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
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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)
Panizza et al.: BDD Electrodes for Wastewater Treatment
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
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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
Panizza et al.: BDD Electrodes for Wastewater Treatment
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-
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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
Panizza et al.: BDD Electrodes for Wastewater Treatment
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
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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
Panizza et al.: BDD Electrodes for Wastewater Treatment
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
J. Environ. Eng. Manage., 18(3), 139-153 (2008)
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.
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Discussions of this paper may appear in the discussion section of a future issue. All discussions should
be submitted to the Editor-in-Chief within six months
of publication.
Manuscript Received: November 6, 2007
Revision Received: December 23, 2007
and Accepted: December 30, 2007