application of boron-doped diamond electrodes for

J. Environ. Eng. Manage., 18(3), 139-153 (2008)

APPLICATION OF BORON-DOPED DIAMOND ELECTRODES FOR WASTEWATER TREATMENT Marco Panizza,1,* Enric Brillas2 and Christos Comninellis3 1

Department of Chemical and Process Engineering University of Genoa Genoa 16129, Italy 2 Laboratori d’Electroquímica dels Materials i del Medi Ambient Departament de Química Física, Facultat de Química Universitat de Barcelona Barcelona 08028, Spain 3 Institute of Chemical Sciences and Engineering Ecole Polytechnique Fédérale de Lausanne (EPFL) Lausanne CH-1015, Switzerland

Key Words: Boron-doped diamond, electro-Fenton, photoelectro-Fenton, electro-oxidation ABSTRACT Boron-doped diamond (BDD) thin film is a new electrode material that has received great attention recently because it possesses several technologically important characteristics such as an inert surface with low adsorption properties, remarkable corrosion stability, even in strong acidic media, and an extremely wide potential window in aqueous and non-aqueous electrolytes. Due to these properties, diamond electrodes are promising anodes for electrochemical treatment of wastewater containing organic pollutants. The objective of this review is to summarise and discuss the recent results available in the literature concerning the application of diamond electrodes to environmentally-oriented electrochemistry. A mechanism of the electrochemical incineration and a kinetic model for organics oxidation on BDD is also presented. Moreover, fundamentals and applications to wastewater treatment of the powerful electro-oxidation method using an undivided cell with BDD anode and the cathodic electrogeneration of hydrogen peroxide by electro-Fenton or photoelectro-Fenton processes are discussed. In these processes, organics are effectively destroyed by the large amounts of hydroxyl radical (•OH) produced on the BDD surface by water oxidation and by Fenton’s reaction between added Fe2+ and H2O2 electrogenerated at the cathode. The presence of Cu2+ as co-catalyst with UVA irradiation enhances the degradation of organics, making the photoelectro-Fenton with BDD anode viable for industrial application. INTRODUCTION As many industries produce wastewater containing toxic organic pollutants, there has been a notable increase in both research and the number of businesses concerned with the treatment of such industrial effluents and, nowadays, many new technologies are available, including biological, physical and chemical processes. In this field, oxidative electrochemical technologies, providing versatility, energy efficiency, amenability to automation and environmental compatibility have reached a promising stage of development and can now be effectively used for the destruction of toxic or biorefractory organics [1]. *Corresponding author Email: [email protected]

The overall performance of the electrochemical processes is determined by the complex interplay of parameters that may be optimized to obtain an effective and economical incineration of pollutants. The principal factors determining the electrolysis performance are electrode potential and current density, mass transport regime, cell design, electrolysis medium and, above all, electrode materials. The ideal electrode material for the degradation of organic pollutants should be totally stable in the electrolysis medium, cheap and exhibit high activity towards organic oxidation and low activity towards secondary reactions (e.g. oxygen evolution). Consequently, many anodic materials have been tested to find the optimum one. According to the

140


model proposed in previous works [2-4] anode materials can be divided into two extreme classes as follows: (i) Active anodes, which have low oxygen evolution overpotential and consequently are good electrocatalysts for the oxygen evolution reaction and only favour partial and selective oxidation (i.e. conversion), include carbon and graphite, platinum-based anodes, iridium-based oxides and ruthenium-based oxides. And (ii) Non-active anodes have high oxygen evolution overpotential and consequently are poor electrocatalysts for the oxygen evolution reaction but enable the complete and non-selective oxidation of organics to CO2 by electrogenerated hydroxyl radicals, such as antimony-doped tin dioxide, lead dioxide and borondoped diamond (BDD). Among the non-active anodes, BDD exhibits several technologically important properties that distinguish it from conventional electrodes, such as: (1) An extremely wide potential window in aqueous and non-aqueous electrolytes: in the case of high-quality diamond, hydrogen evolution commences at about –1.25 V vs. SHE and oxygen evolution at +2.3 V vs. SHE, then the potential window may exceed 3 V [5]; (2) Corrosion stability in very aggressive media: the morphology of diamond electrodes is stable during long-term cycling from hydrogen to oxygen evolution even in acidic fluoride media [6]; (3) An inert surface with low adsorption properties and a strong tendency to resist deactivation: the voltammetric response towards ferri/ferrocyanide is remarkably stable for up to two weeks of continuous potential cycling [7]; and (4) Very low double-layer capacitance and background current: the diamond-electrolyte interface is ideally polarisable and the current between –1000 and +1000 mV vs. SCE is < 50 μA cm-2. The double-layer capacitance is up to one order of magnitude lower than that of glassy carbon [8]. Due to these properties, conducting diamond seems to be a promising electrode material and so, in the last decades, it has been widely studied with the goal of developing applications in the electrochemical oxidation of organics for wastewater treatment [9,10]. In fact, during electrolysis in the potential region of water discharge, BDD anodes involve the production of weakly adsorbed hydroxyl radicals that unselectively and mineralise organic pollutants with high current efficiency [11]. In the last years, several indirect electrooxidation methods based on the cathodic electrogeneration of hydrogen peroxide like electro-Fenton (EF) and photoelectro-Fenton (PEF) are being developed for the remediation of acidic wastewaters containing toxic and non-biodegradable organic pollutants. They are environmentally friendly techniques since the main oxidant of organics is the in situ electrogenerated hydroxyl radical, which is a very strong oxidizing species due to its high standard potential (Eº = 2.80 V vs. SHE). This radical is able to non-selectively react

with most organic pollutants yielding dehydrogenated or hydroxylated products up to total mineralization. When an undivided electrolytic cell is used, the EF method involves the production of hydroxyl radical in the bulk solution from the catalytic Fenton’s reaction between Fe2+ with electrogenerated H2O2, simultaneously to the generation of radical •OH from water oxidation at the surface of the anode. This method thus profits the oxidation ability of both anode and cathode reactions and then, it is expected to be more efficient to destroy organics than direct anodic oxidation (AO). Its oxidation power depends on the electrodes and metallic ions (Fe2+, Fe3+, Cu2+, etc.) used as catalyst. The effectiveness of EF can be enhanced under irradiation of the solution with UVA light in the PEF treatment, which can be even more efficient using sunlight as energy source in the solar PEF (SPEF) process. The aim of this review is to elucidate the basic principles of electro-oxidation of organics with BDD anode and to discuss the recent progress dealing with the application of diamond electrodes in the electrochemical treatment of wastewater containing organic pollutants. A kinetic model for the prediction of the evolution of COD (chemical oxygen demand) during organics incineration on BDD is also presented. Finally, fundamentals and applications of the AO on BDD, as well as of the EF or PEF processes using the same anode with an undivided cell, will also be discussed herein. The influence of applied current density and metallic ions used as catalyst (Fe2+ and/or Cu2+) in the EF and PEF processes is also explored. MECHANISM OF THE ELECTROCHEMICAL INCINERATION In electrochemical incineration (EI) reactions oxygen is transferred from water to the organic pollutant using electrical energy. This is the so-called electrochemical oxygen transfer reaction (EOTR). A typical example of EOTR is the anodic EI of phenol (Eq. 1). C6H5OH + 11H2O → 6CO2 + 28H+ + 28e- (1) In this reaction water is the source of oxygen atoms for the complete oxidation of phenol to CO2 at the anode of the electrolytic cell. The liberated protons in this reaction are discharged at the cathode to dihydrogen (Eq. 2). 2H+ + 2e- → H2 (2) According to the generally accepted mechanism of the EI, water is firstly discharged at the anode active sites M producing adsorbed hydroxyl radicals (Eq. 3). H2O + M → M(•OH)ads + H+ + e(3) These electrogenerated hydroxyl radicals are involved in the incineration of organic pollutants R (present in an aqueous solution) (Eq. 4)

Panizza et al.: BDD Electrodes for Wastewater Treatment

R(aq) + x M(•OH)ads → xM + Incineration product + yH+ + ye- (4) where x and y are the stoichiometric coefficients. This reaction (Eq. 4) is in competition with the side reaction of the anodic discharge of these radicals to dioxygen (Eq. 5). M(•OH)ads → M +

1 O2 + H+ + e2

(5)

KINETIC MODEL OF ORGANICS MINERALIZATION ON BDD ANODE In this section a kinetic model of electrochemical mineralization of organics (R) on BDD anodes under electrolysis regime is presented. In this regime, as reported in the previous paragraph, electrogenerated hydroxyl radicals (Eq. 6) are the intermediates for both the main reaction of organics oxidation (Eq. 7) and the side reaction of oxygen evolution (Eq. 8). BDD + H2O → BDD(•OH) + H+ + e-

(6)

R + x BDD(•OH) → xBDD + Incineration product + yH+ + ye- (7) BDD(•OH)ads → BDD + 1/2O2 + H+ + e- (8) Considering this simplified reaction scheme (Eqs. 6-8) a kinetic model is proposed based on following assumptions: (i) adsorption of the organic compounds at the electrode surface is negligible; (ii) all organics have the same diffusion coefficient D; (iii) the rate of the electrochemical mineralization of organics is a fast reaction and it is controlled by mass transport of organics to the anode surface. The consequence of this last assumption is that the rate of the mineralization reaction is independent on the chemical nature of the organic compound present in the electrolyte. Under these conditions, the limiting current density for the electrochemical mineralization of an organic compound (or a mixture of organics) under given hydrodynamic conditions can be written as : ilim = n F km Corg (9) where ilim is the limiting current density for organics mineralization (A m-2), n is the number of electrons involved in organics mineralization reaction, F is the Faraday constant (96,487 C mol-1), km is the mass transport coefficient (m s-1) and Corg is the concentration of organics in solution (mM). For the electrochemical mineralization of a generic organic compound, it is possible to calculate the number of exchanged electrons, from the following electrochemical reaction: CxHyOz + (2x-z)H2O → xCO2 + (4x+y-2z)H+ + (4x+y-2z)e- (10) Replacing the value of n = (4x+y-2z) in Eq. 9 we obtain: ilim = (4x+y-2z) F km Corg

(11)

141

From the equation of the chemical mineralization of the organic compound (CxHyOz), y ⎛ 4x + y - 2z ⎞ Cx H yOz + ⎜ ⎟O 2 = xCO 2 + H 2 O 4 2 ⎝ ⎠ (12) It is possible to obtain the relation between the organics concentration (Corg in mM) and the COD (in mM): ⎛ ⎞ 4 ⎟⎟COD Corg = ⎜⎜ (13) ⎝ 4x + y - 2z ⎠ From Eqs. 11 and 12 and at given time t during electrolysis, we can relate the limiting current density of the electrochemical mineralization of organics with the COD of the electrolyte (Eq. 14): ilim(t) = 4F km COD(t) (14) At the beginning of electrolysis, at time t = 0, the initial limiting current density (ilim,0) is given by: (15) ilim,0(t) = 4F km COD0 where COD0 is the initial COD. We define a characteristic parameter α of the electrolysis process (Eq. 16). Working under galvanostatic conditions α is constant, and it is possible to identify two different operating regimes: at α < 1 the electrolysis is controlled by the applied current, while at α > 1 it is controlled by the mass transport control.

α=

i i lim,0

(16)

1. Electrolysis under Current Limited Control (α < 1)

In this operating regime (i < ilim), the current efficiency is 100% and the rate of COD removal (mol m-2 s-1) is constant and can be written as: r =α

i lim,0

(17) 4F Using Eq. 17, the rate of COD removal (Eq. 18) can be given by: r = α km COD0 (18) It is necessary to consider the mass-balances over the electrochemical cell and the reservoir to describe the temporal evolution of COD in the batch recirculation reactor system given in Fig. 1. Considering that the volume of the electrochemical reactor VE (m3) is much smaller than the reservoir volume VR (m3), we can obtain from the mass-balance on COD for the electrochemical cell the following relation: Q CODout = Q CODin - α km A COD0 (19) 3 -1 where Q is the flow-rate (m s ) through the electrochemical cell, CODin and CODout are COD (mM) at the inlet and at the outlet of the electrochemical cell, respectively, and A is the anode area (m2). For the


142

C O2

W1 R3

O2

W2 F1

F2

P3

E

R5

P1

dCOD COD A k m =− dt VR

P4

_

+

R1

R4

(26)

Integration of this equation from t = tcr to t, and COD = αCOD0 to COD(t) leads to:

R2

⎛ A km 1- α ⎞ ⎟ (27) COD(t) = αCOD 0 exp ⎜⎜ − t+ V α ⎟⎠ R ⎝

R6

P2

stantaneous current efficiency (ICE). In this case, the COD mass balances on the anodic compartment of the electrochemical cell E and the reservoir R2 (Fig. 1) can be expressed as:

The ICE can be defined as: Fig. 1. Scheme of the two-compartments electrochemical flow cell; R – reservoirs; P - pumps; E electrochemical cell with membrane; W - heat exchangers, F - gas flow controllers; Electrode surface 64 cm2; Volume of reservoir R1 and R2 250 mL.

well-mixed reservoir (Fig. 1) the mass balance on COD can be expressed as: dCOD Q(COD out - CODin ) = VR (20) dt Combining Eqs. 19 and 20 and replacing CODin by the temporal evolution of COD, we obtain: COD 0 A k m dCOD = −α (21) dt VR Integrating this equation subject to the initial condition COD = COD0 at t = 0 gives the evolution of COD(t) with time in this operating regime (I < ilim): ⎛ A km COD(t) = COD 0 ⎜⎜1 − α VR ⎝

⎞ t ⎟⎟ ⎠

(22)

This behaviour persists until a critical time (tcr), at which the applied current density is equal to the limiting current density, which corresponds to: CODcr = αCOD0

(23)

Substituting Eq. 23 in Eq. 22, it is possible to calculate the critical time:

ICE =

i lim COD(t) =− i αCOD 0

(28)

And from Eqs. 27 and 28, ICE is now given by: ⎛ A km 1- α ⎞ ⎟ ICE = exp ⎜⎜ − t+ α ⎟⎠ ⎝ VR

(29)

3. Influence of Organics Concentration and Current Density

A graphical representation of the proposed kinetic model is given in Fig. 2. In order to verify the validity of this model, the AO of various aromatic compounds in acidic solution has been performed varying organics concentration and current density. 3.1 Influence of the nature of organic pollutants Figure 3 shows both the experimental and predicted values (continuous line) of both ICE and COD evolution with the specific electrical charge passed during the AO of different classes of organic compounds (acetic acid, isopropanol, phenol, 4chlorophenol, 2-naphtol). This figure demonstrates that the electrochemical treatment is independent on the chemical nature of the organic compound. Furthermore, there is an excellent agreement between the experimental data and predicted values from proposed model.

2. Electrolysis under Mass Transport Control (α > 1)

3.2 Influence of organic concentration Figure 4a presents both ICE and COD evolution with the specific electrical charge passed during the galvanostatic oxidation (238 A m-2) of 2-naphtol (2-9 mM) in 1 M H2SO4 . As predicted from the model, the critical specific charge Qcr (Eq. 25) increases with increase of the initial organic concentration (COD°). Again, there is an excellent agreement between the experimental and predicted values.

When the applied current exceeds the limiting one (i > ilim), secondary reactions (such as oxygen evolution) commence resulting in a decrease of the in-

3.3 Influence of applied current density The influence of current density on both ICE and COD evolution with the specific electrical charge

t cr =

1 - α VR α Ak m

(24)

or in term of critical specific charge (Ah m-3): Q cr = i lim,0

4F COD 0 (1 - α) 1- α = k m 3600 3600

(25)


3500

B ⎛ Ak m COD(t) = COD 0 ⎜⎜1 − α VR ⎝

⎞ t ⎟⎟ ⎠

(a)

1.0

3000

lCE /(-) ICE -

A COD0

2500

COD cr =

COD (ppm) / ppm COD

COD

⎛ Ak m 1- α ⎞ ⎟ COD(t) = αCOD 0 exp ⎜⎜ − t+ α ⎟⎠ ⎝ VR

i appl 4Fk m

0.8 0.6 0.4 0.2 0.0

2000

0

5

10

15 -1

-1 Specific Specificcharge charge (Ah / AhLL )

1500 1000

0 t

500

b) 100 ICE (%)

0 Qcr = i lim,0

⎛ Ak m 1- α ⎞ ⎟ ICE = exp ⎜⎜ − t+ α ⎟⎠ ⎝ VR

1- α km

(b)

2000

1.0 0.8

lCE ICE(-) /-

a)

143

0 t

Fig. 2. Evolution of a) COD and b) ICE in function of time (or specific charge); (A) represents the charge transport control; (B) represents the mass transport control.

CO(ppm) D / pp m COD

1500

0.6 0.4 0.2 0.0

1000

0

5

10

15 -1

-1 Specific Specificcharge charge(Ah / AhLL )

500

(-) RendemlCE eICE nt e /n- c ourant / -

0

2500

COD (ppm) / ppm COD

2000 1500

0 1.0

0.6 0.4 0.2 0.0 5 10 -1-1-1 Specific charge/ Ah (Ah Charge / Ah Specificvolumique charge L LL )

15

0 5

15

0.8

500

0

10 -1

0

1000

5

Specific Ah L -1) Specificcharge charge/(Ah

10

Fig. 4. Influence of (a) the initial 2-naphtol concentration, (×) 9 mM, ({) 5 mM, (◊) 2mM and (b) the applied current density, (×) 119 A m-2, ({) 238 A m-2, (◊) 476 A m-2 in 5 mM 2-naphtol on the evolution of COD and ICE (inset) during electrolysis on BDD; i = 238 A m-2; T = 25 °C; Electrolyte: 1 M H2SO4; The solid line represents model prediction.

15

Specific charge//(Ah charge Ah LL-1-1-1) Charge volumique Ah

Fig. 3. Evolution of COD and ICE (inset) in function of specific charge for different organic compounds: (×) acetic acid, () isopropanol, ({) phenol, (Δ) 4-chlorophenol, (◊) 2-naphtol; i = 238 A m-2; T = 25 °C; Electrolyte: 1M H2SO4; The solid line represents model prediction.

passed during the galvanostatic oxidation of a 5 mM 2-naphtol in 1 M H2SO4 at different current densities (119-476 A m-2) is shown in Fig. 4b. As previously noted, an excellent agreement between the experimental and predicted values is observed. DEGRADATION OF ORGANICS BY AO ON BDD As mentioned above, thanks to its unique properties, during electrolysis in the region of water discharge, BDD anode produces a large quantity of •OH that is weakly adsorbed on its surface, and consequently it has high reactivity for organic oxidation, providing the possibility of efficient application to wastewater treatment. So far, many papers have reported that during electrolysis at BDD electrodes at high potentials, a

large number of organic pollutants are completely mineralised by the reaction with electrogenerated •OH radicals [12-34]. Some examples are summarised in Table 1. Comninellis and co-workers investigated the behaviour of Si/BDD anodes in an acidic solution for the oxidation of a wide range of pollutants, and they observed that complete mineralisation was obtained with all the experimental conditions studied. In particular, they found that the oxidation is controlled by the diffusion of the pollutants towards the electrode surface, where the hydroxyl radicals are produced, and the current efficiency is favoured by a high mass-transport coefficient, high organic concentration and low current density [35-42]. Performing electrolysis in optimised conditions, without diffusion limitation, the current efficiency approaches 100%. The electrochemical oxidation of different phenolic compounds (phenol, chlorophenols, nitrophenols) and carboxylic acids on BDD anodes was also studied by Canizares et al. [43-50]. They reported that the organic compounds were completely mineralised regardless of the characteristics of the wastewater (initial concentration, pH and supporting media) and operating conditions (temperature and current density) used. They also observed that the organics were oxi-

144


Table 1. Some examples of organic compounds oxidised on diamond electrodes. Pollutant Carboxylic acids Cyanides

Carwash wastewater Naphthol

Experimental conditions i = 30 mA cm-2; T = 30 °C; 1 M H2SO4 7.7 mM of cyanide, 0.05 M Na2SO4, pH of 11. i = 4-20 mA cm-2; initial TOC = 15 mg L-1 i = 15-60 mA cm-2 i = 15-60 mA cm-2; 1 M H2SO4

Triazines

i = 50 mA cm-2; impinging cell

Mixture of phenols

i = 30 mA cm-2; 0.1 M Na2CO3

Industrial wastewaters

i = 7-36 mA cm-2; initial COD 15008000 mg L-1 Amaranth, Basic Yellow, Reactive Black, Alizarin Red, Methyl Red, Orange II Chloromethylphenoxy herbicide, clofibric acid COD 6000 mg L-1, TOC 1600 mg L-1, pH = 6

Surfactants

Synthetic dyes

Herbicides Effluent of a fine chemical plant

dised on both the electrode surface by reaction with hydroxyl radicals and in the bulk of the solution by inorganic oxidants electrogenerated on the BDD anodes, such as hydrogen peroxide, peroxodisulfates, peroxodiphosphates or percarbonates. Some investigations have also tried to compare the behaviour of BDD with other electrodes, such as SnO2, PbO2, IrO2, for the oxidation of organic pollutants. Chen an co-workers [51] reported that the current efficiency obtained with Ti/BDD in oxidizing acetic acid, maleic acid, phenol, and dyes was 1.6-4.3fold higher than that obtained with the typical Ti/Sb2O5-SnO2 electrode. Other papers have demonstrated that Si/BDD electrodes are able to achieve faster oxidation and better incineration efficiency than Ti/PbO2 in the treatment of naphthol [18], 4chlorophenol [35], chloranilic acid [52] and synthetic dyes [53]. FUNDAMENTALS OF INDIRECT ELECTROOXIDATION METHODS WITH H2O2 ELECTROGENERATION Hydrogen peroxide is a green chemical that decomposes to O2 and water. It is a hazardous product due to its comburant character, easily being decomposed under the action of metallic ions, UVA light and high temperature [54,55]. For these reasons, the on-site production of H2O2 for environmental applications is desirable for avoiding its dangerous transport and storage. However, the direct treatment of wastewaters with H2O2 is limited by its low oxidation power since it can only attack reduced sulfur com-

Remarks Average current efficiency: 70-90% Energy consumptions of 20-70 kWh m−3 to remove 70-80% of initial cyanides Average current efficiency of 6% and 12% for anionic and cationic surfactants Initial current efficiency of about 40% Oxidation favoured by the formation of peroxodisulfuric acid The degradation of triazines follows a pseudo first order kinetic Model proposed for the degradation of mixture of organics Current efficiency of 85-100%

Ref. [12,13] [14]

Complete decolourization of the solutions

[24-29]

Complete TOC removal

[30-33]

Energy consumption of about 7 kWh m-3 for complete mineralization

[34]

[15,16] [17] [19] [20,21] [22] [23]

pounds, cyanides, chlorine and certain organics like aldehydes, formic acid and some nitro-organic and sulfo-organic compounds. It is known since 1882 that hydrogen peroxide can be produced from the cathodic reduction of dissolved O2 at high surface area carbon electrodes [55,56]. In the last years the electrogeneration of H2O2 in acid medium has been applied to wastewater remediation by means of indirect electro-oxidation methods. These techniques are based on the continuous supply of H2O2 to the contaminated solution from the two-electron reduction of O2 gas usually at carbon-felt [57-66] and carbon-polytetrafluoroethylene O2diffusion [65-76] cathodes: O2(g) + 2H+ + 2e− → H2O2

(30)

In an undivided electrolytic cell with a Pt [70] or BDD [67] anode, H2O2 is anodically oxidized to O2 yielding hydroperoxyl radical (HO2•) as intermediate, a much weaker oxidant than •OH: H2O2 → HO2• + H+ + e−

(31)

HO2• → O2 + H+ + e− (32) Hydrogen peroxide is then accumulated in the solution up to attain a steady concentration directly proportional to the applied current, just when the rates of reactions become equal. Direct application of this technique to wastewater treatment, so-called AO with electrogenerated H2O2 (AO-H2O2), involves the destruction of organics mainly with Pt(•OH) or BDD(•OH) formed on Pt or BDD, respectively. In the first case using a Pt/O2 cell, pollutants can also react more slowly with H2O2 and HO2• and in the second


one with a BDD/O2 cell, they can be oxidized with other weaker oxidants like ozone, peroxodisulfate, peroxodicarbonate or peroxodiphosphate competetively generated at the BDD anode depending on the electrolyte composition, as stated above for AO. The oxidation power of H2O2 can be strongly enhanced by using the EF method, where a small quantity of Fe2+ is added as catalyst to the acidic contaminated solution to generate Fe3+ and •OH from Fenton’s reaction [77]: (33) Fe2+ + H2O2 → Fe3+ + •OH + OH− Reaction 33 can be propagated due to the catalytic behaviour of the Fe3+/Fe2+ system. Thus, Fe2+ is continuously regenerated from the reduction of Fe3+ at the cathode by Eq. 34 or in the medium with electrogenerated H2O2 by Eq. 35, with hydroperoxyl radical (HO2•) by Eq. 36 and/or with organic radical intermediates R• by Eq. 37: Fe3+ + e− → Fe2+

(34)

Fe3+ + H2O2 → Fe2+ + H+ + HO2•

(35)

Fe3+ + HO2• → Fe2+ + H+ + O2

(36)

Fe + R• → Fe + R (37) However, a part of •OH produced in the medium is lost by oxidation of Fe2+ from Eq. 38 and by decomposition of H2O2 from Eq. 39: 3+

2+

+

Fe2+ + •OH → Fe3+ + OH−

(38)

H2O2 + •OH → HO2• + H2O

(39)

The rate of Fenton’s reaction can also decay by oxidation of Fe2+ in the bulk with HO2• by Eq. 40 and at the anode by Eq. 41 when an undivided cell is used: Fe2+ + HO2• → Fe3+ + HO2− −

(40)

Fe → Fe + e (41) The EF method in a BDD/O2 cell, for example, involves the simultaneous oxidation of pollutants with BDD(•OH) produced from Eq. 33 and by •OH formed from Eq. 41 (Fenton’s reaction), although a slower degradation with weaker oxidants (H2O2, HO2•, O3, S2O82−, etc.) is feasible. In the PEF process the solution treated under EF conditions is exposed to UVA illumination. The action of this irradiation is complex and can be accounted for by: (i) a greater production of •OH from photoreduction of Fe(OH)2+, the predominant Fe3+ species in acid medium [77], by Eq. 42 and (ii) the photodecarboxylation of complexes of Fe(III) with generated carboxylic acids, as shown in Eq. 43 for oxalic acid [78]. This acid is generated during the oxidation of most organics and the fast photolysis of Fe(III)-oxalate complexes (Fe(C2O4)+, Fe(C2O4)2− and Fe(C2O4)33−) favours the decontamination process [6871]. 2+

3+

145

Fe(OH)2+ + hν → Fe2+ + •OH

(42)

2Fe(C2O4)n(3-2n) + hν → 2Fe2+ + (2n-1)C2O42- + 2CO2

(43)

An interesting possibility in PEF is the use of sunlight as an inexpensive source of UVA light. This so-called SPEF can also enhance photolytic processes that are extended in the visible region, as in the case of Eq. 43 that occurs between 300 and 480 nm. DEGRADATION OF ORGANICS BY INDIRECT ELECTRO-OXIDATION METHODS USING A BDD/O2 CELL A small number of papers have been devoted to the destruction of aromatic pollutants by indirect methods with a BDD/O2 cell. Only some common herbicides [72,73], pharmaceuticals [65,66,76] and dyes [74] have been degraded in small undivided electrolytic cells containing a 3 cm2 BDD thin layer deposited on a conductive Si sheet from Centre Suisse d'Electronique et de Microtechnique S.A. and a 3 cm2 O2-diffusion cathode for H2O2 electrogeneration by reaction (30). The contaminated solutions contained 0.05 M Na2SO4 as background electrolyte and its pH was adjusted in the range 2.0-6.0 with H2SO4. Electrolyses were performed at constant current between 100 and 450 mA and temperature of 25 or 35 ºC. EF with a BDD anode (EF-BDD) treatments were carried out by adding 0.2-2 mM FeSO4 to initial solutions, whereas a 6 W UVA light was used to illuminate the solution in PEF-BDD. Figure 5a illustrates the comparative degradation behaviour of indirect methods when 100 mL of a 179 mg L-1 clofibric acid solutions of pH 3.0 were treated by AO-H2O2 with a BDD anode (AO-H2O2-BDD) without and with UVA irradiation, EF-BDD and PEFBDD at 100 mA cm-2 [20]. Their total organic carbon (TOC) always falls with applied specific charge (Q, in Ah L-1), although the time required for overall mineralization (tTM) depends on the procedure used. Both AO-H2O2-BDD methods yield a slow and similar TOC decay up to total decontamination at Q = 18 Ah L-1 (tTM = 6 h), as expected if all organics are destroyed by BDD(•OH) without significant photolysis by UVA light. In contrast, the destruction of organics is significantly accelerated in the early stages of the EF-BDD process due to their quicker reaction with the great amount of •OH formed from Fenton’s Eq. 33. In this method, however, less TOC is gradually removed with prolonging electrolysis due to the formation of complexes of Fe(III) with final oxalic acid that are hardly oxidized with BDD(•OH) and •OH, and then, the solution is totally mineralized at tTM = 6 h, similar to that AO-H2O2-BDD treatments. The fast photolysis of such complexes with UVA light causes the much quicker TOC removal by PEF-BDD, where solu-


146

120 (a)

TOC / mg TOC / mg L)-1l TOC (mg L-1

-1

100 80 60 40 20 0 0

3

6 9 12 15 -1) -1-1 Specific charge (Ah Specific charge / Ah L specific charge l

18

21

25 (b) 20 -1

-

[Cl-]] (mg / mgLL-1) [Cl

-1

/ mg ]l − [Cl

15 10 5 0 0

60

120

180 240 Time Time(min) / min

300

360

420

Fig. 5. (a) TOC abatement vs. specific charge and (b) chloride ion concentration vs. electrolysis time for the treatment of 100 mL of solutions containing 179 mg L-1 clofibric acid and 0.05 M Na2SO4 of pH 3.0 using a BDD/O2 cell at 100 mA cm-2 and 35.0 ºC by: ({) AO with electrogenerated H2O2 (AO-H2O2-BDD), (z) AO-H2O2-BDD with UVA irradiation, () electro-Fenton (EF-BDD) with 1.0 mM Fe2+, (S) photoelectro-Fenton (PEF-BDD) with 1.0 mM Fe2+ [76].

tion is decontaminated at Q = 12 Ah L-1 (tTM = 4 h). Mineralization of clofibric acid is accompanied by the release of chloride ion. As can be seen in Fig. 5b, this ion is accumulated and totally removed in 300-360 min in all cases, indicating that it is oxidized to Cl2 on BDD. The slow accumulation of Cl− in AOH2O2-BDD confirms the slow reaction of chloroorganics with BDD(•OH), whereas its much faster release in EF-BDD and PEF-BDD corroborates the quick destruction of these pollutants with •OH. Chlorophenoxy herbicides also show the same degradation behaviour rising the oxidation power of the methods in the sequence: AO-H2O2-BDD < EFBDD < PEF-BDD [72,73]. For all chloroaromatic pollutants, AO-BDD with and without generated H2O2 yields similar degradation rate, as expected by the small oxidation ability of this compound. In all indirect methods an increase in current causes a faster decontamination due to the production of more amounts

of oxidant radicals, but the specific charge for total mineralization becomes greater since larger proportions of reactive BDD(•OH) and •OH are wasted by non-oxidizing reactions involving for example, the AO of BDD(•OH) to O2 and recombination of •OH to H2O2. Moreover, a high current accelerates the formation of weaker oxidants like peroxodisulfate and ozone that also reduce the relative proportion of reactive BDD(•OH). The optimum conditions for EF-BDD and PEF-BDD treatments are found at pH 3.0, near the optimal pH of 2.8 for Fenton’s Eq. 33, and operating with 0.5-1.0 mM Fe2+. The comparative oxidation power of indirect electro-oxidation methods can be better explained from its mineralization current efficiency (MCE, in %), calculated for each treated solution at a given electrolysis time t (h) from the expression [75]: n F VS ΔTOC exp (44) MCE = 4.32 × 105 m I t where n is the number of electrons consumed in the mineralization process, Vs is the solution volume (L), Δ(TOC)exp is the experimental TOC decay (mg L-1), 4.32 × 105 is a conversion factor ( 3,600 s h-1 × 12,000 mg of C mol-1 / 100), m is the number of carbon atoms in the pollutant molecule and I is the applied current (A). Table 2 summarizes the MCE values thus determined for 4-chlorophenoxyacetic acid (4-CPA), 4chloro-2-methylphenoxyacetic acid (MCPA), 2,4dichlorophenoxyacetic acid (2,4-D) and 2,4,5trichlorophenoxyacetic acid (2,4,5-T) using Pt/O2 and BDD/O2 cells under comparable conditions. In the first system a 10 cm2 Pt sheet of 99.99% purity was used as anode and at 3 h of electrolysis efficiencies of 3-6% for AO-H2O2-Pt, 14-20% for EF-Pt and 25-29% for PEF-Pt, corresponding to 12-19, 53-66 and 9096% mineralization, are obtained. The first two methods with Pt do not allow total mineralization and are much less efficient than AO-H2O2-BDD and EF-BDD, respectively, which decontaminate completely all solutions practically at the same tTM. This confirms the much greater oxidation power of BDD than Pt, producing enough amount of reactive BDD(•OH) to destroy the initial contaminants and their by-products. Overall mineralization can only be attained by PEF-Pt due to the efficient photodecarboxylation of Fe(III)oxalate complexes as ultimate by-products under the action of UVA light [68-71]. The decay kinetics for the above herbicides was followed by reversed-phase HPLC chromatography showing that they undergo a pseudo-first-order reaction with the different hydroxyl radicals. The last column of Table 3 shows a similar pseudo-first-order rate constant (k1) for the AO-H2O2 methods, independent of the anode used, indicating that the initial chloroaromatics are removed at similar rate by BDD(•OH) and Pt(•OH). A much higher and similar k1-value can


147

Table 2. Percent of TOC removal and mineralization current efficiency (MCE) determined after 3 h of electrolysis of 100 mL of solutions of chlorophenoxyacetic acid herbicides with a concentration equivalent to 100 mg L-1 TOC at pH 3.0 by indirect electro-oxidation methods at 100 mA. The last column gives the rate constant determined for their pseudo-first-order decay with hydroxyl radical [68-72]. Herbicide 4-CPA

Methoda AO-H2O2-Pt AO-H2O2-BDD EF-Pt EF-BDD PEF-Pt

MCPA

AO-H2O2-Pt AO-H2O2-BDD EF-Pt EF-BDD PEF-Pt

2,4-D

AO-H2O2-Pt AO-H2O2-BDD EF-Pt EF-BDD PEF-Pt

T (ºC) 35

TOC removal (%) 19 54 66 75 96

MCE 5.6 16 20 22 29

k1 (s-1) 8.2×10-5 9.0×10-5 2.0×10-3 4.5×10-3 2.7×10-3

19 58 65 76 91

6.0 18 20 24 29

9.8×10-5 1.2×10-4 2.0×10-3 2.3×10-3 2.3×10-3

13 57 57 78 90

3.6 16 16 22 25

1.3×10-4 1.2×10-4 3.0×10-3 5.2×10-3 3.8×10-3

25 35 25 35 25

2.0×10-4 3.1 35 12 AO-H2O2-Pt 15 59 AO-H2O2-BDD 1.1×10-4 14 53 EF-Pt 1.8×10-3 21 80 EF-BDD 4.0×10-3 26 99 PEF-Pt 2.0×10-3 a AO-H2O2-Pt: anodic oxidation in a Pt/O2 cell, AO-H2O2-BDD: anodic oxidation in a BDD/O2 cell, EF-Pt: electro-Fenton with 1 mM Fe2+ in a Pt/O2 cell, EF-BDD: electro-Fenton with 1 mM Fe2+ in a BDD/O2 cell, PEF-Pt: photoelectro-Fenton with 1 mM Fe2+ in a Pt/O2 cell. 2,4,5-T

Table 3. Percent of TOC removal and mineralization current efficiency (MCE) after 2 h of treatment and time for total mineralization (tTM) and energy cost for total mineralization (ETM) of 2.5 L of p-cresol solutions of pH 3.0 by solar photoelectro-Fenton using a flow reactor with a 20 cm2 BDD anode and a 20 cm2 O2-diffusion cathode, at 30 ºC and liquid flow rate of 180 L h-1 [75]. [p-cresol]0 (mg L-1) 128

256 512 1024 a Not determined

[Fe2+]0 (mM) 0.25 1.0 0.25 1.0 0.25 1.0 1.0 1.0

Current density (mA cm-2) 25 25 50 50 100 50 50 50

TOC removal (%) 42 56 75 87 89 55 36 34

be seen for EF-Pt, EF-BDD and PEF-Pt, thus confirming that in these treatments the herbicides react much more rapidly with •OH. The greater oxidation ability of indirect methods using a BDD/O2 cell in comparison to a Pt/O2 is then due to the faster mineralization of final by-products with reactive BDD(•OH). GC-MS and HPLC analyses of chlorophenoxyacetic acid solutions electrolyzed in a BDD/O2 cell revealed the formation of primary phenol intermediates such as 4-chlorophenol for 4-CPA, 4-chloro-o-cresol

MCE (%) 119 158 104 118 60 148 195 373

tTM (min) 300 240 240 180 180 330 450 –a

ETM (kWh m-3) 8.3 6.6 20 15 51 27 37 –a

for MCPA, 2,4-dichlorophenol for 2,4-D and 2,4,5trichlorophenol for 2,4,5-T [72]. These species react rapidly with the same oxidant since they were uniquely detected while starting herbicides are destroyed. HPLC analysis of generated carboxylic acids revealed the large persistence of the ultimate oxalic acid in AO-H2O2-BDD or its Fe(III) complexes in EFBDD, since these species can only be converted to CO2 by BDD(•OH). This explains the long tTM for chloroaromatic pollutants in EF-BDD. In contrast, the


148

120

-1 -1 -1 TOC mg TOC TOC/ (mg / mgLLl )

100 80 60 40 20 0 0

2

4

6

8

10

-1 -1 -1

Specific charge / Ah specific charge / Ah Specific charge (Ah L L)l

Fig. 6. Dependence of TOC on specific charge for the mineralization of 100 mL of a 220 mg L-1 indigo carmine solution in 0.05 M Na2SO4 of pH 3.0 at 33 mA cm-2 and 35.0 ºC using a Pt/O2 or BDD/O2 cell [74]. (z) EF-Pt with 1.0 mM Fe2+, () PEF-Pt with 1.0 mM Fe2+, (¡) EF-BDD with 1.0 mM Fe2+, (S) PEF-Pt with 1.0 mM Fe2+ + 0.25 mM Cu2+. 120 (a) -1 -1 -1 [Oxalic mg [Oxamic acid] //(mg [oxalic acid] mg LL l )

100 80 60 40 20 0 60

(b) 50 -1 [Oxamic [Oxamicacid] acid](mg / mgLL-1)

-1

/ mg l [ox am ic aci d]

40 30 20 10 0

0

120

240

360

480

600

720

840

Time(min) / min Time

Fig. 7. Evolution of the concentration of: (a) oxalic acid and (b) oxamic acid during the degradation of a 220 mg L-1 indigo carmine solution under the conditions given in Fig. 6.

much quicker mineralization found for PEF-BDD (Fig. 5a) is related to the fast photolysis of Fe(III)-oxalate complexes [76]. The high oxidation power of the EF-BDD method has been confirmed in the comparative treat-

ment of acidic aqueous solutions containing up to 0.9 g L-1 of the dye indigo carmine by EF and PEF using BDD/O2 and Pt/O2 cells [74]. As illustrates by Fig. 6 for 220 mg L-1 of the dye at pH 3.0, the use of EF-Pt with 1.0 mM Fe2+ leads to a fast degradation up to attain 46% mineralization at Q = 3 Ah L-1 (3 h), but at longer time decontamination becomes so slow that TOC is only reduced by 49% at Q = 9 Ah L-1 (9 h). Surprisingly, the PEF-Pt method only yields a partial mineralization of 84% at the end of electrolysis, as expected if some Fe(III) complexes different to those oxalic acid can not be photodecomposed by UVA light. A quicker TOC destruction takes place in EFBDD with 1.0 mM Fe2+, where 91% of TOC is removed at 9 h and complete mineralization is reached at tTM = 13 h. That means that all complexes of Fe(III) with final carboxylic acids are efficiently oxidized with BDD(•OH). Figure 6 shows that overall destruction of all by-products is also feasible by PEF-Pt if 1.0 mM Fe2+ and 0.25 mM Cu2+ are combined as cocatalysts. For all procedures tested, a higher mineralization rate was found with increasing current density and initial dye content. The initial nitrogen of indigo carmine was mainly released as ammonium ion, along with a small fraction of nitrate ion. The indigo carmine decay always followed a pseudo-zero-order reaction. GC-MS and HPLC analyses of electrolyzed solutions evidenced its conversion into isatin-5-sulfonic acid, indigo and isatin, which are degraded to oxalic and oxamic acids. The different oxidation power of the EF and PEF treatments of the dye can be satisfactorily explained from the evolution of the Fe(III) and/or Cu(II) complexes of these final by-products, as can be seen in Figs. 7a and 7b. Thus, Fe(III)-oxalate and Fe(III)-oxamate remain stable in EF-Pt because they can not be oxidized either with Pt(•OH) or with •OH, but the latter complexes can be photolyzed under the action of UVA light in PEF-Pt. The opposite behaviour can be observed for EF-BDD where both complexes can be completely removed with BDD(•OH). When Cu2+ is used as co-catalyst in PEF-Pt, Cu(II)-oxalate and Cu(II)-oxamate are competitively produced and efficiently destroyed by Pt(•OH) and/or •OH. This evidences the positive catalytic action of combining Fe2+, Cu2+ and UVA light for the treatment of wastewaters with aromatic pollutants containing nitrogen if oxamic acid is formed as by-product. The studies carried out for the mineralization of aromatic compounds with a BDD/O2 cell have clearly demonstrated the fastest removal of all pollutants by PEF-BDD. Recently, the alternative use of sunlight in this method has been assessed using a flow plant with a filter-press electrochemical cell coupled to a solar photoreactor [67,75]. This system treated 2.5 L of a contaminated solution operating in batch and under steady conditions. The liquid circulated continuously through the BDD/O2 cell containing electrodes of 20


E TM =

V I tTM VS

(45)

where V is the average cell voltage (V). An inspection of Table 3 shows that ETM decreases strongly as less current density is applied due to the drop in cell voltage and the gradual increase in MCE. This parameter falls with rising Fe2+ content from 0.25 to 1.0 mM due to the faster photodecarboxylation of Fe(III)-oxalate complexes, but increases with rising initial pollutant

120 (a) 100 80 60 40 -1 TOC TOC (mg/ mg L-1)L

cm2 area and the solar photoreactor, which was a polycarbonate box (irradiated volume 600 mL) with a mirror at the bottom and inclined 30º from the horizontal to collect better the direct sun rays. Figure 8a presents the TOC-time plots obtained for the degradation of 128 mg L-1 of o-cresol, m-cresol or p-cresol (corresponding to 100 mg L-1 TOC), 0.25 mM Fe2+ and 0.05 M Na2SO4 of pH 3.0 in the flow plant by EF-BDD and SPEF-BDD at 50 mA cm-2 [75]. A fast destruction of the o-cresol solution for 120 min of EF-BDD with 50% TOC reduction can be observed, but at longer time its mineralization becomes so slow that at 180 min only 54% decontamination is attained due to the hard oxidation of final Fe(III)-oxalate complexes with BDD(•OH) and/or •OH. In contrast, the same o-cresol solution, as well as the m-cresol and pcresol solutions, are much more rapidly degraded by SPEF-BDD, being completely decontaminated at tTM =180 min (Q = 1.20 Ah L-1) due to the quick and efficient photodecarboxylation of Fe(III)-oxalate complexes by the incident UVA light supplied by solar irradiation. Figure 8b evidences a gradual rise in degradation rate of the p-cresol solution with increasing current density of 25, 50 and 100 mA cm-2, yielding total mineralization at decreasing tTM of 240, 180 and 160 min, respectively, corresponding to increasing Q of 0.80, 1.20 and 1.41 Ah L-1. Accordingly, Table 3 shows a higher percent of TOC removal at 2 h of electrolysis, along with the concomitant fall in efficiency, for these experiments. This corroborates the greater production of more amounts of BDD(•OH) and •OH when current density rises, but with an acceleration of their non-oxidizing reactions in larger extent. Results of Table 3 also reveal an enhancement of the mineralization as the contaminant content rises up to 1 g L-1, as expected if the waste reactions of the above oxidants are gradually less significant. A MCE value as high as ca. 400% is found after 2 h treatment of the most concentrated solution, indicating the high efficiency of sunlight to photolyze final Fe(III)-oxalate complexes. In addition, p-cresol solutions are more rapidly destroyed with 1.0 mM Fe2+ than with 0.25 mM Fe2+ by the quicker photolysis of great amounts of such complexes. The energy cost for total mineralization (ETM, in kWh m-3) of p-cresol solutions at time tTM (h) in the flow plant was determined as follows:

149

20 0

0

30

60

90

120

150

180

210

120 (b) 100 80 60 40 20 0

0

30

60

90

120

150

180

210

240

270

Time /(min) min Time

Fig. 8. TOC decay with electrolysis time for the degradation in a flow reactor of 2.5 L of solutions with 128 mg L-1 of cresols and 0.25 mM Fe2+ in 0.05 M Na2SO4 at pH 3.0, 30 ºC and liquid flow rate of 180 l h-1 [75]. In plot (a), () EF-BDD of o-cresol and SPEF-BDD of ({) ocresol, () m-cresol, (S) p-cresol at 50 mA cm-2. In plot (b), SPEF-BDD treatment of p-cresol at (z) 25 mA cm-2, (S) 50 mA cm-2, (¡) 100 mA cm-2.

concentration because longer tTM is needed for greater contents of p-cresol. The lowest ETM value of 6.6 kWh m-3 is found when treating a 128 mg L-1 p-cresol solution with 1.0 mM Fe2+ at 25 mA cm-2. These results confirm the high efficiency and low energy cost required for the degradation of cresols by SPEF-BDD, making this technique viable for industrial application. CONCLUSIONS This paper summarizes recent researches carried out on electrochemical oxidation of organic pollutants for wastewater treatment using diamond electrodes and indirect electro-oxidation methods with a BDD/O2 cell. During electrolysis in the potential region of water discharge, BDD anodes involve the production of weakly adsorbed hydroxyl radicals that unselectively and completely mineralise organic pollutants with high current efficiency. The oxidation rate on BDD is independent on the chemical nature of organic pollutants but depends only on the operating conditions. In particular higher

150


current efficiency, near 100%, were obtained at high organic concentration and low current density. A theoretical kinetic model of organic mineralization on boron-doped diamond anodes has been presented for the prediction of the evolution of chemical oxygen demand and current efficiency during organic oxidation. The experimental verification of the model at several levels (influence of the nature of the organic pollutants, organic concentration and applied current density) shows excellent agreement with theory in all investigated cases. In AO with electrogenerated H2O2, organics are effectively destroyed by BDD(•OH). The EF-BDD treatment removes more rapidly the aromatic pollutants because they mainly react with •OH formed from Fenton’s reaction. However, complexes of Fe(III) with final carboxylic acids like oxalic and oxamic can only be slowly oxidized with BDD(•OH). The PEFBDD method is more efficient due to the parallel and quicker photodecarboxylation of final Fe(III)-oxalate complexes by UVA light. When oxamic acid is produced, its Fe(III) complexes can not be photolyzed and the use of Cu2+ as co-catalyst enhances the mineralization process since Cu(II)-oxamate complexes are quickly removed with hydroxyl radicals. SPEF can be a viable technique for the treatment of industrial wastewaters due to its high efficiency and low operational cost. REFERENCES 1. Chen, G., Electrochemical technologies in wastewater treatment. Sep. Purif. Technol., 38(1), 11-41 (2004). 2. Foti, G., D. Gandini and C. Comninellis, Anodic oxidation of organics on thermally prepared oxide electrodes. Curr. Top. Electrochem., 5, 71-91 (1997). 3. Simond, O., V. Schaller and C. Comninellis, Theoretical model for the anodic oxidation of organics on metal oxide electrodes. Electrochim. Acta, 42(13-14), 2009-2012 (1997). 4. Comninellis, C., Electrocatalysis in the electrochemical conversion/combustion of organic pollutants for waste water treatment. Electrochim. Acta, 39(11-12), 1857-1862 (1994). 5. Martin, H.B., A. Argoitia, U. Landau, A.B. Anderson and J.C. Angus, Hydrogen and oxygen evolution on boron-doped diamond electrodes. J. Electrochem. Soc., 143(6), L133-L136 (1996). 6. Ramesham, R. and M.F. Rose, Electrochemical characterization of doped and undoped CVD diamond deposited by microwave plasma. Diam. Relat. Mater., 6(1), 17-27 (1997). 7. Swain, G.M., The use of CVD diamond thin film in electrochemical system. Adv. Mater., 6(5), 388-

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