity of pesticide solutions can be reduced or diminishedthrough treatment with various Fenton-type AOPs. However,formation of more toxic intermediates has been suggested dur-ing the advanced oxidation treatment of some pesticides (e.g.,octachlorodibenzo-p-dioxin formation during photo Fentontreatment of PCP, toxic intermediate formation during photoFenton treatment of methomyl, oxon of organophosphates).Thus, monitoring of the evolution of toxic intermediates as wellas toxicity during treatment is recommended to ensure the qual-ity of treated water and wastewater. In addition to the acutetoxicity, the potential endocrine disrupting activities of pesti-cide degradation products may need to be assessed because themajority of the synthetic pesticides currently marketed are aro-matic compounds, which often mimic estrogens in biologicalsystems (Kojima et al. 2004). Development of fast, sensitive,and reliable detection methods for estrogenic activity has beenan active research area in recent years (Diel et al. 1999; Combes2000); therefore, employing such techniques may be a goodstrategy to ensure the safety of treated water and wastewatercontaminated with pesticides as well as to validate the treat-ment processes.
Acknowledgements
The authors thank the Natural Sciences and Engineering Re-search Council of Canada (NSERC) andAlberta Ingenuity Fund(AIF) for their financial support.
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