J. Electrochem. Sci. Eng. 4(4) 2014

ISSN: 1847-9286 Open Access Journal www.jese-online.org

Journal of Electrochemical

Science and Engineering

J. Electrochem. Sci. Eng. 4(4) 2014, 135-326

Volume 4 (2014) No. 04 pp. 135-326

IAPC

http://www.jese-online.org/

J. Electrochem. Sci. Eng. 4(4) (2014) 135-236 Published: December 6, 2014

Open Access : : ISSN 1847-9286

www.jESE-online.org

Contents

Mahbobeh Moazampour, Fahimeh Tahernejad-Javazmi, Maryam Salimi-Amiri, Hassan Karimi-Maleh, Mehdi Hatami Voltammetric determination of hydroxylamine in water and waste water samples using a NiO nanoparticle/new catechol derivative modified carbon paste electrode .............................................. 135

Abel I. Balbín Tamayo, Ana M. Esteva Guas, Juan J. Piña Leyte-Vidal Marcelo Maccini Analytical method for heavy metal determination in algae and turtle eggs from Guanahacabibes Protected Sea Park ............................................................................................................................. 145

Irena Ciglenečki, Marija Marguš, Elvira Bura-Nakić, Ivana Milanović Electroanalytical methods in characterization of sulfur species in aqueous environment ..................... 155

Ana Carolina O. Santana, Erica F. Southgate, João Paulo B. G. Mendes, Jo Dweck, Eliana Mosse Alhadeff, Ninoska Isabel Bojorge Ramirez Characterization of an hrp-aox-polyaniline-graphite composite biosensor ........................................... 165

Paul-Cristinel Verestiuc, Igor Cretescu, Oana-Maria Tucaliuc, Iuliana-Gabriela Breaban, Gheorghe Nemtoi Voltammetric determination of hydroxylamine in water and waste water samples using a NiO nanoparticle/new catechol derivative modified carbon paste electrode .............................................. 177

Ramakrishnan Kamaraj, Pandian Ganesan, Subramanyan Vasudevan Use of hydrous titanium dioxide as potential sorbent for the removal of manganese from water......... 187

Annabel Fernandes, Catarina Oliveira, Maria J Pacheco, Lurdes Ciríaco, Ana Lopes Anodic oxidation of oxytetracycline: Influence of the experimental conditions on the degradation rate and mechanism .............................. 203

Marijana Kraljić Roković, Mario Čubrić, Ozren Wittine Phenolic compounds removal from mimosa tannin model water and olive mill wastewater by energy-efficient electrocoagulation process ........................................................................................ 215

Mani Nandhini, Balasubramanian Suchithra, Ramanujam Saravanathamizhan, Dhakshinamoorthy Gnana Prakash Optimization of parameters for dye removal by electro-oxidation using Taguchi Design ..................... 227

Camilo González-Vargas, Ricardo Salazar, Ignasi Sirés Electrochemical treatment of Acid Red 1 by electro-Fenton and photoelectro-Fenton processes .......... 235

María I. León, Zaira G. Aguilar, José L. Nava Electrochemical combustion of indigo at ternary oxide coated titanium anodes .................................. 247


Jéssica Pires de Paiva Barreto, Elisama Vieira dos Santos, Mariana Medeiros Oliveira, Djalma Ribeiro da Silva, João Fernandes de Souza, Carlos A. Martínez-Huitle Electrochemical mediated oxidation of phenol using Ti/IrO2 and Ti/Pt-SnO2-Sb2O5 electrodes ............. 259

Djamel Ghernaout, Abdulaziz Ibraheem Al-Ghonamy, Mohamed Wahib Naceur, Noureddine Ait Messaoudene, Mohamed Aichouni Influence of operating parameters on electrocoagulation of C.I. disperse yellow 3............................... 271

Carlos E. Barrera-Díaz, Gabriela Roa-Morales, Patricia Balderas Hernández, Carmen María Fernandez-Marchante, Manuel Andrés Rodrigo Enhanced electrocoagulation: New approaches to improve the electrochemical process (Review) ....... 285

Mohammed A. Karim Electrokinetics and soil decontamination: concepts and overview (Review) ......................................... 297

Anil N. Ghadge, Mypati Sreemannarayana, Narcis Duteanu, Makarand M. Ghangrekar Influence of ceramic separator’s characteristics on microbial fuel cell performance ............................. 315

J. Electrochem. Sci. Eng. 4(4) (2014)


www.jESE-online.org

EDITORIAL Special Issue on New achievements and methodologies of electrochemistry and electrochemical engineering in the environmental protection and pollution control

During the last decades, many applications of Electrochemistry and Electrochemical Engineering

have arisen for the characterization and remediation of environmental problems. As a result, now-

adays this subject has become one of the most interesting areas of research in applied

electrochemistry, with hundreds of papers published every year and many applications already

available in the market. This special issue contains sixteen very valuable contributions on these

topics, written by highly recognized authors and covering the most relevant areas of interest

within the topic.

Environmental monitoring is a matter of the major importance because it helps to prevent and

remediate pollution with the development of novel warning detection systems. For this reason,

the first sets of contributions are related to characterization of environmental issues with electro-

chemical methods and it contains valuable information about new tools for the characterization of

organics, heavy metals and sulphur.

Treatment of industrial wastes is one of the more stimulating environmental applications

nowadays. Water is extensively used in industry not only as a heat exchanger fluid or a cleaning

agent, but also for the production of many chemicals. As a consequence, significant volumes of

wastewater are produced every day in our industries and they get into the environment after their

treatment with technologies which are not always completely effective. An electrochemically-

based solution to this problem is faced in this special issue with exciting contributions on

electrolysis, electro-Fenton oxidation and electrocoagulation of wastewater, in which technologies

for the efficient removal of dyes, persistent chemicals and inorganic pollutants are evaluated.

Finally, the last set of papers included in this special issue focusses on soil remediation and bio-

electrochemical treatments. Electrokinetic soil remediation (EKSR) is one of the most motivating

topics of research for electrochemical and environmental engineering in our time. Many

applications are currently working at the full scale and in this issue, an authoritative review is

included, in which the fundamentals and applications of the technology are clearly described.


J. Electrochem. Sci. Eng. 4(4) (2014) EDITORIAL

To conclude, trying to save energy, one of the more exciting and innovative areas of research is

the production of electricity from bio-electrochemical processes. Research on this topic is still at a

very early stage but results are promising and the concept of producing energy directly from waste

is an out breaking idea as it is explained in the last contribution of this special issue.

As a conclusion, this special issue is a very good summary of the most exciting research on

electrochemistry and electrochemical engineering in the environmental protection and pollution

control and, for sure, it will become a reference for many researchers in the near future.

Manuel Andrés Rodrigo Rodrigo

doi: 10.5599/jese.2013.0049 135

J. Electrochem. Sci. Eng. 4(4) (2014) 135-144; doi: 10.5599/jese.2014.0049

Open Access: ISSN 1847-9286

www.jESE-online.org

Original scientific paper

Voltammetric determination of hydroxylamine in water and waste water samples using a NiO nanoparticle/new catechol derivative modified carbon paste electrode

Mahbobeh Moazampour, Fahimeh Tahernejad-Javazmi, Maryam Salimi-Amiri*, Hassan Karimi-Maleh and Mehdi Hatami**

Department of Chemistry, Graduate University of Advanced Technology, Kerman, Iran *Department of Physics, Sari Branch, Islamic Azad University, Sari, Iran **Polymer Research Laboratory, University of Bonab, Bonab, Iran

Corresponding author: E-mail: [email protected] Tel.: +989112540112

Received: February 21, 2014; Revised: March 22, 2014; Published: December 6, 2014

Abstract A (9,10-dihydro-9,10-ethanoanthracene-11,12-dicarboximido)-4-ethylbenzene-1,2-diol (DED) mo-dified NiO/NPs carbon paste electrode “(DED/NiO nanoparticle (NiO/NPs)/CPE) was constructed for determination of hydroxylamine (HX). The cyclic voltammogram showed that the electro-catalytic oxidation of HX at the surface of DED/NiO/NPs/CPE occurs at a potential of about 800 mV less positive than with an unmodified electrode. Square-wave voltammetry results presented that the electrocatalytic oxidation peak currents of HX in pH 8.0 had two linear dynamic ranges in the range of 0.1 to 2.0 and 2.0 to 400.0 µM HX, with a detection limit of 0.07 µM. The kinetic

parameters such as electron transfer coefficient (0.47) and rate constant (2.454 × 103 M-1 s-1) were determined for the chemical reaction between HX and DED. Finally, this method was evaluated for the determination of HX in water and waste water samples.

Keywords Hydroxylamine; NiO nanoparticle; water and waste water analysis; sensor; voltammetry

Introduction

Hydroxylamine (HX) is known as a type of reducing agent and is widely used in industrial and

pharmaceutical applications. It has been identified as a key intermediate in nitrogen cycles and

nitrous oxide production [1]. The quantitative determination of HX is very important in both

studies of biological processes and for industrial purposes. It has been confirmed that HX is

produced during the reduction of nitrates by Escherichia coli and Torula yeast [2].


mailto:[email protected]

J. Electrochem. Sci. Eng. 4(4) (2014) 1350-144 VOLTAMMETRIC DETERMINATION OF HYDROXYLAMINE

136

Electrochemical analysis is gaining significance within industrial process control, environmental

monitoring and various pharmaceutical and biotechnology applications [3-7]. The use of

unmodified electrodes for electrochemical detection has a number of limitations, such as low

selectivity and sensitivity, poor reproducibility, slow electron transfer reaction, low stability over a

wide range of solution compositions and the high overpotential at which the electron transfer

process occurs [8-10]. Chemical modification of inert substrate electrodes with redox active thin

films offers significant advantages in the design and development of electrochemical sensors. In

operation, the redox active sites shuttle electrons between the analyte and the electrodes with a

significant reduction in activation overpotential [11]. A further advantage of chemically modified

electrodes is that they are less prone to surface fouling and oxide formation, compared to inert

substrate electrodes [12-14]. A wide variety of compounds have been used as electron transfer

mediators for the modification of electrode surfaces in various procedures [15-17].

Nanotechnology has become one of the most interesting disciplines in science and technology

today. The intense interest in nanotechnology is being driven by various interesting fields and is

creating a new industrial revolution [18]. Nano-materials such as nanoparticles, carbon nanotubes

or nanocomposite connected with biomolecules are being used for several bioanalytical

applications [19-21]. Electroanalysis is taking advantage of all the possibilities offered by

nanomaterials that are easy to detect using conventional electrochemical methods.

Nanocomposite of a variety of shapes, sizes and compositions continues to change the field of

bioanalytical measurement.

In the present work, we describe the preparation and suitability of a DED modified NiO/NPs

carbon paste electrode as a new electrode for electrocatalysis and determination of HX in an

aqueous buffer solution. To demonstrate the catalytic ability of the modified electrode toward the

electrooxidation of HX in real samples, we examined the utility of this method for the

voltammetric determination of HX in water and waste water samples.

Experiment

Chemicals

All chemicals used were of analytical reagent grade purchased from Merck (Darmstadt,

Germany), unless otherwise stated. Doubly distilled water was used throughout.

1.0×10–2 mol L–1 HX solution was prepared daily by dissolving 0.064 g HX in water and the

solution was diluted to 100 mL with water in a 100 mL volumetric flask. The solution was kept in a

refrigerator at 4oC and in the dark. Further dilution was made with water.

Phosphate buffer solutions (sodium dihydrogen phosphate and disodium monohydrogen

phosphate, plus sodium hydroxide, 0.1 mol L–1) (PBS) with different pH values were used.

High viscosity paraffin (d = 0.88 kg L–1) from Merck was used as the pasting liquid for the

preparation of the carbon paste electrode. Spectrally pure graphite powder (particle size <50 µm)

from Merck was used as the substrate for the preparation of the carbon paste electrode as a

working electrode.

Apparatus

Cyclic voltammetry (CV), chronoamperometry and square wave voltammetry (SWV) were

performed using an analytical system, Autolab, with PGSTAT 302N (Eco Chemie, The Netherlands).

The system was run on a PC using GPES software. A conventional three-electrode cell assembly

consisting of a platinum wire as an auxiliary electrode and an Ag/AgCl (KClsat) electrode as a

S. Kaushal at al. J. Electrochem. Sci. Eng. 4(4) (2014) 1350-144

doi: 10.5599/jese.2014.0049 137

reference electrode was used. The working electrode was either an unmodified carbon paste

electrode (CPE) or a DED/NiO/NPs/CPE. X-ray powder diffraction studies were carried out using a

STOE diffractometer with Cu–K radiation (l = 1.54 Å).

Preparation of the modified electrode

To prepare the modified electrode, 150.0 mg of NiO/NPs and 70.0 mg of DED was hand mixed

with 780.0 mg of graphite powder using a mortar and pestle. Using a syringe, 15 drops of paraffin

were added to the mixture and mixed well for 55 min until a uniformly wetted paste was obtained.

The paste was then packed into a glass tube. By pushing a copper wire down the glass tube into

the back of the mixture, electrical contact was created. When necessary, a new surface was

obtained by pushing an excess of the paste out of the tube and polishing it on weighing paper. The

unmodified carbon paste electrode (CPE) was prepared in the same way without NiO/NPs and DED

to the mixture, to be used for comparison purposes.

Preparation of real samples

Water samples were stored in a refrigerator immediately after collection. Ten millilitres of the

sample was centrifuged for 15 min at 1500 rpm. The supernatant was filtered using a 0.45 µm

filter and then diluted three times with the PBS pH 8.0. The solution was transferred into the

voltammetric cell to be analysed without any further pre-treatment. The standard addition

method was used for the determination of HX in real samples.

Results and discussion

NiO/NPs characterisation

NiO/NPs were analysed by XRD analyses. The XRD pattern of NiO/NPs nanopowders in the 2

range of 10-80° is shown in Fig. 1.

Figure 1. XRD patterns of as-synthesised NiO/NPs nanoparticles.


138

Figure 1 clearly proves the presence of NiO/NPs. An average diameter of as-synthesised

NiO/NPs was calculated from the broadness peak (2 = 44°) by using the Scherrer equation

D = Kλ/ cos , and measured about 25.0 nm.

Electrochemical investigation

Figure 2 depicts the cyclic voltammetry responses from the electrochemical oxidation of 400

µM HX at DED/NiO/NPs/CPE (curve c), DED/CPE (curve b), NiO/NPs/CPE (curve d) and unmodified

CPE (curve e). As shown, the anodic peak potential for HX oxidation at DED/NiO/NPs/CPE (curve c)

and at DED/CPE (curve b) was about 200 mV, while at NiO/NPs/CPE (curve d); the peak potential

was about 1000 mV. At the unmodified CPE, the peak potential of HX was about 1050 mV

(curve e). From these results, it was concluded that the best electrocatalytic effect for HX

oxidation was observed at DED/NiO/NPs/CPE (curve c).

Figure 2. Cyclic voltammograms of (a) the buffer solution at DED/NiO/NPs/CPE; (b) 400 µM HX at DED/CPE; (c) 400 µM HX at DED/NiO/NPs/CPE; (d) 400. µM HX at NiO/NPs/CPE; (e) 400 µM

HX at CPE. Conditions: 0.1 mol L-1 PBS (pH 8.0), scan rate of 20 mV s-1.

For example, the results show that the peak potential of HX oxidation at DED/NiO/NPs/CPE

(curve c) shifted by about 800 and 850 mV toward less positive values when compared with

NiO/NPs/CPE (curve d) and unmodified CPE (curve e), respectively. Additionally, DED/NiO/NPs/CPE

showed higher anodic peak current for the oxidation of HX compared to DED/CPE, indicating that

the combination of NiO/NPs and the mediator significantly improved the performance of the

electrode toward HX oxidation. In fact, DED/NiO/NPs/CPE in the absence of HX exhibited a well-

behaved redox reaction (Figure 2a) in the buffer solution (pH 8.0). However, there was a drastic

increase in the anodic peak current in the presence of 400 µM HX (curve c), which can be related

to the electrocatalytic role of DED/NiO/NPs/CPE towards oxidation of HX.

We observed a linear variation of the peak current with the square root of scan rate (ν1/2) at

scan rates ranging from 2-12 mV s–1 at pH 8.0 (Figure 3). This result clearly indicates a diffusion-

controlled electrooxidation process [22].


doi: 10.5599/jese.2014.0049 139

Figure 3 Plot of Ipa versus ν1/2 for the oxidation of 100.0 µM HX at various scan rates of (a) 2.0; (b) 3.0; (c) 5.0; (d) 8.0; (e) 12.0 mV s−1 in 0.1 mol L−1 phosphate buffer solution (pH 8.0) at

NiO/NPs/DED/CPE. Inset: Cyclic voltammograms of 100.0 μM HX at various scans.

To obtain information about the rate-determining step, a Tafel plot was drawn, derived from

points in the Tafel region of the linear sweep voltammogram (Figure 4). The slope of the Tafel plot

was equal to n(1−α)F/2.3RT, which resulted in 0.1115 V decade-1 [23]. Therefore, we obtained the

value of α being equal to 0.47.

For further investigations, the value of α was calculated for the oxidation of HX at pH 8.0 for

both the modified and unmodified paste electrodes using one other method (see equation 1):

αnα = 0.048/(EP–EP/2) (1)

where EP/2 is the potential corresponding to IP/2. The values for αnα were found to be 0.47 and 0.12

at the surface of DED/NiO/NPs/CPE and CPE, respectively. This result was also confirmed by the

larger Ipa values recorded during linear seep voltammetry at DED/NiO/NPs/CPE.

Chronoamperometric measurements of HX at DED/NiO/NPs/CPE were carried out for various

concentrations of HX in buffered aqueous solutions (pH 8.0) by setting the working electrode

potentials at 0.0 mV and 400 mV vs. Ag/AgCl/KClsat (Figure 5A). For an electroactive material (HX,

in this case) with a diffusion coefficient of D, the current observed for the electrochemical reaction

at the mass transport limited condition was described using the Cottrell equation. Experimental

plots of I vs. t-1/2 were employed, with the best fits for 300 µM of HX (Figure 5B). The slope of the

resulting straight line was then plotted against HX concentration (Figure 5B). From the resulting

slope and Cottrell equation, the mean value of the D was found to be 2.1×10−6 cm2 s-1 [24].


140

Figure 4. Tafel plot for DED/NiO/NPs/CPE in 0.1 mol L−1 PBS (pH 8.0) with a scan rate of 8 in the

presence of 100.0 µM HX.

Figure 5. A – Chronoamperograms obtained at DED/NiO/NPs/CPE (a) in the absence

and in the presence of (b) 300 μM HX at pH 8.0; B – Cottrell’s plot for the data from the chronoamperograms; C – Dependence of IC/IL on t1/2 derived from the chronoamperogram data;

D – The charge-time curves (a′) for curve (a) and (b′) for curve (b).

The rate constant for the chemical reaction between HX and redox sites in DED/NiO/NPs/CPE,

kh, can be evaluated by chronoamperometry according to the method set out by Galus [25]:

IC/IL = π1/2 γ1/2 = π1/2 (kCbt)1/2 (2)

Where IC is the catalytic current of HX at DED/NiO/NPs/CPE, IL the limited current in the

absence of HX and t is the time elapsed (s). The above equation can be used to calculate the rate


doi: 10.5599/jese.2014.0049 141

constant of the catalytic process kh. Based on the slope of the IC/IL versus t1/2 plots (Figure 5C), kh

can be obtained for a given HX concentration. From the values of the slopes, an average value of

kh was found to be kh = 2.454×103 M–1 s–1. The value of kh also explains the sharp feature of the

catalytic peak observed for catalytic oxidation of HX at the surface of DED/NiO/NPs/CPE.

Double potential step chronocoulometry, as well as other electrochemical methods, was in

addition employed for the investigation of the electrode processes at DED/NiO/NPs/CPE. Forward

and backward potential step chronocoulometry on the modified electrode in a blank buffer

solution showed very symmetrical chronocoulograms. These had about an equal charge consumed

for both oxidation and reduction of the DEDRed/DEDOx redox system in DED/NiO/NPs/CPE.

However, in the presence of HX, the charge value associated with forward chronocoulometry was

significantly greater than that observed for the backward chronocoulometry (Figure 5D). This

behaviour is typical of that expected for electrocatalysis at a chemically modified electrode [26].

Stability and reproducibility

The repeatability and stability of modified electrode was investigated using CV measurements

of 400.0 µM HX in a buffer solution. The relative standard deviation (RSD) for five successive

assays was 1.4 %. When seven different DED/NiO/NPs/CPEs were used, the RSD for ten

measurements was 2.1 %. When the electrode was stored in the laboratory, the modified

electrode retained 95 % of its initial response after a week and 92 % after 30 days (see Figure 6).

These results indicate that DED/NiO/NPs/CPE has good stability and reproducibility, and could be

used for HX measurements.

Determination of HX individually

Square wave voltammetry (SWV) was used to determine the concentration of HX. Since square

wave voltammetry has a much higher current sensitivity and better resolution than cyclic voltam-

metry, the SWV was used for the determination of HX (Figure 7 inset). The plot of peak current vs.

the HX concentration consisted of two linear segments with slopes of 5.9035 and 0.1498 µA/µM at

the concentration ranges of 0.1-2.0 µM and 2.0-400.0 µM, respectively (Fig. 7).

Figure 6. Cyclic voltammograms of 300 μM HX at a surface of DED/NiO/NPs/CPE in a 0.1 mol

L−1 phosphate buffer solution (pH 8.0) at different times.


142

The decreasing of sensitivity (slope) of the second linear segment was likely due to kinetic

limitation. The detection limit was determined as 0.07 µM for HX based on YLOD = YB+3σ.

Interference study and real sample studies

Analytical selectivity was one of the important parameters that affected the accuracy of the

analysis. In order to evaluate the selectivity of the proposed method for the determination of HX,

the influence of various substances as potentially interfering compounds, which can be present in

the water and waste water samples with the determination of HX, were studied under optimum

conditions with 1.0 µM HX at pH 8.0. The tolerance limit was taken as the maximum concentration

of the foreign substances, which caused an approximate 3% relative error (in potential or current)

in the determination. The result of interfering studied for some of the various substances in

oxidation current and oxidation peak potential of HX showed that 1000-fold of Ni2+, CN–, Ca2+,

Mg2+, Mn+2, K+, Na+, Cl- and SCN–, 800-fold of glucose, sucrose, lactose and fructose did not affect

the selectivity.

Figure 7. The plots of the electrocatalytic peak current as a function of HX concentration. Inset

shows the SWVs of DED/NiO/NPs/CPE in a 0.1 mol L−1 phosphate buffer solution (pH 8.0) containing different concentrations of HX. From bottom-up corresponds to 0.0, 0.1, 0.5, 1.0,

2.0, 50.0, 100.0, 150.0, 200.0, 250.0, 300.0, 350.0 and 400.0 μM of HX.

In order to demonstrate the applicability of the new sensor in determining HX in real samples,

we used the new sensor in determining HX in tap water, river water, wastewater and well water.

The determinations of HX in samples were carried out using the standard addition method

(Table 1). The accuracy of the method was examined by comparing the results obtained from this

method with published methods for the determination of HX [2]. The results from the statistical

calculation indicated good agreement between them for the mean values (t-test) and the precision

(F-test) in the determination of HX in real samples for the three analyses. It was clear that a

modified electrode was capable of voltammetric determination of HX, with high selectivity and

good reproducibility.


doi: 10.5599/jese.2014.0049 143

Table 1: Determination of HX in water samples (n=3).

Sample Added, µM Founded, µM Published method, µM Fex Ftab tex ttab

Tap water 1.00 1.05±0.10 0.98±0.11 6.3 19.0 1.2 3.8 Well water 5.0 4.85±0.16 5.22±0.32 7.8 19.0 1.5 3.8 River watera 10.0 10.45±0.65 10.55±0.76 9.6 19.0 2.1 3.8 Waste water 30.0 29.73±0.75 30.75±1.2 13.5 19.0 3.2 3.8 a

Tejan River, Sari, Iran Fex - calculated F value; Ftab - reported F value from F-test table with 95 % confidence level and 2/2 degree of freedom; tex - calculated t; ttab - reported t value from student t-test table with 98 % confidence level.

Conclusion

A carbon paste electrode modified with NiO/NPs and DED was used for electrocatalytic

determination of HX. The results showed that the oxidation of HX was catalysed at pH 8.0,

whereas the peak potential of hydrazine was shifted by 800 mV to a less positive potential at the

surface of DED/NiO/NPs/CPE. In addition, it was shown that HX can be determined using the

square wave voltammetry technique. The detection limit (3σ) was 0.07 according to the SWV

method. The kinetic parameters, such as electron transfer coefficient, α (0.47) and a rate constant

for the chemical reaction between HX and redox sites in DED/NiO/NPs/CPE, kh (2.454×103 M–1 s–1)

were also determined using electrochemical approaches. Finally, the electrocatalytic oxidation of

HX at the surface of this modified electrode can be employed as a new method for the

voltammetric determination of HX in real samples such as tap water, river water, wastewater and

well water.

Acknowledgements: The authors wish to thank Gradate University of Advanced Technology, for their support.

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© 2014 by the authors; licensee IAPC, Zagreb, Croatia. This article is an open-access article distributed under the terms and conditions of the Creative Commons Attribution license

(http://creativecommons.org/licenses/by/4.0/)

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doi: 10.5599/jese.2014.0051 145



www.jESE-online.org


Analytical method for heavy metal determination in algae and turtle eggs from Guanahacabibes Protected Sea Park

Abel I. Balbín Tamayo, Ana M. Esteva Guas, Juan J. Piña Leyte-Vidal and Marcelo Maccini*

Department of Analytical Chemistry, Faculty of Chemistry, University of Havana, Havana, 10400, Cuba *Department of Food Science, University of Teramo, 64023, Teramo, Italy

Corresponding author: E-mail: [email protected]; Tel.: +53 873 82 22

Received: February 4, 2014; Revised: April 4, 2014; Published: December 6, 2014

Abstract A standard digestion method coupled to electrochemical detection for the monitoring of heavy metals in biological samples has been used for the simultaneous analysis of the target analytes. Square wave anodic stripping voltammetry (SWASV) coupled to disposable screen-printed electrodes (SPEs) was employed as a fast and sensitive electroanalytical method for the detection of heavy metals. The aim of our study was to determine Cd, Pb and Cu by SWASV in brown algae (Sargasum natan) and green turtle eggs (Chelonia mydas) using screen-printed electrodes. The method proved useful for the simultaneous analysis of these metals by comparison between two different procedures for preparing the samples. Two different approaches in digestion protocols were assessed. The study was focused on Guanahacabibes brown algae and green turtle eggs because the metal concentrations recorded in this area may be used for intraspecific comparison within the Guanahacabibes Protected Sea Park area, a body of water for which information is still very scarce. The best results were obtained by digesting biological samples with the EPA 3050B method. This treatment allowed the fast and quantitative extraction from brown algae and green turtle eggs of the target analytes, with high sensitivity and avoiding organic residues, eventually affecting electrochemical measurements.

Keywords Cadmium; copper; lead; Sargasum Natan; Chelonia Mydas eggs; square-wave anodic stripping

Introduction

Marine contaminations by anthropogenic chemicals pose one of the worst problems to coastal

and estuarine ecosystems around the word. Certain heavy metals have gained great significance in

chemical and toxicological studies of the environment. Among those heavy metals are Cd and Pb,



J. Electrochem. Sci. Eng. 4(4) (2014) 145-154 HEAVY METAL DETERMINE IN ALGAE AND TURTLES EGGS

146

which are generally toxic even at very low levels, and potentially toxic metals, e.g. Cu which also

has indispensable essential properties with different threshold levels in different types of plants

and organisms, including man. Therefore, the evaluation of heavy metal concentrations in marine

organisms constitutes an important area of research [1-4].

The use of marine organisms (algae, turtles eggs, fish, etc.) as bioindicators to trace metal

pollution is very common these days [5-7].

Macroalgae are able to accumulate trace metals, reaching concentrations that are thousands of

times higher than the corresponding concentrations in sea water. Algae accumulate only free

metal ions, the concentrations of which depend on the nature of suspended particulate matter [8],

which, in turn, is formed by both organic and inorganic complexes. Moreover, algae satisfy all of

the basic requirements of bioindicators: they are sedentary, their dimensions are suitable, they

are easy to identify and collect, they are widely distributed, and they accumulate metals to a

satisfactory degree [9].

On the other hand, many investigations have reported the accumulation of heavy metals in

marine sea turtle having a long lifespan and occupying high trophic levels in the marine food

chains, and showed the utility of this specie as a biological indicator of heavy metal pollution. The

intentional killing of any living sea turtle is prohibited, except for research purposes, for which only

very limited samples are available. Hence, it is possible to estimate the concentration of heavy

metals in the tissue of nesting female sea turtles by using their eggs [10-15]. For that reason, the

eggs are a useful non-lethal indicator for monitoring heavy metals in the body of sea turtles.

A wet-digestion procedure can be applied to all types of biological materials. In this procedure,

small amounts of nitric and perchloric acids are added to the sample material. The overall

reliability of the digestion method will follow the adequate mineralisation of samples, i.e. the

levels of the heavy metals. If any metals were linked in their insoluble form, they are not of

relevance for pollution control [16].

The digestion method involves the liberation of the analyte (metal) of interest from an

interfering matrix using a reagent (mineral/oxidising acids or fusion flux) and/or heat. The

utilisation of reagents (acids) and external heat sources can then cause problems. In elemental

analysis, these problems are particularly focused on the risk of contamination and loss of

analytes [17-19].

Considering the low content of heavy metals in environmental samples, sensitive analytical

methods are required. The heavy metal determination in organic samples can usually be carried

out by atomic spectrometry: inductively-coupled plasma optical emission (ICP-OES) or

electrothermal and flame atomic absorption spectrometry (ETAAS and FAAS), although the

detection limits are not sufficient when the concentrations are too low. However, many pre-

concentration techniques have been employed for analysing complex matrices and samples with

low levels of metals. Hence, square-wave anodic stripping voltammetry (SWASV) includes a pre-

concentration step in situ; for this reason, this is an electroanalytical technique used for the

analysis of traces metals in solution [20-27]. Such a combination of an effective accumulation step

with an advanced measurement procedure results in a very low detection limit, and makes

stripping analysis one of the most important techniques in trace analysis.

The coupling of disposable screen-printed electrodes with stripping techniques is a revolution in

comparison with conventional stripping analysis: the design and operation are greatly simplified,

in accordance with the requirements of a decentralised assay. The greater proportion of articles

A. I. Balbín Tamayo at al. J. Electrochem. Sci. Eng. 4(4) (2014) 145-154

doi: 10.5599/jese.2014.0051 147

have utilised the technique of stripping voltammetry, gaining detection limits in the low ng/mL

(ppb) region [28].

Screen-printed electrodes are planar devices realised by printing layers of different

electroconductive and insulating inks with controlled thickness and shapes on a plastic substrate.

In this work, the carbon surface of the screen-printed working electrode was employed as a

substrate for a thin mercury film (TMF) [29].

The aim of our study was to apply a digestion method (EPA 3050B) to determine Cd, Pb and Cu

by square-wave anodic stripping voltammetry in brown algae (Sargasum natan) and green turtle

(Chelonia mydas) eggs, using a screen-printed electrode, and it demonstrated the usefulness of

this method for the simultaneous analysis of Cd, Pb and Cu by comparison between two different

procedures for preparing the samples.

Experimental

Collecting and treatment of samples

The study area was located in Antonio beach Guanahacabibes Protected Sea Park. This site is

situated in an area characterised by low anthropogenic activity [30].

The sample of brown algae (Sargasum natan) was handpicked in the subtidal zone at a depth of

about 2-3 m. Care was taken to choose the sample to ensure that all were at a similar stage of

development. The samples were washed in seawater at the sampling site and transferred to the

laboratory in pre-cleaned polyethylene bags under refrigerated conditions. Upon arrival at the

laboratory, they were thoroughly cleaned and any sediment was carefully removed with nylon

brushes under tap water for a few seconds. Algal material was rapidly rinsed in deionised water

(Milli-Q, Millipore Corp) to minimise any possible metal loss during the procedure and was then

pulverised. Finally, the samples were frozen and stored (4 °C) until analysis.

The green turtle (Chelonia mydas) egg samples were collected in the nesting area of this

species. All samples were stored at 4 °C until chemical analysis, and then the eggshell, the

albumen and the yolk were subsequently separated. The separation was carried out quickly to

prevent thawing.

Samples were digested by two separate digestion procedures in order to select the simplest,

which in turn would provide suitable analytical results:

a) General acid digestion: A 1 g dried sample was placed in a Teflon beaker; the acid digestion

reagent (concentrated HNO3) was added and the mixture was allowed to stand overnight.

The sample was heated until the production of red NO2 fumes had ceased. This mixture was

digested via the addition of HClO4 and was heated until it had evaporated to a small volume.

The samples were brought to an appropriate volume with a dilute acid solution (0.01 mol L-1

HCl).

b) The method EPA 3050B [17,18,31] was used to produce a transparent solution. This is a very

strong acid digestion that will dissolve almost all elements that could become

“environmentally available”. For the digestion of samples, a representative 1 g (dry weight)

sample was digested with the repeated addition of nitric acid (HNO3) and hydrogen peroxide

(H2O2). The resultant solutions were diluted to a known volume with 0.01 mol L-1 HCl.

For each analytical batch of samples processed, blanks were carried throughout the entire

sample preparation and analytical process. These blanks will be useful for determining whether

samples are contaminated, and are necessary to provide a realistic estimate of interferences that

could be encountered in the analysis of test samples.


148

Heavy metal determination

All experiments were carried out using a PalmSens portable electrochemical analyser (Palmsens

BV, Houten, The Netherlands). The conditions for square wave voltammetry striping onto a screen-

printed electrode of carbon modified by plated Hg films were:

Cd(II), Pb(II) and Cu(II) analysis: conditioning potential (Econd) - 0.3 V for 60 s, deposition

potential (Edep) − 1.0 V for 300 s, equilibration time (teq) 30 s, SW amplitude (Eamp) 28 mV, step

potential (Estep) 3 mV, frequency (f) 15 Hz.

Electrodes were serigraphically screen-printed with a shape similar to that reported by Palchet-

ti [29]. They consisted of a round-shaped working electrode (diameter 3 mm), a graphite counter

electrode and a silver pseudo-reference electrode. In addition, the silver electrical contacts were

covered by a graphite layer in order to prevent oxidation phenomena during storage.

Graphite-based Hg-modified screen-printed electrodes were used as the working electrode.

These are based on the use of a special coating cellulose-derivative film deposited onto the

graphite working electrodes containing a Hg(II) salt, as reported by Meucci [32]. Hg(II) is reduced

from the salt to the metallic form and the modified sensor can be then used for heavy metal

accumulation and stripping. The use of this strategy allows the use of large amounts of Hg

solutions to be avoided, whilst retaining the high sensitivity which characterises mercury-coated

electrodes [32].

Each sensor was pre-treated in 0.1 mol L-1 HCl before being used for the first time, by applying

ten cycles of square wave voltammetry (SWV) using the following conditions: potential initial 1 V,

potential final 0 V, scan rate 50 mV s-1, SW amplitude (Eamp) 28 mV, step potential (Estep) 3 mV,

frequency (f) 15 Hz. This step is necessary to obtain a stable baseline.

Then, 0.1 mol L-1 HCl was used as the supporting electrolyte. All measurements were performed

without removing oxygen from the solution. The measurements were performed by immersing

the sensor in 5.0 ml of solution, with magnetic stirring during the conditioning and accumulation

steps, whereas the square wave scan was performed without stirring.

Suprapure grade hydrochloric acid was purchased from Merck. The water used for the

preparation of solutions was from a Milli-Q System (Millipore). The working standard solution of

Cd, Pb and Cu was prepared by diluting standard 1 g L-1 metal solutions with 0.01 mol L-1 HCl.

Statistical analysis

For the statistical treatment, the experimental results followed the recommendations proposed

by Miller [33]. Determinations of means, standard deviations, coefficients of variation and

percentage recovered were performed using statistical software.

LOD: The limit of detection, expressed as the concentration cL, or the quantity qL, is derived

from the smallest measure xL, that can be detected with reasonable certainty for a given analytical

procedure. The value of xL is given by equation (1):

xL = xbl + ksbl 1

LOQ: The lowest concentration of an analyte that can be determined with acceptable precision

(repeatability) and accuracy under the stated conditions of the test.

The ability to quantify is generally expressed in terms of the signal or analyte (true) value that

will produce estimates with a specified relative standard deviation (RSD), which is commonly 10%.


doi: 10.5599/jese.2014.0051 149

Results and Discussion

When measurements are made at low analyte levels, e.g. in trace analysis, it is important to

determine the lowest concentration of the analyte or property value that can be confidently

detected by the method, and the lowest concentration of analyte that can be determined with an

acceptable level of repeatability, precision and trueness. The importance in determining this, and

the problems associated with it, arises from the fact that the probability of detection does not

suddenly change from zero to unity as some threshold is crossed. The detection and quantification

limits for the general acid digestion and EPA 3050B by square wave voltammetry striping methods

are shown in Table 1.

With these procedures for preparing the samples, tiny, clear, well-separated signals

corresponding to the different metals were recorded by SWASV; no matrix effect and reproducible

peaks and linear standard addition plots were observed in digested reagent blanks.

The mean calculated detection limits method (based on three times the standard deviation of

the blank signal) and quantification limits (based on ten times the standard deviation of the blank

signal) for Cd, Pb and Cu showed a marked improvement over those reported by Wang, Locatelli

and Palchetti [25,34,35] .

Taking into account the low detection limits, quantification limits and coefficient of variation

(CV) in Table 1, the general acid digestion and EPA 3050B using square wave voltammetry anodic

striping methods give good estimations for the metals analysed.

Table 1. Detection limits, quantification limits and coefficient of variation for Cd, Pb and Cu for general acid digestion and method EPA 3050B by square wave voltammetry anodic striping.

Metal General acid digestion Method EPA 3050B

LOD, 10-4 µg/gdry LOQ, 10-4 µg/gdry CV, % LOD, 10-4 µg/gdry) LOQ, 10-4 µg/gdry CV, %

Cd 12.5 15 1.6 13 19 1.5

Pb 310 350 3.2 150 200 6.6

Cu 210 240 4.7 118 130 3.4

Determination of heavy metals in brown algae (Sargasum natan) and green turtle (Chelonia mydas) eggs

All metal contents reported in this work refer to the initial dry mass. Mean metal

concentrations are reported as values with standard deviations. Cd, Pb and Cu concentrations in

brown algae and green turtle eggs are shown in Table 2 and 3 for the general acid digestion and

EPA 3050B methods, respectively. The standard deviations of pooled samples refer to the

variability within different replicates.

For these procedures for preparing the samples, SWASV recorded tiny, clear, well-separated

signals corresponding to the different metals (Figs. 1-4); no matrix effect, reproducible peaks and

linear standard addition plots were observed in the digested biological matrix.

Different concentrations of Cd, Pb and Cu were used to perform linear regression analysis for

the utilised screen-printed electrodes. The linear regression analysis, generated by plotting the

height of the peaks obtained for each concentration, gave the following equations:

General acid digestion of brown algae

for Cd: ip = 0.24 + 10.7 cCd, for Pb: ip = 6 + 383 cPb for Cu: ip = 10.8 + 356 cCu

General acid digestion of green turtle eggs


150

for Cd: ip = 0.198 + 9 cCd, for Pb: ip = 3.22 + 42 cPb for Cu: ip = 3.5 + 44 cCu

EPA 3050B of brown algae

for Cd: ip = 0.43 + 16 cCd, for Pb: ip = 3.38 + 227 cPb for Cu: ip = 3.95 + 59 cCu

EPA 3050B of green turtle eggs

for Cd: ip = 0.27 + 6.6 cCd, for Pb: ip = 7.64 + 84 cPb for Cu: ip = 7.9 + 88 cCu

E / V vs. Ag/AgCl

Figure 1. Signals corresponding to standard addition the different metals concentration to brown algae samples digest by method General acid digestion. +0.02 ppm and +0.04 ppm

of multistandard of Cd(II), Pb(II), Cu (II)

E / V vs. Ag/AgCl

Figure 2. Signals corresponding to standard addition the different metals concentration to green turtles eggs samples digest by method General acid digestion +0.01 pm and +0.02 ppm of

multistandard of Cd(II), Pb(II), Cu (II).

i p /

A

i p

/

A


doi: 10.5599/jese.2014.0051 151

E / V vs. Ag/AgCl

Figure 3. Signals corresponding to standard addition the different metals concentration to brown algae samples digest by method EPA 3050B. +0.01 ppm and +0.02 ppm of multistandard

of Cd(II), Pb(II), Cu (II)

E / V vs. Ag/AgCl

Figure 4. Signals corresponding to standard addition the different metals concentration to green turtles eggs samples digest by method EPA 3050B +0.02 ppm and +0.04 ppm of multistandard of

Cd(II), Pb(II), Cu (II)

In all cases, linearity ranging from 10 ppm to 30 ppm was obtained with a correlation ≥ 0.98 for

all of the metals analysed.

The values determined in the brown alga and turtle eggs generally had low values of standard

deviation and a coefficient of variation of 5 % for both moist digestion techniques used to digest

the brown seaweed and green turtle eggs; this is indicative of analytical quality results.

i p /

A

i p

/

A


152

The samples were analysed by the method under validation both in its original state and after

the addition (spiking) of a known mass of the analyte to the test sample. In the absence of

reference materials, bias was investigated by spiking and recovery [22,35].

Spiking/recovery studies are very strongly subjective; the recoveries that are significantly

different from unity indicate that bias is affecting the method. Better spiking/recovery data were

obtained by the EPA 3050B method, even though the poor recovery by general acid digestion was

certainly an indication of a lack of trueness.

The variation between the spiking/recovery data using different digestion methodologies may

not only be due to volatility during digestion but could also be linked to the way in which these

elements are attached to the biological matrix.

On the other hand, the bioaccumulation of these metals in brown algae is influenced by

numerous factors: pH, ligand concentration and type, and various sediment components [36-38].

Trace metal concentrations reported in this study were of the same order of magnitude as those

measured by other authors in uncontaminated sites [36,39]. These analysed metals in brown algae

showed low anthropogenic activity.

Table 2. Concentration of Heavy metals in brown algae (Sargasum natan) and in green turtle eggs (Chelonia mydas) by general acid digestion

Metal

Brown algae (Sargasum natan) Green turtle eggs (Chelonia mydas)

Means ± SD, µg/gdry

CV, % Recuperation surrogate

recovery, %* Means ± SD,

µg/gdry CV, %

Recuperation surrogate recovery, %*

Cd 1.16 ± 0.06 5.0 89 2.2 ± 0.1 4.5 89

Pb 0.79 ± 0.01 5.6 92 7.6 ± 0.2 2.6 90

Cu 1.52 ± 0.01 1.0 78 7.7 ± 0.2 2.6 100

* Recoveries for standard solutions are calculated by dividing the observed value by the expected value. The result is multiplied by 100 to give a percent recovery

Table 3. Concentration of Heavy metals in brown algae (Sargasum natan) and in green turtle eggs (Chelonia mydas) by EPA 3050b

Metal

Brown algae (Sargasum natan) Green turtle eggs (Chelonia mydas)

Means ± SD, µg/gdry

CV, % Recuperation surrogate

recovery, %* Means ± SD,

µg/gdry CV, %

Recuperation surrogate recovery, %*

Cd 0.528 ± 0.004 2.9 94.3 4.6±0.2 4.3 100

Pb 1.49 ± 0.01 4.3 100.0 8.9±0.3 3.1 101

Cu 6.7 ± 0.1 1.5 100.0 8.7±0.2 1.8 101

* Recoveries for standard solutions are calculated by dividing the observed value by the expected value. The result is multiplied by 100 to give a percent recovery

Green turtles are herbivorous and feed on macroalgae, occupying a trophic level lower than

carnivorous turtles. In this specie, Cu plays a crucial role in oxygen transport, energy production

and enzyme activity, and it can be freely transferred from the mother to the egg [10]; however,

although Cd and Pb do not perform any known role in biological systems, limited amounts were

transferred, which suggests the relation of metal concentrations in egg. Trace metal

concentrations in turtle eggs reported in this study were the largest compared to those measured

by other authors in sea turtles [10,12,15]. We hypothesise that the concentrations of Cd, Pb and


doi: 10.5599/jese.2014.0051 153

Cu in turtle eggs could depend mainly on their feeding habits, as already suggested by other

authors [10,40]. In addition to diet composition, age and gender could be important factors

affecting metal excretion in egg.

Conclusions

Digestion techniques studied for the treatment of samples suggest that the EPA3050B

technique should be used for the simultaneous analysis of Cd, Pb and Cu in in brown algae and sea

turtle eggs by anodic stripping voltammetry with square wave, as it showed the highest

percentage recovery values with low coefficients of variation. The concentrations in brown algae

(Sargasum natan) confirm that pollution in Antonio beach Guanahacabibes Protected Sea Park

area is low, while the concentration in green turtle (Chelonia mydas) eggs suggest that it depends

on their feeding habitats within the Caribbean Sea.

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www.jESE-online.org


Electroanalytical methods in characterization of sulfur species in aqueous environment

Irena Ciglenečki, Marija Marguš, Elvira Bura-Nakić and Ivana Milanović

Division for Marine and Environmental Research, Ruđer Bošković Institute, Bijenička 54, 10 000 Zagreb, Croatia

Corresponding author: E-mail: [email protected]; Tel: 0038514561105; Fax: 0038514680242

Received: April 10, 2014; Revised: June 13, 2014; Published: December 6, 2014

Abstract Electroanalytical (voltammetric, polarographic, chronoamperometric) methods on an Hg electrode were applied for studying of different sulfur compounds in model and natural water systems (anoxic lakes, waste water, rain precipitation, sea-aerosols). In all investigated samples typical HgS reduction voltammetric peak, characteristic for many different reduced sulfur species (RSS: sulfide, elemental sulfur, polysulfide, labile metal sulfide and organosulfur species) was recorded at about -0.6 V vs. Ag/AgCl reference electrode. In addition, in anoxic waters which are enriched with sulfide and iron species, voltammetric peaks characteristic for the presence of free Fe(II) and FeS nanoparticles (NPs) were recorded at -1.4 V and around -0.45 V, respectively. Depending on the used electroanalytical method and experimental conditions (varying deposition potential, varying time of oxidative and/or reductive accumulation, sample pretreatment i.e. acidification followed by purging) it is possible to distinguish between different sulfur species. This work clearly shows a large potential of the electrochemistry as a powerful analytical technique for screening water quality regarding presence of different reduced sulfur species and their speciation between dissolved and colloidal/nanoparticle phases.

Keywords Voltammetry; chronoamperometry; speciation; reduced sulfur species; metal sulfide nanoparticles; Hg electrode; anoxic water samples

Electrochemical measurements are along with ICP-MS, the most used but challenging approaches

in essential elements analysis and speciation in complex natural samples. There is a wide range of

electroanalytical techniques for qualitative and quantitative determination of essential and poten-

tially toxic elements in natural waters [1,2]. Some examples include: potentiometry, polarography,

voltammetry, chronopotentiometry, chronoamperometry, etc. These electrochemical methods,

especially voltammetry, have appropriate features to be used as monitoring methods (early

warning tools) for assessment of water quality in aqueous systems in general and will be key



J. Electrochem. Sci. Eng. 4(4) (2014) 155-163 ELECTROANALYSIS IN SULFUR SPECIATION

156

methods for trace pollutant analyses (sulfur species [3-17], organic compounds [18-20], trace

metals [2-4, 8, 21-26], engineered and natural nanoparticles [27-34]).

Working electrodes, so called voltammetric sensors, have many embodiments that make them

specific for detection of above listed natural and anthropogenically introduced compounds in

natural environment, enabling their quantitative determination. Electrochemical techniques offer

increasing degree of accuracy, decreasing detection limits, simplicity, prompt response, ect. It

involves dramatically lower costs than other techniques to reach same sensibility and with

automated, portable instrumentation is suitable for fieldwork. In addition, many substances that

are analyzed by other techniques use electrochemical detectors.

Voltammetry is the only technique allowing speciation and determination of the truly dissolved

metal species without many sample handling [2,21-26]. Speciation of a metal affects its

biogeochemical cycling processes and its biological impacts. Thus, electrochemical measurements

in natural waters are essential in order to obtain more complete speciation information and to

fully understand the geochemical cycling and bioavailability (toxicity) of trace metals.

EU water quality guidelines are searching for new innovative methods for water quality mon-

itoring, and electrochemistry in comparison with Inductively Coupled Plasma Mass Spectrometry

(ICP-MS) and/or Inductively coupled plasma/optical emission spectrometry (ICP-OES) and diffusive

gradients in thin-films (DGT) approach was found as preferable choice. Besides, new investigations

showed that voltammetry has a potential to be used in determination of metal NPs, metal sulfide

(MS) NPs and aquatic colloids in natural waters [27-32]. Growing evidence implies that MS NPs of

natural and anthropogenic origin exist in aquatic environments. These NPs could play important

role as mediators of the trace metal nutrition and toxicity. Using different electrochemical

methods it is possible to measure a variety of soluble and particulate sulfur compounds [3-17,32].

In this work voltammetric, polarographic and chronoamperometric measurements on a Hg

electrode were used for characterization and speciation of dissolved and particulate sulfur species,

including thiols, HS-, S0, MS NPs (FeS, PbS), Sx2- in different contrasting aqueous natural samples

such as oxic/anoxic systems, rain precipitation and aerosols.

Experimental

Materials

All chemicals used were reagent grade and were not further purified. Stock solutions of sulfide,

polysulfide, suspensions containing NPs of FeS and PbS were prepared as previously described [6,

7, 10-14, 32]. All measurements were performed in NaCl (Chemica, Croatia) electrolyte solutions

with ionic strengths ranging from 0.11 to 0.55 M NaCl. In some experiments the NaCl electrolyte

was buffered with 0.03 M NaHCO3 (Chemica, Croatia)..

Instrumentation

Electrochemical measurements were performed with a BAS-100-A chemical analyser, µ-Autolab

Electrochemical Instruments (Eco Chemie) and PGSTAT 128 N (Metrohm, Switzerland) connected

to pencil like HMDE and 663 VA Stand Metrohm Electrode (Metrohm, Switzerland) as a working

electrodes, respectively. The reference electrode was an Ag/AgCl (3 M KCl) electrode connected to

the solution via an electrolyte bridge, and a platinum electrode served as an auxiliary electrode.

Reduced sulfur species (RSS) were determined by linear sweep and cyclic voltammetry (LSV, CV)

[6,7,13] and by polarographic measurements [3] in fresh nonfiltered samples. In the case of CV

and LSV the accumulation (ta = 0-120 s) of RSS on the Hg electrode surface with stirring was

I. Ciglenečki at al. J. Electrochem. Sci. Eng. 4(4) (2014) 155-163

doi: 10.5599/jese.2014.0053 157

performed at the deposition potential of E =-0.20 V (vs. Ag/AgCl). After accumulation the potential

was shifted in the negative direction (to E= -1.70 V vs. Ag/ AgCl) with a scan rate of 100 mV/s and

HgS reduction peak at around -0.6 V, characteristic of many RSS were recorded [6,7,13]. In the

same cycle reduction peaks characteristics for the presence of metal sulfide layers and NPs from

the bulk of the solution were recorded at potential more negative than -0.6 V [10,11,17]. Next, the

solution was acidified with 30 μL of concentrated HCl (Chemica, Croatia) to pH ~2 and purged for 5

min. After restoring the original pH with NaOH (Chemica, Croatia) the accumulation and scan steps

were repeated. The result of the first measurement, prior to acidification, is assigned as total redu-

ced sulfur species, RSST = H2S/HS- + S0 and the result of the second measurement is assigned to

elemental sulfur, S0 as model representative for non-volatile reduced sulfur species, RSSNV [6,7,13].

For detection of S0 and S2- presence in polysulfide (Alfa Aesar, USA) containing solutions

sampled DC polarography (SDC) or voltammetry at the Hg electrode was performed with step

potential of 0.0051 V, starting from -0.4 V (vs. Ag/AgCl) and shifting to more negative values.

In chronoamperometry the detection potential at which current was measured as a function of

time (I-t curves) was selected depending on the potential at which reduction of the NPs from the

bulk of the solution is proceeding [11,17,32]. In the case of PbS the used potential was -1.5 V. The

scan lasted for 30 s and the sampling time was 0.1 s. The suspension of PbS NPs was prepared by

mixing the equimolar concentrations of Pb2+ and HS- directly in the electrochemical cell [11].

During the ageing process the suspension was not stirred and recorded changes in the NPs sizes

were only due to aggregation caused by Brownian motion.

Results and discusion

Typical voltammetric signal which can be found in an anoxic sulfide rich environment is

presented in Fig. 1. The obtained peak, usually in our papers designated as C2, represents the well-

known dissolution/reduction of HgS layer on the Hg electrode surface [5-7,9-17]:

E / V (vs. Ag/AgCl)

-1,0-0,8-0,6-0,4-0,2

I /

nA

-120

-100

-80

-60

C21 2

E / V (vs. Ag/AgCl)

-1,0-0,8-0,6-0,4-0,2

I /

mA

-30

-20

-10

0

C21

2

a b

Figure 1. LSV obtained from Rogoznica Lake water in the oxic (a) and anoxic bottom water layer (b), before 1) and 2) after acidification and purging with N2; (E = -0.2 V, ta = 120 s). The C2 peak increases with

either sulfide or with S0 addition and corresponds to 6.5 nM RSSNV in a) and to mM RSSv in b. The shift of this peak to a more negative potential after the acid-purge-base treatment is due to a final pH which is higher

than the original pH. Carbonate buffering in the sample is destroyed by acidification, so it is difficult to return the sample exactly to the original pH.


158

HgS + H+ + 2e-→ HS- + Hg0 (1)

This peak usually is taken as a measure for “free” and labile sulfur species content (H2S/HS-/S2-, S0,

SnS2-, thiols, labile metal sulfide complexes and nanoparticles). In oxic water layers, in addition to

C2, the peak at more positive potentials than -0.5 V can be frequently revealed. This peak usually

corresponds to the presence of different organosulfur species (DMS, 3-mercaptopropionat, thio-

compounds) which at used experimental conditions oxidize the Hg electrode but do not deposit

HgS layer on its surface, therefore their peak appears more positively than C2 [9].

In cases when sample solution contains sulfide and metal ions (M2+) which are present in an

excess, depending on the electrochemical conditions (deposition potential, accumulation time)

[15,17], the peak marked as C3 (Figure 2) appears. This peak corresponds to formation of metal

sulfide (MS) deposit (layers), in the given case PbS, due to electrochemical exchange reaction

between Hg2+ from a HgS layer and the free M2+ (Pb2+) ion from the solution [15,17]:

HgSlayer + M2+ + 2e- MSlayer + Hg0 (2)

C3 might be easily misrepresented for the dissolved organosulfur species represented by C2

peak in Figure 1.

The MS layer stays on the Hg surface without desorption up to potentials that are more

negative than C2 peak (usually up to -1.6 V and more negative potentials) [14,17,32]. Reduction of

the MS (PbS) layers formed at C3 peak potential usually occurs at potentials of the peak C5

(Figure 2, reaction 3). The potentials of both electrode reactions, and the formation and reduction

of the MS layer, are shown to be directly controlled by the MS solubility products [15,17].

PbSlayer +2e- + H+ → Pb0 + HS- (3)

Figure 2. CV of solution containing 4x10-5 M Pb2+ and 3x10-5 M HS- in 0.55 M NaCl/0.03 M NaHCO3 electrolyte (E= -0.2 V; ta=60 s; v=100 mV/s).

C4/A4 peaks in the Figure 2, correspond to reduction/oxidation of the free metal on the Hg, i.e.

to the reduction of Pb2+ to Pb0, while C6 in accordance with our previous work was ascribed to

reduction of PbS nanoparticles (NPs) from the bulk of the studied solutions [10,11,17]. These

particles usually do not form MS layers. In the case of studied CdS, PbS, Ag2S, Cu2S, HgS

E / V (vs. Ag/AgCl)

-1,6-1,4-1,2-1,0-0,8-0,6-0,4-0,2

I / nA

-50

0

50

100

150

4.0x10-5

M Pb2+

and 3.0x10-5

M HS-

C2C4

A4

C3

C5C6


doi: 10.5599/jese.2014.0053 159

suspensions, depending on the solution conditions (concentration and ratio between metal and

sulfide species, ionic strength, pH) larger NPs will form and result in the appearance of the peaks

similar to recorded C6 reduction peak. Direct reduction of the formed NPs, which is placed more

negative than reduction process of the relevant MS layers [17] and/or reduction processes which

occur on the NPs surface upon collision with the Hg electrode, and the potential where this

process is occurring on the Hg surface is successfully used as a background for further NPs

characterization by chronoamperometric measurements [32], as shown here later.

In samples of anoxic seawater lake Rogoznica Lake (Croatia), shown in Figure 1, the peak at -0.5

V corresponds to the presence of organosulfur species (RSSNV) which do not deposit HgS, and peak

at -0.6 V corresponds to presence of RSStotal (all RSS that deposit HgS). The major difference

between oxic and anoxic Rogoznica lake water layers is in the existence of volatile sulfide species

(RSSV) which are present in mM concentration in anoxic part mainly in the form of sulfide (HS-) and

RSSNV which presence is determined to be around 10 nM in oxic and 1-10 µM in anoxic water

layers. The RSSv can be removed by acidification and purging while nonvolatile species during

acidification and purging procedure will remain in the sample and contribute to the C2 peak

(Figure 1b) [9,12,13,16].

With use of polarographic measurements on the Hg electrode it is possible to distinguish

further between detected RSS on the polysulfide (Sx2-), elemental S (S0) and/or HS- species without

any pretreatment of the samples [3] (Figure 3).

Figure 3. SDC polarographic curves recorded in: a) 0.55 M NaCl solution containing 7 x 10-6 M S4

2- and b) anoxic sample od Rogoznica lake taken at 11 m depth. The voltammograms were recorded

between –0.4 and -0.8 V vs. Ag/AgCl with potential steps of 1 mV.

The voltammograms measured in the tetrasulfide containing solution (Figure 3a) were charac-

terized by the anodic and cathodic currents in the potential range from –0.5 V to –0.8 V, respect-

tively. The cathodic current is assigned to a reduction process given by the equation (4) [3,33,34],

and it is a measure for the presence of S0 in the polysulfide molecule:

22S 2 1 e H O HS OHn n n n n (4)

Anodic current recorded at potentials more positive than –0.6 V is assigned to the well-known

oxidation of the Hg by HS- according to equation (5) [3-17,33], and it can be taken as a measure for

the S2- presence in the molecule of polysulfide:

HS Hg HgS 2e H (5)

E / V (vs. Ag/AgCl)

-0,8-0,7-0,6-0,5

I / nA

-20

-15

-10

-5

0

5

3x S0

1x S2- 7 x 10-6 M S42-

electrolyte

a

E / V (vs. Ag/AgCl)

-0,9-0,8-0,7-0,6-0,5-0,4-0,3

I / µ

A

-0,2

-0,1

0,0

0,1

0,2

0,3

0,4

0,5

sulfide

elemental sulfur

electrolyte

Rogoznica Lake sample,11m depth b


160

Ratio between cathodic and anodic currents in the studied case of tetrasulfide solution was

roughly 3:1 indicating 3 S0 and 1 S2- in the molecule of S42-. In Rogoznica Lake sample this ratio was

much lower (1:6.5), pointing to a high excess of the free sulfide in the sample. Common ratio

between sulfide and elemental sulfur in the anoxic Rogoznica Lake samples is 10-15 to 1 in favour

of sulfide [6,13].

Similar voltammetric curves to Rogoznica Lake samples with revealed C2 - RSSNV peak can be

found in rain precipitation and aerosols. Usually in these samples RSSNV are detected in much

lower concentration range from 1-10 nM, while industrial waste samples could contain total RSS

from 10 µM up to mM concentration.

In sulfide and iron rich natural samples such is anoxic water column of freshwater Pavin Lake in

France, voltammetric curves similar to curves obtained in MS suspensions could be recorded

(Figure 4) [12,16]. Besides C2 peak in such samples other relevant peaks can be seen:

1) C3/A3 peak couple can be attributed to transformation reaction of FeS to the HgS [14,16,35]:

FeSlayer + Hg0 → HgSlayer + Fe2+ + 2e- (6)

and can be taken as an indication and rough measure for nanoparticulate FeS [15,16].

2) C4 is the well-known reduction peak of Fe(II) aqua ion, which is irreversibly reduced on the

Hg near -1.4 V [4,8,14,16,25]:

Fe2+ + 2e- → Fe0 (7)

3) The C1/A11 peaks arises from the reduction of Fe2+ or its labile complexes on the Hg

electrode surface modified by a FeS layer [14,16,35 and references therein]. Fe0 deposited at the

C1 would be oxidized in reverse scan at the A11 peak.

Figure 4. Typical voltammetric curves of anoxic sulfide and iron rich samples taken from crater

Pavin Lake, France at 71 m depth (E = -0.2 V, ta = 0 s for a, and E = -0.75 V, ta = 30 s for b. A3 peak is a rough measure for FeS nanoparticles presence.

In order to better characterize and possibly quantify concentration of the formed MS NPs and

to estimate its size ranges in water samples, additional chronoamperometric measurements were

employed for characterization of FeS NPs recently. It was shown that recorded chronoampere-

metric signals are carrying FeS NPs size, possibly charge and concentration information [32].

Similar approach was adopted here for PbS NPs characterization (Figure 5).

The PbS is chosen as one of the models because its redox chemistry is relative simple and not

governed by multiple relevant redox states in comparison with Cu and Fe and in several anoxic

E / V (vs. Ag/ AgCl)

-1.8-1.6-1.4-1.2-1.0-0.8-0.6-0.4-0.2

I / µ

A

-6

-5

-4

-3

-2

-1

0

1

C3

C1

C4

A11A3

aPavin Lake sample, 71 m depth

E / V (vs. Ag/AgCl)

-0.8 -0.7 -0.6 -0.5 -0.4 -0.3 -0.2 -0.1 0.0

I / µ

A

-0.6

-0.5

-0.4

-0.3

-0.2

-0.1

0.0

0.1

0.2A3

C2

b

Pavin Lake sample71 mdepth


doi: 10.5599/jese.2014.0053 161

samples occasionally voltammetric peak at -1.2 V similar as shown in Figure 2, was detected. In

accordance with our previous studies [11,17] this peak is attributed to the reduction of the PbS

NPs according to reaction (8):

PbS + 2e- Pb(Hg) + HS (7)

Consequently, the chronoamperometric measurements were started in the area of the C5 peak

potential (Figure 2) and the highest frequency of spike like signals were detected at -1.5 V. The

recorded chronoamperometric curves at -1.5 V were characterized by the sharp reduction current

transients with duration lasting from 100 ms and higher and peak heights in the range of 10-10 to

10-8 A. It is assumed that each spike represents a reduction of the PbS NPs at the Hg surface during

collision according to reaction 8.

Figure 5. Chronoamperograms for 5x10-5 M Pb2+ and HS- in 0.11 M NaCl/0.03 M NaHCO3

recorded at -1.5 V (vs. Ag/AgCl) 5 min (black curve), 10 min (yellow curve) and 30 min (green curve) after mixing of Pb2+ and HS-.

Recorded spikes appear to be sensitive on concentration of the PbS NPs in the solution, pH

changes, ionic composition and ageing time, similarly as obtained with FeS NPs. Please be aware

that in the case of FeS NPs, recorded spike like signals were caused by catalytic processes that

occurred on the FeS NPs surface during collision with the Hg electrode [32]. In Figure 5, it is

evident decrease in the signals frequency and increase of signals charge with ageing time of the

PbS suspension. In the given time, size of the formed PbS NPs, monitored by dynamic light

scattering measurements were changed from 140 to 480 nm according to number size

distribution. All above mentioned parameters (concentration, pH, suspension composition, and

ageing time) highly influence physico-chemical properties of the formed NPs, indicating a great

potential of the chronoamperometric measurements for the characterization, quantification and

sizing of all chalcogenides and other NPs which behave similarly at the Hg surface.

t / s

0 10 20 30 40 50 60

-40

-30

-20

-10

5 min after mixing Pb2+

and HS-


and HS-


and HS-

I / nA


162

Conclusion

In this work it is clearly shown how electrochemistry by choosing appropriate methodology and

experimental conditions can be successfully used for characterization, speciation and

deterimantion of different dissolved and colloidal sulfur species in natural waters including rain

precipitation and aerosols. Further, it appears that today, in the time of growing nanotechnology

and production of different NPs and nanomaterials, electroanalytical methods due to its relative

simplicity and prompt response, low cost and relatively high sensitivity and selectivity, might be a

good alternative analytical tools for characterization and possibly quantification of different NPs in

natural waters. This work is still challenging for the future.

Acknowledgements: This work is supported by the Ministry of Science and Technology of the Republic of Croatia Projects: ‘Nature of organic matter, interaction with traces and surfaces in environment’ (number 098-0982934-2717) and ‘Nanoparticles in aqueous environment: electrochemical, nanogravimetric,STM and AFM studies’, a Unity through Knowledge Fund, UKF project.

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doi: 10.5599/jese.2014.0057 165

J. Electrochem. Sci. Eng. 4(4) (20YY) 165-175; doi: 10.5599/jese.2014.0057


www.jESE-online.org


Characterization of an hrp-aox-polyaniline-graphite composite biosensor

Ana Carolina O. Santana, Erica F. Southgate, João Paulo B. G. Mendes*, Jo Dweck, Eliana Mosse Alhadeff and Ninoska Isabel Bojorge Ramirez**,

Escola de Química, Universidade Federal do Rio de Janeiro, Av. Horácio Macedo, 2.030, Centro de Tecnologia, Bloco E, E-203, Cidade Universitária,CEP 21941-909, Rio de Janeiro, Brasil *Instituto de Química, Universidade Federal do Rio de Janeiro, Av. Horácio Macedo, 2.030, Centro de Tecnologia, Bloco A, A-302, Cidade Universitária, CEP 21941-909, Rio de Janeiro, Brasil **Universidade Federal Fluminense, Dep. Engenharia Química e de Petróleo, R. Passo da Pátria, 156, Bl E-226, São Domingos , Niterói, CEP 24210-240, Rio de Janeiro, Brasil

Corresponding author: E-mail: [email protected]; Tel.: +55-21-26295598

Received: March 23, 2014; Revised: June 7, 2014; Published: December 6, 2014

Abstract Nowadays there is an increasing demand to develop new and robust biosensors in order to detect low concentrations of different chemicals, in practical and small devices, giving fast and confident responses. The electrode material was a polyaniline-graphite-epoxy composite (PANI/GEC). Alcohol oxidase (AOX) and horseradish peroxidase (HRP) enzymes were immobilized and the responses were tested by cyclic voltammetry. The conductivities for the composites of graphite/polyaniline were determined. The cyclic voltammograms allowed detecting ethanol in pure diluted samples in a range from 0.036 to 2.62 M. Differential scanning calorimetry (DSC) and thermal gravimetry analysis (TGA) were used to verify the thermal characteristics of the composites (0, 10, 20, 30 and 100 % of graphite). The Imax value was determined for the dual enzyme biosensor

(0.0724 A), and the Kapp

m as 1.41 M (with R2 =0.9912).

Keywords Cyclic voltammetry; ethanol; immobilized enzymes; PANI/GCE

Introduction

Ethanol is the most frequently analyzed aliphatic alcohol and several methods have been

developed for its quantitative determination [1-3]. Measurement of alcohol levels in liquors and

alcoholic drinks is a common necessity as is clinical analysis of patient tissue samples. The method



J. Electrochem. Sci. Eng. X(4) (2014) 165-175 ELECTROKINETICS AND SOIL DECONTAMINATION: AN OVERVIEW

166

approved by the Association of Official Analytical Chemists [4] for quantitative volumetric

determination of alcohol in beer, wine and distilled spirits is pycnometry, which is the most

common method for determining solution density. This method is considered as reference and has

the advantages of accuracy and no need for comparison against a standard solution. The principal

disadvantage is that the methodology is laborious, requiring a significant amount of time for its

performance. Another disadvantage is that it requires pre-distillation, generally regarded as the

first step in which error is introduced during the process of quantitative detection and analysis [5].

Other more accurate analytical methods include spectrophotometry and chromatographic

techniques: gas chromatography (GC) or high performance liquid chromatography (HPLC).

However, these methodologies are often less favorable due to high equipment prices and the

need for well-trained operators. There is currently a movement toward replacing these methods

with low-cost, fast and reliable electrodes working in conjunction with immobilized enzymes [6].

There is a growing need for the development of disposable devices for clinical and/or

environmental monitoring. This need has stimulated the development of new technologies and

methodologies that can efficiently monitor an increasing number of analytes on site in the

environmental field or support clinical diagnoses as quickly and as cheaply as possible; offering

even the possibility of on-site field monitoring. Besides selectivity, an analytical device must also

be sensitive. In this respect, biosensors have shown great potential in recent years and thus

appear to be useful components of effective analytical tools [7-8].

Biosensors that link enzyme catalyzed chemical reactions with amperometric detectors are

having a great impact on fields such as environmental monitoring [9-10], analysis of the quality of

food and beverages [11-12], biomedical monitoring process [13-14] and biomedicine [15]. These

analytical tools, prepared by immobilization of enzymes on an electrode surface, are simple,

sensitive and offer a fast response. The main problem that appears in the operation of these

devices is the transfer of electrons from the active site of the enzyme to the electrode.

Immobilization of the biological material on the electrode surface constitutes a crucial step in

development of the biosensor, since the enzyme’s structure must be maintained in order to

enable its action on the sample of interest [16-18]. Horseradish peroxidase (HRP) is widely used in

enzyme-linked biosensors. However, there are at least two main drawbacks shown by this

enzyme: (1) It exhibits a very broad specificity to reduce substrates [19-20], which results in low

selectivity of the biosensor; (2) although it displays good stability at room temperature, it is

unstable at high temperatures [21-22]. The co-immobilization of alcohol oxidase with horseradish

peroxidase is expected to increase the selectivity and amplify the sensitivity of the biosensor for

the quantitative determination of ethanol [23-24].

Immobilization of dual enzymes provides an excellent basis for increasing the selectivity,

sensitivity and the thermal stability of the biosensor, depending on the strategy adopted for

immobilizing the enzymes [15]. The immobilized enzymes may be reused several times or

employed in an economical continuous flow path. Dual enzyme-linked sensors are amenable to

automation for analytical measurements, scale up of enzymatic biotransformation reactors, or to

recover a product with greater purity [19, 25]. Xie, et al. [26] reported recent advances in enzyme

immobilization technologies that enhance enzyme properties such as activity, stability, specificity

and reduced inhibition effects. The authors suggest that in the future multi-enzyme sensors based

on co-immobilization would be the solution to many of the applications for the biotechnology

industry and analytical devices.

A. C. O. Santana et al. J. Electrochem. Sci. Eng. 4(4) (2014) 165-175

doi: 10.5599/jese.2014.0057 167

The objective of this study was to characterize a composite-based on PANI / epoxy / graphite

and evaluate its performance as a substrate for horseradish peroxidase (HRP) and alcohol oxidase

(AOX) enzymes immobilized to an electrode creating a biosensor for ethanol detection. The

development of such a method for the immobilization of multiple enzymes is highly attractive,

especially for economic reasons because as enzymatic activity decays the support can be

regenerated and reloaded with fresh enzyme. In fact, the cost of support is often a primary factor

in the overall cost of the immobilized catalyst. In order to build the biosensor a composite

prepared with graphite and an electron conductor polymer as polyaniline was studied and

characterized in terms of its electrochemical conductance capacity and thermal stability [18,27].

We find critical compositions of the material that works with improved sensitivity over a relatively

broad range of ethanol concentrations.

Experimental

Materials

Horseradish peroxidase (HRP; EC 1.11.1.7) was purchased from Toyobo, Brazil and alcohol

oxidase (AOX, EC 1.1.1.1, specific activity of 200 units/mg of protein), graphite powder and

polyaniline (emeraldine salt) were purchased from Sigma-Aldrich. For the incorporation of

enzymatic solutions, a 2.5 % (v/v) of glutaraldehyde (Sigma-Aldrich) and 1 mg/mL of protein

albumin were used. The ethanol standard solutions were prepared with 0.1 M mono potassium

phosphate buffer (pH 7.0). All reagents were of analytical-reagent grade. All solutions were

prepared with distilled water.

Apparatus

Amperometric measurements were carried out using an AUTOLAB PGSTAT12 (Ecochemie)

connected to a personal computer via a serial RS232 port for data acquisition. The obtained

amperometric alcohol dual enzyme sensors were evaluated by means of cyclic voltammetry in a

three-electrode configuration with Ag/AgCl/KCl (3M) reference electrode and Pt-wire counter

electrode. When not in use, the electrode was stored dry at 4 °C in a refrigerator.

The thermal properties (thermogravimetric analysis (TGA) and differential scanning calorimetry

(DSC)) for the composites of graphite : PANI prepared with 0, 10, 20, 30 and 100 % of graphite

were performed by TA Instruments SDT Q600. Analyses were conducted in a 30 mL/min flow rate

of air atmosphere, with a ramp of 5 °C/min from 30 to 800 °C.

The electrical conductivities of the composites pellets were evaluated using the techniques of

two electrodes, between which a pellet of known composition of the composite was fixed with the

aid of a sleeve of Teflon. The tests were done using a bench meter ICEL Manaus MD - 6700

coupled to a computer. The disks pellets prepared with 0% and 100% of pure graphite and

composites with 1, 3, 5, 10, 20, 30, 50, 70, 100% of graphite mixed with PANI were measured. Pure

samples of PANI and graphite were also determined.

Preparation and evaluation of the AOX–HRP-based biosensors

HRP (3.60 g) was dissolved in 30 mL of 50 mM phosphate buffer (PB, pH 7.0). After filtration

and dialysis steps, a 0.133 mg/mL of HRP solution was mixed with AOX (47 mg/mL) in buffer

pH 7.0. The 10 % (w/v) of bovine serum albumin and 2.5% (v/v) of glutaraldehyde were also

prepared in 50 mM PB (pH 7.0) solution. A 10 μL volume of bovine serum albumin and 10 μL of

glutaraldehyde were deposited on the electrode surface sequentially. The excess of glutar-


168

aldehyde was rinsed off with water. Teflon cylindrical electrodes (5.0 × 0.7 cm and 0.13 cm inside

orifice) were used to construct the working dual enzyme biosensor with a 20 mL electrolytic cell

and an Ag/AgCl reference electrode and a platinum counter-electrode.

Procedure for immobilization

The methodology used was the ionic immobilization of AOX and HRP enzymes on the electrode

surface constructed using a graphite matrix with polyaniline and an epoxy resin. During the

immobilization step, a solution containing 2.5 % (v/v) glutaraldehyde, 0.5 % (v/v) BSA and

97 % (v/v) enzyme solution containing 1100 μL HRP and 15 μL of AOX was deposited on the

electrode surface. The electrode was left at 4 °C for 24 hours [28].

Measurement procedure

Cyclic voltammetry (CV) measurements on the electrode were performed in a 3-electrode

system containing a Ag/AgCl/KCl 3M (Microquímica®) reference electrode, coiled platinum wire

(99.99 % pure) mounted at the end of a chemically-resistant epoxy rod as counter electrode in

addition to the modified working electrode based on PANI/GCE. The potential was cycled between

–400 and 400 mV vs. Ag/AgCl.

Determination of ethanol in samples

Ethanol (95 %) samples (0.15 mL) were diluted in a 10 mL flask with 0.1 M mono potassium

phosphate buffer solution (pH 7.0). Voltammetric determination was carried out by applying the

standard addition method. Diluted sample and standard ethanol solution (0.15 μL) were added to

the voltammetric cell containing 10 mL of 0.1 M mono potassium phosphate buffer solution

(pH 7.0).


Study of differential scanning calorimetry and thermal gravimetry analysis

The thermal stability of the graphite composite samples was analyzed by TG, derivatived

thermogravimetric analysis (DTG) and DSC. Results of TG and DTG analyses are presented in Figure

1. The curves in Figure 1 follow the mass as a function of temperature of composite samples

containing 100 % graphite, 20 % graphite and 0% graphite (100 % polyaniline), respectively. The

curve for the composite of Graphite/PANI (red dashed line) shows an intermediate stability

between pure samples of Graphite and PANI polymer. The presence of the PANI introduces four

decomposition steps. In the first stage, beginning at 150 °C, there is a slow weight loss associated

to the release of trapped water or organic solvents in the polymer structure. The second stage of

weight loss is observed from 270 °C to 550 °C and is attributed to decomposition of the oligomers.

The third decay, from 350 °C to 450 °C, was assigned to the thermal decomposition of the PANI

chains. The DTG curves fully support the above mentioned losses. The pure graphite sample

decomposes above 600 °C, whereas the 20 % graphite composite presents four degradation steps

(three attributed to the pure polyaniline and one to the pure graphite). Similar results were found

by Kowner, et al. [29], Bourdo, et al. [30] and Mo, et al. [31].

A linear fitting between the data of polymer content estimated by TG and composition on a dry

basis of raw materials in the composites was proposed, showing a good correlation coefficient (R2

= 0.9811). The difference between the values estimated from the correlation with those of the

components in the composite has an average value of -0.1 % with standard deviation of 1.86 %.


doi: 10.5599/jese.2014.0057 169

Fig. 1. TG and DTG curves for the samples prepared with 100, 20 and 0 % of graphite.

Conductivity of graphite/PANI composite

Song and Choi [32] have reported that the most conductive form of PANI is the fully

protonated, half-oxidized emeraldine salt form. A decrease in conductivity was observed when the

polymer was deprotonated or either fully oxidized or reduced. This work intends to develop a

prototype composite-biosensor based on typical PANI that maintains the conductivity.

Conductivity values were determined for different concentrations of graphite : polyaniline

composites (1, 3, 5, 10, 20, 30, 50, 70, 100 % of graphite) and for the pure alcohol oxidase. As

shown in Figure 2 an increase in the conductivity of the composite samples was observed.

Fig. 2. Electrical conductivities of PANI/GEC as a function of graphite concentration in the

composites. Also shown are samples with 100 % PANI (green bar) and 100 % graphite (red bar).


170

This is probably due to synergistic effects of mixing the conductive polymer PANI and graphite

powder. The mixture better supports electron transfer and, consequently, displays enhanced

electrical conductivity. The conductivity values determined for the pure samples of graphite and

PANI were 1.82×10-3 S / cm and 4.64×10-4 S / cm, respectively. The maximum value was obtained

with the 70 : 30 (graphite:PANI) composite; higher than the value measured for the 100 % graphite

sample. For the composites prepared with 1 to 10 % of graphite, there were not any significant

variations in conductivity. A linear relationship between graphite content and conductivity was

observed from 20 to 70 % with the conductivity values for 50 to 70 % of graphite surpassing those

for the 100 % graphite sample. Mo et al. [31] has detected an increase in the electrical

conductivity as a function of graphite nanosheet content in a composite prepared with graphite

nanosheets and PANI. Bourdo et al. [30] also found similar behavior for pure PANI and graphite

samples and for PANI/graphite composites. In the present study, the 30 % PANI composite

compound was employed due to the improved performance of its electrical response.

Electrochemical behaviour of the biosensor

The biosensor bi-enzymatic HRP/AOX was characterized using cyclic voltammetry to

demonstrate the electrochemical performance of the system. Figure 3 shows the cyclic

voltammograms obtained from 5 to 150 mV s-1 in a solution of 1mM K4Fe(CN)6 mixture in

0.1 M KCl and phosphate buffer pH 7.0. The peaks currents of the CVs indicating quasi-reversible

processes between Fe(CN)64-/Fe(CN)6

3- couple and the electrodes at the faster scan rates Each

curve has the same form but it is apparent that the total current increases with increasing scan

rate. This again can be rationalized by considering the size of the diffusion layer and the time taken

to record the scan. Clearly the voltammogram will be slower to record as the scan rate is

decreased. Hence the size of the diffusion layer above the electrode surface will be different

depending upon the voltage scan rate used. In spite of that, working with lower scan rates a well-

-defined cathodic peak and a small anodic could be identified, and the scan rate of 10 mV s-1

applied to analyze the ethanol samples. So, the best quality voltammogram was obtained working

with a scan rate of 10 mV s-1. Therefore, that was the scan rate applied to analyze the ethanol

samples

Fig. 3. Cyclic voltammograms of the AOX/HRP/Graphite/PANI in 0.1 M PBS, pH 7.0 at various

scan rates (from inner to outer curves: 5, 10, 20, 50, 100, 150 mV s−1).


doi: 10.5599/jese.2014.0057 171

Figure 4 shows shows the current intensity for the calibration curve changed between 0.61 mA

(0.316 M) to 0.25 mA (2.62 M) The concentration range of standard ethanol solutions used in the

electrochemical measurements was 0.316 - 2.62 M (R2 = 0.991). This clearly demonstrates that the

current density reduces linearly with increased ethanol concentration in the samples. This is

attributed to an inhibition of the enzyme. The effect is especially evident at the higher ethanol

concentrations, probably due to the reaction end-products (acetaldehyde) or external mass

transfer limitations. A similar behavior was reported for AOX and HRP that was covalently

immobilized on controlled pore glass [23]. However, that study showed that all the supports

exhibited less than 20 % of the specific activity of the free enzyme, as a consequence of

conformational changes in the 3-D structure of the protein caused by the covalent binding of AOX

to the supports. In this work the enzymes were immobilized by adsorption, which is less aggressive

than the covalent immobilization. However, the enzyme may be coupled to the support in a way

that hinders the access of substrates to the active center, promoting the mass transfer limitations.

The amperometric response exhibited by the different immobilized AOX preparations was also

very similar although the highest value was obtained when the support was activated using

glutaraldehyde in phosphate buffer pH 7. Sirkar, et al. [33] observed an increase in the current

density (60 %) of the electrochemical biosensor response for a multilayer nanocomposite thin

film using glutaraldehyde as a crosslinking agent in a trial for stabilizing the structure. The authors

proposed that arginine and lysine residues of the enzyme react with amines present on the redox

polymer and, as a consequence, the activity was maintained near 100 % for three weeks.

Fig. 4. Ethanol biosensor calibration curve. Scan rate 10 mV s-1. (n = four measurements).

Wu, et al. [34] reported an inverse calibration curve for oxygen consumption by a sensitive

ethanol biosensor nanocomposite of carbon nanofiber with immobilized ADH. They observed

decreased oxygen consumption with the increase of ethanol concentration in the sample.

Chronoamperometric curves showed a decreasing response, upon addition of ethanol aliquots

(0 - 112 μM) to static air-saturated pH 7.0 phosphate buffer saline. Wen, et al. [35] reported an

ethanol biosensor constructed with alcohol oxidase/chitosan immobilized eggshell membrane and

a commercial oxygen sensor. Those measurements were based on the depletion of dissolved


172

oxygen upon exposure to ethanol solution (0.15 – 0.75 mM). Al-Mhanna and Hueber [36] reported

an economic system that worked with one enzyme in a differential pH measurement device for

alcohol oxidase and -nicotinamide adenine dinucleotide (NADH+) reaction and obtained a

logarithmic curve for ethanol concentrations against change in pH for standard samples. These

authors described an inverse correlation between the signal response and the analyte

concentration for the indirect detection measurements working with a wide range of ethanol

standard concentration solutions (17.14 μM – 17.14 M).

Mackey, et al. [37] optimized the proportion of dual enzyme horseradish peroxidase:glucose

oxidase biosensor working with ratios of 1 : 7 to 7 : 1, immobilized on a polyaniline-

polyvinylsulphonate modified screen-printed carbon paste electrode and identified the proportion

that produced the best response signal was 1 : 1. Rondeau, et al. [38] identified the optimal

proportion of two enzymes in the biosensor composite by monitoring electrical response signals to

establish idealized conditions for glucose oxidase : horseradish peroxidase immobilized with a

modified carbon paste for in order to increase the selectivity, sensitivity, accuracy and stability

[38]. The intensity of the electrochemical signal response was analyzed by Alpat and Telefoncu

[39] who measured the amount of alcohol dehydrogenase immobilized on the electrode surface

(47.1 to 200 U cm-2) and found that the linear response was between 0.01 mM and 0.04 mM for

117.6 U cm-2.

Nicell and Wright [40] reported the dependence of horseradish peroxidase activity over a wide

range of hydrogen peroxide concentrations. They observed an increase in the inhibitory effect on

the enzyme catalytic activity. The static procedures of the electrochemical measurements of this

work for the ethanol concentration solutions (0.330 – 2.62 M) probably promotes an increase of

the peroxide hydrogen concentration in the electrolytic cell, and hence the inhibition of the HRP.

Yotova and Medhat [41] reported the inhibition effect in a multi-enzyme immobilized biosensor

system constructed to analyze residue from pesticides with acetylcholinesterase and choline

oxidase. The relative inhibition percentage of each measurement was calculated using the

following equation:

0

0

, % 100Ipc Ipc

I =Ipc

(1)

where I is the relative inhibition; Ipc0 is the initial inhibited cathode current intensity measured for

the lower ethanol concentration and Ipc the inhibited cathode current intensity determined for

each sample. Assuming a possible inhibition effect on the cyclic voltammetric response signal with

the increase of the ethanol concentration in the sample, this treatment was adopted for this work.

A linear correlation was observed, confirming the inhibitory effect of the ethanol on the enzyme.

Amine, et al. [42] published a review that discusses horseradish peroxidase among the enzymes

that could be used for inhibition-based biosensors applied for food safety and environmental

monitoring. Kuusk and Rinken [43] classified the carbaril inhibition of tyrosinase biosensor by

excess substrate and considered the reasons behind their inability to determine low carbaryl

concentrations by a classical steady state kinetic approach. The Km and Imax kinetics parameters

were calculated from Lineweaver–Burk plots by using the relative inhibition values as described in

equation 2:

appm

max ethanol max

1 1 1K= +

RI RI C RI (2)


doi: 10.5599/jese.2014.0057 173

where 1/Cethanol is the concentration of the ethanol in the solution sample, RI and RImax represent

the initial and the maximum relative inhibition current, respectively, and Kapp

m is the apparent

Michaelis constant.

The Lineweaver-Burk plot for the dual enzyme AOX-HRP biosensor showing 1/I versus 1/Cethanol

is illustrated in Figure 5.

The Imax value determined considering the inhibition effect on the dual enzyme biosensor was

0.0724 μA, and the Kapp

m was 1.41 M (R2=0.9912).

Fig. 5. Line-weaver-Burk plot for the bienzimatic AOD-HRP biosensor for different ethanol concentration

Table 1 shows the analytical performance of the proposed ethanol biosensor towards ethanol

detection compared with various electrochemical biosensors modified for dual enzymes that also

reported Kapp

m and Imax. Despite the low affinity for substrate observed in this work, the sensitivity

was higher when compared with those determined for both redox hydrogel dual enzyme films

previously reported in the literature [44-45]. This suggests that the linear range and detection limit

of the proposed ethanol biosensor mentioned above appear to be beneficial compared to other

previously reported modified electrodes.

Table 1. Comparison of analytical characteristics of ethanol dual enzyme biosensors.

Film/Composite/Enzymes I.D. / cm Kapp

m / mM Imax / nA Sensitivity, nA/M Reference

HRP+AOX+PVI-Os 0.305 4.71 813.95 0.17 [44]

HRP/PVI10-Os/PEG-DGE/AOX/CP5 0.305 9.6 ± 0.3 572 ± 7 0.06 [45]

PANI-GEC/HRP/BSA/AOX 0.130 1.410 72.4 51.3.

This work

I.D. - internal diameter; PVI - Poly(vinyl-imidazole; PVI10-Os - redox hydrogel synthesized; PEG-DGE - Poly(ethylene glycol) (400) diglycidyl ether; CP5 - electrodeposition polymer; Os - complex: redox polymers synthesized (4,4'dimethylbipyridine); PANI-GEC: polyaniline in Graphite epoxy composite; BSA - Bovine serum albumin. The applied potentials for all configurations are –50 mV vs. Ag/AgCl.


174

Conclusions

The composite material prepared from differing proportions of graphite and PANI displayed

enhancement in the conductivity for compositions of less than 20 % graphite and a synergistic

effect that increased its response for mixtures with more than 50 % of graphite. The thermal

analysis techniques applied to characterize the prepared composites showed a good agreement

with the original proposed formula composition. The electrochemical results confirm that it is

possible to detect ethanol with this biosensor in the ethanol concentration range of 0.316 to

2.62 mol L-1 limited by a significant inhibition effect observed in the enzyme.

Acknowledgements: Thanks to Toyobo of Brazil (enzyme horseradish peroxidase) and CNPq support from the Announcement Universal - 2008/2010 and PIBIC.

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doi: 10.5599/jese.2014.0068 177



www.jESE-online.org


Voltammetric studies on mercury behavior in different aqueous solutions for further development of a warning system designed for environmental monitoring

Paul-Cristinel Verestiuc, Igor Cretescu*,, Oana-Maria Tucaliuc, Iuliana-Gabriela Breaban and Gheorghe Nemtoi**

Faculty of Geography and Geology, Al. I. Cuza University of Iasi, 20 A. Carol I Bd., Iasi, 700505, Romania *Faculty of Chemical Engineering and Environmental Protection, Gheorghe Asachi Technical University of Iasi, 73, D. Mangeron Street, Iasi, 700050, Romania **Faculty of Chemistry, Al. I. Cuza University of Iasi, 11, Carol I Bd., Iasi, 700506, Romania

Corresponding author: E-mail: [email protected]; Tel.: +40-741-914-342.

Received: October 5, 2014; Revised: October 31, 2014; Published: December 6, 2014

Abstract This article presents some results concerning the electrochemical detection of mercury in different aqueous solutions, using the following electrodes: platinum-disk electrode (PDE), carbon paste electrode (CPE) and glass carbon electrode (GCE). Using the voltam-metric technique applied on the above mentioned electrodes, the experimental conditi-ons were established in order to obtain the maximum current peaks, in terms of the best analytical characteristics for mercury analyses. The dependence equations of cathodic current intensity on the scan rate were established in the case of mercury ion discharge in each prepared solution of 0.984 mM HgCl2 in different electrolyte background: 0.1 M KCl, 0.1 M H2SO4 and 0.9 % NaCl. Among the three investigated electrodes, the carbon paste electrode presented the highest detection sensitivity toward mercury ions in the aqueous solution. It was observed that, at a low scanning rate, the pH had an insi-gnificant influence over the current peak intensity; however, the quantification of this in-fluence was achieved using a quadratic polynomial equation, which could prevent the er-rors in mercury detection in case of industrial waste stream pH changes. The calibration curves for mercury in 0.9 % NaCl solution and in the tap water respectively were carried out.

Keywords Cathodic linear voltammetry; platinum electrode; carbon paste electrode; glass carbon

electrode; mercuric ion; flow electrochemical cell.



J. Electrochem. Sci. Eng. 4(4) (2014) 177-186 MERCURY BEHAVIOR IN SYSTEM FOR ENVIRONMENTAL MONITORING

178 178

Introduction

The ubiquitous presence of mercury in the environment is due to both natural geological

activities as well as due to increasing anthropogenic pollution. Because of its unique electronic

configuration, mercury behaves similarly to noble gas elements, but the physical and chemical

properties of mercury such as high surface tension, high specific gravity, low electrical resistance,

and a constant volume of expansion over the entire temperature range in liquid state can rapidly

transform this element into a hazardous air pollutant [1,2].

The Water Framework Directive (2000/60/EC) classified mercury as a priority hazardous

substance, establishing that from 2015, no more mercury from production processes can be

discharged [3, 4].

Mercury exists in a large number of forms, i.e. as “elemental” Hg(0), monovalent or divalent

mercury Hg(I) and Hg(II), and in inorganic and organic compounds. Metallic mercury Hg(0) and

most mercury compounds present high toxicity, acting as a bioaccumulative neurotoxin [5,6].

Mercury ions are strongly adsorbed by soils or sediments in acid medium and are slowly

desorbed, due to the content of clay minerals and/or organic matter, which are responsible for its

behavior. The reaction products resulting from the methylation of inorganic mercury forms impose

a significant risk to humans and wildlife due to tendency to accumulate in the food chain, and their

ability to act as neurotoxins [6,7]. Exposure to various forms of mercury will harm human health.

Moderate and repeated exposure to organic forms (lower than a few mg m-3 Hg, but higher than

0.05 mg m-3 Hg) causes symptoms of poisoning such as: lack of coordination of movement,

impairment of peripheral vision, speech, hearing or walking as well as muscle weakness. Inhaled or

physical contact with inorganic forms cause: tremors, emotional or neuromuscular changes,

insomnia, headaches, disturbances in sensation, changes in nerve responses, and performance

deficits on tests of cognitive function. With prolonged or high concentration exposure, kidney

effects, respiratory failure and death may occur [8,9].

Mercuric chloride (HgCl2) is used as a depolarizer in electric batteries and as a reactant in

organic synthesis and analytical chemistry [10]. The presence of this element in different

environmental components could be considered as harmful to human health well environmentally

dangerous due to the mercury content.

Taking into consideration all the above mentioned aspects, the detection of mercuric ion has

become a priority for environmental safety and human health. The determination of trace

amounts of mercury, has led to some analytical problems because it can be found in several

chemical forms [11]. For an accurate determination of mercury at trace and ultra-trace levels,

analytical methods with high sensitivity and selectivity are needed. There are a number of

analytical methods for mercury detection which require expensive instruments, well-controlled

experimental conditions, sample preparation and relatively large sample volumes [12].

Electrochemical detection of trace metals offers important advantages, such as remarkable

sensitivity, inherent miniaturization and portability, remote monitoring and decentralized

measurements, low cost and compatibility with turbid samples [13,14]. Therefore, electrochemical

methods are less costly and require no sophisticated equipment.

Chemically modified electrodes have received increasing attention, which has led to

improvements in the sensitivity and selectivity of electrochemical analysis techniques in the recent

decades. The determination of Hg(II) ions using chemically modified electrodes has been investi-

gated recently, using plants [14], gold film [15], polymer films [16] or organic compounds with

chelating groups [17].

P.-C. Verestiuc et al. J. Electrochem. Sci. Eng. 4(4) (2014) 177-186

doi: 10.5599/jese.2014.0068 179

The behavior of electrodes has also been studied using modified a carbon paste electrode [18],

modified glass carbon electrode [17] or modified gold nanoelectrode ensembles [19].

A relatively recent review of electrochemical sensors and detectors has been done by Bak-

ker [20], characterizing this area of research as being one of the most fruitful and interdisciplinary

areas of research in analytical chemistry. A recent paper by Pujol [21] analyzed the actual state of

art concerning the sensors and devices for heavy metals detection in water, but the mercury

detection from environmental samples was not studied in details.

For this reason in this study it was investigated the cathodic discharge of the mercuric ion on

different type of electrodes such as: platinum-disk electrode, carbon paste electrode and glass

carbon electrode, in respectively different aqueous solutions (0.1 M KCl, 0.1 M H2SO4, 0.9 % NaCl)

by the voltammetric method in order to find the most suitable conditions for analytical purpose.

EXPERIMENTAL

In order to simulate the wastewater from the industrial stream, aqueous solutions of HgCl2 in

0.1 M H2SO4 at a pH value of 0.99, 0.9 % NaCl at a pH value of 5.66 and 0.1 M KCl at a pH value of

7.43 was prepared, using analytical purity reagents and double distilled water. The investigation of

mercury behavior in these solutions was carried out using voltammetric measurements [22-26]

with the above mentioned electrodes (GCE, CPE and PDE) individually. These electrodes play the

role of working electrode (WE) in a flow electrochemical cell (as is presented in Fig. 1) through

which the samples from the industrial stream are passed using a peristaltic pump. A saturated

calomel electrode (SCE) was used as the reference electrode (RE), while a platinum electrode was

used as the counter (auxiliary) electrode (CE) at the temperature of 25 °C. Also, pH and tempera-

ture sensors were located in the electrochemical cell in order to provide corrections in the case of

possible changes in the above mentioned parameters which should be kept constant at the same

values as were used during the calibration procedure.

Figure 1. Experimental setup for study of the voltammetric detection of mercury in aqueous solution

The pH corrections were achieved using software (the pH dependence of the current peak was

determined by a simple equation), while the temperature corrections were first achieved by the

thermostatic system for sample processing and finally by the software corrections for the fine pH

adjustments.


180 180

The voltammograms were recorded using the potentiostat VoltaLab 32 (Radiometer Analytical)

[27-29] after stopping the flow of liquid samples through the electrochemical cell (using a solenoid

mini-valve) and nitrogen bubbling in the investigated solutions, in order to remove the dissolved

oxygen.

The investigated electrodes were used as purchased from Radiometer Analytical, except for the

carbon paste electrode, which was prepared according to methods previously presented in the

literature [30-31]. The electrode surfaces were: 3.14·10-2 cm2 for PDE; 7.07 10-2 cm2 for GCE and

19.63 10-2 cm2 for CPE.


In Figures 2, 3 and 4, the cathodic linear voltammograms are presented, corresponding to the

reduction of Hg2+ on the PDE (Φ = 2mm) at different scanning rates in different aqueous media.

Figure 2. Linear cathodic voltammograms on PDE in 0.1 M H2SO4 aqueous solution

recorded at different scanning rates

Figure 3. Linear cathodic voltammograms on PDE in 0.1 M KCl aqueous solution recorded

at different scanning rates

Figure 4. Linear cathodic voltammograms on PDE in 0.9 % NaCl aqueous solution recorded at different

scanning rates.

Figure 5. Linear cathodic voltammograms of mercury ion discharge on PDE in different solutions at

constant scanning rate of 50 mV s-1


doi: 10.5599/jese.2014.0068 181

Concerning the electrochemical behavior of mercuric ions in the three different aqueous

solutions shown in Figures 2-4, it was observed that there was a simultaneous increase in PDE

sensitivity with the scanning rate, at a concentration of 0.984 mM, as was expected.

This behavior leads to the following dependence equations for the cathodic peaks, which

express the discharge of the mercuric ion on the platinum electrode, in all investigated solutions:

a. 0.1 M H2SO4: I = -2.114 – 0.058 v + 3.474×10-5 v2 (R = 0.99952) (1)

b. 0.1 M KCl: I = -4.285 – 0.072 v + 2.168 ×10-5 v2 (R = 0.99757) (2)

c. physiological serum (0.9 % NaCl): I =-6.815 - 0.140v - 6.611×10-5 v2 (R = 0.99827) (3)

where: I = intensity current (mA), v = scanning rate (mV s-1), R = correlation coefficient

According to the experimental results, it was noted that the highest sensitivity of PDE was

obtained in physiological serum, followed by the solution prepared based on sulfuric acid and

potassium chloride, respectively (Figure 5).

Figure 6. Linear cathodic voltammograms on CGE in 0.1 M H2SO4 aqueous solution recorded at different

scanning rates.

Figure 7. Linear cathodic voltammograms on CGE in 0.1 M KCl aqueous solution recorded at different

scanning rates

Figure 8. Linear cathodic voltammograms on CGE in 0.9 % NaCl aqueous solution recorded at different

scanning rates

Figure 9. Linear cathodic voltammograms of mercury ion discharge on GCE in different aqueous

solution at constant scanning rate of 50 mV/s


182 182

Figures 6-8 present the cathodic linear voltammograms for each of the aqueous solutions

recorded on the GCE at different scanning rates.

In this case, the following dependence equations for the cathodic peak were established, which

express the discharge of the mercuric ion (0.984 mM) on the platinum disc electrode, in all three

investigated solutions:

a. 0.1 M H2SO4: I = -5.7573 - 0.06488+1.3139×10-4 v2 (R = 0.99404) (4)

b. 0.1 M KCl: I = -11.895 - 0.239 v + 1.079×10-4 v2 (R = 0.998) (5)

c. physiological serum (0.9 % NaCl): I = -3.745 - 0.347 v + 1.355×10-4 v2 (R = 0.99918) (6)

where: I = current intensity (mA), v = scanning rate (mV s-1), R = correlation coefficient

In accordance with the Figure 9, the sensitivity of the GCE toward the mercuric ion was highest

in KCl aqueous solution and lowest in sulfuric acid.

The entire cathodic process of mercuric ion discharge on the CPE in all of the three investigated

solutions is presented in detail in linear voltammograms (Figures 10-12).

Figure 10. Linear cathodic voltammograms

on CPE in 0.1 M H2SO4 aqueous solution recorded at different scanning rates.

Figure 11. Linear cathodic voltammograms on CPE in 0.1 M KCl aqueous solution recorded


Figure 12. Linear cathodic voltammograms on CPE in 0.9 % NaCl aqueous solution recorded


Figure 13. Linear cathodic voltammograms of mercury ion discharge on CPE in different solutions

at constant scanning rate of 50mV s-1


doi: 10.5599/jese.2014.0068 183

Data obtained using the CPE revealed the same behavior as with the GCE regarding higher

sensitivity, as follows: 0.1 M KCl, physiological serum and 0.1 M H2SO4 (Figure 13).

The measurements performed on the CPE, in each investigated solution led to the following

equations:

a. 0.1M H2SO4: I = -38.137 - 1.282 v + 9.944×10-4 v2 (R = 0.99581) (7)

b. 0.1 M KCl: I = -80.923 - 2.168 v + 1.27×10-3 v2 (R = 0.99808) (8)

c. physiological serum (0.9 % NaCl): I = -34.086 - 3.25 v + 2.23×10-3 v2 (R=0.99867) (9)

where: I = current intensity (mA), v = scanning rate (mV s-1), R = correlation coefficient

These equations point out the dependence of the current intensity (corresponding to the

discharge of mercuric ions in aqueous solutions at a concentration of 0.984 mM HgCl2) on scanning

rates.

Increasing the scanning rate during investigations into mercuric ion discharge for each studied

electrode in the aqueous solutions led to a change in the cathodic potential values, which were

shifted to negative values, suggesting a quasi-reversible process.

Among the electrodes used in this study, the carbon paste electrode was the most sensitive

with respect to the cathodic discharge of mercuric ions in all the investigated aqueous solutions.

Taking into consideration this aspect, the pH dependence of the current peak (Figure 14) and the

calibration curves (Figure 15) in the NaCl solution and tap water, respectively, were presented only

for the CPE.

Figure 14. Influence of solution pH on the cathodic peak on CPE in 0.9 % NaCl aqueous solution

recorded at two scanning rates

Figure 15. Calibration curves for determination of HgCl2 using CPE in 0.9 % NaCl aqueous solution and

tap water, respectively

As presented in Figure 14, a low pH dependence of the current peak was observed for a

scanning rate of 50 mV s-1, but if this value was multiplied by three fold, the dependence became

more evident, especially in the acidic region. These equations are presented as follows:

I50 = -0.12847 - 0.09001 pH + 0.00774 pH2 (10)

I150 = -0.10381 - 0.0407 pH + 0.00419 pH2 (11)


184 184

The calibration curves presented in Figure 15 pointed out an approximately linear dependence

for both calibration curves, which are separated into two regions. However, these calibration

curves could be very well fitted following quadratic polynomial equations:

X = -lg[HgCl2]

INaCl = -242.414 + 257.620 X - 102.678 X2 + 18.175 X3 - 1.205 X4 (12)

Itap water = -121.207 + 128.825 X -51.339 X2+9.087 X3 - 0.602 X4 (13)

The decreased sensitivity of the voltammetric method for mercury detection in real samples

(tap water provided from the Iasi drinking water treatment plant) was evident, but the analytical

signal was sufficiently high to be discriminated from the global noise determined in such a

complex matrix.

The response limit of the concentration peak for Hg (II) detection was in the range of 10-5 M to

10-3 M. The optimal operating conditions were established after a period of three minutes of

preconcentration.

Conclusions

The following conclusions can be drawn for the voltammetric studies regarding the behavior of

mercuric ions in the presence of different supporting electrolytes in aqueous solutions using

different electrodes:

The cathodic discharge of mercury ions from HgCl2 aqueous solutions in different electro-

lyte support media such as 0.1 M KCl, physiological serum (0.9 % NaCl), and 0.1 M H2SO4,

using different electrode materials (platinum, glass carbon and carbon paste) was eviden-

ced by a relevant current peak, whose height (in terms of the analytical method sensitivity)

is dependent on the potential scanning rate;

The cathodic discharge of mercuric ions, in all investigated aqueous solutions, was a slow

and quasi-reversible process, on all three of the investigated electrode materials;

The best electrode was the carbon paste electrode, in terms of the sensitivity toward

mercuric ions, in concordance with the equations describing the dependence between the

cathodic peak intensity and the potential scanning rate. It was pointed out that the CPE

presented the highest detection sensitivity toward mercuric ions in an aqueous solution,

working in a wide pH range and having a short response time. Besides these aspects, the

CPE is very easily prepared and has good reproducibility and repeatability for mercury

analysis, which recommend it for this analytical application;

Based on the results, a miniaturized voltammetric flow cell will be developed as an

integrated component of the equipment used for mercury detection/monitoring in

industrial waste streams. Based on the mercury detection capacity, a warning system will be

designed to avoid mercury discharge into surface waters in the case of accidentally

increasing the mercury concentration.

Due to the increased capacity for mercury detection by this proposed monitoring system,

based on a voltammetric method and an electrochemical flow cell, its versatility could be

extended using some auxiliary equipment to other environmental components such as air

and soil.

At low scanning rate, the pH had an insignificant influence over the current peak intensity in

the case of the CPE in 0.9 % NaCl solution.


doi: 10.5599/jese.2014.0068 185

In order to avoid errors in mercury detection in cases of industrial stream pH changes, the

quantification of the pH influence was achieved using a quadratic polynomial equation.

The calibration curves for mercury in a 0.9 % NaCl solution and in tap water had similar

shapes (which could be divided into two linear sections) but with decreased the sensitivity

in tap water.

Acknowledgements: This work was supported by the strategic grant POSDRU/159/1.5/S/133652, co-financed by the European Social Fund within the Sectorial Operational Program Human Resources Development 2007 – 2013.

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doi: 10.5599/jese.2014.0067 187



www.jESE-online.org


Use of hydrous titanium dioxide as potential sorbent for the removal of manganese from water

Ramakrishnan Kamaraj, Pandian Ganesan and Subramanyan Vasudevan

CSIR-Central Electrochemical Research Institute, Karaikudi - 630 006, India

Corresponding Author E-mail: [email protected]; Tel.: +91-4565-241278

Received: October 4, 2014; Revised: October 30, 2014; Published: December 6, 2014

Abstract This research article deals with an electrosynthesis of hydrous titanium dioxide by anodic dissolution of titanium sacrificial anodes and their application for the adsorption of manganese from aqueous solution. Titanium sheet was used as the sacrificial anode and galvanized iron sheet was used as the cathode. The optimization of different experimental parameters like initial ion concentration, current density, pH, temperature, etc., on the removal efficiency of manganese was carried out. The maximum removal efficiency of 97.55 % was achieved at a current density of 0.08 A dm-2 and pH of 7.0. The Langmuir, Freundlich and Redlich Peterson isotherm models were applied to describe the equilibrium isotherms and the isotherm constants were determined. The adsorption of manganese preferably followed the Langmuir adsorption isotherm. The adsorption kinetics was modelled by first- and second- order rate models and the adsorption kinetic studies showed that the adsorption of manganese was best described using the second-order kinetic model. Thermodynamic parameters indicate that the adsorption of manganese on hydrous titanium dioxide was feasible, spontaneous and exothermic.

Keywords: Titanium dioxide; manganese; adsorption; thermodynamics; isotherm; kinetics.

Introduction

Manganese (Mn) is a naturally occurring element found in the air, soil, and water. It can exist in

seven oxidation states ranging from -2 to +7. It is rarely found in its elemental state and is

therefore a component of over 100 minerals and exists mainly as oxides, carbonates, and silicates.

Its most common mineral is pyrolusite (MnO2). In ground water, manganese is a common

contaminant and its presence is due to leaching processes and varies widely depending on the

rock type. Also, manganese has a variety of applications such as in ceramics, metallurgical



J. Electrochem. Sci. Eng. 4(4) (2015) 187-201 TiO2 AS SORBENT FOR THE REMOVAL OF Mn FROM WATER

188 188

processes, mining, dry cell batteries, pigments and paints which all can be the sources of

underground pollution [1]. In addition to the disposal of untreated discharge from above the

applications into water, another major source of pollution of the manganese is burning of coal and

oil [2]. Manganese is an essential metal for the human system and many enzymes are activated by

manganese. The manganese contaminant in ground water affects the intelligent quotient (IQ) of

children. Intake of higher concentrations of manganese causes neuro toxic disease like

Parkinsonism and manganese psychosis, an irreversible neurological disorder [3–5]. The prolonged

over intake potentially affects the central nervous system, lungs, also causes diseases of disturbed

speech called prognosis, also cause bronchitis and pneumonia [6,7]. The World Health

Organization (WHO) prescribed the permissible limit for the manganese in the ground water is

0.05 mg L-1. For this reason, there is great interest in the development of environmentally clean

methods to destroy such compounds in aqueous medium for avoiding their dangerous

accumulation in the aquatic environment.

Because of its high solubility over a wide pH range, toxicity and non-degradable nature, it is

notoriously difficult to remove manganese from contaminated water [8,9]. Hence the researchers

in the world have carried out significant work on their removal from aqueous solutions and

industrial effluents [10-16]. The usual method for removing toxic metals from water include

electrodialysis, chemical coagulation, reverse osmosis, co-precipitation, complexation, solvent

extraction, ion exchange, electrochemical treatment and adsorption. Physical methods like ion

exchange, reverse osmosis and electrodialysis have proven to be either too expensive or

inefficient to remove manganese from water. At present, chemical treatments are not used due to

disadvantages like high costs of maintenance, problems of sludge handling and its disposal, and

neutralization of effluent. In this scenario, the electrochemical technologies have received great

attention for the prevention of pollution problems, as reported in several reviews [17-19].

In the recent decade, electrodissolution process, where the coagulants generated in-situ, has

been increasingly used in the world for treating the industrial wastewater, ground water and

surface water and many studies conducted to optimize this process for specific problems [19]. The

sacrificial anodic electrodes, commonly consisting of iron and aluminum, are used to continuously

supply metallic ions as the source of coagulants, which can hydrolyze near the anode to form a

series of metallic hydroxides capable of destabilizing dispersed particles. This process generates

large quantities of iron and aluminum salt coagulated sludge, which inhibits efficient water

treatment. From the generated coagulant, nothing may be recovered or reused, and require

further incineration and landfill treatment. Furthermore, the appearance of dissolved iron in

aquatic suspensions can lead to visual, odor and taste problems resulting from later growth of iron

bacteria [20]. Even aluminum salts are suspected to be harmful to human and living things [21].

Consequently, a coagulant that is safer and produces more reusable coagulated sludge could offer

a novel solution to many environmental and economic problems associated with sludge handling.

However, reports on novel electrodes materials remain very scarce in the literature for the

generation of reusable and environmentally friendly coagulant. Removal of metal contaminants by

the chemically synthesized, different forms of, titanium dioxide was widely reported [22-25] and

was similar to that of the most widely used iron and aluminum salt flocculation. Furthermore,

long-term toxicological studies have not found titanium salt in water to have any adverse effects.

All the above factors suggest that the titanium salt can be used as an alternative coagulant [26].

In this investigation, titanium was used instead of iron and aluminum as a novel alternative

sacrificial anode, and the removal of manganese from water by titanium-based electrocoagulation

R. Kamaraj et al. J. Electrochem. Sci. Eng. 4(4) (2015) 187-201

doi: 10.5599/jese.2014.0067 189

was investigated. To optimize the maximum removal efficiency of manganese, different

parameters like current density, pH, and temperature, inter electrode distance and co-existing

ions were studied. In doing so, the equilibrium adsorption behavior is analyzed by fitting models of

Langmuir, Freundlich and Redlich Peterson. The adsorption kinetics was modeled by first- and

second- order rate models. Activation energy is evaluated to study the nature of adsorption.

Experimental

Chemicals

Manganese nitrate [Mn(NO3)2] of analytical grade was purchased from MERCK. Hydrochloric

acid (HCl) and sodium hydroxide (NaOH) used for pH adjustment were of analytical grade from

MERCK. Sodium chloride (NaCl) used for better conductivity of electrolyte of analytical grade from

MERCK. Sodium phosphate, sodium silicate, sodium carbonate and sodium fluoride used as co-

existing ions were of analytical grade and purchased from MERCK.

Electrolytic system and electrolysis

The experiments were carried out in a monopolar batch reactor using 1000 mL Plexiglas vessel

that was fitted with a polycarbonate cell cover with slots to introduce the electrodes, pH sensor, a

thermometer and the electrolytes. Titanium (Alfa Aesar, UK) of surface area (0.02 m2) acted as the

anode. The cathode was galvanized iron (commercial grade, India) sheets of the same size as the

anode is placed at an inter-electrode distance of 3 cm. The temperature of the electrolyte was

controlled to the desired value with a variation of ± 2 K by adjusting the rate of flow of thermo-

statically controlled water through an external glass-cooling spiral. A regulated direct current (DC)

was supplied from a rectifier (10 A, 0-25 V; Aplab model).

The required concentration of manganese was prepared using Milli Q water. In all the

experiments 3 g L-1 of sodium chloride was used for better conductivity. The solution volume of 900

mL was used for each experiment as the electrolyte. The pH of the electrolyte was adjusted and

measured initially and during the electrolysis by a pH meter (DKK-TOC, Japan). The pH was

adjusted using either 0.1 M NaOH or 0.1 M HCl as necessary. After adjusting the initial solution pH

to the desired value (3 to 9), the current density was set. The solution was stirred at 250 rpm to

ensure good mixing and transport of reactants. Temperature studies were carried at varying

temperature (323-343 K) to determine the type of reaction.

Analytical procedures

The concentration of manganese was determined using UV-visible Spectrophotometer with

manganese kits (MERCK, Pharo 300, Germany). The SEM image of titanium dioxide was analyzed

with a Scanning Electron Microscope (SEM) made by Hitachi (model s-3000h). The constituents of

the titanium dioxide were analyzed by X-Ray Fluorescence (XRF) made by Horiba (model XGT-

2700). The Fourier transform infrared spectrum of titanium dioxide was obtained using Nexus 670

FTIR spectrometer (Thermo Electron Corporation, USA) and X-ray diffraction (XRD) patterns of

titanium dioxide was analyzed using an X’per PRO X-ray diffractometer (PANalytical, USA). TGA of

titanium dioxide was carried out in the Thermal Analyzer (TA Instruments; Model SDT Q600). The

concentration of carbonate, silicate, and phosphate were determined using UV-Visible

spectrophotometer with respective standard ion kits supplied by MERCK (MERCK, Pharo 300).


190 190


Effect of current density on the removal efficiency

The current density is one of the prominent factors which strongly influence the performance

of electrodissolution process. The current density not only determines the coagulant dosage and

bubble production rate but also the size and growth of the flocks, which can influence the

treatment efficiency. Therefore, the effect of current density on the removal of manganese was

investigated. Applying a constant current to the titanium effectively dissolved Ti according to

Ti→Ti2+→TiO2+. TiO2+ combined easily with OH− to form TiO2·H2O or Ti(OH)4. Ti(OH)4 is unstable

substance, which changes gradually into TiO2·H2O by dehydration. The reaction equations were,

TiO2+ + 6OH- + 2e- → Ti(OH)4 + H2 + 3O2- (1)

and

TiO2+ + 2OH- → TiO2 . H2O (2)

and at the cathode the following reaction is taking place,

2H2O + 2e- → H2 (g) + 2OH- (3)

The amount of manganese removal depends upon the quantity of adsorbent (hydrous titanium

dioxide) generated, which is related to the time and current density [27]. The amount of adsorbent

was determined from Faraday’s law. With the increase in current density the amount of hydrous

titanium dioxide generation also increases. To investigate the effect of current density on the man-

ganese removal, a series of experiments were carried out by solutions containing a constant pol-

lutants loading of 2 mg L-1, at a pH 7.0, with current density being varied from 0.02 to 0.1 A dm-2.

The removal efficiencies are 60.25, 82.54, 90.10, 97.55 and 97.90 % for 0.02, 0.04, 0.06, 0.08 and

0.1 A dm-2 respectively. From the results, it was found that very small raise in removal efficiency

was observed for current densities 0.08 and 0.1 A dm-2. Hence, the further experiments were

carried out at 0.08 A dm-2.

Effect of pH on the removal efficiency

It is believed that the initial pH is an important operating factor influencing the performance of

electrodissolution process. To explain this effect, a series of experiments were carried out using

2 mg L-1 of manganese containing solutions, by adjusting the initial pH in the interval from 3 to 9.

The removal efficiencies for the pH 3, 4, 5, 6, 7, 8 and 9 are 60, 79, 82.5, 95.50, 97.55, 97.56 and

97.65 % respectively. It is well know that the titanium dioxide adsorption is pH dependent. At

acidic pH it is positively charged while at alkaline pH it negatively charged. Mostly point of zero

charge for titanium dioxide is approximately pH 6 to 8 [28,29]. The results agreed well with earlier

results from the literature. Further experiments were carried out at pH 7.

Effect of electrolyte concentration

In order to evaluate the effect of initial concentration of manganese, experiments were

conducted with varying initial concentration from 0.25-2.0 mg L-1. Figure 1 shows that the uptake

of manganese (mg g-1) increased with increase in manganese concentration and remained nearly

constant after equilibrium time. The equilibrium time was found to be 180 min for all

concentration studied. The amount of manganese adsorbed (qe) increased from 0.248 to 1.982 mg

g-1 as the concentration increased from 0.25-2.0 mg L-1. From the Figure 1 it is found that the plots


doi: 10.5599/jese.2014.0067 191

are single, smooth and continuous curves leading to saturation, suggesting the possible monolayer

coverage to manganese on the surface of the adsorbent.

0 100 200 300 400 500

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

0.25 mg / L

0.5 mg / L

1.0 mg / L

1.5 mg / L

2.0 mg / L

qe /

mg

g-1

Time, min Figure 1. Effect of time and initial concentration of manganese for the adsorption

on hydrous titanium dioxide, pH 7.0, T = 303K.

Effect of competing ions

Carbonate

Effect of carbonate on manganese removal was evaluated by increasing the carbonate

concentration from 0 to 250 mg L-1 in the electrolyte. The removal efficiencies are 97.55, 95.3,

72.8, 50.7, 38, and 19 % for the carbonate concentration of 0, 2, 5, 65, 150 and 250 mg L-1,

respectively. From the results it is found that the removal efficiency of the manganese is not

affected by the presence of carbonate below 2 mg L-1. Significant reduction in removal efficiency

was observed above 5 mg L-1 of carbonate concentration is due to the passivation of anode

resulting, the hindering of the dissolution process of anode [30].

Phosphate

The concentration of phosphate ion was increased from 0 to 50 mg L-1, the contaminant

range of phosphate in the ground water. The removal efficiency for manganese was 97.55,

91.3, 60.7, 35.5 and 29.2 % for 0, 2, 5, 25 and 50 mg L-1 of phosphate ion, respectively. There was

no change in removal efficiency of manganese below 2 mg L-1 of phosphate in the water. At

higher concentrations (at and above 5 mg/L) of phosphate, the removal efficiency decreased

drastically. This was due to the preferential adsorption of phosphate over manganese as the

concentration of phosphate increased.


192 192

Arsenic

The concentration of arsenic was gradually increased from 0 to 5 mg L-1. From the results it

was found that the efficiency decrease for manganese was 97.55, 90.7, 78.5, 68.6 and 44.6 % by

increasing the concentration of arsenate from 0, 0.2, 0.5, 2.5 and 5.0 mg L-1, respectively. This was

due to the preferential adsorption of arsenic over manganese as the concentration of arsenate

increases. So, when arsenic ions are present in the water to be treated arsenic ions compete

greatly with manganese ions for the binding sites.

Silicate

From the results it is found that no significant change in manganese removal was observed,

when the silicate concentration was increased from 0 to 2 mg L-1. The respective efficiencies

for 0, 2, 5, 10 and 15 mg L-1 of silicate are 97.55, 80.2, 72.4, 51.6 and 43.8 %. In addition to

preferential adsorption, silicate can interact with titanium dioxide to form soluble and highly

dispersed colloids that are not removed by normal filtration [30].

Adsorption kinetic modeling

The kinetic studies predict the progress of adsorption; however, the determination of the

adsorption mechanism is also important for design purposes. In this research investigation, first-

and second order kinetic models were tested at different concentration (0.25 to 2.0 mg L-1) at a

current density of 0.08 A dm-2.

First order kinetic model

The first order kinetic model is generally expressed as follows [31],

dqt/dt = k1 (qe-qt) (4)

where qe / mg g-1 and qt / mg g-1 are the adsorption capacities at equilibrium and at time t / min

respectively, and k1/ min-1 is a rate constant of first order adsorption. The integrated form of the

above equation with the boundary conditions t = 0 to t = t and qt = 0 to qt = qt is rearranged to

obtain the following time dependence function,

log(qe-qt) = log qe – k1t / 2.303 (5)

The experimental data were analyzed initially with first order model. The plot of log (qe-qt) vs. t

should give the linear relationship from which k1 and qe can be determined by the slope and

intercept, respectively Eq. (5). The computed results are presented in Table 1. The results show

that the theoretical qe (cal) value doesn’t agree to the experimental qe (exp) values at all

concentrations studied with poor correlation coefficient. This result indicated that the adsorption

system do not follow a first-order reaction. So, further the experimental data were fitted with

second order model.

Second order kinetic model

The second order kinetic model is expressed as [32],

dqt/dt = k2 (qe-qt)2 (6)

The integrated form of Eq. (6) with the boundary condition t = 0 to t = t and qt = 0 to qt = qt is,

1/(qe-qt) = 1/qe+ k2t (7)

Eq. (7) can be rearranged and linearized as,


doi: 10.5599/jese.2014.0067 193

t/qt = 1/k2 qe2 + t/qe (8)

where, qe / mg g-1 and qt / mg g-1 are the amount of manganese adsorbed on hydrous titanium

dioxide at equilibrium and at time t / min, respectively, and k2 is the rate constant for the second

order kinetic model.

Table 1 Comparison between the experimental and calculated qe values at different concentrations in first order and second order adsorption kinetics at a current density of 0.08 A dm-2.

C / mg L-1 qe / mg g-1

(exp)

Pseudo first order adsorption Pseudo Second order adsorption

qe / mg g-1

(cal) k1 / min-1 R2 qe / mg g-1

(cal) k2 / (g mg-1) min-1 R2

0.25

0.50

1.0

1.5

2.0

0.248

0.463

0.951

1.470

1.982

35.78

22.84

1.543

1.337

5.780

0.0244

0.0212

0.0168

0.0124

0.0178

0.0069

0.0204

0.6154

0.3462

0.3425

0.248

0.461

0.950

1.470

1.980

1.6715

0.1013

0.1071

0.0874

0.0343

0.992

0.991

0.995

0.986

0.987

The kinetic data were fitted to the second order model Eq. (8). The equilibrium adsorption

capacity, qe (cal) and k2 were determined from the slope and intercept of plot of t/qt versus t and

are compiled in Table 1. Figure 2 shows the plot of t/qt versus t for manganese adsorption and the

plots were found to be linear. The theoretical qe (cal) value also agreed very well with the

experimental qe value, indicating the pseudo second-order kinetics. In addition, the correlation

coefficient for the second-order kinetic model was 0.99, which suggest the applicability of this

kinetic equation and the second-order nature of the sorption process of manganese on hydrous

titanium dioxide.

50 100 150 200 250 300 350 400 450

0

200

400

600

800

1000

1200

1400

1600

1800

(t/q

t) /

min

(m

g/g

)-1

Time, min

0.25 mg / L

0.5 mg / L

1.0 mg / L

1.5 mg / L

2.0 mg / L

Figure 2. Second-order kinetic model plots for adsorption of manganese at different

concentrations, pH of the electrolyte: 7.0, temperature: 303 K, current density: 0.08 A dm-2

The computed results obtained from first order and second order models were depicted in

Table 1. From the tables, it was found that the correlation coefficient values are in the order of


194 194

second order > first order. This indicates that the adsorption follows the second order model.

Further, the calculated qe values well agree with the experimental qe values for second order

kinetics model. These results indicate that the second-order kinetic model can be applied suitably

to predict the manganese adsorption process onto hydrous titanium dioxide.

Isotherm modeling

In order to explain the mechanism of the adsorption process, it is important to establish the

most appropriate correlation for the equilibrium curves. In this study, three adsorption isotherms

viz., Freundlich, Langmuir and Redlich isotherm models were applied to establish the relationship

between the amounts of manganese adsorbed onto the hydrous titanium hydroxide and its

equilibrium concentration in the electrolyte containing contaminant ions.

Freundlich Isotherm

The Freundlich adsorption isotherm typically fits the experimental data over a wide range of

concentrations. This empirical model includes considerations of surface heterogeneity and

exponential distribution of the active sites and their energies. The isotherm is adopted to describe

reversible adsorption and is not restricted to monolayer formation. The linearised in logarithmic

form and the Freundlich constants can be expressed as [33],

log qe = log kf + n log Ce (9)

where, kf is the Freundlich constant related to adsorption capacity, n is the energy or intensity of

adsorption, Ce is the equilibrium concentration of manganese (mg L-1).

In testing the isotherm, the manganese concentration used was 0.25 to 2.0 mg L-1, current

density of 0.08 A dm-2 and at an initial pH 7. The adsorption data is plotted as log qe versus log Ce

by equation (9) should result in a straight line with slope n and intercept kf. The intercept and the

slope are indicators of adsorption capacity and adsorption intensity, respectively. The value of n

falling in the range of 1-10 indicates favorable sorption. Freundlich constant (kf) and n values were

listed in Table 2. From the analysis of the results it is found that the Freundlich plots fit only

satisfactorily with the experimental data obtained in the present study which is shown in the

Figure. 3 (a).

Table 2 Constant parameters and correlation coefficient for different adsorption isotherm models for manganese adsorption at 0.25- 2.0 mg L-1 at a current density of 0.08 A dm-2

Isotherm Parameters Concentration of Mn, mg L-1

0.25 0.5 1.5 1.5 2.0

Langmuir qm / mg g-1 0.2383 0.4597 0.9564 1.4616 1.9491

b / L mg-1 0.1113 0.1102 0.1099 0.1014 0.0948

R2 0.9943 0.9987 0.9954 0.9962 0.9991

RL 0.9729 0.9479 0.8998 0.8569 0.8179

Freundlich

kf / mg g-1 0.5803 0.5512 0.5174 0.4897 0.4613

n / L mg-1) 2.1786 2.0257 1.9457 1.8798 1.7259

R2 0.9812 0.9789 0.9881 0.9836 0.9820

Redlich

Peterson

KF / L g-1 0.9978 0.9981 0.9968 0.9990 0.9891

0.9764 0.9854 0.9817 0.9897 0.9789

aR / L mmol-(1-1/β ) 27.412 28.417 29.648 30.568 32.516


doi: 10.5599/jese.2014.0067 195

log (Ce / mg L

-1) Ce

-1 / g mg

-1

Figure 3. (a) Frendlich plot (log qe vs log Ce) for adsorption of manganese, pH of the electrolyte: 7.0, current density: 0.08 A dm-2, concentration: 2.0 mg L-1, (b) Langmuir plot (1/qe vs. 1/Ce) for adsorption of

manganese, pH of the electrolyte: 7.0, current density: 0.08 A dm-2, concentration: 2.0 mg L-1

Langmuir Isotherm

This model assumes a monolayer deposition on a surface with a finite number of identical sites.

It is well known that the Langmuir equation is valid for a homogeneous surface. The linearized

form of Langmuir adsorption isotherm model is [34],

Ce/qe=1/qmb+Ce/qm (10)

where, qe is amount adsorbed at equilibrium, Ce is the equilibrium concentration, qm is the

Langmuir constant representing maximum monolayer adsorption capacity and b is the Langmuir

constant related to energy of adsorption. The essential characteristics of the Langmuir isotherm

can be expressed as the dimensionless constant RL.

RL=1 / (1+bCo) (11)

where RL is the equilibrium constant it indicates the type of adsorption, b, is the Langmuir

constant. Co is various concentration of manganese solution. The RL values between 0 and 1

indicate the favorable adsorption.

Langmuir isotherm was tested from Eq. (10). The plots of 1/qe as a function of 1/Ce for the

adsorption of manganese on hydrous titanium dioxide are shown in Figure 3 (b). The plots were

found linear with good correlation coefficients (>0.99) indicating the applicability of Langmuir

model in the present study. The values of monolayer capacity (qm) and Langmuir constant (b) is

presented in Table 2. The values of qm calculated by the Langmuir isotherm were all close to

experimental values at given experimental conditions. These facts suggest that manganese is

adsorbed in the form of monolayer coverage on the surface of the adsorbent. The dimensionless

constant RL was calculated from Eq.(11). The RL values were found to be between 0 and 1 for all the

concentration of manganese studied. The correlation co-efficient values of Langmuir and

Freundlich isotherm models are presented in Table 2.

Redlich Peterson isotherm

It is a three parameter hybrid isotherm. It is having features of both the Langmuir and Frendlich

isotherms. This model has linear dependence in the numerator component and exponential

component in denominator of non-linear form [35,36].

log

(qe /

mg

g-1)

(t /

qe)

/ m

in (

mg

/ g)

-

1


196 196

β

eR

eFe

1 Ca

CKq

(12)

where qe / mmol g-1 is the solid-phase sorbate concentration at equilibrium, Ce / mmol L-1 is the

concentration of adsorbate in equilibrium with liquid phase, KF / L g-1 and aR / L mmol-(1-1/β) are the

Redlich-Peterson isotherm constants, and is the exponent, which lies between 1 and 0. If the

tends to 0 then the adsorption follows the Frendlich isotherm and if the value tends to one it fits

with the Langmuir isotherm. In order to verify our investigation regarding the monolayer or

multilayer adsorption, the linear form of Redlich-Peterson is used. It is little bit complicated

compared to the other isotherms. The strategy to find these parameters is based in the

maximization of correlation coefficients (R2) from the linear fit to the data. In this way

the KF values are modified until obtain the best fit of the data. The linear form of the equation for

this model is

ln (KF(Ce/qe-1) = ln aR + ln Ce (13)

Plotting of ln (KF(Ce/qe-1) vs. ln Ce by the Eq. (13) gives the Redlich Peterson equation. This

isotherm is a three parameters isotherm in which KF values are indirectly obtained by plotting the

graph with maximum correlation coefficient by justifying the values of KF. By that KF, and aR are

in the Table 2 for all concentrations. Here the values are above the 0.95. So the adsorption

favors Langmuir isotherm rather than Frendlich isotherm.

Adsorption thermodynamics

To understand the effect of temperature on adsorption process, thermodynamic parameters

should be determined at various temperatures. The energy of activation for adsorption of manga-

nese can be determined by the second order rate constant is expressed in Arrhenius form [37],

ln k2 = ln ko - E/RT (14)

where ko is the constant of the equation (g mg-1) min-1), E is the energy of activation (J mol-1), R is

the gas constant (8.314 J mol-1 K-1) and T is the temperature (K). Figure 4(a) shows that the rate

constants vary with temperature according to Eq.(14) giving an activation energy of 21.01 kJ mol-1

for manganese from the slope of the fitted equation. The free energy change is obtained using the

following relationship

∆G = -RT ln Kc (15)

where ∆G is the free energy (kJ mol-1), Kc is the equilibrium constant, R is the universal gas

constant and T is the temperature in K. The values of Kc and ∆G are presented in Table 3. The

negative value of ∆G indicates the spontaneous nature of adsorption. Other thermodynamic

parameters such as entropy change (ΔS) and enthalpy change (ΔH) were determined using the

van’t Hoff equation:

RT

H

R

SK

cln (16)

The enthalpy change (∆H = -60.57 J mol-1) and entropy change (∆S = -0.047 J mol-1 K-1) were

obtained from the slopes and intercepts of the van't Hoff linear plots of ln Kc versus 1/T (Figure.

4(b)) Eq.(16). Negative value of enthalpy change (∆H) indicates that the adsorption process is

exothermic in nature, and the negative value of change in internal energy (∆G) show the

spontaneous adsorption of manganese on the adsorbent. Negative values of entropy change show


doi: 10.5599/jese.2014.0067 197

the increased randomness of the solution interface which gains heat from the surroundings during

adsorption of manganese on the adsorbent [38] (Table 3). Negative enthalpy and negative entropy

shows that the adsorption is more favorable at low temperature. This is due to the decrement of

pore size as temperature increases.

T

-1 / K

-1 T

-1 / K

-1 Figure 4. (a) Plot of log k2 vs. T-1; (b) Plot of ln Kc vs. T-1: pH 7.0; j = 0.08 A dm-2, C = 2 mg L-1

The enthalpy change (∆H = -60.57 J mol-1) and entropy change (∆S = -0.047 J mol-1 K-1) were

obtained from the slopes and intercepts of the van't Hoff linear plots of ln Kc versus 1/T (Figure.

4(b)) Eq.(16). Negative value of enthalpy change (∆H) indicates that the adsorption process is

exothermic in nature, and the negative value of change in internal energy (∆G) show the

spontaneous adsorption of manganese on the adsorbent. Negative values of entropy change show

the increased randomness of the solution interface which gains heat from the surroundings during

adsorption of manganese on the adsorbent [38] (Table 3). Negative enthalpy and negative entropy

shows that the adsorption is more favorable at low temperature. This is due to the decrement of

pore size as temperature increases.

Table 3 Thermodynamics parameters for adsorption of manganese.

Temperature, K Kc ∆Go / kJ mol-1 ∆Ho / Jmol-1 ∆So / J mol-1 K-1

323

333

343

1.0433

1.0406

1.0382

-0.0464

-0.0450

-0.0436

-60.57

-0.0470

Table 4. Comparison between the experimental and calculated qe values at different temperatures in first and second order adsorption kinetics of manganese: C = 2.0 mg L-1, pH 7.0, j = 0.08 A dm-2

T / K qe / mg g-1

(exp)

First order adsorption Second order adsorption

qe / mg g-1

(cal) k1 / min-1 R2 qe / mg g-1

(cal) k2 / (g mg-1) min-1 R2

323

333

343

2.048

2.160

2.165

1.112

0.987

0.874

-0.0061

-0.0068

-0.0071

0.5430

0.4569

0.4037

2.01

2.13

2.14

0.0320

0.0095

0.0097

0.990

0.986

0.991

Using Lagergren rate equation, pseudo second order rate constants and correlation co-efficient

were calculated for different temperatures (323-343 K). The calculated qe values obtained from

the second order kinetics agrees with the experimental qe values better than the first order

log

(qe /

mg

g-1)

ln (

k 2 /

(g

mg-1

) m

in-1

)


198 198

kinetics model, indicating adsorption following second order kinetics. Table 4 depicts the

computed results obtained from pseudo first and pseudo second order kinetic models.

Characterization of hydrous titanium dioxide

XRF studies

The contents of titanium dioxide were analyzed by XRF. The titanium dioxide sample was dried

in a drying chamber at 100 °C were ground in the agate mortar. As shown in Table 5, 95.2 % of the

sample, by weight, was titanium dioxide. Sodium chloride was came from the electrolyte, and

calcium and barium elements came from the impurity in the titanium electrode.

XRD studies

The crystal structure of hydrous titanium dioxide nanoparticles was analyzed by X-ray powder

diffractometer operating with CuKα radiation source filtered with a graphite monochromator.

Figure. 5(a) shows the X-ray diffraction pattern of hydrous titanium dioxide nanoparticles. From

the figure it is found that, most diffraction peaks belong to the anatase phase (JCPDS Card Number

73-1764), and minor peaks from the brookite phase (JCPDS Card Number 76-1936) could also be

observed. The crystallite size D was determined from the broadening of corresponding strongest

X-ray diffraction peaks by using Scherrer's formula [39]:

cos

9.0D (17)

where D is the crystalline size, λ is the average wavelength of the X-ray radiation (λ = 1.5418 Å),

is the line-width at half-maximum peak position, and is the diffracting angle (2 = 25.4°). The

average crystallite size of the hydrous titanium dioxide is 4.3-8.4 nm.

FT-IR spectrum

The hydrous titanium dioxide was analyzed using FTIR and results are presented in Figure 5(b).

The strong peak at 3381 cm−1 is attributed to the stretching vibrations of surface and interlayer

water molecules and hydroxyl groups. This is related to the formation of hydrogen bonds of inter-

layer water with guest anions as well as with hydroxide groups of layers. At 1630.18 cm-1, there is

a strong adsorption peak for hydroxyl bending vibration belonging to physically adsorbed H2O.

One small adsorption peak could also be identified at 1371.37 cm-1, which represents the

coordinated hydroxyl groups. These observations demonstrate that these hydrous titanium

dioxide nanoparticles have high adsorption capacities to H2O and hydroxyl groups exist on their

surfaces [40].

SEM and EDAX analysis

Figure 5(c) shows the SEM images of hydrous titanium dioxide. The SEM images show different

size, shape and dimension and these nanoparticles are aggregated into micro-sized particles. The

volume median diameter value of these nanoparticles in distilled water was determined at

approximately 9.8 nm by the dynamic light scattering technique, which is in accordance with the

SEM observation. This type of aggregation of nanoparticles is beneficial to their removal from

aqueous environment after the treatment process.

Energy-dispersive analysis of X-rays was used to analyze the elemental constituents of titanium

dioxide generated during the electro dissolution process and the results are presented in

Figure 5(d). The figure indicates that the titanium dioxide was composed mainly of Ti an O, which

affirms that the titanium dioxide was generated by anodic dissolution.


doi: 10.5599/jese.2014.0067 199

a

b

Wavenummber, cm-1

c

d

Figure 5. (a) X-ray diffraction pattern of hydrous TiO2, (b) FTIR pattern of hydrous TiO2, (c) SEM image of the hydrous TiO2, (d) EDAX image of the hydrous TiO2

TGA analysis

TGA analysis (figure not shown) of hydrous titanium dioxide was carried out. From the results

we found that the weight loss of 13.0 % where observed when the samples were heated from the

32 - 800°C. The entire range will be divided into three stages viz., first, second and third stage. In

the first stage (32 – 122 °C) weight loss (6.9%) could be attributed to the elimination of physically

absorbed water. In the second stage (122 to 438 °C) weight loss (6.0 %) could be contributed to

the loss of surface hydroxyl groups. In the third stage (438 to 800 °C) no exothermic peak was

observed and the weight loss is around 0.1 %.

Conclusions

The maximum removal efficiency of 97.55 % was achieved with titanium as sacrificial anode at a

current density of 0.08 A dm-2, pH 7.0. The results indicate that the hydrous titanium dioxide, by

electro-dissolution of sacrificial anodes, efficiently adsorbs the manganese from water. Hence this

process can be used as an effective process for the removal of manganese contaminated water

resources. The results indicate that, the second-order kinetic model accurately described the

adsorption kinetics. The adsorption mechanism was found to be chemisorption and the rate-


200 200

limiting step was mainly surface adsorption. The Langmuir isotherm showed a better fit than the

Freundlich and Redlich isotherms, thus, indicating the applicability of monolayer coverage of

manganese on hydrous titanium dioxide.

The thermodynamic parameters like ΔG, ΔH and ΔS were determined. Their values indicated

that the adsorption process was favorable, spontaneous, and exothermic in nature. As the

temperature increased ΔG became less negative, indicating a stronger driving force, resulting in a

greater adsorption capacity at higher temperatures. The negative value of ΔH confirmed that the

process was exothermic. Negative values of entropy change show the increased randomness of

the solution interface which gains heat from the surroundings during adsorption of manganese on

the adsorbent. EDAX analysis confirmed that manganese was adsorbed on to the hydrous titanium

dioxide.

Acknowledgments: The authors wish to express their gratitude to Dr. Vijayamohanan K. Pillai, Director, CSIR-Central Electrochemical Research Institute, Karaikudi to publish this article.

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doi: 10.5599/jese.2014.0062 203



www.jESE-online.org


Anodic oxidation of oxytetracycline: Influence of the experimental conditions on the degradation rate and mechanism

Annabel Fernandes, Catarina Oliveira, Maria J Pacheco, Lurdes Ciríaco and Ana Lopes

UMTP and Department of Chemistry, University of Beira Interior, 6201-001 Covilhã, Portugal

Corresponding author: E-mail: [email protected]; Tel.: +351-275-329-259; Fax: +351-275-319-730

Received: August 19, 2014; Revised: September 1, 2014; Published: December 6, 2014

Abstract The anodic oxidation of oxytetracycline was performed with success using as anode a boron-doped diamond electrode. The experiments were conducted in batch mode, using two different electrochemical cells: an up-flow cell, with recirculation, that was used to evaluate the influence of recirculation flow rate; and a stirred cell, used to determine the influence of the applied current density. Besides oxytetracyclin electrodegradation rate and mineralization extent, oxidation by-products were also assessed. Both the flow rate and the applied current density have shown positive influence on the oxytetracycline oxidation rate. On the other hand, the mineralization degree presented the highest values at the lowest flow rate and the lowest current density tested. The main oxidation by-products detected were oxalic, oxamic and maleic acids.

Keywords Tetracyclines; BDD; antibiotics; pharmaceutical compounds; electrochemical degradation

Introduction

The increasing use of drugs has become a new environmental problem, which has aroused

great concern in recent years. Although these compounds are found in very low concentrations in

the environment, there is still a lack of information about the long-term risks that the presence of

a wide variety of drugs can bring to ecosystems and to human health. These drugs are

continuously introduced into the environment due to their domestic, hospital and veterinary use,

and its presence has been detected in wastewaters [1-4]. Its potential biological activity associated



J. Electrochem. Sci. Eng. 4(4) (2014) 203-213 ANODIC OXIDATION OF OXYTETRACYCLINE

204

with low removal during conventional wastewater treatment processes, can lead to adverse

environmental effects, including the contamination of soil and water resources [5,6].

Among these drugs, tetracyclines are one of the most widely used in the prophylaxis and

therapy of human and animal infections and also at subtherapeutic levels in animal feed as growth

promoters [7,8]. Tetracyclines are considered bacteriostatic antibiotics, although they have also

various non-antibiotic properties. They are characterized by a four ring structure with a

carboxamide functional group and by several ionizable functional groups [8,9]. As a result of the

waste disposal, the drug is transferred to different environmental compartments (water,

sediment, soil) and can contaminate trophic network, transitioning into food and cause negative

effects on natural resources, including the effects on microbial community structure and selection

of strains with antibiotic resistance [10-12]. The presence of tetracyclines in several environmental

matrices has been investigated and the evidence of their existence has been reported [8,13-15].

This is due to the fact that antibiotics are very resistant to biodegradation..

In recent years, there has been a growing awareness about pollution caused by pharmaceutical

wastes, including antibiotics [16,17]. Biological processes, the most economical for wastewater

treatment, have been extensively studied, but they are ineffective in the removal of recalcitrant

compounds with poor biodegradability [17,18]. Some physical and physicochemical techniques,

such as coagulation, flocculation, adsorption, ultrafiltration and reverse osmosis, have been

successfully employed to remove recalcitrant pollutants. However, these conventional treatments

simply transfer pollutants from one phase to another, resulting in secondary pollution [17,19].

New technologies based on the application of advanced oxidation processes have been

reported in the treatment of effluents containing tetracycline [20-37]. Tetracycline degradation

efficiency obtained by photo-Fenton processes in the treatment of wastewaters and surface

waters was reported by Bautitz and Nogueira [21]. Results showed that the photo-Fenton process

under solar radiation can be applied in the degradation of tetracycline present in surface water

samples or even in more complex samples, such as effluent from sewage treatment plants.

However, the processes that involve photolysis should only be applied to bleached effluents, since

the color prevents the efficient propagation of radiation. Li et al. [23] studied the effect of

different pH values on oxytetracycline degradation by ozonation in aqueous solution and found

that this technique had potential to be used as a partial step in combined treatment of

pharmaceutical effluents containing high concentrations of oxytetracycline. The removal rates

increased as a result of high decomposition rates, favored by pH increase. However,

bioluminescence results indicate that after partial ozonation, byproducts of oxytetracycline have a

higher toxicity than the parent compound. Jiao et al. [22] and Shaojun et al. [24] studied the

degradation of tetracyclines by photolysis and reported high COD removals, of 80 %, but very low

TOC removal, about 14 %, which indicated the production of intermediate compounds. On their

studies, it was also found that the toxicity of the treated effluent was higher than in the original

effluent. Photocatalysis studies applied to the treatment of waters with low loads of organic

matter, such as rivers, ground water and drinking water, containing tetracyclines have been

performed by several authors and high removals were obtained, indicating that this method is

promising for this type of waters [20,35].

Electrochemical processes have shown to be effective for the treatment of effluents containing

refractory and toxic organic pollutants [38-41]. Furthermore, they are an effective, versatile, easy

and clean technology [39,42]. For all these reasons, this technology has been applied to remove

tetracycline under different experimental conditions. Removals above 90 % were achieved using

A. Fernandes et al. J. Electrochem. Sci. Eng. 4(4) (2014) 203-213

doi: 10.5599/jese.2014.0062 205

Ti/RuO2-IrO2 anodes [25] and Ti/IrO2 anodes [29]. Despite these good results in tetracyclines

removal, no significant levels of mineralization were achieved using these active anodes. In the last

decade, BDD anode, a non-active anode, is being widely used. It has several advantages, namely

good chemical and electrochemical stability, extended lifetime and a high overpotential for water

decomposition, being known for its ability to promote complete mineralization of a wide range of

organic pollutants [39,43-47] due to the hydroxyl radicals formed from water decomposition on

the electrode surface. Brinzila et al. [30] reported an electrodegradation study of tetracycline on

BDD anode where removals of 93 % and 87 % were obtained for COD and TOC, respectively. These

authors also studied the influence on the degradation rate of initial pH, applied current intensity

and electrolyte added [37]. It was observed that an increase in current density leads to a decrease

in the current efficiency of the process and the complete removal of tetracycline was much faster

in the presence of chloride ions that promoted the complete degradation of this antibiotic in 30

min. The effect of different anode materials in both electrochemical oxidation and electro-Fenton

processes on the oxidation of tetracycline was investigated by Oturan et al. [36]. They have

reported that processes using BDD anode demonstrated superior oxidation/mineralization power.

Almost total mineralization (TOC removal up to 98 %) of 100 mg L−1 of tetracycline solutions was

achieved after 6 h treatment with BDD anode.

Considering the good results obtained with the BDD anode, in the present study it is proposed

its use in the electrochemical degradation of oxytetracycline, an antibiotic from the tetracycline

family, widely used in intensive animal husbandry to treat enteric and respiratory diseases. The

influence of the hydrodynamics inside the electrochemical cell on the rate of electrodegradation

and mineralization of oxytetracycline was studied and the oxidation by-products were also

assessed, in order to establish the degradation mechanism.

Experimental

Oxytetracycline (OTC) used in this study was purchased from Sigma Aldrich (purity 99 %), with

the chemical formula C22H24N2O9.2H2O (Table 1), and used without further purification.

Oxytetracycline degradation experiments were conducted in batch mode using two different

electrochemical cells. The first set of assays was performed in an up-flow electrochemical cell, with

recirculation, composed by a BDD anode with an area of 20 cm2 and a stainless steel cathode with

identical area, using 250 mL of solution. The recirculation of the solution was enabled by a

centrifugal pump, Pan World Magnet, Model: NH-30PX, Pan World Co., Ltd. Tokyo, Japan, which

allowed the evaluation of different flow rates: 37, 75, 100 and 120 L h-1. The applied current

density was kept constant at 20 mA cm-2. The second set of assays was conducted in batch mode,

with stirring, in a cell containing a BDD anode, with an immersed area of 10 cm2, and a stainless

steel cathode, with identical area. 200 mL of solution were used in each run. In order to study the

oxidation mechanism and identify the by-products, assays were performed applying different

current densities: 2.5, 5, 7.5, 10, 20 and 30 mA cm-2. The experimental conditions used are

summarized in Table 1.

For all the experiments performed, the initial oxytetracycline concentration was 100±10 mg L-1.

The assays were conducted at room temperature (25±2 °C), adding as support electrolyte

anhydrous sodium sulfate (Merck, 99.5 %), in a concentration of 5 g L-1. A GW, Lab DC, model

GPS-3030D (0–30 V, 0–3 A), was used as power supply. The assays were performed in duplicate,

and the values presented for the parameters used to follow the assays are the mean values.


206

Degradation tests were followed by total organic carbon (TOC) and total nitrogen (TN),

measured in a Shimadzu TOC-V CPH analyzer combined with a TNM-1 unit, by chemical oxygen

demand (COD), performed using closed reflux and titrimetric method, and by ammonia nitrogen

(AN), using a Vapodest 20s distillation system from Gerhardt, according to standard procedures

[48]. UV–Visible absorption spectra were also performed, with measurements made between 200

and 800 nm, using a Shimatzu UV-1800 spectrophotometer. High performance liquid chromate-

graphy (HPLC) was performed using a Shimadzu 20A Prominence HPLC system equipped with a

diode array detector SPD-M20A, a column oven CTO-20AC and a pump LC-20AD SP. For

oxytetracycline determination a RP-18 reversed phase Purospher STAR column (250 × 4 mm (i.d.),

5 µm) was used and the elution was performed isocratically with an oxalic acid aqueous solution

(10 mM): acetonitrile, 70:30 (v/v), mixture at a flow rate of 1 mL min-1 and 30 ºC. The carboxylic

acids determination was made by ion-exclusion chromatography using a Biorad Aminex HPX-87H

column (300 × 7.8 mm (i.d.)) and the elution was performed isocratically with a sulfuric acid

aqueous solution (4 mM) at a flow rate of 0.6 mL min-1 and 35 °C. The selected wavelength was

354 nm for oxytetracycline and 210 nm for carboxylic acids. The reagents used were HPLC grade

and supplied by Sigma-Aldrich. All the solutions for HPLC were prepared with ultrapure water

obtained with Milli-Q system. Measurements of pH were carried out with a Mettler-Toledo pH-

meter. Conductivity was determined using a conductivity meter Mettler Toledo (SevenEasy S30K).

Table 1. OTC chemical structure and experimental conditions used in the OTC oxidation assays.

OTC chemical structure

Operating mode

Flow rate, L h-1 Stirring speed, rpm [OTC]0 / mg L-1 j / mA cm-2

OH

CH3

OH

CH3

CH3

N

OH

NH2

O

O

OH

OH

O

OH

Batch with recirculation

37

- 100 20 75

100

120

Batch with stirring

- 100 100

2.5

5.0

7.5

10

20

30


Figure 1 presents variation in time of COD, TOC and OTC concentration for the first set of assays

performed in batch with recirculation conditions at different flow rates. Initial COD and TOC values

are slightly different for the various experiments performed, since fresh solutions were prepared

for each assay to avoid OTC photodegradation, and the values presented are the mean values

obtained for the different replicas. Up to 4 h, there is a regular decay in time of COD and TOC.

After that, particularly at higher flow rates, there is a decrease in the organic load removal rate.

However, after 8 h degradation, for the flow rates tested (37, 75, 100 and 120 L h-1) the remaining

CODs were 17, 12, 15 and 17 %, and the remaining TOCs were 8, 9 11 and 9 %, respectively,

meaning that the flow rates used almost didn’t interfere with the organic load removal. Similar

behavior was already observed by other authors [49,50]. On the other hand, if data for the first 4

hours assay is used to calculate the TOC/COD ratios (insets of Figure 1), different slopes can be


doi: 10.5599/jese.2014.0062 207

obtained for the different flow rates, showing the influence of this parameter on the degradation

mechanism. In fact, TOC vs. COD slope decreases with the increase in flow rate, showing that the

increase in flow rate has a negative impact on the OTC mineralization degree at the earlier stages

of the electrodegradation assay. This happens because the increase in flow rate decreases the dif-

fusion layer width, favoring the counter diffusion of the reaction intermediate products, avoiding

their complete mineralization. After 4 h assay, the TOC vs. COD slope changes, since the products

in solution are other than OTC, as can be seen by the OTC concentration determined by HPLC.

Regarding the OTC concentration decay (Figure 1), the electrodegradation process kinetics is

almost independent of the imposed flow rate. It presents a pseudo-first order kinetic and only the

assay performed at the lowest flow rate shows a lower kinetic constant (1.18 h-1), probably due to

some diffusional hindrance at low flux and to cathodic reactions that can contribute to the overall

kinetic process (including COD and TOC kinetic rates).

Figure 1. Variation of COD, TOC and [OTC] with time for the electrodegradation assays

performed in batch with recirculation mode for the different flow rates tested. Insets: Variation of TOC with COD along time.

The samples collected during the assays were also used to run UV-Vis absorption spectra and

results for the flow rates of 37 and 120 L h-1 are presented as the absorbance variation in time me-

asured at 276 and 355 nm, the two major OTC characteristic absorption bands (Figure 2, a and b).

Similarly to COD and TOC, the absorbance decay at both wavelengths increase with flow rate,

being the decay at 355 nm faster than at 276 nm. This means that in the OTC molecule the ring

containing the N-groups (responsible for the absorbance at 276 nm) is not so easily opened as the

other rings, or the intermediate products formed also absorb at this wavelength, thus contributing

to increase the absorbance at 276 nm.


208

Regarding nitrogen removal from solution (Figure 2c), there is only a very slight decay in the

total nitrogen amount and an increase followed by a decrease in the ammonium nitrogen

concentration during the 8 h assays. This means that the organic nitrogen is slowly converted into

ammonium and, only when the organic load is very small, ammonium is oxidized to nitrogen

volatile species, being this conversion higher for higher flow rates, probably because COD and TOC

removal rates also increase with flow rate.

Figure 2. Variation with time of absorbance, measured at (a) 276 and (b) 355 nm, (c) AN and

TN for the electrodegradation assays performed in batch mode with recirculation for two different flow rates tested: 37 and 120 L h-1.

The influence of the current density on the OTC degradation rate was studied in a stirred batch

system and results for the decays in COD, TOC and absorbance, measured at 276 and 355 nm, are

presented in Figure 3. The absolute COD and TOC removals increased with current density, mainly

due to the increase in indirect oxidation promoted by the hydroxyl radicals, formed when the

applied current exceeds the limiting current corresponding to the organic load of the solution.

Simultaneously, the formation of persulfate radicals can also happen, since sulfate was the chosen

electrolyte.

The absorbance variation measured at the OTC characteristic absorption bands presents a

behavior that seems dependent on the applied current density. For the lowest applied current,

after 4 h assay there is a divergence between the absorbance curves at 276 and 355 nm, with an

increase in the absorbance at 276 nm, meaning that products that absorb at this wavelength are

being formed. These products must be resistant to oxidation, since the curve related with

absorbance at 355 nm suffers a sudden decay that must be related with an increase in the

degradation of the remaining OTC, since at this applied current density the OTC decay is slower, as

will be discussed below.

For higher applied current densities, there is a separation between the absorbance curves

measured at the two characteristic wavelengths that increases with the applied current density,

due to reasons already discussed. This fact must also be a consequence of the different minera-

lization degree for the different experimental conditions, as can be observed in Figure 4. In fact,

the mineralization degree decreases with applied current densities between 2.5 and 7.5 mA cm-2,

showing a smaller increase for higher current densities. This behavior must be related with


doi: 10.5599/jese.2014.0062 209

hydroxyl radicals’ indirect oxidation and the diffusion hindrance promoted by the oxygen

evolution at higher applied current densities.

Figure 3. COD and TOC decays with time for the electrodegradation assays performed in batch with stirring mode for the different applied current densities tested. Insets: Relative absorbance

decays in time, measured at 276 and 355 nm.

Regarding the electrodegradation kinetics (Figure 5), the OTC concentration decays show a

dependence on the applied current density. As previously observed for the assays run in batch

mode with recirculation, it presents a pseudo-first order kinetic, with the kinetic constant (Figure

5, symmetric of the slopes of the adjusted equations) increasing with current density.


210

Figure 4. TOC vs COD variation for the electrodegradation assays performed in batch mode

with stirring at different applied current densities.

Figure 5. Variation of OTC concentration with time for the electrodegradation assays

performed in batch mode with stirring at different applied current intensities.

In order to compare the values obtained in both operating systems used, the slope obtained for

20 mA cm-2 was converted to the same units as those in Figure 1, giving 0.71 h-1, showing that in

the applied experimental conditions batch with stirring operating mode is less efficient than batch

with recirculation mode, probably due to the importance of the OTC diffusion to the reaction zone

in the degradation process.

The concentration of the main reaction intermediate products was also followed by HPLC.

Besides the main carboxylic acids detected, oxalic, oxamic and maleic acids and vestiges of formic

acid were also identified, particularly at higher current densities. In Figure 6, the variation in time

of the oxalic, oxamic and maleic acids is presented for the current densities of 2.5, 10 and

30 mA cm-2, for the two first current densities between 0 and 8 h and for 30 mA cm-2 between 0

and 12 h. When OTC is the main organic compound in solution, there is an increase in those

intermediates concentration. After that, their concentration start to decrease and they are

mineralized. The existence of only small dicarboxylic acids with conjugated double bonds points to

a degradation mechanism characteristic of indirect oxidation, where the parent molecule is

attacked in many different places, leaving unchanged double conjugated bond systems, less easily

oxidized.


doi: 10.5599/jese.2014.0062 211

Figure 6. Influence of the current density on the variation with time of the concentration of

some of the intermediate products formed during the OTC electrodegradation, performed in batch with stirring mode at different applied current intensities.

Conclusions

In this study oxytetracycline was successfully degraded through anodic oxidation with a BDD

anode. The investigation of the influence of the experimental conditions on the OTC degradation

allowed drawing the following conclusions:

- Degradation mechanism occurred mainly through indirect oxidation.

- COD and TOC removals increased with current density, being almost independent of the flow

rates tested, meaning that the current density plays an important role in the OTC oxidation

rate, mainly due to the formation of hydroxyl and persulfate radicals that are the main species

responsible for the indirect OTC oxidation.

- The increase in flow rate has a negative impact on the OTC mineralization degree.

- The organic nitrogen is slowly converted into ammonium and, only when the organic load is

very small, ammonium is oxidized to nitrogen volatile species, being this conversion higher for

higher flow rates.

- The electrodegradation process presents pseudo-first order kinetic and the kinetic constant

increases with current density. For flow rates higher than 75 L h-1, the process kinetics is almost


212

independent of the imposed flow rate. However, for lower flow rate the diffusion hindrance

leads to lower OTC removal rates.

- The main by-products detected were oxalic, oxamic and maleic acids, whose concentration

increased while OTC was the main organic compound in solution. After that, the by-products

concentration started to decrease, indicating their mineralization.

Acknowledgements: Financial support from FEDER, Programa Operacional Factores de Competitividade – COMPETE, and FCT, for the project PEst-OE/CTM/UI0195/2011 of the MTP Unit and for the grant awarded to A. Fernandes SFRH/BD/81368/2011.

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www.jESE-online.org


Phenolic compounds removal from mimosa tannin model water and olive mill wastewater by energy-efficient electrocoagulation process

Marijana Kraljić Roković, Mario Čubrić and Ozren Wittine

Faculty of Chemical Engineering and Technology, University of Zagreb, Marulićev trg 19, Croatia

Corresponding Author E-mail: [email protected]; Tel.: +-385-1-4597112; Fax: +-385-1-4597139

Received: August 1, 2014; Revised: August 26, 2014; Published: December 6, 2014

Abstract

The objective of this work was to study the influence of NaCl concentration, time, and current density on the removal efficiency of phenolic compounds by electrocoagulation process, as well as to compare the specific energy consumption (SEC) of these processes under different experimental conditions. Electrocoagulation was carried out on two different samples of water: model water of mimosa tannin and olive mill wastewater (OMW). Low carbon steel electrodes were used in the experiments. The properties of the treated effluent were determined using UV/Vis spectroscopy and by measuring total organic carbon (TOC). Percentage of removal increased with time, current density, and NaCl concentration. SEC value increased with increased time and current density but it was decreased significantly by NaCl additions (0-29 g L-1). It was found that electro-coagulation treatment of effluents containing phenolic compounds involves complex formation between ferrous/ferric and phenolic compounds present in treated effluent, which has significant impact on the efficiency of the process.

Keywords

Complexation; NaCl; low carbon steel, UV/Vis spectroscopy, total organic carbon (TOC).

Introduction

High amounts of waste water are generated in the Mediterranean area each year during the

short periods of the olive oil production process, from October to December. Its direct disposal in

nature is not acceptable since it contains dark colour, organic materials, and emulsion oils.

Additionally, such waters contain phenolic compounds that have a negative impact on vegetation



J. Electrochem. Sci. Eng. 4(4) (2014) 215-225 PHENOLIC COMPOUNDS REMOVAL BY ELECTROCOAGULATION

216

and microorganisms, and therefore must be treated in order to remove organic and toxic

pollutants. The amount of these compounds is usually in the range of 5-25 g L-1 depending on

climate, variety of olive fruit, cultivation, ripeness at harvest, as well as on extraction process [1,2].

Different methods have been developed for olive mill wastewater (OMW) treatment such as

biological treatment [3-5], physico-chemical treatments [6], photocatalytic oxidation [7,8],

electrooxidation [9], electrocoagulation [10-13], and a number of combined treatments [14,15].

Phenolic compounds extracted from OMWs could possibly be utilized in the cosmetic,

pharmaceutical, or food industries [16], or even as potential source of natural dyes [17]. However,

none of treatments used were completely satisfying.

Electrocoagulation is a well-known remediation technique that can be used alone [10-18] or in

combination with other techniques [19]. It is a simple, effective, and low cost process, easily

adaptable to other systems. Furthermore, in remote areas without electricity, it could be directly

powered by a photovoltaic system in order to achieve a self-sustainable unit [20-22].

The goal of this work was to study the influence of NaCl concentration, current density and

time on the removal efficiency of phenolic compounds by the electrocoagulation process, as well

as to compare specific energy consumption (SEC) of the processes under different experimental

conditions. Electrocoagulation was carried out for two different samples of water: model water of

mimosa tannin and olive mill wastewater (OMW). Mimosa tannin is a phenolic compound similar

to those phenolic compounds present in OMW but it is somewhat more complex. In this work it

was used in order to examine the influence of pure phenolic compounds on the removal efficiency

of the electrocoagulation process. After detailed analysis of the results obtained in model water,

experiments were also carried out in OMW. Electrocoagulation processes are usually carried out

without or with only small amounts of NaCl to avoid its presence in discharge water. However,

since most olive oil production plants are situated on the Mediterranean coastline, additions of

NaCl and the disposal of the treated effluent containing NaCl into the sea is acceptable.

Additionally, it is well known that the amount of salt is the crucial parameter for energy

consumption in the electrolysis process, and for the feasibility of this technique.

Experimental

All the chemicals used in this research were of analytical grade. Mimosa tannin (Mimosa

Central Co-operative Ltd., South Africa) solutions were prepared using bi-distilled water with the

addition of an appropriate amount of NaCl.

The olive mill wastewater (OMW) used in this study was obtained from a local olive oil

manufacturer (Croatia). It was stored in an open puddle for two months before the sampling. Prior

to the experiments the OMW was filtrated to remove suspended solids.

Conductivities and pH values were measured using a conductivity meter (Oakton PCD650) and a

pH meter (Radiometer, PHM 220).

At the end of the electrocoagulation process of the mimosa tannin model water, the formed

colloids were left during the night to settle and afterwards the sludge was separated from the

effluent on a Buchner funnel using a water aspirator. The sludge was dried in open air for three

days and then the mass of the sludge was determined. Treated effluents were analysed by

different techniques (pH meter, conductivity meter, UV/Vis spectroscopy, and TOC). Before the

UV/Vis spectroscopy and TOC analysis were carried out, the solutions was centrifuged using

9000 rpm (Hettich Universal, Mikro 12-24 centrifuge).

M. Kraljić Roković et al. J. Electrochem. Sci. Eng. 4(4) (2014) 215-225

doi: 10.5599/jese.2014.0066 217

At the end of electrocoagulation process of the OMW, the water was immediately filtrated in a

Buchner funnel and analysed using the same techniques as in the case of mimosa tannin solution.

The phenolic compounds concentration was measured at the wavelength corresponding to the

maximum absorbance using a spectrophotometer (Ocean Optics 200, UV light source Analytical

Instrument Systems Inc., Model D 1000 CE) connected to a computer, in 1 cm path-length cells.

The equation used to calculate the phenol removal efficiency in the experiments was:

0

0

/ % = 100R

(1)

where γ0 and γ are defined as the concentration before and after electrocoagulation process. A

similar equation was used to calculate total organic carbon (TOC) removal efficiency (RTOC / %).

The TOC measurements were done on a Shimadzu Analyser TOC V-CSN using the NPOC

method.

The electrocoagulation experiments were carried out in a glassy electrolytic cell with

dimensions of 7 x 7 x 6 cm. Parallel plate electrodes were immersed in the cell with a working

electrode situated between two counter electrodes in order to achieve a good current and

potential distribution, and a uniform electrode dissolution. Low carbon steel (composition:

0.06 % C, 0.015 % P, 0.008 % S, 0.007 % Si and 0.35 % Mn) was used for all electrodes. Before the

experiment the electrodes were polished using 600 grit emery paper, they were washed with bi-

distilled water, and finally, with ethanol. The working electrode had a total immersed area of

10 cm2, and the counter electrodes had a total immersed area of 20 cm2. The distance between

the electrodes was 1 cm. Before the experiment, the appropriate amount of NaCl was dissolved in

the treated solutions. During the experiment constant stirring speeds of 600 rpm and DC Power

Supply (Iskra, MA 4165; 1.5 A; 25 V) were used. All the experiments were carried out at room temperature (23±1 oC).

RESULTS AND DISCUSSIONS

Treatment of the model water containing mimosa tannin

The mimosa tannin (γ = 1 g L-1) solution prepared in 0.5 mol dm-3 NaCl has pH 5.1 and

κ = 38.3 mS cm-1. In order to find out the removal efficiency of mimosa tannin over different

durations, the process was carried out for 5, 15, and 35 minutes using current densities of 10 mA

cm-2. The characteristic voltage of this process was 1.25 V. The influence of the duration of the

experiment on the electrocoagulation efficiency is presented in Table 1.

Within the 5 min removal period removal efficiency reached 92.2 %, while further treatment

resulted only in its slight increase (35 min, 96.7 %). Energy consumption during the experiment

increased proportionally with time since the current density and voltage were constant

throughout the experiments.

To explain the mechanism of the electrocoagulation process one must consider the reactions

occurring on both electrodes. During the anodic process iron is oxidized and dissolved as Fe2+.

Under the experimental conditions used in this work (pH = 5.1) it does not undergo hydrolysis.

However, in aerated conditions it is further oxidized by dissolved oxygen to Fe3+, which is

susceptible to hydrolysis, resulting in different aqua and hydroxyl complexes, such as Fe(OH)2+,

Fe(OH)+, Fe(OH)3 under acid conditions, or Fe(OH)63- and Fe(OH)4

- under alkaline conditions [23].

Since the pH values registered before and after electrocoagulation process varied from 5-7,


218

positively charged or neutral particles were expected under the given experimental conditions.

Furthermore, negative ions present in water (in the case of NaCl addition it is Cl-) will surround a

positive charge, forming a diffuse layer which makes the particle neutrally charged. These surface

properties make the particles unstable and agglomeration takes place. As a result of agglome-

ration, particles form flocks precipitating or floating by the bubbles of hydrogen (Figure 1.). The

stability of the particles and their agglomeration depends on the type and concentration of ions in

the solution. The formed flocks can effectively remove pollutants by adsorption or enmeshment in

a precipitate.

The aim of this work was to remove mimosa tannin from model water. It is well known from

the literature that dissolved iron in the presence of tannin forms ferrous/ferric tannates. These

reactions are pretty complex since both ions, ferrous and ferric, can participate in the reaction,

and in addition there is also a possibility of Fe3+ reduction by mimosa tannins. According to the

author’s knowledge, the formation of the complex during phenolic compounds removal by

electrocoagulation technique was not considered before, although it might be of a great

importance for its progress. It can influence the amount of ferrous/ferric ions required for the

coagulation, since Fe2+/Fe3+are consumed by the complex formation. Additionally, it can also

impact the mechanism of the reaction because generated complex will be involved in the

adsorption or enmeshment by formed flocks instead of mimosa tannin. It could also be important

for the electrode reaction kinetics since tannin inhibits dissolution of iron [24,25]. Therefore it can

be concluded that complex formation should not be ignored when considering the

electrocoagulation process.

Figure 1. Illustration of flocks precipitating due to gravity, and floating due to the bubbles of gas

An important parameter of the electrocoagulation process is the effluent pH value at the end of

the process. In these experiments pH values decreased for processes conducted over 5 min, while

they increased for prolonged treatments (15 and 35 min) (Table 1). The increase of pH values were

caused due to the intensive hydrogen evolution at the cathode and the generation of OH- ions.

The generated OH- ions were consumed during the hydrolysis of Fe3+, but the overall reaction

obviously resulted in the excess of OH- ions. The final pH value depended on the equilibrium of

each reaction in the process. Additionally, the increase of pH value can be explained by mimosa

tannin removal due to its acidic behaviour.


doi: 10.5599/jese.2014.0066 219

Table 1. Results of treatments at different times (j = 10 mA cm-2, U = 1.25 V, γ (NaCl) = 29.22 mol g L-1, κ = 38.3 mS cm-1, pH = 5.1, γ0(tannin)=1 g L-1, V = 0.1 L).

T / min R / % SEC/ kW h kg-1 m(precipitate) / g pH(after EC)

5 92.21 0.113 0.143 4.81

15 95.54 0.327 0.197 5.74

35 96.68 0.754 0.263 6.31

Since the concentration of ferrous/ferric ions is dependent on current density, the efficiency of

the process is dependent on current density as well. In this work the electrocoagulation process

was conducted for 15 min using different current values (Figures 2 and 3). When the current

increased from 1 to 10 mA cm-2, removal efficiency increased from 10 to 20 % in 0.58 g L-1 NaCl,

from 47 to 93 % in 5.84 g L-1 NaCl, and from 58 to 96 % in 29.22 mol dm-3 NaCl (Figure 2). However,

specific energy consumption per mass of mimosa tannin increased more significantly from 0.302

to 7.311 kW h kg-1 in 0.58 g L-1 NaCl, from 0.034 to 0.469 kW h kg-1 in 5.84 g L-1 NaCl, and from

0.021 to 0.327 kW h kg-1 in 29.22 g L-1 NaCl (Table 2). As evident, the highest removal efficiency

was obtained at high current density. The obtained results also show that the highest removal

efficiency was registered in the presence of high NaCl concentrations, where the lowest specific

energy consumption is required. It is supported by Figure 4, where the colour of the solution

suggests more difficult precipitation of flocks for small additions of NaCl causing reduced removal

efficiency. This is in accordance with theory that coagulation process depends on type and

concentration of ions present in solution. It is obvious that optimal process conditions were

obtained in the presence of high amounts of NaCl. These results pointed out the importance of the

NaCl concentration as a key parameter for an efficient and low cost process.

NaCl’s presence is important because of the two effects: (a) it decreased the applied voltage

and energy power demand [26] and (b) it changed the ionic strength that affected the coagulation

process as evident from Figures 2 and 4. The influence of ionic concentration and zeta potential on

the electrocoagulation process was reported previously [27, 28].

Figure 2. Influence of NaCl concentration and current density on removal efficiency.


220

The main factor influencing energy consumption is applied voltage. The overall voltage is

dependent on equilibrium potential difference (Er,k-Er,a), anode and cathode over-potentials

(ηa, ηk), and ohmic potential drop in the solution (ηIR ) according to the equation (2):

er r,k r,a a k IRU E E (2)

Ohmic potential drop in the solution is dependent on cell configuration, electrode area (A / m2),

and the distance between electrodes (d / m), as well on the conductivity of solution (κ / S cm-1):

IR

dI

A

(3)

Table 2. Results of treatments at different NaCl concentration and current density (t = 15 min, pH = 5.1, γ(tannin) = 1 g L-1, V = 0.1 L).

γ (NaCl) / g L-1 J / mA cm-2 U / V SEC / kW h kg-1 m(precipitate) / g

29.22 (κ = 38.3 mS cm-1)

10 1.25 0.327 0.197

5 0.95 0.146 0.134

1 0.50 0.021 0.103

5.84 (κ = 10.11 mS cm-1)

10 1.75 0.469 0.140

5 1.15 0.188 0.128

1 0.65 0.034 0.094

0.58 (κ = 0.98 mS cm-1)

10 5.75 7.311 0.080

5 3.65 3.061 0.044

1 1.25 0.302 0.038

Figure 3. Results of treatments at different current densities in the case of mimosa tannin:

(a) 10 mA cm-2; (b) 5 mA cm-2; (c) 1 mA cm-2 (t = 15 min; 29.22 g L-1NaCl).

Figure 4. Results of treatments at different NaCl concentrations in the case of mimosa tannin: (a) 29.22 g L-1 NaCl; (b) 5.84 g L-1NaCl; (c) 0.58 g L-1NaCl (j = 10 mA cm-2; t = 15).

a

a c b

c b


doi: 10.5599/jese.2014.0066 221

SEC is dependent on current value (I), applied voltage (U) and time (t) and it is expressed as

energy consumption per mass of removed tannin:

0

SEC IUt

tannin tannin V

(4)

where γ is the mass concentration of tannin or phenolic compound, V is volume of treated

solution.

Another important parameter for SEC is the distance between the electrodes, which in most of

the previous reports ranged from 0.3-3.0 cm, while the distance in this work was 1 cm. A small

distance is preferable to decrease potential drop, but the electrodes should be adequately

separated in order to enable unhindered movement of flocks between them.

Conductivity of the solution i.e. salt concentration also plays important role for SEC value and

according to our knowledge its value was quite different in different reports.

Kobaya et al. [29] treated textile wastewater by electrocoagulation process and it was shown

that an addition of NaCl (κ = 1000-4000 S cm-1) did not influence process efficiency but energy

consumption decreased with increased wastewater conductivity (2.2-0.75 kW h kg-1 (COD)).

Sengil et al. [30] have used electrocoagulation for decolourization of Reactive red and it was found

that small additions (0.5-2.0 g L-1) of NaCl increase efficiency while further addition did not have

any impact. The optimal conditions were found to be 2.3 g L-1 NaCl and 4.54 kW h kg-1 (dye). B. K.

Nandi et al. [31] varied NaCl concentration from 0.1-1.0 g L-1 and it was found that efficiency had

increased from 97-100 % and energy consumption had decreased from 17-3 kW h kg-1 (Fe). X.

Chen et al. [32] separated pollutants from restaurant wastewater by electrocoagulation process

without the addition of NaCl when energy consumption was in the range of 0.2-1.4 kW h m-3 dep-

ending on the solution conductivity that varied from 770-227 S cm-1. The additions of NaCl chang-

ed conductivity from 443-2850 S cm-1 and energy consumption as well, from 0.32-0.29 kW h m-3;

however, it did not change the efficiency of the process.

From the previous results it follows that NaCl concentration can influence specific energy

consumption drastically, which will have an impact on operating cost. However, the dependence

of removal efficiency on NaCl concentration is not completely clear and it depends on the type of

pollutant and its concentration.

The results of this paper confirm that NaCl addition decreases specific energy consumption in

accordance with the previous results. Furthermore, it was shown that efficiency of the phenolic

compound removal can also be improved considerably by the addition of NaCl in the range from

0.58 g L-1 to 29.22 g L-1.

Treatment of OMW

The starting OMW solution had the following characteristics: pH 5.37, concentration of phenolic

compounds, γ0 = 0.613 g L-1 (mimosa tannin equivalent), and the TOC value was 1376 mg L-1. These

values were somewhat lower in comparison to the values frequently found in the literature

[10-12]. This can be explained by the fact that the OMW was kept in an open puddle for 2 months.

During OMW treatment by electrocoagulation process in previous investigations the solution as

received was used or it was diluted with water. The conductivity of the pure OMW sample was

11 mS cm-1 [10-12] and for diluted OMW (1:5) conductivity was 3.6 S cm-1 [13]. The SEC value

obtained during the OMW treatment was found to be 4 kW h kg-1 (COD) [13] or 20-300 kW h m-3

(volume of treated solution) [12], which were quite similar considering the characteristic COD


222

value for OMW. Also, it was shown that small additions of NaCl improve removal efficiency while

additions higher than 2 g L-1 decrease removal efficiency. Energy consumption has decreased upon

NaCl addition.

From the results of the treatment of mimosa tannin, the current density of 10 mA cm-2 was

chosen for the OMW treatment. The process was carried out with different additions of NaCl from

0-20 g L-1 during 35 or 60 min. Depending on NaCl addition conductivity of the solutions was

changed from the value similar to the previously reported values (2.3 S cm-1) to the value higher

than previously reported (23.7 S cm-1). The SEC value was changed from 8.5-1.6 kW h kg-1 (mass of

phenolic compounds) during 35 min or it was changed from 8.2-2.6 kW h kg-1 (mass of phenolic

compounds) during 60 min.

Similarly as in the case of model water, addition of NaCl had positive impact on removal

efficiency (Figures 5 and 6) and energy consumption (Table 3). Furthermore, better efficiency was

obtained by prolonged process time while SEC was not increased significantly. At the end of the

electrocoagulation process pH value was close to 7, which was acceptable for discharge, while

conductivity increased only slightly.

Table 3. Results of treatments at different NaCl concentration and process times (j= 10 mA cm-2. pH = 5.37. V= 0.1 L, γ0 = 0.613 g L-1).

t/ min γ(NaCl) / g L-1

U / V SEC/ kW h kg-1

pHafter EC κ(OMW)before EC / mS cm-1

κ(OMW)after EC / mS cm-1

0 3.8 8.492 6.65 2.32 3.43

35 5 1.9 2.677 6.85 8.23 9.48

10 1.5 1.998 6.88 13.77 14.75

20 1.4 1.604 6.9 23.72 24.79

0 3.9 8.226 6.7 2.32 3.26

60 5 2.1 4.26 6.81 5.99 6.55

10 1.6 2.941 6.94 10.12 10.98

20 1.5 2.562 6.98 20.44 21.75

Figure 5. Dependence of removal efficiency of phenolic compound in OMW treated effluent on NaCl concentration and process time.


doi: 10.5599/jese.2014.0066 223

Figure 6. Dependence of removal efficiency of TOC in OMW treated effluent on NaCl

concentration and process time.

From Figures 5 and 6 it is evident that the removal of overall organic loading (TOC) is lower

(30 - 70 %) in comparison to phenolic compounds (40 - 90 %). Better efficiency in the case of

phenolic compounds could be the consequence of complex formation between ferrous/ferric and

phenolic compounds present in the OMW. It is also evident that removal efficiency of phenolic

compounds in OMW was lower compared to the removal efficiency of mimosa tannin, although

longer times were used (Figures 2 and 5). It is not surprising since this solution, apart from the

phenolic compounds, contains some other constituents such as oil, sugar, and pulp suspension [2].

Therefore, the capacity of produced sludge for phenolic compounds removal is reduced in the case

of OMW compared to the model water.

Conclusions

The results obtained in this paper show that it is possible to obtain high removal efficiency of

mimosa tannin and phenolic compounds from OMW by electrocoagulation process. In the case of

mimosa tannin, electrocoagulation was able to reduce the phenolic content up to 96 %, while in

the case of OMW electrocoagulation wasa able to reduce the phenolic content up to 92 %. The

percentage of removal was increased with increased time, current density, and NaCl

concentration. Apart from the increasing removal efficiency of the process, an improvement in

energy demand was also obtained with the addition of NaCl. Therefore, it can be concluded that

an addition of NaCl can significantly improve the electrocoagulation process. Furthermore,

additions of NaCl and the disposal of treated effluent containing NaCl are acceptable for the

production plants located close to the coast. At the end of process pH value was close to 7, which

is acceptable for discharging.

It was shown that complex formation between ferrous/ferric and phenolic compounds present

in treated effluent could change the efficiency of the process. Thus, due to the complexation,

removal of phenolic compounds was higher in comparison to removal of overall organic loading

(TOC). It was also observed that the removal efficiency of mimosa tannin is higher compared to

the removal efficiency of phenolic compounds from OMW although longer times were used. It is

explained by the decreased capacity of produced flocks for phenolic compounds removal, in the

case of OMW, due to the presence of other organic constituents.


224

Acknowledgements: Financial support by Ministry of Science, Education and Sports of Republic of Croatia (project 125-1252973-2576) is gratefully acknowledged. The authors express their - gratitude to Višnja Pavić for providing mimosa tannin.

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Optimization of parameters for dye removal by electro-oxidation using Taguchi Design

Mani Nandhini, Balasubramanian Suchithra,

Ramanujam Saravanathamizhan and Dhakshinamoorthy Gnana Prakash

Department of Chemical Engineering, SSN College of Engineering, Kalavakkam, Chennai 603110, India

Corresponding author: E-mail: [email protected]; Tel. +91 44 27469700; Fax +91 44 27469772

Received: April 18, 2014; Revised: May 9, 2014; Published: December 6, 2014

Abstract The aim of the present investigation is to treat the dye house effluent using electro-oxidation and to analyse the result using Taguchi method. L16 orthogonal array was applied as an experimental design to analyse the results and to determine optimum conditions for acid fast red dye removal from aqueous solution. Various operating parameters were selected to study the electro-oxidation for the colour removal of the effluent. The operating parameter such as dye concentration, reaction time, solution pH and current density were studied and the significance of the variables was analysed using Taguchi method. Taguchi method is suitable for the experimental design and for the optimization of process variables for the dye removal.

Keywords Electro-oxidation; Taguchi design; colour removal; acid fast red; optimization

Introduction

Effluents discharged from textile industries have high intensity of colour, which leads to

pollution. Highly coloured wastewater can be treated by different methods such as biological

treatment, chemical coagulation, activated carbon adsorption, ultrafiltration, ozonation, wet

oxidation, photocatalysis, electrochemical methods etc. [1-5]. Among these methods the electro-

chemical treatment has been receiving greater attention in recent years due to its unique features

such as versatility, energy efficiency, automation and cost effectiveness [6]. The electrochemical

technique offers high removal efficiencies and the main reagent is the electron called ‘clean

reagent’ which degrades all the organics present in the effluent without generating any secondary

pollutant or by-product/sludge.



J. Electrochem. Sci. Eng. 4(4) (2014) 227-234 DYE REMOVAL USING TAGUCHI DESIGN

228

Industrial wastewaters have been treated by electro-oxidation techniques and the operating

parameters have been optimized by different techniques. The widely used technique is response

surface methodology for the experimental design and process optimization. Taguchi method is

another method in the experimental design, proposed by Genichi Taguchi, contains system design,

parameter design, and tolerance design. Taguchi method was effectively used to improve the

product or process effectiveness by using a loss function to attain the product quality in terms of

the parameter design [7]. Taguchi is the preferable technique among statistical experimental

design methods since it uses a special design of an orthogonal array to study the effective para-

meters with a minimum number of experiments. This method helps researchers to determine the

possible combinations of factors and identify the best combination. However, in industrial set-

tings, it is extremely costly to run a number of experiments to test all combinations. The parame-

ter design using Taguchi method minimizes the time and experimental runs. In the design, ‘signal’

and ‘noise’ (S/N) represent the desirable and undesirable values for the output characteristics,

respectively and the ratio is a measure of the quality characteristic deviating from the desired

value. Further an analysis of variance (ANOVA) is used to determine the significant parameters.

Recently Taguchi’s designs have been applied to various chemical and environmental

engineering studies for the experimental design and for the optimization of the process variables.

Asghari et al. [8] used Taguchi method to determine the optimum conditions for methylene blue

dye removal from aqueous solutions using electrocoagulation. The authors found that the amount

of electrolyte was the most significant parameter for the colour removal. Srivastava et al. [9] used

Taguchi method to determine the optimum conditions for the orange-G dye removal from

aqueous solution by electrocoagulation using iron plate electrodes. Author also applied this

methodology to optimize the process variables for the multi-component adsorption of metal ions

onto bagasse fly ash and rice husk ash [10]. Kaminari et al. [11] used Taguchi method to determine

the optimum condition of process variables and find the influencing parameters for the recovery

of heavy metals from acidified aqueous solutions using electrochemical reactor. Kim et al. [12]

optimized the experimental process variable using Taguchi technique for the nano-sized silver

particles production by chemical reduction method. In another study Mohammadi et al. [13] used

Taguchi method to determine the optimal experimental conditions for the separation of copper

ions from a solution. Moghaddam et al. [14] used Taguchi method to design the experimental runs

and optimize the parameters for the ammonium carbonate leaching of nonsulphide zinc ores.

Maria et al. [15] designed the experimental runs and optimized the process parameters using

Taguchi method for the adsorption of acid orange 7 dyes on guava seed. In that way Taguchi

method was used to study the electro oxidation of dye removal.

The objective of the present study is to treat the acid fast red dye using electro-oxidation. The

effect of experimental parameters such as initial dye concentration, reaction time, solution pH and

current density on colour removal was investigated using an L16 orthogonal array. The Taguchi

experimental design has been used to determine the optimum conditions for the maximum colour

removal of the dye from aqueous solutions.

Materials and Methods

All the chemicals used in this study were AR grade (Merck). Acid fast red was prepared by

dissolving definite quantity of dye in the distilled water. To increase the conductivity of the

solution of 1 mg L-1 NaCl was used as supporting electrolyte for all experimental runs. The initial

M. Nandhini et al. J. Electrochem. Sci. Eng. 4(4) (2014) 227-234

doi: 10.5599/jese.2014.0056 229

solution pH was adjusted by 0.1 M HCl or 0.1 M sodium hydroxide solution. Experiments were

repeated twice to minimize the experimental error.

Experimental setup

The experimental setup of the batch reactor is shown in the Figure 1. The volume of the reactor

was 250 ml and electrodes used were fixed inside the reactor with 1 cm space between them.

Stainless steel sheet cathode and mesh type Ruthenium oxide coated Titanium anode were used.

The void fraction of the mesh type anode accounts 20 % by area, which resulted in an effective

anode area of 28 cm2 (7×5 cm). The electrodes were connected to a 5 A, 10 V DC regulated power

supply, through an ammeter and a voltmeter. The solution was constantly stirred at 200 rpm using

a magnetic stirrer in order to maintain a uniform concentration. DC power supply was given to the

electrodes according to the required current density and the experiments were carried out under

constant current conditions. The samples were analysed for the colour removal using UV-Vis

spectrophotometer (Jasco, V-570). The percentage colour removal was calculated by:

Colour removal, % = i t

i

100Abs Abs

Abs

(1)

where Absi and Abst are absorbance of initial and at time t at the corresponding wavelength max.

Figure 1. Schematic diagram of the experimental setup

1. Magnetic stirrer 2. Anode 3. Cathode 4. DC Power supply

Taguchi design

The following procedure was adopted for the parameter design.

1. Planning of experiment

i. Determine the experimental responses of the process.

ii. Determine the levels of each variable.

iii. Select a suitable orthogonal array table. The selection based on the number of variables

and number of levels.

iv. Transform the data from the experiments into a proper S/N ratio.

2. Implementing the experiment, based on design table.


230

3. Analyzing and examining the result

i. ANOVA analysis to determine the significant parameters in the process.

ii. Draw the main effect plot, S/N ratio plot, mean plot to analysis the optimal level of the

control variables.

The factors and levels chosen for the present experiment are shown in the Table 1. L16

orthogonal array design was selected for the four variables with four different levels for the each

factor. Table 2 shows the Taguchi design for the electro oxidation of acid fast red dye. Each row

represents one experimental run. Based on Taguchi design the experiments were carried out and

the percentage colour removal was observed as response. The proposed design was an orthogonal

array, for which each pair of the columns had all the possible combinations of levels. The S/N ratio

characteristics can be divided into three categories when the characteristic is continuous:

(i) Nominal is the best characteristic

2y

10 logS y

N s (2)

Table 1. Variables and their values corresponding to their levels investigated in the experiments

Variables Level

1 2 3 4

A Dye concentration, mg l-1 25 50 75 100

B Time, min 15 30 45 60

C pH 2 5 7 10

D Current density, mA cm-2 5 7.5 10 12.5

Table 2. Experimental variables, their levels and results of conducted experiments corresponding to L16 experimental plan

S. No Levels Run 1 Run2

A B C D Colour removal, %

1 1 1 1 1 78.50 78.40

2 1 2 2 2 81.82 81.20

3 1 3 3 3 86.42 86.40

4 1 4 4 4 89.75 90.00

5 2 1 2 3 70.08 69.98

6 2 2 1 4 84.41 84.15

7 2 3 4 1 77.20 78.01

8 2 4 3 2 91.53 91.50

9 3 1 3 4 71.55 71.54

10 3 2 4 3 68.21 68.12

11 3 3 1 2 86.78 86.78

12 3 4 2 1 83.44 83.58

13 4 1 4 2 52.05 52.19

14 4 2 3 1 56.70 57.25

15 4 3 2 4 91.66 91.89

16 4 4 1 3 96.32 96.80


doi: 10.5599/jese.2014.0056 231

(ii) Smaller the better characteristics

2110 log

Sy

N n (3)

(iii) Larger the better characteristics

2

1 110 log

S

N n y (4)

Where, y is the average of observed data, 2ys the variance of y, n is the number of

observations, and y the observed data. For each type of the characteristics, with the above S/N

ratio transformation, the higher the S/N ratio the better is the result. The experimental data were

analysed using MINITAB 14 (PA, USA) [16].

Result and discussion

Main effect plot

Main effect plot for the percentage colour removal using electro oxidation of acid fast red dye

is shown in the Figure 2. The plot is used to visualize the relationship between the variables and

output response. The effect of initial dye concentration on colour removal is shown by the factor

‘A’. The percentage colour removal of the dye increases with decrease in dye concentration. This

is due to the fact that at higher initial dye concentrations, the intermediate products formed due

to the degradation of dyes increase the resistance of current flow by blocking the electrode active

sites, and thus, decrease colour removal. The effect of electrolysis time on colour removal is

shown in the figure by a factor ‘B’. The percentage colour removal depends on the electrolysis

time. When time increases the generation of OCl- radical increases due to electro-oxidation which

results increase the percentage colour removal. The effect of pH on the mean colour removal is

represented by the factor ‘C’. As it is observed from the figure, colour removal increases with the

decrease of pH. This is due the fact that the hydroxyl radical generation is high at acidic pH which

results to increase the rate of colour removal. The effect of current density on the removal of dye

is shown by the factor ’D’. The increase of current density increases the OCl- generation hence the

percentage colour removal increases. High concentrations of chloride ions and salts in water can

improve the performance and effectiveness of the electro-oxidation process. Various levels (1, 2,

3, 4) of the operating parameter (A, B, C, D) and their mean colour removal is shown in the Table

3. It is observed form the table, Level ‘1’ shows highest colour removal of 84.06 %, 84.37 % for the

initial dye concentration and pH, respectively. Level ‘4’ shows highest colour removal of 90.37 %,

86.52 % for electrolysis time and current, respectively.

Table 3. Mean colour removal for electro oxidation of acid fast red

Level Mean colour removal, %

A B C D

1 84.06 68.04 86.52 74.14

2 80.86 72.74 81.71 77.98

3 77.50 85.64 76.62 80.29

4 74.36 90.37 71.94 84.37


232

Figure 2. Main effect plot for the percentage colour removal of acid fast red

A: Dye concentration parameter; B: Time parameter; C: pH parameter; D: Current density parameter

Signal to noise (S/N) ratio

Taguchi method was used to identify the optimal conditions and most influencing parameters

on colour removal. In the Taguchi method, the terms ‘signal’ to ‘noise’ ratio represent the

desirable and undesirable values for the output response, respectively. The S/N ratios are different

according to the type of output response. In the present case, larger S/N ratio is better for high

colour removal. Figure 3 shows the S/N ratio of dye removal using electro-oxidation.

Figure 3. S/N ratio plot for the percentage colour removal of acid fast red .

A: Dye concentration parameter; B: Time parameter; C: pH parameter; D: Current density parameter

Mea

n c

olo

ur

rem

iva

l, %


doi: 10.5599/jese.2014.0056 233

It can be noticed that at higher S/N ratio better level for colour removal was achieved. It is

observed that factor ‘A’, initial dye concentration, and factor ‘C’, pH, are required in a lower level

and electrolysis time ‘B’ and applied current density ‘D’ are required at higher level for the

maximum colour removal.

Optimization of process parameters based on S/N ratio

The process parameters were optimized based on S/N ratio. Lager the S/N ratio higher the

percentage colour removal and vice versa. The values of the S/N ratios for the operating

parameters are shown in the Table 4. It is observed from the Table that the S/N ratio of level ‘1’

38.48, 38.72 shows the higher value for the factor ‘A’ and ‘C’, respectively. It shows that level ‘1’

gives the higher colour removal. The larger S/N ratios of factor ‘B’ and ‘D’ for the level ‘4’ are

39.11, 38.48, respectively. It shows that level ‘4’ gives the higher colour removal.

Table 4. S/N ratio for electro oxidation of acid fast red

Level S/N ratio of electro oxidation

A B C D

1 38.48 36.56 38.72 37.31

2 38.11 37.13 38.20 37.64

3 37.74 38.64 37.54 38.00

4 37.10 39.11 36.97 38.48

Analysis of variance (ANOVA)

Analysis of variance (ANOVA) was performed to see whether the process parameters were

statistically significant or not. The F-test is a tool to check which process parameters have a

significant effect on the colour removal. The P value less than 0.05 shows that the parameter is

significant. ANOVA table for the colour removal of the dye is shown in the Table 5. It is observed

form the table the most influential factor was electrolysis time because the P value is 0.049 with

the corresponding sum of the square is higher compared to other variable in the Table. Then the

less significant variables are pH, current density and initial dye concentration.

Table 5. ANOVA table for the electro oxidation of acid fast red

Source Degree of freedom Sum of the squares Mean of squares F P

A 3 210.9 70.29 1.50 0.374

B 3 1330.5 443.48 9.46 0.049

C 3 476.8 158.94 3.39 0.171

D 3 220.2 73.40 1.56 0.361

Residual Error 3 140.7 46.90

Total 15 2379

Conclusion

Taguchi experimental design was used to determine the optimum operating conditions of the

dye removal from aqueous solutions using electro oxidation. The significant variables were

identified for the colour removal process. Optimum levels for operating parameters can be simul-

taneously identified with the Taguchi method. The advantage of the Taguchi method is the

reduction in time and minimization of the number of experimental runs. For the present study


234

level ‘1’ is best for initial dye concentration (25 g L-1) and initial pH (2), level ’4’ is best for

electrolysis time (60 min) and current density (12.5 mA cm-2) for the highest colour removal. It can

be concluded that Taguchi method is suitable for the experimental design and to optimize the

process variable for the colour removal of the dye effluent.

Acknowledgement: The authors are grateful to the SSN Trust for the financial support of this work.

References

[1] M. C. Gutierrez, M. Crespi, Journal of the Society of Dyers and Colorists 115 (1999) 342–345. [2] M. Panizza, G. Cerisola, Chemical Reviews 109 (2009) 6541–6569. [3] C. A. Martinez-Huitle, E. Brillas, Applied Catalysis B: Environmental 87 (2009) 105-145. [4] L. C. Davies, I. S. Pedro, J. M. Novais, S. Martins-Dias, Water Research 40 (2006) 2055 - 2063. [5] G. Chen, Separation & Purification Technology 38 (2004) 11–41. [6] F. C. Walsh, Pure and Applied Chemistry 73 (2001) 1819–1837. [7] J. A. Ghani, I. A. Choudhury, H. H. Hassan, Journal of Materials Processing Technology 145

(2004) 84–92. [8] A. Asghari, M. Kamalabadi, H. Farzinia, Chemical and Biochemical Engineering Quarterly 26

(2012) 145–154. [9] V. C. Srivastava, I. D. Mall, I. M. Mishra, Chemical Engineering Journal 140 (2008) 136–144

[10] V. C. Srivastava, I. D. Mall, I. M. Mishra, Industrial & Engineering Chemistry Research 46 (2007) 5697-5706.

[11] N. M. S. Kaminari, D. R. Schultz, M. J. J. S. Ponte, H. A. Ponte, C. E. B. Marino, A. C. Neto, Chemical Engineering Journal 126 (2007)139-146.

[12] K. D. Kim, D. N. Han, H. T. Kim, Chemical Engineering Journal 104 (2004) 55-61. [13] T. Mohammadi, A. Moheb, M. Sadrzadeh, A. Razmi, Desalination 169 (2004) 21-31. [14] J. Moghaddam, R. Sarraf-Mamoory, M. Abdollahy, Y. Yamini, Separation & Purification

Technology 51 (2006)157-164. [15] M. P. Elizalde-Gonzalez, V. Hernandez-Montoya, Journal of Hazardous Materials 168 (2009)

515–522. [16] K. Ravikumar, S. Ramalingam, S. Krishnan, K. Balu, Dyes and Pigments 70 (2006) 18-26.




doi: 10.5599/jese.2014.0058 235

J. Electrochem. Sci. Eng. 4(4) (2014) 235-245 ; doi: 10.5599/jese.2014.0058


www.jESE-online.org


Electrochemical treatment of Acid Red 1 by electro-Fenton and photoelectro-Fenton processes

Camilo González-Vargas, Ricardo Salazar and Ignasi Sirés*,

Laboratorio de Electroquímica Medio Ambiental, LEQMA, Departamento de Ciencias del Ambiente, Facultad de Química y Biología, Universidad de Santiago de Chile, USACh, Casilla 40, Correo 33, Santiago, Chile *Laboratori d’Electroquímica dels Materials i del Medi Ambient, Departament de Química Física, Facultat de Química, Universitat de Barcelona, Martí i Franquès 1-11, 08028 Barcelona, Spain

Corresponding Author: E-mail: [email protected]; Tel.: +56-2-27181134 Corresponding Author: E-mail: [email protected]; Tel.: +34-93-4039243; Fax: +34-93-4021231

Received: July 23, 2014; Published: December 6, 2014

Abstract Small volumes (100 mL) of acidic aqueous solutions with 30-200 mg L-1 TOC of the toxic azo dye Acid Red 1 (AR1) have been comparatively treated by various electrochemical advanced oxidation processes (EAOPs). The electrolytic system consisted of a BDD anode

able to produce OH and an air-diffusion cathode that generated H2O2, which

subsequently reacted with added Fe2+ to yield additional OH from Fenton’s reaction. Under optimized conditions (i.e., 1.0 mM Fe2+, 60 mA cm-2, pH 3.0, 35 ºC), the analysis of the initial rates for decolourization and AR1 decay assuming a pseudo-first-order kinetics

revealed a much higher rate constant for photoelectro-Fenton (PEF, ~ 2.7x10-3 s-1)

compared to electro-Fenton (EF, ~ 0.6x10-3 s-1). Mineralization after 180 min was also greater in the former treatment (90 % vs 63 %). The use of UV radiation in PEF contributed to Fe(III) photoreduction as well as to photodecarboxylation of refractory intermediates, yielding a mineralization current efficiency as high as 85% during the treatment of solutions of 200 mg L-1 TOC. Primary reaction intermediates included three aromatic derivatives with the initial naphthalenic structure and four molecules only featuring benzenic rings, which were totally mineralized in PEF.

Keywords Air-diffusion cathode; Azophloxine; boron-doped diamond (BDD); E128; EAOPs; decolourization; food azo dye; mineralization; Red 2G




J. Electrochem. Sci. Eng. 4(4) (2014) 235-245 ACID RED 1 BY ELECTRO-FENTON AND PHOTOELECTRO-FENTON

236

Introduction

In recent years, great attention has been paid to the electrochemical advanced oxidation

processes (EAOPs) based on Fenton’s reaction chemistry for water decontamination [1]. Among

them, electro-Fenton (EF) and photoelectro-Fenton (PEF) have been the most studied methods for

the treatment of waters polluted by organic pollutants such as pharmaceuticals [2-4], pesticides

[5-7] and synthetic dyes [8-11] due to their outstanding performance. In EF and PEF, H2O2 is

continuously electrogenerated in the reaction cell from the two-electron reduction of injected or

flushed gaseous O2 as follows [1]:

O2 + 2 H+ + 2 e → H2O2 (1)

The use of carbonaceous cathodes, particularly in the form of gas-diffusion electrodes (GDEs),

promotes the production of a high yield of H2O2 at high rate. Therefore, the current efficiency for

reaction (1) on GDEs turns out to be the highest among all tested materials. The activation of H2O2,

which is a weak oxidant, is achieved in the presence of a small amount of metal catalyst, especially

Fe2+, at pH around 3. Under such conditions, OH is produced in the bulk from Fenton’s reac-

tion (2) [1]:

Fe2+ + H2O2 → Fe3+ + OH + OH (2)

In PEF, the contaminated acidic solution is irradiated with artificial UV light, thus causing: (i) the

photoreduction of Fe(OH)2+, which is the main Fe3+ species at pH 3, via photo-Fenton’s reaction (3),

and (ii) the photodecarboxylation of some Fe(III)-carboxylate complexes, which are quite refractory

to oxidation by OH formed from Fenton’s reaction, via reaction (4) [1]. While reaction (3)

contributes to form additional OH as well as to maintain the Fenton’s cycle by continuously

regenerating Fe2+, reaction (4) makes it possible to reach great mineralization degrees.

Fe(OH)2+ + hν → Fe2+ + OH (3)

[Fe(OOCR)2+] + hν → Fe2+ + CO2 + R (4)

For some pollutants, the performance of EF and PEF can be further enhanced by using a large

O2-overpotential anode such as boron-doped diamond (BDD). This material favours the electro-

oxidation (EO) of the organic molecules by action of OH adsorbed on the anode surface via

reaction (5) [12]:

BDD + H2O → BDD(OH) + H+ + e (5)

Acid Red 1 (AR1, also called Amido Naphtol Red G, Red 2G, Food Red 10 or Azophloxine, see

chemical structure in Table 1) is a synthetic monoazo dye. Azo dyes constitute the most

comprehensive and varied family among all synthetic organic dyes available in the industry for

dyeing all kinds of fabrics [13]. Until 2007, AR1 was the preferred dye for baby food colouring due

to the resulting intense red colour, but the Scientific Panel of the European Food Safety Authority

(EFSA) on food additives, flavourings, processing aids and materials in contact with food was then

asked to re-evaluate this dye due to food safety concerns related to its high toxicity and

carcinogenic effects [14]. Based on this report, the European Union has agreed with its suspension

as food colouring, as published in the Official Journal [15]. Not surprisingly, it is also banned in

other countries outside Europe.

At present, it is known that AR1 is extensively metabolized to aniline, it may interfere with

blood haemoglobin and it is suspected of being a carcinogen [16]. It is therefore one of the dyes

C. González-Vargas et al. J. Electrochem. Sci. Eng. 4(4) (2014) 235-245

doi: 10.5599/jese.2014.0058 237

that the hyperactive children’s support group recommends to be eliminated from the diet of

children. In contrast, AR1 has become one the most used dyes in Chile and over the world for

dyeing polyester-, nylon-, cellulose- and acrylic-based fibers. Consequently, if such industrial

wastewaters are not conveniently treated before discharge, their impact on the surrounding

ecosystems may be dramatic.

Some few studies have focused on the fate of AR1 upon application of non-electrochemical

advanced oxidation processes (AOPs) [17-19]. Regarding the electrochemical technology, only its

electrochemical reduction on an activated carbon fiber cathode has been reported [20], but as far

as we are concerned its removal by EF and PEF with BDD or other anodes has not been reported

yet. These latter processes are of great interest for the removal of food azo dyes, as some of us

recently demonstrated for the treatment of Tartrazine solutions [21].

In this work, the decolourization, AR1 decay and mineralization profiles resulting from the

treatment of synthetic aqueous solutions of 100 mL of AR1 by several EAOPs have been

investigated. BDD and GDE, both of 2.5 cm2, have been used at constant current upon addition of

50 mM Na2SO4 as supporting electrolyte at pH 3. The optimization of iron catalyst concentration

and applied current has preceded the treatment of solutions with up to 200 mg L-1 TOC content by

PEF. In addition, chromatographic analyses allowed the identification of aliphatic and cyclic by-

products formed during the cleavage of the AR1 structure.

Experimental

Chemicals

AR1 (disodium 8-acetamido-1-hydroxy-2-phenylazonaphthalene-3,6-disulfonate, C18H13N3Na2O8S2,

CI 18050, 60% purity) was purchased from Sigma-Aldrich. Anhydrous sodium sulfate used as

background electrolyte and iron(II) sulfate heptahydrate used as catalyst in EF and PEF were of

analytical grade from Merck. Solutions were prepared with bidistilled water and their pH was adjusted

to 3 before the electrolyses with analytical grade sodium hydroxide or sulfuric acid from Merck. Other

chemicals were obtained from Merck and Sigma-Aldrich.

Electrolytic system

The electrolyses were performed in an open, undivided cell containing a 100 mL solution and

featuring a double jacket for circulation of external thermostated water at 35 ºC (WiseCircu®

WCB-11 water bath). The solution was stirred with a magnetic bar at 800 rpm to ensure good

mixing and transport of reactants. The cell contained a 2.5 cm2 Si/BDD thin-film electrode from

Adamant® (500 ppm B) as the anode and a 2.5 cm2 carbon-PTFE air-diffusion cathode from

Electrocell®. The cathode was fed with compressed air flowing at 1 L min-1 for H2O2 generation.

The trials were performed at constant current provided by an MCP M10-QD305 power supply. In

PEF, the solution was irradiated with a Black Ray B100AP lamp.

Equipment and analytical procedures

The solution pH was measured on an Extech 321990 pH-meter. Samples withdrawn at regular

time intervals from electrolyzed solutions were neutralized at pH 7-8 to stop the degradation

process and filtered with 0.45 μm PTFE filters from Whatman before analysis. The decolourization

of AR1 solutions was monitored from the absorbance (A) decay at the maximum visible wave-

length (max) of 520 nm, measured from the spectra recorded on a Cary 1E UV/Vis

spectrophotometer (Varian). The percentage of colour removal or decolourization efficiency was

then determined as follows [21]:


238

Colour removal, % = 0 t

0

100A A

A

(6)

where A0 and At denote the absorbance at initial time and after an electrolysis time t, respectively.

The mineralization of solutions was monitored from their total organic carbon (TOC)

abatement, determined on a Vario TOC Select analyzer. From these data, the mineralization

current efficiency (MCE) at a given current (I / A) and electrolysis time (t / h) were estimated as

follows [1]:

MCE, % = s exp

7

( )100

4.32 10

nFV TOC

mIt

(7)

where F is the Faraday constant (96487 C mol-1), Vs is the solution volume (L), ∆(TOC)exp is the

experimental TOC decay (mg L-1), 4.32107 is a conversion factor (3600 s h-1 12000 mg mol-1) and

m is the number of carbon atoms of AR1 (18 atoms). The number of electrons (n) consumed per

each dye molecule was taken as 98 considering that its mineralization leads to carbon dioxide and

nitrate and sulfate ions as follows:

C18H13N3Na2O8S2 + 45 H2O 18 CO2 + 3 NO3 + 2 SO4

2 + 2 Na+ + 103 H+ + 98 e (8)

AR1 decay was followed by reversed-phase high performance liquid chromatography (HPLC)

with a Waters 625 LC fitted with a Hibar® RP-18e 5 µm, 150 4 mm, column at 25 ºC and coupled

with a photodiode array detector set at = 520 nm. The mobile phase was a 70:30 (v/v) acetonitri-

le/1.0 mM ammonium acetate (pH 4) mixture at 0.5 mL min-1. Generated carboxylic acids were de-

tected by ion-exclusion HPLC using the same LC fitted with a Bio-Rad Aminex HPX 87H,

300 × 7.8 mm, column at 25 °C and setting the array detector at = 210 nm. The isocratic elution

at 0.6 mL min-1 with 4 mM H2SO4 as the mobile phase yielded good peaks for maleic (tr = 8.8 min),

oxamic (tr = 10.6 min), malic (tr = 11.3 min), formic (tr = 14.9 min) and acetic (tr = 16.1 min).

The cyclic and/or aromatic intermediates were analyzed by gas chromatography coupled to

mass spectrometry (GC-MS). Several electrolyses were carried out under different experimental

conditions for short and long times. The final solutions were collected together until reaching

500 mL, which were then extracted three times with 30 mL CH2Cl2. The resulting organic solution

(90 mL) was dried with anhydrous Na2SO4, then filtered and completely evaporated in a rotary to

obtain a pale yellow solid that was further analyzed.


Influence of the experimental parameters on the degradation of Acid Red 1 by EF process

Solutions containing 300 mg L-1 AR1 (i.e., 0.59 mM AR1 or 100 mg L-1 TOC) were electrolyzed at

60 mA cm-2 in the presence of different amounts of Fe2+ as catalyst. As can be seen in Fig. 1, the

absence of Fe2+ (so-called EO process) caused the slowest decolourization and TOC abatement,

only reaching 70 % colour removal after 70 min and 25 % TOC decay after 180 min. Under the

present EO conditions, given the weak oxidation power of H2O2, the organic matter can be mainly

degraded by BDD(OH) formed via reaction (5). This radical tends to be very active towards the

initial pollutants and their by-products because it is weakly physisorbed on the anode surface and

it is generated at a very positive potential. Moreover, it is known that hydroxyl radicals can react

at high rate with all double bonds in the aromatic rings and, especially, with the –N=N– bond.

However, since BDD(OH) is confined to the anode vicinity, the degradation process becomes


doi: 10.5599/jese.2014.0058 239

severely limited by mass transport, thus being needed a much longer electrolysis time to

effectively destroy the molecules in a batch system without recirculation like the one tested here.

In contrast to the previous finding, within the same time period the presence of Fe2+ allowed

the complete decolourization as well as a greater mineralization in all cases, which can be account-

ed for by the crucial contribution of OH formed in the bulk from Fenton’s reaction (2). As can be

seen in Fig. 1a, the increase in Fe2+ concentration from 0.1 to 0.5 and then to 1.0 mM clearly

accelerated the colour removal, being necessary 70, 60 and 40 min, respectively, to get colourless

solutions.

Figure 1. Effect of Fe2+ concentration on (a) decolourization efficiency at 520 nm and

(b) TOC removal with electrolysis time for the electro-Fenton treatment of solutions of 300 mg L-1 AR1 in 0.05 M Na2SO4 at pH 3.0, 35 °C and 60 mA cm-2.

The rising catalyst content had a positive effect on the mineralization profiles as well, since 47,

54 and 63 % TOC removal was attained after 180 min using 0.1, 0.5 and 1.0 mM Fe2+. In contrast,

further increase to 1.5 mM was detrimental, eventually leading to slower colour removal and only

59 % TOC abatement at 180 min. This phenomenon can be mainly explained by the larger extent

of parasitic reactions causing the consumption of OH, particularly by Fe2+. It must be noted that

the mineralization was always partial, with a tendency to reach a plateau owing to the plausible


240

accumulation of refractory intermediates (as confirmed later on) that could not be oxidized by OH

in the bulk and were very slowly destroyed by BDD(OH). In conclusion, 1.0 mM was chosen as the

optimum Fe2+ concentration for subsequent tests.

The effect of current density within the range 8-80 mA cm-2, carried out under conditions

shown in Fig. 1 with 1 mM Fe2+ as the optimized amount, is depicted in Fig. 2. These trials aimed at

exploring the possibility of enhancing the decolourization and mineralization kinetics, which is

based on the fact that the applied current determines the yield of BDD(OH) formed via reaction

(5) as well as that of OH via reaction (2) because it depends on the H2O2 formation rate and Fe2+

regeneration rate.

Figure 2. Effect of current density on (a) decolourization efficiency at 520 nm and (b) TOC

removal with electrolysis time for the electro-Fenton treatment of solutions of 300 mg L-1 AR1 in 0.05 M Na2SO4 with 1.0 mM Fe2+ at pH 3.0 and 35 ºC.

A progressive increase in current density from 8 to 60 mA cm-2 caused the acceleration of both,

decolourization and mineralization. This can be easily explained by the faster generation of

BDD(OH) on the anode and OH in the bulk. Note that even the lowest current densities were able

to yield the complete decolourization at long electrolysis time. However, further increase to

80 mA cm-2 was detrimental since it caused a slower colour removal and led to a lower TOC


doi: 10.5599/jese.2014.0058 241

removal. This negative effect arises from the shift in cathode potential to unfavourable values that

promoted the reduction of O2 to H2O over reaction (1) and hindered the conversion of Fe3+ to Fe2+.

As a matter of fact, H2O2 concentration analyzed during the electrolyses at 60 and 80 mA cm-2

reached 25 mM and 15 mM, respectively. Thus, 60 mA cm-2 was chosen as the optimum value.

Effect of UVA light

As discussed in the Introduction, for most of the contaminated solutions studied in the past it

was possible to enhance the degradation process by irradiating them with UV light, which

favoured the oxidation of pollutants and their by-products due to the action of reaction (3) and

(4). In the present study, no direct photolysis of AR1 by UV light was observed, since its peak

during the HPLC analyses remained unchanged. This ensured the photostability of AR1 during the

electrolyses run under PEF conditions. For this purpose, the AR1 solutions were treated as done in

the previous EF experiments but incorporating the UV lamp near the cell.

As shown in Fig. 3a, solutions of 300 mg L-1 AR1 treated by PEF under the optimized conditions

described before (i.e., 1.0 mM Fe2+ and 60 mA cm-2) were more quickly decolourized compared to

EF trials, being required 50 min instead of 60 min to become colourless (see Fig. 1a for

comparison). The key contribution of photoreduction reaction (3) favoured the faster regeneration

of Fe2+, which then was able to accelerate the production of OH from reaction (2). On the other

hand, PEF also yielded a much larger mineralization after 180 min, reaching 90 % owing to

photodecarboxylation reaction (4). As reported elsewhere [1], some of the reaction intermediates

can form stable complexes with Fe(III) that can be effectively degraded only upon action of UV

photons since OH and BDD(OH) are much less effective. Fig. 3a also depicts the decay of AR1

monitored by HPLC during the same experiment. Its profile is quite similar to colour removal

profile, which means that no other coloured by-products were formed during the treatment.

Assuming a pseudo-first-order kinetics, an apparent rate constant (kapp) of 2.74×10-3 s-1 for AR1

decay was determined. The decay of the dye was also similar to the colour removal trend for EF

treatment (not shown), revealing a much smaller kapp = 0.59×10-3 s-1. Therefore, the beneficial

synergy between OH, BDD(OH) and UV light for the decontamination of AR1 solutions is

demonstrated.

Due to production needs, actual wastewaters may present a significant variation in the dye

content over time and thus, it is mandatory that the water treatment technology is flexible enough

to be adapted to such changes. The effect of AR1 concentration on the mineralization profile vs.

time is shown in Fig. 3b. A similar TOC removal was attained after 180 min for solutions containing

30 and 100 mg L-1 TOC. In contrast, only 75 % mineralization was reached for solutions with

200 mg L-1 TOC, which is simply due to the much larger number of organic molecules to be

degraded in the latter case. But, an important feature to be highlighted is the progressively

increasing slope of the curves (i.e., larger mineralization rate) upon increase of AR1 concentration,

which can be related to the more efficient reaction between OH/BDD(OH) and the organic

molecules. Indeed, low AR1 concentrations cause the waste of radicals in self-destruction and

other side reactions, whereas high organic contents lead to effective oxidation reactions. This is

clearly demonstrated in Fig. 3c, which compares the evolution of MCE vs time for several EAOPs.

For solutions with 100 mg L-1 TOC, the efficiency increases in the sequence EO < EF < PEF, with

maximum values of 10, 25 and 55 %, respectively. As discussed before, this can be related to the

more favorable synergy between different oxidants in the case of PEF. In addition, a greater MCE

resulted from the treatment of larger AR1 concentrations in PEF, reaching 85 % during the


242

treatment of 200 mg L-1 TOC, thus confirming the lower extent of parasitic reactions that cause

radical waste. Note that MCE tends to decrease at long electrolyses time, which can be explained

by (i) the formation of more resistant intermediates and (ii) the mass transport limitations related

to low organic loads.

Figure 3. (a) Decolourization efficiency at 520 nm and percentage of AR1 removal with

electrolysis time for the photoelectro-Fenton treatment of solutions of 300 mg L-1 AR1 in 0.05 M Na2SO4 with 1.0 mM Fe2+ at pH 3.0, 35 °C and 60 mA cm-2. (b) TOC abatement vs. time for the same experiment compared to trials at different AR1 concentrations. (c) MCE for trials shown in (b) compared to EO (0 mM Fe2+) and EF (1.0 mM Fe2+) treatments shown in Fig. 1b.


doi: 10.5599/jese.2014.0058 243

Identification of reaction intermediates

Table 1 summarizes the seven aromatic intermediates identified during PEF treatments.

Table 1. Structures of Acid Red 1 and its degradation intermediates identified by GC-MS analysis.

Chemical name Structure m/z

Acid Red 1

509.42

Naphthalene derivatives

Disodium 8-acetamido-1-hydroxy-2-(4-hydroxyphenylazo)naphthalene-3,6-disulfonate

524.12

Disodium 1,2-dihydroxy-3-phenylazonaphthalene-4,7-disulfonate

468.37

(5-Phenylazo-3,4,6-trihydroxynaphthalene)sulfonic acid

362.33

Benzene derivatives

Sodium (3,5-dihydroxy-2-phenylazo)benzenesulfonate

316.26

Sodium (1-hydroxy-2-[4-hydroxyphenylazo]-2-oxo)ethanesulfonate

282.21

2-(3,5-dihydroxyphenylazo)-1-hydroxy-2-oxo-ethanesulfonic acid

276.22

Hydroquinone

110.11


244

As can be seen, OH and BDD(OH) led to the hydroxylation of the benzenic and naphthalenic

rings of AR1 to yield three naphthalene derivatives. The subsequent action of radicals onto AR1

and/or onto those intermediates led to the formation of four benzene derivatives. The

accumulation of these aromatic intermediates would be dangerous due to their inherent high

toxicity and thus, PEF treatments had to be prolonged until their complete disappearance.

The progressive cleavage and oxidation of the aromatic intermediates gave rise to the

formation of short-chain aliphatic carboxylic acids, as also described elsewhere for other

pollutants [22]. Five C1-C4 acids were identified by ion-exclusion HPLC, namely maleic, oxamic,

malic, formic and acetic. As explained from Fig. 3, PEF ensured the almost complete removal of all

these acids since the TOC in the final solutions was <10 % after 180 min.

Conclusions

PEF technology is confirmed as a very powerful alternative for giving response to environmental

concerns related to water contamination by organic pollutants. This process allows a much faster

destruction of AR1 as well as a more significant and efficient TOC removal compared to EO and EF,

thus becoming a promising technology for the treatment of industrial wastewaters containing this

azo dye. The toxic intermediates formed during the first degradation stages are completely

transformed into aliphatic molecules, which are slowly converted to CO2 and H2O. The use of

renewable energy such us sunlight in sunny countries like Chile and Spain, giving rise to the so-

called solar photoelectro-Fenton (SPEF) process, would be an interesting feature for real-scale

application.

Acknowledgements: The authors thank CONICYT (Chile) for support under FONDECYT grant 1130391 and DICYT-USACh, as well as for the PhD fellowship N° 21130071 awarded to C. González-Vargas.

References

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Bioanalytical Chemistry 400 (2011) 353–360 [3] A. El-Ghenymy, N. Oturan, M. A. Oturan, J.A. Garrido, P. L. Cabot, F. Centellas, R. M.

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Electroanalytical Chemistry 702 (2013) 31–36 [5] N. Oturan, M. Zhou, M. A. Oturan, Journal of Physical Chemistry A 114 (2010) 10605–10611 [6] R. Salazar, M. S. Ureta-Zañartu, Water Air Soil Pollution 223 (2012) 4199–4207 [7] J. Urzúa, C. González-Vargas, F. Sepúlveda, M. S. Ureta-Zañartu, R. Salazar, Chemosphere 93

(2013) 2774–2781 [8] M. Zhou, Q. Yu, L. Lei, Dyes and Pigments 77 (2008) 129–136 [9] M. Panizza, M. A. Oturan, Electrochimica Acta 56 (2011) 7084–7087

[10] E. J. Ruiz , A. Hernández-Ramírez, J. M. Peralta-Hernández, C. Arias, E. Brillas, Chemical Engineering Journal 171 (2011) 385–392

[11] R. Salazar, E. Brillas, I. Sirés, Applied Catalysis B: Environmental 115-116 (2012) 107–116 [12] M. Panizza, G. Cerisola, Chemical Reviews 109 (2009) 6541–6569 [13] C. A. Martínez-Huitle, E. Brillas, Applied Catalysis B: Environmental 87(2009) 105–145 [14] European Food Safety Authority, The EFSA Journal 515 (2007) 1-28

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[15] Official Journal of the European Union L 195/8, 27.7.2007. Commission regulation (EC) No 884/2007 of July 2007 on emergency measures suspending the use of E 128 Red 2G as food colour

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doi: 10.5599/jese.2014.0061 247



www.jESE-online.org


Electrochemical combustion of indigo at ternary oxide coated titanium anodes

María I. León, Zaira G. Aguilar and José L. Nava

Departamento de Ingeniería Geomática e Hidráulica, Universidad de Guanajuato, Av. Juárez 77, Zona Centro, C.P. 36000, Guanajuato, Guanajuato, Mexico

Corresponding Author: E-mail: [email protected]; Tel.: +52-473-1020100 ext. 2289; Fax: +52-473-1020100 ext. 2209

Received: August 04, 2014; Revised: August 22, 2014; Published: December 6, 2014

Abstract The film of iridium and tin dioxides doped with antimony (IrO2-SnO2–Sb2O5) deposited on

a Ti substrate (mesh) obtained by Pechini method was used for the formation of OH

radicals by water discharge. Detection of OH radicals was followed by the use of the N,N-dimethyl-p-nitrosoaniline (RNO) as a spin trap. The electrode surface morphology and composition was characterized by SEM-EDS. The ternary oxide coating was used for the electrochemical combustion of indigo textile dye as a model organic compound in chloride medium. Bulk electrolyses were then carried out at different volumetric flow

rates under galvanostatic conditions using a filterpress flow cell. The galvanostatic tests using RNO confirmed that Ti/IrO2-SnO2-Sb2O5 favor the hydroxyl radical formation

at current densities between 5 and 7 mA cm2, while at current density of 10 mA cm2 the oxygen evolution reaction occurs. The indigo was totally decolorized and mineralized

via reactive oxygen species, such as (OH, H2O2, O3 and active chlorine) formed in-situ at

the Ti/IrO2-SnO2-Sb2O5 surface at volumetric flow rates between 0.10.4 L min-1 and at fixed current density of 7 mA cm-2. The mineralization of indigo carried out at 0.2 L min-1 achieved values of 100 %, with current efficiencies of 80 % and energy consumption of 1.78 KWh m-3.

Keywords Dimensionally stable anodes; electrochemical degradation of organics; Pechini method; textile effluents; indigo textile dye



J. Electrochem. Sci. Eng. 4(4) (2014) 247-258 ELECTROCHEMICAL COMBUSTION OF INDIGO TEXTILE DYE

248 248

Introduction

Textile processing industries nowadays are widespread sectors in many countries. This industry

is one of the most polluting industries in terms of the volume, color and complexity of its effluent

discharge. Textile effluents include dyes that have a complex chemical structure, which most of

the time are disposed on municipal sewers or into surface waters. Residual textile dyes tend to be

transformed into toxic aromatic amines which cannot be degraded by sunlight and, once in the

environment, they exhibit recalcitrant properties [1-3].

Electrochemical incineration [4-10] is a technique that has been found adequate for the

treatment of colored wastewaters. It is important to point out that several color degradation

studies mention systems with platinum electrodes [7] and dimensionally stable anodes (DSA) [6,8],

which have shown mineralization of 50-70 %. Dogan and Turkdemir [7] consider that mineraliza-

tion of indigo dye on Pt is induced by by-products of water and chloride discharge on the platinum

surface; however, the indigo achieved mineralization of 60 %. Similar results in the degradation of

acid red 29 [11], reactive blue 19 [8], mediated by active chlorine (given by the mixture of chlorine

(Cl2), hypochlorous acid (HOCl) and hypochlorite ion (OCl-)), produced on DSA lead to mineraliza-

tion of 56 % and 70 %, respectively. BDD electrodes exhibit a superior performance, since a large

amount of hydroxyl radicals (OH) are formed by water oxidation on the BDD surface [5-6,12-13],

achieving 100 % efficiency in color removal and mineralization. The main problem encountered

with BDD electrode is its high price limiting its industrial application.

For the above it is necessary to develop a DSA of metal oxides as an alternative to oxidize

recalcitrant organic matter similar to a BDD electrode, in other words to produce DSA(OH)

capable to oxidize recalcitrant organic matter. Comninellis and coworkers have developed a DSA

electrode of SnO2–Sb2O5 with an interlayer between supports (Ti) of IrO2 by the spray pyrolysis

technique, capable to produce hydroxyl radicals physisorbed on DSA (Eq. 1), by water dischar-

ge [14]. The interlayer of IrO2 improves useful life of the electrode. These authors put on evidence

that the physisorbed hydroxyl radical DSA(OH) cause predominantly the complete combustion of

organics (R), Eq. (2); for example, these authors demonstrated that DSA(OH) reacts with

p-clorophenol leading to complete combustion. On such electrode, IrO2 acts as a catalyst, SnO2

acts as a dispersing agent and Sb2O5 as a doping agent. Such ternary electrodes are among the

best electrocatalysts for O2 evolution, being able to produce physisorbed hydroxyl radicals on their

surface from water discharge. The high catalytic activity of this ternary oxide electrode has been

recently reported for the electrochemical oxidation of other organic compounds [15,16]. Another

paper by Comninellis put on evidence the convenience of using Ti/SnO2 to oxidize phenol matter

via OH radicals adsorbed onto Ti/SnO2 [17]. However, the main problem encountered with the

Ti/SnO2 anode is its low stability under anodic polarization, which is not the case of the SnO2–

Sb2O5 coating having an IrO2 interlayer between the Ti substrate [18].

-2DSA H O DSA( OH) H 1e (1)

-z 2R DSA( OH) CO zH ze DSA (2)

In a previous paper carried out by our group a film of iridium and tin dioxides doped with

antimony oxide (IrO2-SnO2–Sb2O5) was deposited onto Ti substrate mesh and plate by the Pechini

method [19]. The ternary oxide coating was used for the anodic decolorization of methyl orange

(MO) azo dye via reactive oxygen species, such as (OH, H2O2 and O3) formed in-situ from water

M. I. León et al. J. Electrochem. Sci. Eng. 4(4) (2014) 247-258

doi: 10.5599/jese.2014.0061 249

oxidation at the Ti/IrO2-SnO2-Sb2O5 surface. However, in that paper we did not follow the

formation of OH at DSA surface and the electrochemical combustion of organic matter.

The indirect technique for the detection and identification of low concentration of OH radicals

formed by water discharge at the oxide anodes involves trapping of the OH radical by an addition

reaction (spin trap) to produce a more stable radical (spin adduct). A number of OH radical spin

traps are available in the literature but N,N-dimethyl-p-nitrosoaniline (RNO) has demonstrated to

be effective owing to the selective reaction of RNO with OH radicals, the high rate of the reaction

with OH radicals (k = 1.2×104 M1 s1) and the ease of application as one merely observes the

bleaching of the sensitive absorption band at 440 nm [17, 20].

The goal of this manuscript is to prepare a film of iridium and tin dioxides doped with antimony

(IrO2-SnO2–Sb2O5) onto titanium mesh (expanded metal) to produce OH radicals via water

discharge for the electrochemical combustion of indigo textile dye (which resembles a denim

laundry industrial wastewater). Bulk electrolyses were then carried out at different mean linear

flow velocities and at constant current density using a filterpress flow cell. The integral current

efficiency and the energy consumption of electrolysis were estimated. The detection of OH

radicals formed by water discharge at the oxide anode using RNO as spin trap was also examined.

Experimental

Indigo dye solution was 1 mM indigo textile dye (536 ppm COD) in 0.05 M NaCl (which resem-

bles a denim laundry industrial wastewater). The resulting solution exhibited a conductivity of

5.78 mS cm-1, and a pH of 6.3 at 298 K. The solution was deoxygenated with nitrogen for about

10 minutes before each experiment. All the chemicals employed in this work were reactive grade.

Equipment

A potentiostat-galvanostat model SP-150 coupled to a booster model VMP-3 (20V-10A) both

from Bio-LogicTM with EC-Lab® software were used for the electrolysis experiments. The potentials

were measured versus a saturated calomel reference electrode (SCE), Bio-Logic model

002056RE-2B. All electrode potentials shown in this work are presented with regard to a standard

hydrogen electrode (SHE).

COD analyses were performed using a dry-bath (Lab Line Model 2008), and a Genesys 20 spec-

trophotometer. Chloride volumetric titrations were confirmed by potentiometric measurements

using a silver wire and a SCE, which was inserted in a glassy titration cell. The potential differences

between silver wire and SCE were detected by a high impedance multimeter (Agilent-mo-

del-34401A). The colour removal was registered using a visible spectrophotometer (Genesys 20).

Microelectrolysis experiments

A 100-mL Pyrex electrochemical cell, with a three electrode system and nitrogen inlet was used

for the construction of the anodic polarization curves. The working electrode was

mesh-(IrO2-SnO2–Sb2O5) with 1 cm2 geometric area exposed to the electrolyte. The potentials

were measured vs. SCE and the counter electrode was a glassy carbon. All the potential measure-

ments shown in this work are presented with regard to standard hydrogen electrode (SHE).

A divided cell made of two compartment quartz cells of 3 mL capacity each one for the indirect

detection of OH radicals was used. The anode was in the form of plate (1 cm2) and the cathode

was a vitreous carbon rod (1 cm2). A home-made salt bridge to connect both semi-cells was

employed; this was fabricated with vitreous Pyrex tube of 2 mm diameter sealed with Pt at the

ends; this bridge was filled with phosphate buffer (pH 7.4). The quartz cell used as the anodic


250 250

compartment was collocated into the UV-visible spectrophotometer (Perkin Elmer Lambda 35) to

follow the bleaching (in-situ) of the yellow color of RNO during electrolysis.

Flow cell experiments

The flow cell FM01-LC that includes the turbulence promoter type D was used; the detailed

description of this cell is depicted elsewhere [21]. In this work the spacer was 0.55 cm thick. DSA

anode was a mesh-(IrO2-SnO2–Sb2O5), while platinum coated titanium flat sheet, was used as the

cathode. DSA electrode was prepared by Pechini method described below. The platinum coated

titanium was provided by De Nora. Details on the FM01-LC cell characteristics are given in Table 1.

Table 1. Mesh-(Ti/IrO2-SnO2-Sb2O5) electrode dimensions, experimental details of the FM01-LC electrolyzer.

Electrode length, L 16 cm

Electrode height, B 4 cm

Electrode spacing, S 0.55 cm

Anode area, (Ti/IrO2-SnO2-Sb2O5) 112 cm2

Cathode área, (Ti/platinized) 64 cm2

Overall voidage, (Ti/IrO2-SnO2-Sb2O5) 0.93

Volumetric flow rate, from 0.1 to 0.4 L min-1

Overall voidage is the ratio of the free space in the channel to overall channel volume.

Figure 1. Electrical and flow circuit for the measurement of electrochemical incineration kinetics at FM01-LC electrolyzer.

The undivided FM01-LC cell, with a single electrolyte compartment and the electrolyte flow

circuit, is shown in Figure 1. The electrolyte was contained in a 1 L polycarbonate reservoir. A mag-

netically coupled pump of 1/15 hp March MFG, model MDX-MT-3 was used; the flow rates were

measured by a variable area glass rotameter from Cole Palmer, model F44500. The electrolyte

+ -

Flow meter

Magnetic pump

BDD Anode CathodeN2

FM01

- LC

Reservoir

Potentiostat -Galvanostat

+ -

Flow meter

Magnetic pump

BDD Anode CathodeN2

FM01

- LC

Reservoir

Potentiostat -Galvanostat


doi: 10.5599/jese.2014.0061 251

circuit was constructed from Master Flex tubing, C-Flex 6424-16, of 0.5 inch diameter. The valves

and the three way connectors were made of PVC.

Scanning electron microscopy

Surface characterization of the metallic coating was performed using a SEM Carl Zeiss DSM

940A microscope. The energy of the primary electrons beam employed was 15 keV.

Methodology

Preparation of the DSA material

A ternary oxide (IrO2-SnO2–Sb2O5) film was deposited onto a Ti plate and mesh to be used in the

three electrode cell and in the flow cell (Figure 1) by Pechini method using appropriate molar

ratios of the oxide components. The precursor polymer solution was a mixture of citric acid (CA) in

ethylene glycol (EG) at 60-70 °C. After total dissolution of the CA, H2IrCl6xH2O, SnCl4 and SbCl3

were added to the mixture according to a molar composition of EG:CA:Ir:Sn:Sb as

16:0.12:0.0296:0.0296:0.0004, maintaining the temperature at 60-70 °C for 30 min. This mixture

was then applied with a brush to both sides of the pre-treated Ti support. After the application of

the coating, the electrode was heated at 100 oC for 5 min in a furnace in order to induce the

polymerization of the precursor. This procedure was repeated eight times. After the final coating,

the electrodes were maintained at 550 oC for 1 h in order to calcinate the polymer and form the

ternary oxide (IrO2-SnO2–Sb2O5); XRD analysis confirmed at such at temperature these oxide

phases are obtained [22]. The temperature did not exceed 600 oC to avoid the formation of TiO2

that markedly reduces the electrocatalytic properties of the Ti/IrO2-SnO2–Sb2O5 coating due to

passivation [23].

Microelectrolysis tests

Anodic polarization curves to determine the limits of potential and current density where the

media is oxidized at Ti/IrO2-SnO2–Sb2O5 electrode were performed. These studies were carried out

in the solution containing phosphate buffer (pH 7.4), and in the presence of 2×105 M RNO in the

same buffer at room temperature (298 K). Anodic potential limit of 1.6 V vs. SHE was applied from

open circuit potential (OCP) (0.82 V vs. SHE) using the linear sweep voltammetry technique at

50 mV s1. Based on these polarization curves the detection of hydroxyl radicals was performed.

Detection of hydroxyl radicals

In this paper RNO was used as spin trap for the detection of low concentration of OH radicals

formed by water discharge at the Ti/IrO2-SnO2-Sb2O5 electrode and the bleaching of the yellow

colour was measured during electrolysis [17]. RNO traps the OH radical by an addition reaction to

produce a more stable radical (spin adduct), Eq. (3) [17].

(3)

It is important to mention that RNO is electrochemical inactive at Pt, SnO2 and IrO2 anodes

[17,20]. A divided cell for the indirect detection of OH radicals was used (see microelectrolysis

experiments section). Anodes screening tests were carried out in phosphate buffer (pH=7.4)

containing 2×105 M RNO. Galvanostatic electrolyses at current densities of 5, 7 and 10 mA cm2

applied to the Ti/IrO2-SnO2-Sb2O5 electrode were performed; at the same time the bleaching


252 252

(in-situ) of the yellow color of RNO during electrolysis was followed. The same tests were

performed using Pt plate (1 cm2) as anode for which the surface OH radical concentration is

almost zero [17].

Electrochemical incineration in the filter-press flow cell.

Electrochemical incinerations of indigo were carried out in the FM01-LC cell equipped with

mesh-(Ti/IrO2-SnO2-Sb2O5) at current density of 7 mA cm-2, value determined from microelectro-

lysis studies, at different volumetric flow rates between 0.10.4 L min.1.

Incineration evolution was estimated by COD analysis of samples taken at different times. The

COD values were determined by closed reflux dichromate titration method [24]. It is important to

mention that estimating residual organic matter by COD analysis allowed eliminating any interfe-

rence from chloride species. For this method, an excess of HgSO4 was added and Ag2SO4 in the di-

gestion and catalyst solutions, respectively, with the purpose of eliminating possible interferences

from chloride species during the estimation of the residual organic matter from COD analysis [12].

The chloride concentration was evaluated by volumetric titration using a 0.5 M AgNO3, con-

firmed by potentiometric measurements [12]. In addition, the color removal was determined by

the decrease in absorbance at 639 nm, during electrolyzes.


Characterization of DSA

Figure 2 presents typical scanning electron micrographs for freshly prepared electrode Ti/IrO2-

SnO2-Sb2O5. The surface morphology of the layer is characterized by the presence of crackers and

plates. The presence of plates on the surface is probably due to the drastic heat treatment to

which the sampled was submitted, that promoted the rapid exit of CO2 gas originated from the

decomposition of the organic polymer. EDX analyses focused on several plate structures show

heterogeneous atomic percentage ratio of Sn and Ir (between 1.6 to 2.74), indicating that Sn

segregates from other oxide to form a Sn rich deposit. Moreover, antimony was randomly

detected along the electrode, showing that Sb is not homogeneously distributed along the

electrode surface owing to its low content.

Figure 2. SEM images of Ti/IrO2-SnO2-Sb2O5.

Figure 3 shows typical linear sweep voltammetries obtained on Ti/IrO2-SnO2-Sb2O5 electrode in

the solution containing phosphate buffer (pH 7.4), and in the presence of 2×105 M RNO in the

20 mm


doi: 10.5599/jese.2014.0061 253

same buffer where no differences were detected. The fact that no changes were detected in both

electrolytic solutions suggests the oxidation of water which is found in excess. Tafel slope

performed on Ti/IrO2-SnO2-Sb2O5 from these curves (see inset), gives value of 190 mV dec-1, which

is different to that reported for Ti/IrO2/SnO2-Sb2O5 and Ti/Pt/SnO2-Sb2O4, 120 and 204 mV dec-1

obtained at 298 K, respectively [12,25]; this difference is associated with the electrode

composition and by the method of preparation.

Figure 3. Typical linear sweep voltammetries on Ti/IrO2-SnO2-Sb2O5 anode.

Electrolyte: phosphate buffer (pH 7.4), and phosphate buffer + 2×105 M RNO. The scan rate was 50 mV s1. The inset shows the Tafel plot for J-E curves for phosphate buffer. A = 1 cm2. T = 298 K.

Figure 4. Absorbance spectra of RNO (2×105 M) in phosphate buffer (pH=7.4) obtained at 5 min intervals during galvanostatic electrolyses with Ti/IrO2-SnO2-Sb2O5 (a) and Pt (b) anodes. A = 1 cm2. T=298 K.

For screening tests of anodes we used RNO as spin trap of OH radicals. Figure 4 shows the

absorption spectrum of aqueous solution (2×105 M RNO) in phosphate buffer at pH 7.4 during

0

4

8

12

16

0.8 0.9 1 1.1 1.2 1.3 1.4 1.5 1.6

J /

mA

cm

-2

E / V vs. SHE

Phosphate buffer + RNO

Phosphate buffer

1.1

1.14

1.18

1.22

1.26

1.3

-0.6 -0.4 -0.2 0 0.2 0.4

E / V

vs. S

HE

log J / mA cm-2


254 254

galvanostatic electrolysis at 5, 7 and 10 mA cm2 with Ti/IrO2-SnO2-Sb2O5 and Pt electrode. With Pt

anode, there is no decrease in absorbance at 440 nm, at the three current densities, contrary to

the Ti/IrO2-SnO2-Sb2O5 anode for which there is a rapid decrease in the absorbance at 5 and

7 mA cm2. These results show that there is accumulation of OH radicals at the Ti/IrO2-SnO2-Sb2O5

electrode surface contrary to Pt anode for which the surface OH radical is almost zero. The fact

that the Ti/IrO2-SnO2-Sb2O5 anode at 10 mA cm2 behaves similar to that Pt suggests that at such

current density the accumulation of OH radicals is zero and the oxygen evolution reaction starts

to appear. Therefore, according to the proposed reactions (Eqs. (1) and (2)) [14,17] the Ti/IrO2-

SnO2-Sb2O5 will favor complete combustion of indigo textile dye at 5 and 7 mA cm2.

Electrochemical incineration of indigo textile dye in the FM01-LC using DSA electrode

Figures 5 (a) and (b) show the normalized color (detected at = 639 nm) and COD results

obtained from experiments performed at constant current density (7 mA cm-2) and variable

volumetric flow rates. In these figures, the normalized color decreases faster than COD with the

electrolysis time at different volumetric flow rates. COD kinetic was lower than that obtained for

color decay owing to the slower combustion of by-products. However, color and COD depletion do

not show marked improvement at the elevated volumetric flow rates.

Given that the presence of chloride ions (i.e., 0.05 M in this study) is relevant due to the

possible formation of active chlorine by oxidation at Ti/IrO2-SnO2-Sb2O5, the chloride consumption

at the end of the electrolysis was measured (Figure 6), giving an average conversion between

15-40 %. This value did not show a marked dependence with hydrodynamics. This indicates that,

despite the predominant role of Ti/IrO2-SnO2-Sb2O5 (OH) as oxidant species, indigo and/or its by-

products can be simultaneously destroyed by other oxidants such as dissolved chlorine gas,

hypochlorous acid (HClO) and hypochlorite ion (ClO-), as well as chlorate and perchlorate ions

formed upon electro-oxidation with Ti/IrO2-SnO2-Sb2O5 electrode.

The complete combustion obtained here confirms that the OH radical, in addition to the other

oxidants, are responsible for the oxidation of indigo, which does not occur on platinum electrodes,

where the oxidation of indigo in chloride medium achieved 60 % in terms of COD [7]. The results

obtained here are in agreement with other articles carried out by our group, where we achieved

the complete combustion of indigo mediated by OH and active chlorine (produced on BDD in the

same filter-press flow cell) [12,13].

The fact that hydrodynamics does not improve indigo oxidation and color removal may be

associated with a complex mechanism of indigo degradation. HPLC studies would be helpful in the

identification of possible indigo oxidation by-products; however, these were beyond the scope of

the present work. It is important to point out that all of the electrolyses presented herein were

carried out in the undivided FM01-LC cell, for which reason the degradation of indigo may also

involve reactions at the cathode (Ti/Pt).

With the data obtained from COD for all of the electrolyses at their respective volumetric flow

rates, integral current efficiency and energy consumption were analyzed as a function of

percentage of indigo oxidation, for electrolyses performed at 7 mA cm-2, Figure 7 (a)-(b). The esti-

mation of integral current efficiency and energy consumption were determined using Equations

(4) and (5) [12]:


doi: 10.5599/jese.2014.0061 255

Figure 5. Normalized color ( = 639 nm) (a) and COD (b) decay during the electrolyses of indigo on

(Ti/IrO2-SnO2-Sb2O5) in the FM01-LC electrolyzer. Electrolyte: 1 mM indigo in 0.05 M NaCl; this composition resembles a denim laundry wastewater. A = 112 cm2, j = 7 mA cm-2, T = 298 K.

Volumetric flow rates are shown in the figure.

Figure 6. Normalized concentration of chloride versus volumetric flow rates evaluated at the end of the

electrolyses similar to those from Fig. 5(b). Electrolyte: 1 mM indigo in 0.05 M NaCl. A = 112 cm2, j = 7 mA cm-2, T = 298 K. Volumetric flow rates are shown in the figure.


256 256

4 [ (0) ( )]FV COD COD t

It

(4)

celc

m

4 1

3.6lFE

EV

(5)

where F is the Faraday constant, 96485 C mol-1, V is the solution volume (cm-3), COD(0) and COD(t)

are the chemical oxygen demand initially and at time (t) of the electrolysis, in mol cm-3, I is the ap-

plied current, in A, t is the time of electrolysis (s), Ecell is the cell potential in V, and Vm is the molar

volume in cm3 mol-1. The value of 3.6 is a correction factor which converts Ec to units of KWh m-3.

Figure 7(a) shows that current efficiency surpasses 100 % (theoretical value) at volumetric flow

rates of 0.1 and 0.3 L min1, suggesting those indigo oxidation by-products and/or the processes

taking place at the cathode enhance the degradation of indigo. A similar behavior was obtained in

a previous communication carried out by our group [12], during indigo mineralization process in

the same filter-press reactor. On the other hand, for the volumetric flow rates of 0.2 and

0.4 L min1, the current efficiencies were lower than that obtained for 0.1 and 0.3 L min1. It is

important to remark that at the end of the electrolyses the current efficiency where 80 % for all

volumetric flow rates studied, and there are no marked effects of the hydrodynamics on current

efficiency in the set of electrolyses studied herein.

The analysis of Figure 7(b) shows that the energy consumption is not strongly influenced by

hydrodynamics at 0.2-0.4 L min1. It is important to emphasize that the energy consumption is at

least four times lower than those obtained in a previous paper, carried out by our group using the

FM01-LC electrolyzer equipped with BDD electrodes in the same indigo solution [12]. This savings

in energy consumption is due to the lower electrode polarization obtained using DSA (1.2 V) than

the obtained on BDD (2.4 V), diminishing cell potential.

Figure 7. (a) Integral current efficiency versus percentage of oxidized indigo in the FM01-LC electrolyzer, evaluated from the electrolyses similar to those from Fig. 5(b). (b) Energy consumption versus volumetric

flow rate evaluated at 88 % of degradation from the electrolyses similar to those from Fig. 5(b).


doi: 10.5599/jese.2014.0061 257

The study presented here indicates that, despite the predominant role of

Ti/IrO2-SnO2-Sb2O5(OH) as oxidant species, indigo and/or its by-products can be simultaneously

destroyed by other oxidants such as dissolved chlorine gas, hypochlorous acid (HClO) and

hypochlorite ion (ClO-), as well as chlorate and perchlorate ions formed upon electro-oxidation

with Ti/IrO2-SnO2-Sb2O5 electrode.

Conclusions

The detection of OH radicals formed by water discharge at Ti/IrO2-SnO2-Sb2O5 using

N,N-dimethyl-p-nitrosoaniline (RNO) as a spin trap showed that exits an accumulation of OH

radical at Ti/IrO2-SnO2-Sb2O5 surface. Therefore, the Ti/IrO2-SnO2-Sb2O5 anode favors complete

combustion of indigo by bulk electrolysis.

The galvanostatic tests using RNO as spin trap of OH radicals confirmed that Ti/IrO2-SnO2-Sb2O5

will favor the hydroxyl radical formation at current densities between 5 and 7 mA cm2, while at

current density of 10 mA cm2 the oxygen evolution reaction occurs.

Electrolyses in a FM01-LC flow cell indicates, that despite the predominant role of

Ti/IrO2-SnO2-Sb2O5 (OH) as oxidant species, indigo and/or its by-products can be simultaneously

destroyed by other oxidants such as dissolved chlorine gas, hypochlorous acid (HClO) and

hypochlorite ion (ClO-), as well as chlorate and perchlorate ions formed upon electro-oxidation

with Ti/IrO2-SnO2-Sb2O5 electrode.

The mineralization of indigo carried out at 0.2 L min1 and 7 mA cm2 achieved values of 100 %,

with current efficiencies 80 %, and energy consumption of 1.78 KWh m-3. The FM01-LC equipped

with mesh-(Ti/IrO2-SnO2-Sb2O5) improves space-time yield, allowing better interaction between

mesh-(Ti/IrO2-SnO2-Sb2O5)(OH) and organics, a phenomenon that increases organic mineralization

efficiency.

In this manner, the complete mineralization of indigo with high current efficiency, obtained in

this work is a notable improvement over those reported in the literature by using other DSA

electrode. Additionally, the performance of the FM01-LC electrolyzer equipped with mesh-

(Ti/IrO2-SnO2-Sb2O5) electrodes, demonstrate the convenience of using this electrochemical

reactor as a pre-pilot cell for other water samples containing recalcitrant organic matter.

Acknowledgements: María I. León and Zaira G. Aguilar thank CONACYT for the given grant. Authors are grateful to CONACYT and CONCYTEG for the economic support via the project FOMIX GTO-2012-C04-195057. Authors also acknowledge Universidad de Guanajuato for the financial support.

References

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[10] P. Cañizares, A. Gadri, J. Lobato, B. Nasr, M. A. Rodrigo, C. Saez, Industrial & Engineering Chemistry Research 45 (2006) 3468-3473.

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doi: 10.5599/jese.2014.0069 259



www.jESE-online.org


Electrochemical mediated oxidation of phenol using Ti/IrO2 and Ti/Pt-SnO2-Sb2O5 electrodes

Jéssica Pires de Paiva Barreto, Elisama Vieira dos Santos, Mariana Medeiros Oliveira, Djalma Ribeiro da Silva, João Fernandes de Souza* and Carlos A. Martínez-Huitle

Federal University of Rio Grande do Norte, CCET - Institute of Chemistry, Campus Universitario, Lagoa Nova - CEP 59.072-970, RN, Brazil *Federal University of Rio Grande do Norte, CCET – Department of Chemical Engineering, Campus Universitario, Lagoa Nova - CEP 59.072-970, RN, Brazil

Corresponding Author: E-mail: [email protected]; Tel.: +55-84-9181-7147;

Received: October 13, 2014; Published: December 6, 2014

Abstract The indirect electrochemical oxidation of phenol was studied at Ti/IrO2 and Ti/Pt-SnO2-Sb2O5 electrodes by bulk electrolysis experiments under galvanostatic control. The obtained results clearly shown that the electrode material was an important para-meter for the optimization of such processes determining their mechanism and oxidation products. Different current efficiencies were obtained at Ti/IrO2 and Ti/Pt-SnO2-Sb2O5, depending on the applied current density in the range from 10, 20 and 30 mA cm−2. The effect of the amount of dissolved NaCl was studied also. It was observed that the electrochemical processes (direct/indirect) favor specific oxidation pathways depending on electrocatalytic material. Phenol degradation generates several intermediates eventually leading to complete mineralization, as indicated by the results obtained with the High-performance liquid chromatography (HPLC) technique.

Keywords Phenol; anode material; chlorine active species; indirect electrochemical oxidation.

Introduction

Phenol is an aromatic compound and it is a hygroscopic crystalline solid at ambient temperatu-

re and pressure. When pure, solid phenol is white but is mostly colored due to the presence of im-

purities. Phenol is very soluble in ethyl alcohol, in ether and in several polar solvents, as well as in

hydrocarbons such as benzene. In water it has a limited solubility and behaves as a weak acid [1].

More phenol than is usually found in the environment has been found in surface waters and



J. Electrochem. Sci. Eng. 4(4) (2014) 259-270 ELECTROCHEMICAL OXIDATION OF PHENOL

260 260

surrounding air that were contaminated when phenol was released from industries and

commercial products containing phenol. It has been found in materials released from landfills and

hazardous waste sites, and it has been found in the groundwater near these sites [2,3].

The interest in developing new and more efficient methods for destruction of hazardous waste

such as phenol [4] and the conversion of mixed waste to low-level-toxicity waste has significantly

increased. These wastewaters are difficult to be effectively treated by conventional biological

methods, for this reason, advanced oxidation processes (AOPs) have been developed to treat

these bio-refractory organic wastewaters [1-11]. In this frame, the electrochemical oxidation of

the model substrates has been investigated using several anodic materials, generally metal oxides

like IrO2, PbO2, SnO2 and SnO2–Sb2O5 [5,6]. Dimensionally stable anodes (DSAs) belong to a

particular category of electrodes, constituted of a Ti-support coated by noble metal oxides, which

confer the enhanced catalytic activity towards chlorine evolution and oxygen evolution reaction

(o.e.r). To date, DSAs have been largely employed in the chloro-alkali industry, due to their

excellent catalytic property and service life [12,13]; however, significant performances have been

obtained when these have been used for electrochemical treatment of industrial effluents [14].

In the case of active chlorine, the interest in this oxidant is based on the ubiquitous presence of

chloride ions in a certain number of effluents and natural waters, making possible the involvement

of active chlorine during electrochemical treatment using principally DSA electrodes; and the chemi-

stry and electrochemistry of higher oxidation states for chlorine close to neutral pH [15,16].

The electrochemical treatment of phenol has been already studied by other authors [2-4,9-11,

17,18]. Using NaCl solutions to degradate phenol by direct anodic reaction and/or through the

mediation of active chlorine [19,20]; the possible role of the Cl- ion during the process was not

taken into account by the authors [19,20]. Therefore, the electrochemical degradation of phenol

was studied at Ti/IrO2 and Ti/Pt-SnO2-Sb2O5 electrodes by varying operating conditions such as

current density and Cl- concentration. The intermediate species formed at each one of the anodes

were compared.

Experimental

Chemicals

The compounds used such as phenol, catechol, hydroquinone, oxalic acidic were of analytical

grade. The chromatographic elution solvents were of HPLC grade (Merck). The stock solutions

(1000 mg L-1) of phenol and its degradation products were prepared from certified reference

standards (purity > 98 %), with the dissolution in methanol HPLC grade. All electrolyte solutions

were prepares using purified Milli-Q water system with a conductivity of 0.1 µS cm-1.

Electrochemical measurements

Electrochemical analyses were performed with an Autolab model PGSTAT320N (Metrohm).

Quasi-steady polarization curves were carried out at a scan rate of 2.5 mV s−1 and with a 0.45 mV

step potential, in solutions of NaCl at different concentrations. Experiments were carried out in a

conventional three-electrode system, and measurements were performed in a range from 0.0 V to

3.5 V. Ti/IrO2 and Ti-Pt-SnO2-Sb2O5, with an exposed geometric area of ca. 1.0 cm2, were used as

the working electrode, while a platinum wire and an Ag/AgCl (KCl 3 mol L−1) electrode were

employed as the auxiliary and reference electrodes, respectively.

J. P. de Paiva Barreto et al. J. Electrochem. Sci. Eng. 4(4) (2014) 259-270

doi: 10.5599/jese.2014.0069 261

Electrolytic systems

Bulk oxidations were performed in an electrolytic flow cell with a single-compartment with

parallel plate electrodes [21]. Circular electrodes (Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrode) were

used as anodes exposing to the effluent a nominal surface area of 63.5 cm2. In all cases, a Ti disc

was used as the cathode. The inter-electrode gap was 10 mm. For the electrochemical flow cell,

inlet and outlet were provided for effluent circulation through the reactor; the solution of phenol

was stored in a thermoregulated glass tank (1 L) and circulated through the cell using a peristaltic

pump, at a flow rate of 151 dm3 h−1, which allowed a mass transfer coefficient (determined using

the ferri/ferro-cyanide redox couple) of 2.0×10−5 m s−1 [22]. The oxidation experiments of phenol

were performed under galvanostatic conditions (using a power supply MINIPA-3305M) at 25 °C for

studying the role of applied current density (j = 10, 20 and 30 mA cm-2) adding 20 and 30 mM of Cl-

for studying the effect of Cl-mediated approach.

Analytical methods

The oxidation intermediates, produced during the electrolysis experiments at both anodes,

were analyzed by HPLC. Chromatographic separations were performed on an analytical column

Supelcosil-C18 (5 µm, 25 × 46 mm) at room temperature and with an UV detector at λ = 225 nm.

Generated carboxylic acids were detected and quantified using an Ultimate TMAQ- C18 5 m

(25 × 46 mm) column at room temperature and photodiode array detector set at λ = 210 nm. For

these analyses, a 70:30 (v/v) methanol/water mixture at 0.5 mL min-1 for the elution was used. The

flow rate of the mobile phase was 1.5 mL min-1. Spectrophotometric measurements (UV–Vis) were

also performed using a Shimadzu model UV-160 spectrophotometer. Experimentally, degradation

of phenol was monitored from the abatement of their chemical oxygen demand (COD). Values

were obtained, using a HANNA HI 83099 spectrophotometer after digestion of samples in a

HANNA thermo-reactor, in order to estimate the Total Current Efficiency (TCE), using the following

relationship [23]:

0 fCOD CODTCE, % 100

8FV

I t

(1)

where COD0 and CODf are chemical oxygen demands at times t=0 (initial) and f (final time) in

g O2 dm−3, respectively; I the current (A), F the Faraday constant (96,487 C mol−1), V the electrolyte

volume (dm3), 8 is the oxygen equivalent mass (g eq.−1) and Δt is the total time of electrolysis,

allowing for a global determination of the overall efficiency of the process. Additionally, the

limiting current can be estimated from the value of COD using the equation 2 for electrochemical

oxidation of phenol [24,25].

lim m( ) 4 COD( )I t FAk t (2)

where Ilim(t) is the limiting current (A) at a given time t, 4 the number of exchanged electrons, A

the electrode area (m2), F the Faraday’s constant, km the average mass transport coefficient in the

electrochemical reactor (m s−1) and COD(t) the COD, mol O2 m−3 at a given time t.

The energy consumption (EC) per volume of phenol oxidized was estimated and expressed in

kWh dm-3. The average cell voltage, during the electrolysis, is taken for calculating the energy EC,

as follows:

cEnergy consumption3600

E It

V

(3)


262 262

where t is the time of electrolysis (s); ΔEc / V and I / A are the average cell voltage and the electro-

lysis current, respectively; and V is the sample volume (dm3).


Polarization curves in the presence of halide

Based on the considerations about the possible effect of halide on the oxygen evolution

reaction (o.e.r.) polarization curves were recorded in the absence and in the presence of different

concentrations of Cl−. The results obtained in the presence of chloride ions (10 to 80 mg L-1), at

both anode materials, are shown in Fig. 1. In the case of the Ti/IrO2 anode, in absence of chloride

in solution, the oxygen evolution reaction is attained around 1.7 V. After that, the whole

polarization curve is modestly shifted to less positive potentials (up to 1.5 V), when the

concentration of NaCl is increased. This behavior is due to the increase of the importance of the

Cl2/H2O system [23,26], favoring the production of active chlorine species than the oxygen

evolution reaction. Under these conditions, a fast incineration of a number of organic substrates

can be favored during mediated electrochemical process due to the production of oxychloro-

radicals, often assumed as intermediates in the chlorine evolution reaction.

Figure 1. Current-potential curves in the presence of different amounts of NaCl on Ti/IrO2 and

Ti-Pt-SnO2-Sb2O5 anodes at scan rate of 2.5 mV s−1. Black curve: Water with lower conductivity.

Similar experiments were carried out using Ti-Pt-SnO2-Sb2O5 anode in the presence of Cl−, as

shown in Fig. 1, employing the same range of Cl− concentrations. In that case, at very small NaCl

concentration (10 mg L−1), a relevant shift to less positive potentials was observed of I/E curves.

Above 20 mg L−1, the anode potential becomes increasingly buffered by the halide electroactivity.

This behavior can be attributed to an interaction between anode surface and Cl− to form active

chlorine species (desirable and undesirable, such as Cl, Cl2, ClO2− and ClO3

−, ClO4−, respectively)

with this active material [16]. According to the electrode nominal composition, it suggests a mix-

behavior as active or non-active anode due to the presence of SnO2 in its surface. In fact, the

oxygen evolution reaction is achieved at more positive potentials than that observed at Ti/IrO2

anode. However, the performances of Ti-Pt-SnO2-Sb2O5 anode are not comparable with the

performance of an ideal non-active anode like diamond electrode [26]. It indicates that, the

concentration of halide in solution increases the importance of Cl2/oxy-chloro radicals system

depending on the electrocatalytic material and this behavior plays an important role in relation

with the oxygen evolution reaction, influencing on the efficiency of electrochemical approach

adopted [27-29].


doi: 10.5599/jese.2014.0069 263

Bulk electrolysis

The experiments were performed at 25 °C varying Cl- concentration ions (20 and 30 mM) in

solution and applied different current density (10, 20 and 30 mA cm-2) in order to evaluate the

elimination of organic load. Fig. 2 shows the performances of each one of the electrocatalytic

material used as a function of applied current density and electrolyte concentration during the

COD removal. It was observed that efficiency of COD removal was dependent on the applied

current density and Cl- concentration. In fact, when 10, 20 and 30 mA cm-2 were applied by using

Ti/IrO2, the initial COD (338 mg L-1) was reduced 262 mg L-1; 182 mg L-1 and 140 mg L-1 with 20 mM

of Cl- in solution, while at 30 mM of Cl-, COD decays to 193 mg L-1, 123 mg L-1 and 121 mg L-1 for 10,

20 and 30 mA cm-2, respectively, after 120 min of electrolysis. For Ti-Pt-SnO2-Sb2O5 anode, under

similar conditions, COD concentration was reduced from 338 mg L-1 to 223, 45 and 96 mg L-1 at

20 mM of Cl- and 158; 18 and 98 mg L-1 when 30 mM of Cl- was added by applying 10, 20 and

30 mA cm-2, respectively.

Based on the COD removal reported in Fig. 2, the increase in the applied current density con-

tributes to the degradation of phenol and the intermediates generated during the electrolysis,

thanks to the action of active chlorine species produced on the electrodes surface. However, the

results reveal that at Ti-Pt-SnO2-Sb2O5 there is greater reduction in COD than that achieved at

Ti/IrO2 (Fig. 2), principally at 20 mA cm-2. Conversely, when 30 mA cm-2 was applied, the elimina-

tion of COD decreased due to the promotion of Cl2 rather than the production of active chlorine

species.

Figure 2. Effect of NaCl for COD removal at different concentrations during mediated

electrochemical oxidation of phenol by using (a) Ti/IrO2 and (b) Ti-Pt-SnO2-Sb2O5 anodes. Operational conditions: [Phenol]0= 100 mg L−1, initial COD = 338 mg L-1, T = 25°C.

This behavior suggests that phenol oxidation depends on the nature of the anode material due

to the efficient production of active chlorine species on anode surface [23, 26-29]. Ti/IrO2 and

Ti-Pt-SnO2-Sb2O5 materials are classified as active anodes [5,6] because these electrocatalytical

materials are characterized by strong electrode-hydroxyl radical interaction, resulting in a low

chemical reactivity for organics oxidation. This problem can be avoided when Cl-mediated

approach is used [14]. Generally, under favorable pH conditions and NaCl in solution,

electrochemical oxidation via •OH radicals is not the only oxidation mechanism that occurs on the

DSA anodes [14]. In this case, chlorohydroxyl radicals are also generated on anode surface and

consequently oxidizing organic matter (Equations 4 and 5) [30,31]:


264 264

H2O + M + Cl- → M[•ClOH] + H+ + 2е- (4)

R + M[•ClOH] → M + RO + H+ + Cl- (5)

Reactions between water and radicals near to anode surface can yield molecular oxygen, free

chlorine, and hydrogen peroxide (6, 7 and 8) [32]:

H2O + M[•OH] → M + O2 +3H+ + 3е- (6)

H2O + M[•ClOH] + Cl-→ M + O2 +Cl2 + 3 H++ 4е- (7)

H2O + M[•OH] → M + H2O2 + H++ е- (8)

Furthermore, hypochlorite can be formed as follows (9 and 10) [31]:

Cl2(diss) + H2O → HOCl + H+ + Cl- (Acidic medium) (9)

Cl2(diss) + 2OH- → ClO- + Cl- + H2O (Alkaline medium) (10)

Therefore, indirect oxidation results in reduction of organic pollutants such as phenol thanks to

the participation of active chlorine species electrochemically formed [15,23,31]. Oxidants are quite

stable and migrate in the solution bulk, and then, these indirectly oxidize the effluent, favored by

hydrodynamic configuration of electrochemical cell. The efficiency of indirect oxidation depends

on the diffusion rate of oxidants in the solution, concentration of oxidants, and pH of solution [15].

For the electrochemical flow cell used in this study, the mass transfer coefficient was

2.0×10−5 m s−1, and the limiting current (for both anodes) results in an average value of 0.88 A,

according to Eq. 2. This current is lower than all the currents applied in this work (1.27–1.90 A),

suggesting that the oxidation under these experimental conditions could occur under mass

transport control since 120 min of electrochemical treatment. These assumptions are in

agreement with the studies published by Cañizares and co-workers [33].

UV spectroscopic characteristics of the electrochemical oxidation of phenol

The oxidation of phenol at Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrodes was also monitored by

spectrophotometric measurements, which allow a straightforward way to follow the elimination

of phenol. Thus, UV spectra for phenol degradation at Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrodes, at

20 mM and 30 mM of Cl- in solution at 25°C by applying 20 mA cm−2, are shown in Figure 3. An

inspection of these UV spectra allowed confirming that phenol was more rapidly removed at

Ti/IrO2 (Fig. 3a and 3b, at 20 mM and 30 mM of Cl- in solution by applying 20 mA cm−2) than at Ti-

Pt-SnO2-Sb2O5 electrodes under the same experimental conditions (Fig. 3c and 3d). Moreover, the

bands in Fig 3a and 3b are completely different after 5 minutes of electrolysis due to formation of

reaction intermediates, confirming the fast degradation of phenol. These major changes in UV- vis

bands are not observed in Figs. 3c and 3d. This behavior depends on the effective electrochemical

production of active chlorine species at both anodes. Perhaps, at IrO2, metal cations in the oxide

lattice may reach higher oxidation states under anodic polarization stabilizing •OH radicals and Cl-

ions on its surface [6,32,34], which favors the O2 and Cl2 evolution at the expense of the electro-

chemical incineration reaction. This feature avoids the formation of enough active chlorine species

that promote an efficient degradation of phenol, generating different aromatic intermediaates.

However, it is not possible to reliably ensure that phenol is degraded because the products and/or

intermediates formed during the process may have a higher molar absorptivity () and absorb in

the same wavelength region of phenol.


doi: 10.5599/jese.2014.0069 265

Figure 3. Spectrophotometric measurements, as a function of time, during Cl-mediated electrochemical treatment of phenol (100 mg L-1) by applying 20 mA cm-2 at 25°C. Ti/IrO2: (a) 20 mM of Cl- and (b) 30 mM of Cl-. Ti-Pt-SnO2-Sb2O5: (c) 20 mM of Cl- and (d) 30 mM of Cl-.

Conversely, at Ti-Pt-SnO2-Sb2O5 electrode (Fig. 3c and 3d), the elimination of phenol is attained

by the reactive chlorine species such as chlorine and hypochlorous acid or hypochlorite ion (Cl2,

HClO and OCl−) that react rapidly with organics mainly by the reactions in solution [23,35,36]. In

acidic conditions, free chlorine is the dominant oxidizing agent, while in slightly alkaline conditions

hypochlorite, chloride ions and hydroxyl radicals are all generated in relevant concentrations

[26,31]. In fact, pH typically varies between 5.5 and 6.2 throughout the course of the reaction for

phenol. Then, this pH behavior suggests the participation of active chlorine oxidants, confirming

the increase on COD removal rate when Ti-Pt-SnO2-Sb2O5 electrode was used.

Distribution of by-products of phenol oxidation

As stated above, phenol could be transformed into different intermediates or carbon dioxide

during the process and such differences would not be apparent from the time- or charge-course of

this parameter. For this reason, identification of some intermediates was performed by HPLC. The

change in concentration of phenol (initial concentration of 100 mg L-1) and intermediates in the

course of indirect electrochemical treatment, by applying 20 mA cm-2 at 25°C, is shown in Fig. 4a-

d. As can be observed, the degradation of phenol (main by-product formed) is influenced by the

amount of Cl- as well as the electrocatalytic material used. In fact, 84 % and 88 % of phenol was

removed by using Ti/IrO2 anode after 120 min of electrolysis when 20 and 30 mM of Cl- in solution

were used, respectively. Conversely, the decrease of phenol on Ti-Pt-SnO2-Sb2O5 was about 93 %

and 83 % by applying 20 mA cm-2 for 20 and 30 mM of Cl-, respectively. As shown in Fig. 4, the

intermediates concentration produced on Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrodes are different.


266 266

Hydroquinone, which is aromatic intermediate, was principally found in high concentration at the

experiments performed with Ti/IrO2 anode. HPLC results indicated the transformation of the

phenol by electrolysis to hydroquinone, which was decomposed to other forms. This behavior was

observed at both anodes where similar by-products were formed. However, at Ti-Pt-SnO2-Sb2O5,

all the intermediates are quasi-completely mineralized to CO2 and H2O due to the attack of active

chlorine species.

Figure 4. Intermediates formed during Cl-mediated oxidation of phenol (100 mg L-1) by

applying 20 mA cm-2 using Ti/IrO2 ((a) 20 mM of Cl- and (b) 30 mM of Cl-) and Ti-Pt-SnO2-Sb2O5 ((c) 20 mM of Cl- and (d) 30 mM of Cl-) anodes. HPLC retention times: benzoquinone: 1.9 min,

catechol: 3.05 min; phenol: 3.5 min; hydroquinone: 2.76 min and oxalic acid: 2.3 min.

Moreover, it is found that the amount of oxalic acid produced on Ti-Pt-SnO2-Sb2O5 is larger than

that at Ti/IrO2, suggesting a better grade of mineralization. It is important to indicate that,

unidentified intermediates were produced; however, these by-products were predominantly

generated at Ti/IrO2 as showed by their chromatographic area (Fig. 5). For Ti/IrO2, aromatic

compounds concentration is not reduced with the action of active chlorine, finding difficult to

break the aromatic ring and favoring the conversion of phenol to other aromatic compounds.

Based on the results obtained, the effect of NaCl on Ti-Pt-SnO2-Sb2O5 is much more evident and

efficient to produce strong oxidants that promote the mineralization of phenol and by-products on

the bulk of solution.


doi: 10.5599/jese.2014.0069 267

Figure 5. Chromatographic areas of unidentified intermediates produced during indirect

electrochemical oxidation of phenol at Ti/IrO2 (20 mM of Cl- and 30 mM of Cl-) and Ti-Pt-SnO2-Sb2O5 (20 mM of Cl- and 30 mM of Cl-) anodes.

Kinetic and activation energy analysis

To study the kinetics of the overall reaction involved in the disappearance of phenol and

intermediates by indirect electrochemical oxidation, the decay of phenol concentration under

different NaCl concentration of electrolyte was considered. Results given in Fig. 6a and 6b were

further analyzed using kinetic equations related to different reaction orders. Good linear plots

were fitted to a pseudo-first-order reaction (ln(C0/Ct) vs. time) for Ti/IrO2 and Ti-Pt-SnO2-Sb2O5

electrodes. For Ti/IrO2, apparent constant rates (kapp) of 0.020 min-1 (r2 = 0.97) for 20 mM of

Cl- and 0.015 min-1 (r2 = 0.95) for 30 mM of Cl- were estimated; while for Ti-Pt-SnO2-Sb2O5,

0.027 min-1 (r2 = 0.97) for 20 mM of Cl- and 0.017 min-1 (r2 = 0.98) for 30 mM of Cl- were achieved.

These figures confirm that the indirect electrochemical oxidation is faster at Ti-Pt-SnO2-Sb2O5

anode, respect to the behavior attained at Ti/IrO2, where the process is disfavored when an incre-

ase in Cl- ion is performed. It can be due to the preferential production of Cl2 gas, as discussed on

polarization curves results [26]. In this form, lower concentrations of active chlorine are produced,

favoring the electrochemical conversion of phenol to aromatic compounds. Considering the kinetic

model pseudo-second order ((1/C0)-(1/Ct) vs. time) and concentration Cl- ions in the medium,

these show suitable linearity when compared to pseudo-first order model.

Figure 6. Kinetic analysis for the pseudo-first-order In(C0/Ct) and pseudo-second-order (1/C0)-(1/Ct)) reaction

for phenol decay during its indirect electrochemical oxidation by active chlorine species.


268 268

The coefficients of pseudo-second order were about 7.0110-4 min-1 (r2 = 0.99) and

2.510-4 min-1 (r2 = 0.99) for concentration 20 and 30 mM of Cl-, respectively, using

Ti-Pt-SnO2-Sb2O5. Conversely, values of 4.510-4 min-1 (r2 = 0.97) and 2.410-3 min-1 (r2 = 0.98) at

concentration 20 and 30 mM of Cl- where estimated at Ti/IrO2 electrode. Then, the kinetic data

obtained by pseudo-second order model indicate that the indirect oxidation is controlled by the

rate at which organic molecules are carried from the bulk to the electrode surface, but when the

concentration of final intermediates increase, their rate is limited by diffusion control. However,

the principal step for the electrochemical approach is the production of active chlorine species, as

indicated by pseudo-second order model.

Finally, it is very important to estimate the treatment applicability, and thus, Table 1 reports the

current efficiency and energy consumption (kWh dm−3) at 10, 20 and 30 mA cm−2, after 120 min of

electrochemical treatment. As can be observed, Ti-Pt-SnO2-Sb2O5 consumed relatively less

electrical energy than Ti/IrO2 anode, but different current efficiencies were achieved after 120 min

of electrolysis at both anode materials.

Table 1. Energy requirements and total current efficiency for elimination of phenol at different applied current densities by Cl-mediated electrochemical oxidation.

Current density, mA cm-2 Current total efficiency, % Energy consumption, kWh dm-3

Ti/IrO2 Ti-Pt-SnO2-SbO5 Ti/IrO2 Ti-Pt-SnO2-Sb2O5

10 59 72 36 28

20 39 56 42 40

30 26 49 69 54

Conclusions

The Cl-mediated electrochemical oxidation of phenol was investigated under galvanostatic

conditions at Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrodes, as a function of applied current density and

amount of NaCl dissolved. A partial elimination of the organic pollutant was achieved at Ti/IrO2,

while a quasi-complete electrochemical elimination takes place at Ti-Pt-SnO2-Sb2O5 anode. The in-

fluence of the anode material on the elimination of phenol seems to be very important due to an

efficient production of active chlorine species during electrolysis. In fact, Ti-Pt-SnO2-Sb2O5 showed

good electrocatalytic activity to promote the electrochemical generation of active chlorine, as

indicated by potentiodynamic measurements, and as a consequence of this characteristic,

significant removal efficiency of phenol and its by-products was attained.

The effect of chloride on the electrooxidation of organics with Ti/IrO2 and Ti-Pt-SnO2-Sb2O5

depends mainly on the reaction between electrogenerated •OH and Cl− ions or the conversion of

chloride ion to chlorine which is further hydrolyzed to other active species [26,37]. At the same

time, the production of Cl- to active chlorine is directly related to different experimental con-

ditions, but it is principally dependent on the concentration of free •OH radicals, the increase of

the importance of the Cl2/H2O system and the interaction of Cl- and •OH radicals with the anode

surface. Because, increasing the concentration of Cl- ions in solution, elimination of Cl- from

solution by Cl2 is favored [23,37,38].

Acknowledgements: E. V. dos S. gratefully acknowledges the Programa de Recursos Humanos – PETROBRAS (PBFRH-22) for her PhD fellowship and support provided by the Núcleo de Processamento Primário e Reuso de Água Produzida e Resíduos (NUPPRAR-UFRN) for analyzing the


doi: 10.5599/jese.2014.0069 269

samples electrochemically treated. The authors thank the financial support provided by CNPq and PETROBRAS. They also thank to the Dott. Christian Urgeghe (Industrie De Nora S.p.A. - Milan, Italy) for providing Ti/IrO2 and Ti-Pt-SnO2-Sb2O5 electrodes.

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doi: 10.5599/jese. 2014.0065 271



www.jESE-online.org


Influence of operating parameters on electrocoagulation of C.I. disperse yellow 3

Djamel Ghernaout*,**,, Abdulaziz Ibraheem Al-Ghonamy**, Mohamed Wahib Naceur*, Noureddine Ait Messaoudene** and Mohamed Aichouni**

*Department of Chemical Engineering, University of Blida, PO Box 270, Blida 09000, Algeria

**Binladin Research Chair on Quality and Productivity Improvement in the Construction Industry; College of Engineering, University of Hail, PO Box 2440, Ha’il 81441, Saudi Arabia

Corresponding Author: E-mail: [email protected] Tel.: +213-025-433-631; Fax: +213-025-433-631

Received: July 24, 2014; Revised: August 14, 2014; Published: December 6, 2014

Abstract This work deals with the electrocoagulation (EC) process for an organic dye removal. The chosen organic dye is C.I. disperse yellow 3 (DY) which is used in textile industry. Experiments were performed in batch mode using Al electrodes and for comparison purposes Fe electrodes. The experimental set-up was composed of 1 L beaker, two identical electrodes which are separated 2 cm from each other. The main operating parameters influencing EC process were examined such as pH, supporting electrolyte concentration CNaCl, current density i, and DY concentration. High performance EC process was shown during 45 min for 200 mg/L dye concentration at i = 350 A m-2 (applied voltage 12 V) and CNaCl = 1 g L-1 reaching 98 % for pHs 3 and 10 and 99 % for pH 6. After 10 min, DY was also efficiently removed (86 %) showing that EC process may be conveniently applied for textile industry wastewater treatment. EC using Fe electrodes exhibited slightly lower performance comparing EC using Al electrodes.

Keywords: Dye removal; aluminium; iron; current density; mechanism.

Introduction

The main problem of access to safe drinking water is continuous pollution of water resources by

agriculture, urban waste and industry. In countries where water resources are relatively limited,

treated wastewater reuse in agriculture has become an urgent necessity. The textile industry

consumes considerable amounts of water in the dyeing and finishing. Effluents containing dyes



J. Electrochem. Sci. Eng. 4(4) (2014) 271-283 ELECTROCOAGULATION OF C.I. DISPERSE YELLOW 3

272 272

can be toxic to the environment [1-4]. In addition, their presence in aquatic systems, even at low

concentrations, is very visible. It reduces the penetration of light and has a detrimental effect on

photosynthesis [5-7].

Therefore, the remediation of water contaminated by these chemicals is necessary both to

protect the environment and for future reuse [8-12]. Therefore, several biological, physical and

chemical methods are used for the treatment of industrial effluents with different efficeincies [13-

15]. Electrochemical technologies, such as electrocoagulation technique (EC), seem to be well

adapted to the textile industry wastewaters treatment [16-22].

This work is devoted to the study of the EC process for bleaching synthetic water containing an

azo dye, C.I. disperse yellow 3 (DY), used in the Algerian textile industry and the assessment of its

performance versus certain operating parameters.

Experimental

Experimental set-up

The EC tests were performed using an experimental set-up shown in Figure 1. In a 1000 mL

beaker, filled with 500 mL synthetic dye solution (distilled water + DY + NaCl), two Al (or Fe in

some experiments) 4 × 20 cm electrodes were immersed (active surface S = 4 × 10.5 = 42 cm2). The

anode is connected to the positive pole and the cathode to the negative pole of the direct current

power supply. The interelectrode distance is fixed at 2 cm. When the electric current is applied,

the magnetic stirrer is started at an average velocity agitation.

Figure 1. Photo of the EC experimental set-up.

Electrodes cleaning

Before experiments, the Fe electrodes were prepared to avoid the presence of any impurity as

follows: (1) polishing with abrasive paper; (2) rinsing with distilled water; (3) degreasing by means

of a solution composed of: NaOH (25 g), Na2CO3 (25 g), K2CO3 (25 g) and distilled water

(q.s.p. 1,000 mL); (4) rinsing with distilled water; (5) pickling in a solution of sulphuric acid H2SO4 at

20 % for 20 min at room temperature; and again (6) rinsing with distilled water. For Al electrodes:

(1) rinse with distilled water and polish using abrasive paper, (2) clean in hydrochloric acid solution

(HCl at 20 %) during 10 min, and (4) rinse with distilled water.

D. Ghernaout et al. J. Electrochem. Sci. Eng. 4(4) (2014) 271-283

doi: 10.5599/jese.2014.0065 273

Prepared solutions

To prepare a solution of 200 mg L-1 dye, 0.2 g of the latter was poured into a 1 L flask and

distilled water was added during stirring for better solubilisation. The initial pH was varied using a

solution of 0.1 M HCl (acidic conditions) or NaOH (alkaline medium). The solution conductivity was

increased by sodium chloride addition. All chemicals used were of analytical grade.

Methods

Once the EC test ends, the treated solutions were left to settle for 30 min in order to sediment

the flocs formed. After decantation, and using a pipette, 25 mL of the solution were carefully

collected for analysis.

The analyses done before and after treatment were as follows: pH, conductivity and ultraviolet

(UV) absorbance (Shimadzu 1601, dual beam with 1 cm quartz vessel). The best UV absorbance

long wave was found at 346 nm (UV346). The DY removal was calculated using the relation (1):

i f

f

/ % 100Ab Ab

RAb

(1)

where Abi and Abf were initial and final UV absorbances, respectively. All the tests were conducted

at ambient temperature (20 °C).


The aim of this work was to perform bleaching EC tests on dye synthetic solutions (distilled

water+dye+NaCl) using EC process and evaluate its performance based on certain key parameters.

Influencing parameters on EC process

Common remarks

During EC tests, some common observations were:

- Aluminium dissolution according to Reaction (2):

Al → Al3+ + e- (2)

- Production of H2 gas bubbles at the cathode according to Reaction (3):

2H2O + 2e- → 2OH- + H2(g) (3)

- Production of O2 gas bubbles at the anode according to Reaction (4):

2H2O → 4H+ + O2 + 4e- (4)

- Flocs formation and their fixation on the H2(g) bubbles during their ascension to the solution

surface as a white foam (Figure 2a and b). Indeed, anode dissolution generates coagulant

species which destabilise the dye molecules forming flocs.

Figure 2. Foam formation: (a) face view, (b) top view. (c) Initial and final state of a 200 mg L-1 DY solution

treated by EC during 1 h: from wright to left, initial solution, treated solution at 12, 8 and 4 V, respectively.


274 274

EC time

The EC efficiency is strongly influenced by the time residence in the electrochemical device. To

study its effect, the EC period was varied from 5 to 75 min and the other parameters were kept

constant. The results obtained are illustrated in Figure 3. The H2 and O2 release and flocs

formation increased over time and the foam became thicker. From 30 min, the flocs settled and

the solution became clearer.

As seen in Figure 3, the dye removal efficiency increased with electrolysis time until 45 min.

After this time, EC performance decreased. Moreover, the good EC efficiencies were reached

between 15 and 45 min.

The removal efficiency was directly dependent upon the metal concentration in solution [23-

25]. The positive metallic species were produced by the Al anode neutralising the negative charges

on the polluting molecules [26-30]. When the electrolysis duration was increased, the cationic

species as well as metal hydroxide (Al(OH)3(s)) concentrations increased [30-32]. Consequently, the

pollutant removal increased [33-36].

Figure 3. Dye removal as a function of EC time (pH 6.5; CNaCl = 1 g L-1; CDY = 20 mg L-1; d = 2 cm).

Electric current density

The electric current density is the most important parameter of the electrochemical

process [37]. The electric current density effect on the dye removal was studied. The current

intensities were 120, 250 and 350 mA corresponding to the applied voltages of 4, 8 and 12 V,

respectively. The dye concentration was fixed at 20 and 200 mg/L. The other parameters were

maintained constant (pH 6.5; CNaCl = 1 g/L; d = 2 cm). The obtained results are shown in Figure 4.

For I = 120 mA (i = 29 A/m2), the produced gas bubbles were small and the formed froth was

thin. For I = 250 mA (i = 60 A/m2), the gas emanation was medium and the formed froth became

important. For I = 350 mA (i = 83 A/m2), flocs settling became significant, the gas emanation

became intense and the solution was transformed clear.

As seen in Figure 4, the electric current had a great effect on the dye removal especially for the

first ten minutes. After 20 min, the electric current had a small effect. This is explained by the fact

that the negative charge on the organic dye is neutralised after the Al3+ action on the dye

molecules.


doi: 10.5599/jese.2014.0065 275

Figure 4. Effect of the electric current i on the EC efficiency for DY removal (pH 6.5, d = 2 cm, CNaCl = 1 g L-1)

(a): CDY = 20 mg L-1; (b): CDY = 200 mg L-1; (c): CDY = 200 mg.L-1 ; tEC = 5 min

Initial pH

The solution pH is an important factor influencing the EC performance [37]. This is due to the

fact that pH determinates the metallic ions speciation, the chemical state of other species in the


276 276

solution, and the formed products solubility. In order to examine the pH effect, the solution pH

was adjusted to the values from 3 to 12 maintaining other parameters constant: U = 12 V

(I = 83 A m-2), CNaCl = 1 g L-1, d = 2 cm, tEC = 30 min). The obtained results are shown in Figure 5.

Figure 5. Effect of the pH on the EC efficiency for DY removal

(CDY = 200 mg L-1, U = 12 V (i = 83 A m-2); tEC = 30 min; d = 2 cm).

For the acidic medium, the solution colour became orange after the addition of HCl. From the

cathode, there is an intense emanation of H2 bubbles. From the anode, there is an important

formation of O2 bubbles. The foam and sediment formed are denser at the anode.

For the alkaline medium, there was formation of white sediment at the bottom of the beaker,

and its volume increased with time in comparison with the acidic medium.

As seen in Figure 5, we noted that the dye removal was well performed between pH 3 and 10.

Several researchers found that the best removal efficiency with aluminum electrodes was reached

in the pH range between 3 and 9 [19,34-36].

In Figure 6, we chose four pH values: 3, 4, 6.5 and 10 with other parameters fixed in order to

illustrate the pH effect.

We also followed the change in pH as a function of time; Figure 7 shows the obtained results.

The medium pH changed during the EC process. This change depended on the type of electrode

material and the initial pH of the treated solution.

We note from Figure 7 an increase in pH in the case of solutions with pH < 7. This increase was

probably due to the release of H2 from the cathode and the formation of OH- according to the

Reaction (3). Moreover, a decrease in pH in the case of solutions with pH > 7 was also noticed. This

decrease was affected to the hydroxyl (OH-) consumption according to Reaction (5):

Al(OH)3(aq) + OH-(aq) → Al(OH)4-(aq) (5)

Initial conductivity

Conductivity promotes the performance of electrochemical processes [26]. We chose NaCl as a

supporting electrolyte. In order to determine its effect on the efficiency of bleaching of the

synthetic solutions, we varied the concentration (0.25, 0.5, 1 and 1.5 g L-1) while keeping the other

parameters constant. The results are shown in Figure 8.

We have observed that (1) the gas production becomes higher with the increase in salt

concentration, (2) the conductivity decreases during EC treatment and, (3) the formation of a small

deposit on the anode.


doi: 10.5599/jese.2014.0065 277

Figure 6. DY removal as a function of pH (U = 12 V (i = 83 A m-2), tEC = 30 min; CNaCl = 1 g L-1; d = 2 cm

(a) CDY = 20 mg L-1; (b) CDY = 40 mg L-1; (c) CDY = 200 mg L-1


278 278

Figure 9 shows the solution conductivity as a function of time during EC process. We note that

the conductivity decreased over time and the difference in the changes in pH was different due to

the HCl and NaOH added during the pH adjustment before EC treatment.

Figure 7. Evolution of pH during EC treatment (same conditions as for Figure 6).

(a) CDY = 20 mg L-1; (b) CDY = 40 mg L-1; (c) CDY = 200 mgL-1

Figure 8. Effect of the NaCl concentration on the EC efficiency U = 12 V (i = 83 A m-2); tEC = 30 min; d = 2 cm; CDY = 200 mg L-1

0

2

4

6

8

10

12

0 20 40 60 80

pH

tEC / min

(a)

0

2

4

6

8

10

12

0 20 40 60 80p

H

tEC / min

(b)

0

2

4

6

8

10

12

0 20 40 60 80

pH

tEC / min

(c)

0

20

40

60

80

100

120

0 0.25 0.5 0.75 1 1.25 1.5

R /

%

CNaCl / g L-1


doi: 10.5599/jese.2014.0065 279

Figure 9. Solution conductivity as a function of time during EC process

(CNaCl = 1 g L-1; CDY = 200 mg L-1; d = 2 cm; U = 12 V (i = 83 A-2)

Inter-electrode distance

We varied the distance between the electrodes d = 0.8; 1; 1.5 and 2 cm while fixing the other

factors. The results are shown in Figure 10. When increasing the inter-electrode distance, the EC

efficiency also increased. This can be explained as follows: for d = 2 cm, there would be more

probabilities to generate global flocs that are able to adsorb more dye molecules.

Figure 10. Effect of the inter-electrode distance on the EC performance

(CNaCl = 1 g L-1, CDY = 200 mg L-1, pH = 6.5, U = 12 V (i = 83 A m-2).

DY concentration

The aim of this part is to determine whether the EC method was applicable to solutions with a

range of concentrations from 20 to 500 mg L-1. The solutions were electrolysed, for a treatment

time of 45 min, at a fixed voltage U = 12 V (i = 83 A m-2) and an inter-electrode distance of 2 cm.

The results obtained are shown in Fig. 11.

Figure 11 shows that the EC method is effective in the range of selected concentrations. A yield

of 97 % is reached at a dye concentration of 200 mg L-1. The results obtained can be justified by


280 280

the increased probability of contact with the dye molecules to aluminum hydroxide Al(OH)3 to

form flocs of large sizes; thereby facilitating their separation by their attachment to the bubbles of

released gases at the electrodes.

Figure 11. EC performance as a function of DY concentration

CNaCl = 1 g L-1; pH 6.5; d = 2 cm, U = 12 V (i = 83 Am-2)

EC using Fe electrodes

To check if DY can be removed by EC using iron electrodes, some tests using the Al optimum

conditions are performed. The results obtained are compared with those obtained with aluminum

electrodes (Figure 12).

Figure 12. EC using Al and Fe electrodes: CDY = 200 mg L-1; CNaCl = 1 g L-1; pH 6.5; d = 2 cm, U = 12 V (i = 83 A/m2)

During EC treatment using Fe electrodes, (1) clouds of green flocs came from the anode surface

and settled to the beaker bottom and, (2) on the solution surface, two layers of foam were

observed: the first a red-brown color, and below, the second green.

We find that the rate of reduction of the dye increased with time until a yield of 96.28 % in the

case of iron electrodes (Figure 12). Comparing the test results with aluminum electrodes

(R = 98.96 %), we can say that iron is also effective in removing DY.

91

92

93

94

95

96

97

98

0 100 200 300 400 500 600

R /

%

CDY / mg L-1


doi: 10.5599/jese.2014.0065 281

The reactions involved in the EC using iron electrodes are as follows: The iron, after oxidation in

the electrolytic system, produces iron hydroxide Fe(OH)n(s), with n = 2 or 3; and two mechanisms

have been proposed [6,12,14,15,24,38]:

Mechanism 1 (green coloration, Fe(OH)2(s)):

Fe(s) → Fe2+(aq) + 2e- (6)

Fe2+(aq) + 2OH-

(aq) → Fe(OH)2(s) (7)

Mechanism 2 (brown coloration, Fe(OH)3(s)):

Fe2+(aq) → Fe3+

(aq) + 1e- (8)

Fe3+(aq) + 3OH-

(aq) → Fe(OH)3(s) (9)

The species Fe(OH)n(s) formed (by the two mechanisms) remained in the aqueous phase in the

form of gelatinous suspension which can then remove the pollutants from the water (Figure 13),

either by complexation, or by electrostatic attraction, followed by coagulation and flotation or

sedimentation [24,25,27,28].

Figure 13. Predominance-zone diagrams for (a) Fe(II) and (b) Fe(III) chemical species in aqueous solution. The straight lines represent the solubility equilibrium for insoluble Fe(OH)2 and Fe(OH)3, respectively, and the dotted lines represent the predominance limits between soluble chemical species. (c) Diagram of solubility of Al(III) species as a function of pH [27,28].


282 282

Conclusions

Highly performant EC process is shown in dye removal during 45 min for 200 mg/L dye

concentration at i = 350 A/m2 (applied voltage 12 V) and CNaCl = 1 g/L reaching 98 % for pH 3 and

10 and 99 % for pH 6. For 10 min, DY is also efficiently removed (86 %) showing that EC process

may be well convenient for textile industry wastewater treatment. EC using Fe electrodes is

slightly less performant than EC using Al electrodes.

Acknowledgements: The present research work was undertaken by the Binladin Research Chair on Quality and Productivity Improvement in the Construction Industry funded by the Saudi Binladin Constructions Group; this is gratefully acknowledged. The opinions and conclusions presented in this paper are those of the authors and do not necessarily reflect the views of the sponsoring organisation.

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doi: 10.5599/jese.2014.0060 285



www.jESE-online.org

Review

Enhanced electrocoagulation: New approaches to improve the electrochemical process

Carlos E. Barrera-Díaz, Gabriela Roa-Morales, Patricia Balderas Hernández, Carmen María Fernandez-Marchante* and Manuel Andrés Rodrigo*

Centro Conjunto de Investigación en Química Sustentable UAEM – UNAM, Carretera Toluca-Atlacomulco, km 14.5, Unidad El Rosedal, C.P. 50200, Toluca, Estado de México, México *Departartment of Chemical Engineering. Faculty of Chemical Sciences & Technologies, Enrique Costa Building, Campus Universitario s/n 13071 Ciudad Real, Spain

Corresponding Author: E-mail: [email protected]

Received: August 14, 2014; Published: December 6, 2014

Abstract Electrocoagulation is a promising technology for the removal of colloids from different types of wastewater and it has also demonstrated good efficiencies for the breaking-up of emulsions. It consists of the dissolution of aluminum or iron anodes, promoting the formation of coagulant reagents in wastewater that helps to coagulate pollutants and the formation of bubbles that favors the mixing (electroflocculation) and the removal of suspended solids by flotation (electroflotation). During the recent years, the combination of this technology with other treatment technologies has become a hot topic looking for a synergistic improvement in the efficiencies. This work aims to review some of the more recent works regarding this topic, in particular the combination of electrocoagulation with ozonation, adsorption and ultrasound irradiation.

Keywords Electrocoagulation; ozonation; adsorption; ultrasound irradiation; pulse application

Introduction

Electrochemical treatment techniques have attracted a great deal of attention because of their

versatility and environmental compatibility. Electrochemical reactions take place at the anode and

the cathode of an electrolytic cell when an external direct current voltage is applied. In fact, the

main reagent is the electron, which is a “clean reagent”[1,2] and this fact helps to explain the

lower production of wastes associated to these technologies. Applications studied in the recent

years range from the oxidation of organic pollutants contained in wastewater to the

electroremediation of soils.



J. Electrochem. Sci. Eng. 4(4) (2014) 285-296 ENHANCED ELECTROCOAGULATION

286 286

Electrochemical methods have also been used as coagulation processes to remove color and

cloudiness from turbid industrial wastewater. In this application, the electrochemical process

generated numerous flocculates, achieving high efficiency in clearing the wastewater[3,4]. The

term electrocoagulation involves the in situ generation of coagulants by electrolytic oxidation of

an appropriate sacrificial anode (iron or aluminum), which causes the dissolution of electrode

plates into the effluent. Metal ions, at an appropriate pH, can form wide range of coagulated

species and metal hydroxides that destabilize and aggregate particles or precipitate and adsorb

the dissolved contaminants. Main stages involved in the electrocoagulation process using

aluminum anodes have been previously identified [5,6]. The anodic process involves the oxidative

dissolution of aluminum into aqueous solution as reaction (1) indicates as well as the oxidative

dissociation of water as reaction (2) shows.

Al → Al3+ + 3e− (1)

2H2O → O2(g) + 4H+ + 4e− (2)

In the case of iron or steel anodes, it is not iron (III) but iron (II) the main product of the

electrochemical process (Eq.3) [7]. Then, oxygen is known to be involved for further Fe2+ oxidation

into Fe3+ (Eq. 4)

Fe(s) → Fe2+ + 2e− (3)

4 Fe2+ + 4H+ + O2(g) → 4Fe3+ + 2H2O (4)

Once dissolved iron and aluminum, can participate in many chemical reactions (Eqs. 5-10). In

fact, speciation of iron and aluminum during electrocoagulation is very complex [8,9] and the

description of the interactions between pollutants and the coagulant species is one of the most

relevant topics nowadays in this field [10-13].

M(OH)4- + H+

M(OH)3 + H2O (5)

M(OH)3 + H+ M(OH)2+ + H2O (6)

M(OH)2+ + H+ M(OH)+2 + H2O (7)

M(OH)+2 + H+ M+3 + H2O (8)

M(OH)3(s) M3+ + 3OH- (9)

It is interesting that in electrocoagulation papers little attention has been paid on cathodic

reactions. Regardless of whether iron or aluminum is used, the main reaction that is reported is

water reduction (Eq. 10).

2H2O + 2e− → H2(g) + 2OH−(aq) (10)

However, this reaction has three important implications on the electrocoagulation technology:

a. provides hydroxyl ions which then react in bulk solution with iron or aluminum cations to

form insoluble species and other coagulants (Eqs. 5 to 9);

b. hydrogen gas is produced increasing turbulence. This process contributes in the

destabilization of colloidal particles leading to flocculation (so-called electro-flocculation

process), and

c. contribution to electroflocculation which is a simple process that floats pollutants (or other

substances) by their adhesion onto tiny the bubbles formed by the hydrogen evolution [14]

(so-called electroflotation process)

C. E. Barrera-Díaz et al. J. Electrochem. Sci. Eng. 4(4) (2014) 285-296

doi: 10.5599/jese.2014.0060 287

As a consequence of this complex interaction, the electrochemical cell combines several

processes at the same time in the same reactor and this becomes a significant advantage of this

type of processes as compared with conventional coagulation treatments. In particular, from the

economical point of view they compare favorably with coagulation processes [15-17] in many

applications.

As for coagulation processes, electrocoagulation highly depends on the wastewater pH and it

becomes a critical parameter in the performance of this technology. This parameter determines

the speciation of aluminum and iron and hence the primary coagulation mechanisms occurring in

the electrocoagulation cell. In fact, pH is one of the key differences between coagulation and

electrocoagulation as conventional coagulation acidifies the treated wastewater due to the acidic

properties of the typical coagulants dosed (iron chloride, aluminum sulfate, etc.), which are known

to behave as Lewis acid. These properties make necessary the neutralization of wastewater after

the coagulation treatment and this process implies an undesired increase in the salinity of the

effluent. On the other hand, electrocoagulation typically buffers the pH during the treatment in

values within the range 8-9, which should be a proper value even for direct discharge and no

further neutralization is required [18].

Recent studies shows many promising applications of electrocoagulation in the treatment of

lowland surface water [19], water [6,20,21], metal plating wastes [22], other types of industrial

wastewater [5,23-30], urban wastewater [31-33] and even in disinfection [34,35]. In fact, it is one

of the most promising environmental technologies based on electrochemical engineering [36,37].

Electrochemical methods offer two main advantages over traditional chemical treatment: less

coagulant ion is required and less sludge is formed [19,22,31]. In the recent years, the potential of

this technology is tried to be even further increased by the synergistic combination with other

treatment technologies. The objective of the present manuscript is to review the potential of

electrocoagulation for the treatment of industrial effluents coupling it with four types of

processes: Electrocoagulation-ozone processes

Electrocoagulation- adsorption processes

Electrocoagulation-ultrasound processes

Electrocoagulation-pulses processes

2. Electrocoagulation-ozone processes

Ozonation implies the use of ozone in the treatment of wastewater. Ozone is a strong oxidant

that oxidizes organic pollutants via two pathways: direct oxidation with ozone molecules and/or

the generation of free-radical intermediates, such as the •OH radical, which is a powerful,

effective, and non-selective oxidizing agent [38]. The ozonation process has the advantage of

being able to be applied when the flow rate and/or composition of the effluents are fluctuating.

However, the high cost of equipment and maintenance, as well as energy required to supply the

process, constitutes some of the disadvantages. Moreover, ozonation process requires the

transfer of ozone molecules from gas phase to liquid phase, where the attack on the organic

molecules occurs. Therefore, mass transfer limitations are also a relevant factor to be considered

in the oxidation process involving ozone. In many cases, the ozone consumption rate per unit of

volume can be so high that mass transfer is the limiting step, reducing the process efficiency and

increasing the operating costs [39]. In addition, the ozonation performance is affected by the

presence of organic matter, suspended solids, carbonate, bicarbonate and chlorine ions and also


288 288

by pH and temperature [40]. Some studies using real industrial wastewater have pointed out that

ozone by itself does not achieve high levels of pollutant removal [41]. In particular, the oxidation

of wastewater from molasses fermentation with ozone results in an effective color removal but is

less effective in removing organic matter [42]. Similar results were obtained when ozone was used

to treat textile wastewater, where ozone treatment proves to be very effective for complete color

removal but provides only partial reduction of the chemical oxygen demand (COD) [43]. Also,

previous research on ozone-coupled methods indicates that the ozonation of anaerobically

pretreated wastes enhances significantly the organic removal in comparison to the ozonation of

unpretreated wastes, and substrate conversions in the range of 40–67 % are obtained [44].

This behavior in the reduction of COD can be ascribed to the initial pH value of wastewater,

where the decomposition of ozone in water to form hydroxyl radicals occurs through the following

mechanism [45], where hydroxide ions initiate the reaction (Eqs. 11-16):

O3 + OH− → O2 + HO2− (11)

O3 + HO2− → HO2

. + O3.− (12)

HO2. → H+ + O2

.− (13)

O2.− + O3 → O2 + O3

.− (14)

O3.− + H+ → HO3

. (15)

HO3. → OH. + O2 (16)

According to reactions (11) and (12) the initiation of ozone decomposition can be artificially

accelerated by increasing the pH value. Side reaction (Eq. 17) is a fast process and plays an

important role in waters with low dissolved organic carbon and alkalinity [46] since it can reduce

the oxidative capacity of the system:

OH. + O3 → HO2. + O2 (17)

Regarding the combined process, Table 1 summarizes the main papers found in the literature.

Typically, the iron provided by the electrochemical reactor is not enough to remove all the

pollutants present in aqueous solution. Thus, the ozone contributes importantly to improve the

pollutant removal. Initially, the ozone contribution in the integrated process increases the

oxidation of pollutants that are dissolved in the solution and that cannot be eliminated via

electrocoagulation. An advantage of supplying ozone into the reactor is that it promotes the

mixing between the reactants and also maximizes the organics oxidation, that results in the

decreasing of COD and color [28,47,48]. Furthermore, the ozone provides good mixing thought the

reactor which improves the mass transfer. The ozone action also contributes to reduce the

amount of sludge produced.

In addition to processes coming from the combination of the effects of the single treatment

technologies, the combined process involves an increased hydroxide radical production because

Fe2+ catalyzes ozone decomposition to generate hydroxyl radicals (Eqs. 18-20) in the well-known

Fenton process. This process helps to explain the synergistic effect of the combination of both

technologies and the resulting high efficiencies.

Fe2+ + O3 (FeO)2++ O2 (18)

FeO2+ + H2O Fe3+ + HO· + OH- (19)

FeO2+ + Fe2+ +2H+ 2Fe3+ + H2O (20)


doi: 10.5599/jese.2014.0060 289

Table 1. Pollutant removal using coupled electrocoagulation - ozone processes

Wastewater Process Conditions Poll. Removal Ref.

C.I. Reactive Yellow 84 Ozone flow rate 20 mL min-1,

Iron electrodes; current density 15 mA cm-2

85 % TOC

100 % color [49]

Reactive Blue 19 Ozone flow rate 20 mL min-1,


80 % TOC

96 % Color [50]

Reactive Black 5 Ozone flow rate 20 mL min-1,


60 % COD

94 % Color

[51]

Distillery effluent Ozone flow rate 15 L min−1; initial pH 6

Iron electrodes; current density 3 Adm−2

83 % COD

100 % Color [52]

Industrial wastewater Ozone flow rate 23 L min−1; initial pH 7

Iron electrodes; current density 26 mA cm−2

63 % COD

90 % Turbidity [42]

Red MX-5B Ozone flow rate 0.5 L min−1; initial pH 6.1

Iron electrodes current density of 1.5 mA cm−2 100 % Color [53]

Boat pressure washing

wastewater

Iron and aluminium electrodes

current density 17 mA cm−2

88.46 %,TOC

76.28 % COD [54]

Acid Orange 6 azo

Dye

Ozone concentration 36 mg L−1; initial pH 4.5

Iron electrodes current density 88.6 mA cm−2

50 % TOC

80 % Color [55]

Main challenge for this technology is the scale-up. Most of the studies are at the lab-scale or

the bench-scale and typically efficiencies can be greatly improved if a proper scale up is carried

out. The design of the reactor seems to be a critical point because it fixes the flow patterns and

hence the interaction of the species formed by electrocoagulation with ozone. Another challenge

for this technology is the production of ozone by simultaneous anodic oxidation, taking advantage

of the possibilities of electrochemical technology to produce oxidants[56]. A good possibility could

be the use of cells equipped with bipolar cells such as the recently proposed by Llanos et al. [35]

3. Combined electrocoagulation- adsorption processes

Adsorption is a very well-known water and wastewater treatment process, which is gaining

prominence as a means of reducing metal ion and organic concentrations in industrial efflux-

ents [57]. The biosorbents derived from dead biomass, are considered the cheapest and most

abundant environmentally friendly option [58,59]. Nowadays, the development of inexpensive

adsorbents for the treatment of wastewater is an important area in the environmental

sciences [60,61].

The use of an electrochemical treatment in combination with adsorption as a pre-treatment

step to enhance adsorption capability of biosorbents has been assessed in many cases. However,

the applications must be carefully evaluated, because technical incompatibilities may arise. This

combined technology demonstrates a very good efficiency in the removal of many different

pollutants as it is shown in Table 2. The filtering capacity of the sorbent bed is an efficient

treatment to remove the suspended solids produced by the electrocoagulation process while

simultaneously it helps to remove all soluble pollutants that were not effectively trapped by the

flocs. Most of the studies select aluminum instead of iron as anode because aluminum coagulants

promotes neutralization coagulation processes instead of enmeshment into growing precipitates

which helps avoiding operational problems in the filtering system.


290 290

Table 2. Pollutant removal using coupled electrocoagulation-adsorption processes

Wastewater Process Conditions Pollutant Removal Ref.

Cr(VI) Al electrodes, sorbent red onion skin, pH 3-6

97 % Cr [62]

Cardboard paper mill effluents

Al electrodes, current density 4.41 mA cm-2

sorbent granular activated carbon, pH 5.3

99 % COD [63]

Marine Blue Erionyl MR Al electrodes, sorbent granular activated carbon, pH 6.0

100 % dye [64]

Reactive Black 5 current density 277 A m

-2

sorbent granular activated carbon, pH 7 100 % dye, 100 % COD, 100% Toxicity

[65]

Cr(VI) Al-Fe, current density 26.7 mA cm

-2

Sorbent granular activated carbon 99 % Cr(VI) [66]

Indigo carmine Al electrodes Sorbent granular activated carbon

99 % Colorant [67]

Nakdong River Al electrodes, Al-fiber filter 65 % TOC [20]

Industrial Wastewater Al electrodes, current density 45.45 A m

−2

Sorbent Ectodermis of Opuntia, pH 8

84 % COD, 78 %, BOD5, 97 % color, 98 % turbidity, 99 % fecal coliforms

[68]

Thus, the coupling of electrochemical and adsorption processes might prove a judicious choice

for treating industrial wastewater with mixtures of different types of pollutants including both

organic and inorganic pollutants. This technology has also been studied in systems in which an

adsorbent bed used for the fast removal of pollution from wastewater is continuously regenerated

using electrolysis [69,70]. Efficiencies obtained are high enough to consider this technology as a

promising choice in the treatment of many effluents polluted with organic species. Most studied

found in the literature are carried out at the lab or bench-scale. As for the combination of

electrocoagulation with ozone, it is expected that with a proper scale-up, which develop an

efficient cell from the view point of the filtering, adsorption and electrochemical processes,

efficiencies obtained would be even higher.

4. Combined electrocoagulation-ultrasound processes

The treatment of wastewater in an electrolytic cell by ultrasound irradiation is expected to

improve significantly the kinetics and the effectiveness of the electrode processes taking place in

the cell [71-73]. A number of favorable impacts of using ultrasound in electrocoagulation are the

following: Destruction of the compact layer formed at the electrode surfaces by the products of

electrode reactions.

Decrease in the thickness of the diffuse part of the electrical double layer created at the

electrode surface.

Direct activation of the ions in the reaction zone at the electrodes by ultra-sound waves.

Activation of the electrode surfaces by means of generation of defects in the crystal lattices

of the electrodes.

Local augmentation of the temperature at the electrode surfaces as a result of friction

between the liquid and the surfaces.

However, the ultrasound used may cause a few negative effects directly related to the

purification process, such as the following: Destruction of a part of the obtained colloidal hydroxides by the action of the acoustic

waves. This means a diminution of the solid phase that takes part in the adsorption process

and a diminution of the removed contaminations respectively.


doi: 10.5599/jese.2014.0060 291

Destruction of a part of the formed adsorption layer at the surface of the colloidal particles

and possible return of the adsorbed ions to the liquid phase.

Disorganization of the migration processes in the medium by the ultra-sonic waves.

Main studies found in the literature regarding this combined technology are shown in Table 3.

Table 3. Pollutant removal using coupled electrocoagulation- US irradiation processes

Pollutant Process Conditions Pollutant Removal, % Ref.

Cl - SO4-2

treatment time 60 min; Ultrasonic low-frequency Electrocoagulation Fe, current density 40 mA cm

−2

Media Cl- 500 ppm pH 3.8; Media SO4-2

500 ppm pH 2.8; Mine water

Very important removals with an increase in the amount of sludge at 25Hz.

[71]

Cr(VI)

Ultrasonic and sludge obtained for Electrocoagu-lation Fe of pharmaceutical wastewater ; EC: Conditions: rpm = 150, pH = 7.0 and sludge = 10 g L

-1.; Sono-EC Conditions:

frequency 30 kHz, pH = 7.0 and time = 100 min.

100 % of removal at 275 min when used 200 mg L

-1 of Cr(VI);

100 % of removal at 190 min when used 200 mg L

-1 of Cr(VI)

[72]

Cu(II)

Current 1.0 A. Electrolysis 8 h In the sonicated field, voltage and temperature were constantly increased, in order to maintain the same thermal conditions for non-sonicated solutions. Temperature was adjusted to match those during the sonicated process.

Electrolysis 100 and 200 mg L-1

removed 55 an 63 % of Cu(II) increasing concentration of Cu the removal was of 93 %; Sono-Electrolysis 100 and 200 mg L

-1

removed 94.6 an 95.5 % of Cu(II) increaseing concentration of Cu the removal

[73]

Non-ionic surfactants (SA)

Current density 0.5–2.5 A dm-2

, treatment time 5–40 min, ultrasonic power density 0.5–3.0 W cm

-2.

Frequency 22 ± 1 kHz. Treatment time: 10 min

68 % of AS only CE EC with US 90 %

[74]

car-washing wastewater

I=1.2 A pH= 6 treatment time 20 min

COD 68.77 % and turbidity 96.27 %

[75]

Main results of these studies show that the combined process promotes the flocculation

through vigorous mixing and the oxidation through the formation of radicals that contribute to the

enhancement of the efficiency of electrocoagulation processes by chemical polishing of the

surface of the flocs and by the oxidation of soluble pollutants in the bulk. This fact helps to explain

the high efficiencies reached. In this case, performance of iron electrodes is better than that of

aluminum electrodes. This fact can be explained in terms of the enhanced performance of the

enmeshment of the pollutant into growing metal hydroxide flocs which is much more important

for iron than for aluminum coagulants.

Hence, sono-electrocoagulation treatment has demonstrated superior performances in

treatment of industrial effluents than single electrocoagulation. However, and as it was described

for the other two previous technologies, scale-up should be considered as a major challenge.

5. Electrocoagulation-pulsed processes

Pulsed electrocoagulation technology is a novel method for wastewater treatment. It uses the

interactions of electrochemical technology and polarity reversal in an electrical field to induce

dipole formation in nonpolar particles in the wastewater, thus enabling the formation of micro-

aggregates of insoluble substances.

The aggregates formed are further assisted in forming macro-aggregates. Charge neutralization

of ions or charged materials also takes place in the electrochemical reactor, turning them into

insoluble, suspended substances in the wastewater. The neutralization process enhances the

efficiency of removing electrical conductivity [76]. In Table 4, there are some examples of using


292 292

electrocoagulation pulsed treatments, which show that has been done with wastewater of

different origins. As it can be observed, it has proven effective in the treatment of urban

wastewater and different types of industrial wastes.

Table 4. Pollutant removal using coupled electrocoagulation-pulsed processes

Wastewater Process Conditions Pollutant Removal Ref.

Higher Cr(VI) concentrations

Cr(VI) initial concentrations (50- 1000 mg L-1

) Electrical energy consumption (EEC) range: 4-58 kWh m

-3 wastewater, current density (CD):

56–222 A m-2

, operating time: 20–110 min, pH 3–9 (pHoptimum 5), voltage: 15–25 V.

99 % [47]

Synthetic solutions containing mercury(II)

Hg (II) 2×10−5

M,distance between the electrodes was 3 cm, current density ranging from 2.5 to 3.1 A dm

−2; charge loading 9.33- 15.55 F m

−3 ,

iron and aluminum electrodes, 3 - 7.

99.9 %. With iron, 15 min of electrolysis was sufficient to reach the highest removal; aluminum required 25 min for the same result.

[77]

Solutions of a dye Dianix Yellow CC (DY) and Procion Yellow (PY)

Range pH (4-8), Current density (40-120 A m-2

) Frequency (200-900 Hz

-1),

Operation time (100 min) 99 % [78]

Industrial and municipal wastewater

Pilot plant of electrochemical treatment system (0.3 m

3h

−1). Ti/RuO2–TiO2 anode was larger than

with a platinum anode

The removal of T-N, T-P, NH4–N and COD was approximately 90 %

[79]

Berberine hydrochloride (BH) wastewater

Fe electrodes and Al electrodes. The optimal conditions of reaction time of 3.5 h, pulse duty cycle of 0.3, pulse frequency of 1.0 kHz, current density of 19.44 mA cm

-2, and

electrode distance of 2.0 cm

90.1 % BH and 62.6 % COD

[80]

Dye wastewater Fe electrodes and 1000 mg L

-1 Dye solution in a

15 mins electrolyzing time 99.62% of color removal and 91.15% of COD

[81]

Old corrugated containerboard (OCC)-based Paper Mill Wastewater

Current density of 0 to 240 A m-1

, a hydraulic retention time of 8 to 16 min and a coagulant (anionic polyacrylamide) dosage of 0 to 30 mg L

-1

Electrical conductivity: 47.7 %; Suspended. Solids: 99.3 %;COD: 75 %

[76]

Cooking oil (1800 mg/L, scour (1000 mg/L) and sodium sulfate (1g/L)

Al electrodes, dimensions of 50×110×2 mm; AC power (SMD 30)

Passivation of Al electrodes is not observed

[82]

Electroplating wastewater

Having a pH of 4, voltage 2.5 V, hydraulic retention time of 15 minutes, current density of 25 A m

-2

99.5 % [83]

Oil wastewater Electrode distance of 3.3 cm, pH of 4, current density of 49.38 mAcm

-2, reaction time of 15 min

and pole switching time of 10 s.

96.21% [84]

Conclusions

Electrocoagulation has demonstrated to be a promising technology in the removal of pollutants

from different types of wastewater. However its combination with other technologies can help to

increase efficiency due to synergistic effects such as those derived from the formation of radicals

in the ozonation (by interaction of ozone with iron (II)) or in the US irradiation. Results depend on

the particular application (technology combined and type of wastewater) and should be evaluated

carefully. Scale-up is the major challenge of this technology for the next years, although the very

positive results obtained at the lab and bench scales make these studies very promising.


doi: 10.5599/jese.2014.0060 293

Acknowledgements: The authors wish to acknowledge the support given by the UAEM trough the project 3409/2013 M and financial support from CONACYT in project 153828 is greatly appreciated.

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doi: 10.5599/jese.2014.0054 297



www.jESE-online.org

Review

Electrokinetics and soil decontamination: concepts and overview

Mohammed A. Karim

Department of Civil and Construction Engineering, Southern Polytechnic State University (SPSU), 1100 South Marietta Parkway, Marietta, Georgia 30060, USA

E-mail: [email protected] & [email protected], Phone: (678) 915-3026 (Off.); (804) 986-3120 (Cell)

Received: February 12, 2014; Revised: May 17, 2014; Published: December 6, 2014

Abstract Electrokinetic decontamination and extraction have been proven to be one of the most viable, cost effective and emerging techniques in removing contaminants, especially heavy metals from soils for about last five decades. Basic concepts and an overview of the electrokinetic extraction processes and their potential applications in geotechnical and geoenvironmental engineering have been reviewed based on the literature and presented in this paper. Primarily, theoretical and laboratory experimental studies related to electroreclamation of soils are summarised in brief with basic concepts of electrokinetic processes. The paper has been divided into different sections that include history of electrokinetics, background and concepts, modelling, parameter effects, instrumentation, contaminant extraction, field applications, and summary and recommendation. Based on the review it is obvious that the field application of electrokinetic technology to remediate heavy metal contaminated soils /sediments is very limited and site specific. Additional laboratory studies and more pilot- and full-scale information from field applications are critical to the further understanding of the technology and to customize the process in different field conditions.

Keywords Electrokinetic decontamination; heavy metals; site remediation; soil; EDTA; soil pH; electro-osmosis; electrophoresis; streaming potential; ion migration; sediment potential; zeta potential; electrolysis; electrokinetic modelling

Introduction

Contaminant such as heavy metals removal from solid porous medium such as soils and

sediments has been a technological challenge for engineers and scientists for the past several

decades. A variety of remedial options exist to cleanup a hazardous waste site; however, the




J. Electrochem. Sci. Eng. 4(4) (2014) 297-313 ELECTROKINETICS AND SOIL DECONTAMINATION

298

technological challenge, efficiency, and costs of these options may vary widely. Conventional

ground burial and land disposal are often economical, but they do not provide a permanent

solution, and in some cases they are not necessarily the most effective solutions. For removing

contaminants such as organics and inorganics from solid porous media, the most common ex-situ

methods employed include soil washing, and ligand extraction. Ex-situ methods may not be

technologically challenged that much; however, they suffer from several problems. Apart from the

generic problems of any ex-situ process, i.e., the need to excavate the media and place it in an

external reactor, the above mentioned processes suffer from several disadvatages [1].

Several in situ methods include vacuum extraction, thermal desorption, hydraulic fracturing,

electrokinetic decontamination (including the "Lasagna" process), biotreatment, immobilization by

encapsulation, and placement of barrier systems are already in use to some extent for soil and

sediment remediation and decontamination. Most of these processes are employed for removal of

organics present in soils or sediments. Among these in-situ methods electrokinetic

decontamination (EKD) processes are in use for the past five decades in different applications. The

major advatages of the EKD processes include (a) they can be implemented in-situ with minimal

disruption, (b) they are well suited for fine-grained, heterogenous media, where other processes

can be ineffective, and (c) accelerated rates contaminant extraction and transport may be

achieved. The basic concepts and an overview of the EKD processes and their real life applications,

as of now, in geotechnical and geoenvironmental engineering have been reviewed and presented

in this paper. Primarily, theoretical and laboratory experimental studies related to EKD of soils and

sediments are presented in brief with basic concepts of electrokinetic processes.

History

The movement of water through capilary and pores as a result of the application of electric

potential is known as electrokinetic phenomena and this phenomena was first described by F. F.

Reuss in Russia in 1808. This phenomenon was first treated analytically by Helmholtz in 1879,

which was later modified by Pellat in 1903 and Smoluchowski in 1921. This phenomenon is widely

known as the Helmholtz-Smoluchowski model which relates electro-osmotic velocity of a fluid of

certain viscosity and di-electric constant, through a charged porous medium under an electric

gradient. The Helmholtz-Smoluchowski model is the most common theoretical description of

electro-osmosis and is based on the assumption of fluid transport in the soil or sediment pores

due to transport of the excess positive charge in the diffuse double layer towards the cathode [2].

It applies to systems with pores that are large relative to the size electric diffuse double layer and

provides with reasonable predictions for electro-osmotic flow in most soils. The rate of electro-

osmotic flow is controlled by the coefficient of electro-osmotic permeability of porous media and

the balance between the electrical force on the liquid and the friction between the liquid and the

surface of the particles of the porous media. The first application of electrokinetics was made by

Casagrande in 1939 for consolidation and stabilization of soft fine-grained soils. Numerous

laboratory studies and a very few field applications have been conducted to investigate the

electrokinetic processes to date. The areas in which electrokinetics have been applied successfully

to some extent include increasing pile strength, stability of soil during excavation and

embankments, increasing flow rate of petroleum production, removal of salts from agricultural

soils, removal of metalic objects from the ocean bottom, injection of grouts, microorganisms and

nutrients into the subsoil strata of low permeability, barriers and leak detection systems in clay

liners, dewatering of clayey formations during excavation, control and decontamination of

M. A. Karim J. Electrochem. Sci. Eng. 4(4) (2014) 297-313

doi: 10.5599/jese.2014.0054 299

hazardous wastes, removal of chemical species from saturated and unsaturated porous medium,

removal of gasoline hydrocarbons and trichloroethylene from clay and removal or separation of

inorganic and organic contaminants and radionuclides.

Background and Concepts

Electrokinetic processes are a relatively new and promising technology being investigated for

their potential applications in hazardous waste management specifically in case of high clay

containing soils. United State Environmental Protection Agency (USEPA) has designated

electrokinetic method as a viable in-situ process and interested parties are attempting to apply

this method at contaminated sites which have inherently low permeability soils and otherwise

difficult to decontaminate. Electrokinetic flows occur when an electric gradient is applied on a soil-

fluid-contaminant system due to existence of the diffuse double layer at the soil particle surface –

pore fluid interface. Several electrokinetic phenomena arise in clay when there are couplings

between hydraulic and direct current (DC) electrical driving forces and flows. Those phenomena

can broadly be classified into two pairs by the driving forces causing the relative movement

between the liquid and the solid phases. The first pair consists of electro-osmosis and

electrophoresis, where the liquid or the solid phase moves relative to the other under the

influence of an imposed electrical potential. The second pair consists of streaming potential and

migration or sedimentation potential, where the liquid or the solid phase moves relative to the

other under the influence of hydraulic or gravity force and thus inducing an electrical potential.

Those four electrokinetic phenomena in clay are depicted in Fig. 1 [3].

Fig. 1. Electrokinetic phenomena in clay


300

The detailed description of these flow processes and the associated complicated features

generated by electrochemical reactions are given by several authors [4-23]. The use of

electrokinetics in sealing leaks in geomembrane and compacted clay liners has been explained in

detail by a few authors [24-28]. Potential applications of electrokinetics in geotechnical and

geoenvironmental engineering are described elaborately by multiple authors [21,22,27,29,30-34].

Some of the applications, as appropriate, are reviewed and included in the subsequent sections.

The extraction technique, variably called electrokinetic remediation, electroremediation,

electroreclamation, electrorestoration, electrochemical soil processing or electrochemical decon-

tamination, uses low level constant voltage DC power supply, potential gradients in the range of

20–200 V m-1 [35] or alternatively a constant current density in the range of 0.025–5 A m-2 [31]

between the electrodes placed at the end of the contaminated soil sample. When an electric field

is imposed to a wet soil mass, positive ions are moved toward the cathode (the negative

electrode) and the negative ions toward the anode (positive electrode) as illustrated in Fig. 2 [36].

Because of the isomorphous substitution and the presence of broken bonds in the soil structures,

excess mobile cations are required to balance the negative fixed charges on the soil particle

surfaces. Therefore, mobile cations exert more momentum to the pore fluid than do mobile

anions. As a result there is a net movement of fluid relative to soil particles under the influence of

imposed electric potential gradient which is called electro-osmosis (field-induced convection of

water through a porous medium with a surface charge). Unlike water flow under pressure, electro-

-osmosis depends on the electric current through the soil, the flow resistance of soil, and the

frictional drag exerted by the migrating ions in the water molecule and this flow originates at the

electric double layer of the soil pores. The electrokinetic flow rate qeo in a porous medium of

length L, porosity n, area A and degree of saturation S, may be presented by the following

equation [37]:

d o seo s

D Rq I nAS

L

(1)

where d is the potential at the slipping plane, o is the permeability of free space, D is the

dielectric constant of the pore fluid, is the pore water viscosity, Is is the current carried by

surface conductance and Rs is the surface resistance of the porous medium i.e. soil.

Fig. 2. Concept of electrokinetic extraction of contaminants


doi: 10.5599/jese.2014.0054 301

When the electrokinetic technique is applied without conditioning of the process fluid at the

electrodes, which is termed as unenhanced electrokinetic remediation, the applied electric current

leads to electrolysis reactions at the elctrodes, generating an acidic medium at the anode and an

alkaline medium at the cathode [38]. The electrolysis reactions of the primary electrodes are

presented in the following equations:

Anode Reaction: 2H2O - 4e- O2 + 4H+, Eo = -1.229 V (2)

Cathode Reaction: 2H2O + 2e- H2 + 2OH-, Eo = -0.828 V (3)

where Eo is the standard reduction electrochemical potential, which is a measure of the tendency

of the reactants in their standard states to proceed to products in their standard states. Although

some secondary reactions might occur at the cathode because of their lower electrochemical

potential, the water reduction half reaction (H2O/H2) is dominant at early stages of the process.

Within the first few days of the process, electrolysis reaction drops the pH at the anode below 2

and increases the pH at the cathode above 10, depending the total current applied [9]. The

following are the secondary reactions that may exist depending upon the concentration of

available species:

H+ + e- (1/2) H2 (4)

Mn+ + ne- M (5)

M(OH)n(s) + ne- M + nOH- (6)

where M refers to metals. The acid medium (Eq. 2) generated at the anode advances through the

soil toward the cathode by ionic migration and electro-osmosis due to electrical gradient, pore

fluid flow due to any externally applied or internally generated hydraulic gradient and diffusion

due to the chemical gradients developed in the system. The base developed at the cathode initially

advances toward the anode by diffusion and ionic migration. However, the counterflow due to

electro-osmosis retards the back-diffusion and migration of the base front. The advance of this

front is slower than the advance of the acid front because of the counteracting electro-osmotic

flow and also because the ionic mobility of H+ is about 1.76 times that of OH-. As a result, the acid

front dominates the chemistry across the specimen except for small section of the specimen close

to the cathode, where base front prevails [21,35]. As the acid buffer capacity of soil or sediment is

low, acid front moving through the soil lowers the system pH. Since most heavy metals are soluble

in an acidic environment, this lowering of pH promotes desorption of heavy metals from the soil

and solubilization of metal ions. Ions in dissolved phase can be removed effectively by the

combined actions of electro-osmosis and ion migration. However, the presence of heavy

molecular weight organic matter (humus substances) within the soil pores may reduce the

mobility of the heavy metals due to the formation of organometallic compounds. Under these

circumstances, enhanced electrokinetic remediation could be necessary. Numerous studies have

been conducted to date using different chelating and complexation agents to enhance the

remedial techniques [39-52]. The particular use of the enhancing and conditioning agents are

reviewed and included in the appropriate sections.

Modeling electrokinetics

Electrokinetic modeling is based on the applicability of coupled flow phenomena for fluid,

solute, current and temperature flow through porous media under the influence of hydraulic,


302

electrical, concentration, and thermal gradients, respectively. The governing equations for these

analyses generally have been formulated on the basis of the postulates of irreversible

thermodynamics and the applicability of the Onsager reciprocal relations under the assumption of

isothermal conditions [14,16], although equation formulation on the basis of continuity

considerations has also been shown [53,54]. The state-of-the-art in modeling electrokinetic

remediation is represented by the one-dimensional finite element model for coupled multi-

component, multispicies transport under electrical, chemical and hydraulic gradients described in

a study conducted by Alshawabkeh and Acar [54]. This study compared the predictions of Pb

removal using the model with the results of pilot scale study involving electrokinetic extraction of

Pb from a spiked kaolinite sand mixture. Multidimensional models for multi spices transport have

been developed by several reserachers [55-57]. A study conducted by Haran et al. [58] developed

a mathematical model for decontamination of hexavalent chromium from low surface charged

soils. They simulated the concentration profiles for the movement of ionic species under a

potential field for different time period. The model predicted the sweep of the alkaline front

across the cell due to the transport of OH- ions. A comparison of chromate concentration profiles

with experimental data for 28 days of electrolysis showed a good agreement. A numerical model

of transport and electrochemical processes was extended for the first time to incorporate

complexion and precipitation reactions in a study by Jacobs et al. [59]. Their model confirmed that

the isoelectric focusing could be eliminated and high metal removal efficiencies could be achieved

by washing the cathode. In order to describe the transport and reaction processes in a porous

medium in electrical field, one-dimensional numerical models have been developed by several

authors [60-62]. In several studies, Choi and Lui [63-66] developed a mathematical model for the

elctrokinetic remediation of contaminated soils assuming the contaminants are mostly heavy

metals, water is in excess, the dissociation-association of water into hydrogen and hydroxyl ions is

rapid, and that electro-osmosis is significant when compared to electromigration (field-induced

transport of ions in an electrolyte as defined earlier) as a transport mechanism. The analytical

steady state solutions of electroplating and transport in binary electrolyte arising from

electrochemistry were provided in several articles by several authors [67-70]. Electrolysis and

isoelectric focusing effects were also theoretically analyzed by various researchers [68-71].

Modified finite difference model of electrokinetic transport in porous media was developed and

numerical solutions were provided in studies [60,72]. An assessment of available multispecies

transport model and an investigation of long-time behavior of multi-dimensional electrophoretic

models were done in couple of studies [9,73]. The quantitative determination of potential

distribution in Stern-Gouy double layer model was elaborated by Shang et al. [74]. The analytical

and numerical steady state solutions for electrochemical processes with multiple reacting species

were provided in articles [75,76]. Shackelford [77] summarized the modeling electrokinetic

remediation. In his review he emphasized that the prediction of multi-component, multi-species

transport with chemical reactions through soil medium represents one of the challenging

modeling endeavors in environmental geotechnics. He compared his statement with studies

conducted by Acar and Alshawabkeh [78] and mentioned that this study provided some insight of

the advances along these lines. However, he stressed on the additional effort that is needed in

evaluating the potential limitations in modeling these electrokinetic processes in terms of the

assumptions inherent in the models and field-scale applications.


doi: 10.5599/jese.2014.0054 303

Instrumentation

Electrokinetics has many applications in geo-environmental and geotechnical engineering. For

the measurements of electrokinetic properties of soil and soil remediation processes, individual

researchers have designed their own apparatuses of various shapes, sizes and materials for

different purposes. Some significant experimental apparatuses used for geotechnical and geo-

environmental engineering investigation have been reviewed in detail by Yeung [13]. A number of

important apparatuses that have been used for soil remediation by electrokinetics are mentioned

here. The apparatuses currently available for the purpose of electrokinetic remediation include

those developed at Louisiana State University [31,79], Lehigh University [52,80,81], University of

Texas at Austin [11,82], the University of California at Berkeley [3,83], Massachusetts Institute of

Technology [38, 47, 59], Texas A & M University [6], The Technical University of Denmark [84,85],

Vanderbilt University, Nashville, Tennessee [86], Royal Institute of Technology, Stockholm, Sweden

[87-89], University of South Carolina [58] and many others. A comprehensive review of the

apparatus used in the EKD experiments has been presented by Yeung et al. [6]. However, it is

obvious from the literature that most of these apparatuses are used for the remediation of fine-

grained soils by electro-osmosis. None of them except the last three are used for the

decontamination of course-grained soils such as sandy/salty soils, where the electro-osmosis is

ineffective [90]. It is reported that the last two instruments have been successfully used to

decontaminate sandy soils using electrolysis and electro-migration.

Parameter Effects

The important parameters of EKD processes are electric gradient, system pH, electro-osmotic

flow, ion-migration, zeta potential, electro-osmotic permeability, and current density. All of these

parameters play important role in the process efficiency, soil decontamination, and ultimately the

cost. Therefore, parameter optimization should be an important part the process performance. In

general, the application of electric gradient induces electric current density and promotes the

electrolysis reactions at anode and cathode. Electric current results in generation of protons (H+)

at the anode (Eq. 2) that migrate together with the metal cations to the negatively charged ca-

thode (Fig. 2) for removal and processing. A very low voltage can serve the purpose of electrolytic

reactions and create low pH solution in the anode. So determination of optimum electric gradient

or current density is important as higher electric gradient or current density may increase the cost

of the process and create higher gases in anode and cathode which may require careful watch and

become difficult to maintain experiments. Electro-osmotic flow is the prevalent parameter for the

low permeable soils having high surface charges whereas ion-migration may be the driving force

for high permeable soils having low surface charges. System pH contributes to the dissolution of

metal precipitates and depends on the type of contaminants and their salts present in the soils.

Most of the metal salts may be soluble in a pH range of 2 to 4. Therefore, bringing the soil pH

below 2 may not be necessary to optimize the removal efficiency.

It is reported that the values of hydraulic conductivity of different soils can differ by orders of

magnitude; however, those of coefficients of electro-osmotic conductivity are generally between

1 × 10-5 and 10 × 10-5 cm2 V-1 s-1 and are relatively independent of soil type. Thus, an electric

gradient is much more effective driving force than a hydraulic gradient for moving fluid through

fine-grained soils of low hydraulic conductivity [6,9,83]. Korfiatis et al. [91] used an experimental

approach to assess the relative magnitudes of hydraulic and electro-osmotic permeability under

application of hydraulic or electric gradients or both and to study the extent of pH changes during


304

the electro-osmotic process. The practical and theoretical aspects of ion exchange resins and

membranes have been investigated by Hansen [85,92]. Acar et al. [30] estimated the electro-

osmotic permeability in kaolinite to be in the range of 0.80 × 10-5 to 3.0 × 10-5 cm2 V-1 s-1 which is

within the range reported in the literature.

The zeta potential of most soils, except for quartz, is negative, because soil surfaces carry a

negative charge that causes the electro-osmotic generally from anode to cathode. The pH and

ionic strength of the pore fluid may affect the value of zeta potential and zeta potential is reported

to decrease linearly with logarithm of the pH of the porous medium [2]. High acidic solution causes

the zeta potential to become less negative and even to attain positive values at low pH. As a result

flow rates have been reported to decrease if the pH of the electrolyze is depressed below neutral

and to increase at alkaline pH values [47,93]. The effect of zeta potential on electro-osmotic

permeability has further been investigated by Shang [94].

The steady state and limiting current conditions are investigated by Dzenitis [95]. Influence of

current density and system pH on electro-remediation of kaolinite clay was investigated by

Rahman [45] and Hamed and Bhadra [93] and soil saturation effect on electrorestoration was

investigated by Puppala [46]. The effects of temperature on electrokinetic remediation on low

permeability soils are explored by Penn [96]. The effects of electrokinetics in complex natural

sediments are explained by Grundl and Reese [97]. Shang et al. [98] investigated the effects of

polarization and conduction on clay-water-electrolyte systems. Shri Ranjan and Karthigesu [99]

devised a capillary flow meter for measuring the hydraulic conductivity of clay under the

applications of low gradients. A theoretical and experimental basis on electrokinetic

sedimentation is explained by Shang [5]. Reddy et al. [100] investigated the effects of soil

composition on the electrokinetic extraction of chromium (VI). They used three kinds of soil

minerals such as kaolin, glacial till, and Na-montmorillonite in their study. Their study found that

the adsorption and removal of Cr (VI) are greatly dependent on the compositions of the soil

minerals.

Contaminant extraction

There are some cases where unenhanced electrokinetic extraction is ineffective for soil

remediation. In this situation chelating and conditioning agents are used to enhance the process

which is termed as enhanced electrokinetic remediation. The most commonly used chelating and

conditioning agents are Ethylene diamine tetraacetate (EDTA), HCl, acetic acid, iodine-iodide etc. A

few important studies using enhanced and unenhanced electrokinetic process have been reviewed

and presented below. However, only the studies related to heavy metal removals were reviewed

and reported here.

Heay Metal Removal with Enhanced Process

In a study conducted by Cameselle and Reddy [101] found that electro-osmotic flow under

applied electric potential depends on a number of soil, contaminant and applied electric potential

conditions. Electro-osmotic flow induced in the same direction of metal or complexed metal ions

transport can enhance heavy metal removal. In case of hydrophobic organic contaminants,

periodic voltage application combined with the use of a solubilizing solution is shown to create

sustained electro-osmotic flow and enhanced contaminant removal. The suggested to validate the

optimum conditions determined from laboratory investigations for generating significant electro-

osmotic flow through field pilot-scale demonstrations. Joseph et al. [41] investigated the feasibility


doi: 10.5599/jese.2014.0054 305

of mobilizing precipitate heavy metals from soil by ionic migration using EDTA. They used EDTA

solution to catholyte where it solubulizes the precipitated metals. The resulting complexes are

then transported to the anode. The removal efficiencies were found to be very close to 100 % for

Zn and Pb. A feasibility study of using surfactants and organic acids sequentially and vice versa

during EKD was evaluated by Reddy et al. [102] for removal of both heavy metals and PAHs from

clayey soils. They selected kaolinite as a model clayey soil and spiked it with phenanthrene and

nickel at concentrations of 500 mg kg-1 dry each to simulate typical field mixed contamination.

They performed bench-scale electrokinetic experiments with the sequential anode with 1 M citric

acid followed by 5 % Igepal CA-720, 1 M citric acid followed by 5 % Tween 80, and 5 % Igepal

CA-720 followed by 1 M citric acid. The migration and removal efficiency of panathrene in the first

two sets of tests were found to be very low. But overall the sequential use of 5 % Igepal CA 720

followed by 1 M citric acid appeared to be an effective remedial strategy to remove coexisting

heavy metals and PAHs from clayey soil. The effect of EDTA in removing Pb and Zn from millpond

sludge during EKD was investigated by Karim and Khan [39]. They conducted several experiments

with distilled water and dilute EDTA solutions with strengths of 0.05 M and 0.125 M. The beneficial

effects of using EDTA that were observed in this investigation are EDTA substantially increased the

electro-osmotic flow in the millpond sludge indicating that it could significantly reduce the

duration of EKD, a significantly higher percentage of Pb and Zn removal from the solid phase due

to the complexation of EDTA with these heavy metals, and EDTA was able to prevent the

precipitation of metals near the cathode electrode typically observed in EKD process. Yeung

et al. [103] studied the basic Pb-EDTA complexion reactions and their influence on electrokinetic

extraction process. Their main focus was on EDTA enhanced electrokinetc extraction of lead from

Milwhite and Georgia kaolinite and the acid/base buffer and sorption capacities of these soil

minerals. Their study revealed that more than 90 % of lead was migrated toward the cathode with

a lower voltage applied across the sample within a shorter duration of treatment. Allen and

Chen [48] investigated the extraction of lead from the contaminated New Jersey and Delaware

soils with EDTA. The investigation found almost 100 % extraction of lead from New Jersey soil at a

10-3 M concentration of EDTA and at 10-3 M or lower concentration of EDTA, the recovery of lead

that had been added to the Delaware soil was greater than that of New Jersey soil that had been

previously contaminated at level of pH 4.30.1. Li et al. [88,89] suggested a new approach in

electrokinetic decontamination in which a conductive solution was inserted between the cathode

and the soil to be treated. By this approach, the pH in the soil can be kept low so that no metal

precipitation would occur near the cathode. This would eliminate the isoelectric focusing effect.

Their study found the metal removal efficiencies of more than 96 % for both copper and zinc. A

similar approach was suggested by Shapiro et al. [104] in which acetic acid was used to rinse the

catholyte to reduce the pH near cathode. Cox et al. [105] studied the remediation of mercury from

soils using iodine-iodide as a chelating agent and found it to be very effective. Acar and

Alshawabkeh [78] investigated the feasibility and efficiency of transporting Pb under an electric

field with a constant current. The tests were conducted with a Pb concentration of 856 mg kg-1 and

1,553 mg kg-1 respectively. The third test was conducted on a 1:1 mixture of kaolinite and sand

with Pb concentration of 5,322 mg kg-1. Their study found that 55 % of Pb mobilized inside the soil

precipitated within the last 2 cm close to the cathode, 15% were left in the soil before reaching

this zone, 20 % precipitated on the fabric separating the soil from cathode, and 10 % were

unaccounted. Ellis et al. [106] studied the release of cadmium, chromium, copper, lead, and nickel

from soil collected from a Superfund site near Seattle, Washington. They conducted both batch


306

equilibrium and column studies using EDTA alone and EDTA followed by hydroxylamine

hydrochloride, to reduce iron oxides in the soil. Results of their batch and column tests showed

that EDTA was able to remove more than 90 % Pb and 60 % Cd. Huang et al. [107] found that the

removal of Zn (II) from solids is independent of types of solids. The addition of EDTA resulted in a

shift of maximum Zn (II) adsorption to the acidic pH range, and reduction of zeta potential and

overall Zn (II) removal in presence of EDTA was significantly reduced at alkaline pH range and

slightly enhanced in the acidic range. Klewick and Morgan [108] explored the rates of decompo-

sition of complexes for Manganese in the +III oxidation state as a function of the complexing

ligand, the total ligand: manganese concentration ratio and the pH. Three ligands were chosen,

EDTA was one of them. The rate of appearance of the Mn (III) complex decreased with increasing

pH over the range of 6 to 8. McArdell et al. [109] studied cobalt-EDTA complexation generated on

site at Oak Ridge, TN shallow landfills. Their study confirmed the ability of EDTA to solubilize

mineral surface-bond Co (III). Davis and Singh [110] studied the several chemical washing

procedures for Zn (II) contaminated soil to determine the metal extraction efficiency from using

specific extractants such as acid solution, EDTA, diethylenetriamine pentaacetic acid (DTPA), and

Chlorine. Their study found 79 % removal of Zn(II) with 0.001 M EDTA, 85 % with 0.003 M EDTA for

pH around 2; 79 % with 0.001 M DTPA, 90% with 0.003 M DTPA for a pH of 2, and 85 % with

0.003 M DTPA for a pH of 6. They also found that about 99 % of Zn(II) was in the form of Zn-EDTA

complex at pH level 6. Amrate et al. [111] tested the removal of lead from an Algerian contami-

nated soil (with Pb concentration ≈4.43 mg/g of soil) sited near a battery plant using EDTA at

various concentrations (0.05–0.20 M). They applied a constant voltage corresponding to nominal

electric field strength of 1 V cm-1 for duration of 240 hours. Results of contaminant distribution

across the experimental cell have shown efficient transport of lead toward the anode despite the

presence of calcite (25 %) and the high acid/base buffer capacity of the soil. They modified the cell

by adding extra compartments and inserting cation exchange membranes (Neosepta CMX) to

avoid ligand loss, which would be anodically oxidized. They found simultaneous recovery of EDTA

and lead from their chelated solutions. Reddy et al. [112] conducted batch and electrokinetic

experiments to investigate the removal of three different heavy metals, chromium (VI), nickel (II),

and cadmium (II), from a clayey soil by using EDTA as a complexing agent. Their batch experiments

revealed that high removal of these heavy metals (62–100 %) was possible by using either a 0.1 M

or 0.2 M EDTA concentration over a wide range of pH conditions (2–10). However, the results of

the electrokinetic experiments using EDTA at the cathode showed low heavy metal removal

efficiency. They used EDTA at the cathode along with the pH control at the anode with NaOH

which increased the pH throughout the soil and achieved high (95 %) Cr (VI) removal, but the

removal of Ni (II) and Cd (II) was limited due to the precipitation of these metals near the cathode.

Their finding was that the low mobility of EDTA and its migration direction, which opposed electro-

-osmotic flow, prevented EDTA complexation from occurring. They also found many complicating

factors that affected EDTA-enhanced electrokinetic remediation and suggested further research to

optimize this process to achieve high contaminant removal efficiency.

Heay Metal Removal with Unenhanced Process

A comprehensive treatise on removal of Pb (II) from kaolin is reported by Hamed [34] and

Hamed et al. [79]. The process removed about 75 % to 95 % of Pb (II) at concentrations up to

1500 g g-1 across the test specimen at a energy expenditure of 29–60 kWh m-3 of soil processed.

Li et al. [88] examined the efficiency of electro-migration process in removing Pb (II), Cd (II) and


doi: 10.5599/jese.2014.0054 307

Cr (III) from sandy soils. Their study showed the removal efficiencies more than 90 % for all three

metals. Hamed and Bhadra [93] studied the effect of current density and influent pH on

electrokinetic processing. Their study results revealed that flow rate increases as the current

density increases and the electro-osmotic flow increases gradually between pH of 2 to 10 and

sharply between pH of 10 to 12. Acar and Alshawabkeh [78] investigated the feasibility and

efficiency of transporting lead under electric field conducting three pilot-scale tests with lead-

spiked kaolinite at an electrode spacing of 72 cm. In their tests program, a constant current of

density 133 A cm-2 was applied. Out of three tests, two of them were conducted with a lead

concentration of 856 mg kg-1 and 1,533 mg kg-1 respectively. The third test was conducted on a 1:1

mixture of kaolinite and sand with lead concentration of 5,322 mg kg-1. Their study found that

55 % of lead removal across the soil precipitated within the last 2 cm close to the cathode, 15 %

left in the soil before reaching this zone, 20 % precipitated on the fabric separating the soil from

cathode and 10 % unaccounted. Hansen et al. [84] investigated the removal of Cu, Cr, Hg, Pb and

Zn from sandy loam by electrodialysis. Their study found that decontamination of soil was to an

extent lower than the recommended critical values for metal concentration in soil. The

elctrochemical analysis of ion-exchange membrane with respect to a possible use in electrodialytic

decontamination of soil polluted with heavy metals was also studied by Hansen et al. [85]. Their

study revealed that cation-exchange membranes show the transport number of average 0.97 in

NaCl and CaCl2 solutions and anion-exchange membranes about 0.95 in NaCl, CaCl2 and ZnCl2

solutions. One-dimensional experimental studies were conducted by Yeung et al. [113] and

Darilek et al. [27,28] to examine the feasibility of using electrophoresis to repair in-service leaking

surface impoundment lined by geomembranes. Their studies were concentrated on the effect of

clay type, clay particle concentration in the suspension and the electric field strength on the cake

formation mechanism. Acar et al. [114] investigated the removal of Cd (II) from saturated kaolinite

under the application of electric current and found to remove more than 95 % of Cd (II) within

10 days of experiment. The effect of various sites and operating conditions on the efficacy of metal

removal by electromigration was investigated by Hicks and Tondorf [38] and Pamukcu and

Wittle [81]. Pamukcu et al. [52] investigated the feasibility of electro-osmosis to remove zinc from

soil since it was listed among the 129 priority pollutants by EPA and is known to possess moderate

noncarcinogenic toxicity and is found frequently in the soil in contaminated sites. Their finding was

encouraging in zinc migration to the cathode chamber. Reddy and Chinthamreddy [115] studied

the migration of hexavalent chromium, Cr (VI), nickel, Ni(II), and cadmium, Cd (II), in clayey soils

that contain different reducing agents under an induced electric potential. They conducted bench-

scale electrokinetic experiments using two different clays, kaolin and glacial till, both with and

without a reducing agent. The reducing agent used was either humic acid, ferrous iron, or sulfide,

in a concentration of 1,000 mg kg-1. They spiked the soils with Cr (VI), Ni (II), and Cd (II) in

concentrations of 1000, 500 and 250 mg kg-1, respectively, and tested under an induced electric

potential of 1 V DC cm-1 for duration of over 200 hours. Their study found that the reduction of

chromium from Cr (VI) to Cr (III) occurred prior to electrokinetic treatment and the extent of this

Cr (VI) reduction was found to be dependent on the type and amount of reducing agents present

in the soil. The maximum reduction was found to be occurred in the presence of sulfides, while the

minimum reduction was found to be occurred in the presence of humic acid. Their study

concluded that significant removal of the contaminant from the soils was not achieved and

suggested additional research to determine strategies by which contaminant migration may be

enhanced and ultimately lead to significant contaminant removal. Ricart et al. [116] investigated


308

the feasibility of electrokinetic remediation for the restoration of polluted soil with organic and

inorganic compounds had been development and evaluated using a model soil sample. They

prepared model soil was prepared with kaolinite clay artificially polluted in the laboratory with

chromium (Cr) and an azo dye: Reactive Black 5 (RB5). They focused on the electromigration of Cr

in a spiked kaolinite sample in alkaline conditions. Despite of the high pH registered in the

kaolinite sample (around pH 9.5), they reported that Cr migrated towards the cathode and it was

accumulated in the cathode chamber forming a white precipitate. The removal was not complete,

and only 23 % of the initial Cr was retained into the kaolinite sample close to the cathode side.

They also reported that the electrokinetic treatment of a kaolinite sample polluted with both Cr

and RB5 yielded very good results. The removal of Cr was improved compared to the experiment

where Cr was the only pollutant, and RB5 reached a removal as high as 95 %. RB5 was removed by

electromigration towards the anode, where the dye was degraded upon the surface of the

electrode by electrochemical oxidation. Chromium (Cr) was transported towards the cathode by

electromigration and electro-osmosis. The concluded that the interaction among RB5 and Cr into

the kaolinite sample prevented premature precipitation and allow Cr to migrate and concentrate

in the cathode chamber. The removal of PAH and metal contaminants from a former

manufactured gas plant polluted soil was studied by Reddy et al. [117] and found that the removal

is influenced by the type of flushing solution and application of voltage gradient. Igepal surfactant

was shown to remove PAHs, while EDTA chelant was shown to remove heavy metals. Sequential

application of surfactant and chelant removed both PAHs and heavy metals present in the soil and

the efficacy of the process depends on the order of flushing. Application of voltage gradient is

found to retard the removal of PAHs and enhance the removal of metals from the soil. Their

experiments conducted only for a short duration and suggested to run the experiments for longer

duration to establish this as a potent technology for the remediation of soil contaminated by

mixed wastes. The study suggested that soil composition can have a profound effect on the

contaminant removal; therefore, site-specific soil investigations must be conducted to develop

sequential process that will be effective to remove mixed contaminants from the soil.

It is apparent to say that enhanced electrokinetic removal technology has been more effective

in removing heavy metals from low permeability soils compared to unenhanced electrokinetic

removal technology as the enhanchment agents eliminate the pH jump topwards the cathode

region and be able to break the organometalic complexes in samples where organic matters are

present. EDTA, a chelating agent that is readily available and environmentally benign and does not

interact with soils, seems to be the best enhancing agent, especially to break the organometalic

complexes. Many of the chelating agents other than EDTA are ionic and can, in principle, be

introduced into the soil by ionic migration. Allen and Chen [48] have shown that EDTA is an

excellent solubilizing agent for many metals including Pb and Zn. It is of interest that EDTA has

been used medically to promote removal of lead from the human body and also as an additive to

render floor polishes with zinc binders amenable to detergent washing [41].

EDTA is a tetraprotic acite abbreviated as H4Y where Y denotes the ethylenediamine-

tetraacetate ion EDTA4-. It is slightly solution in water and the four stepwise dissociation constants

of the parent acid to yield H3Y-, H2Y2-, HY3- and Y4- ions are 1.00 × 10-2, 2.16 × 10-3, 6.92 × 10-7 and

5.50 × 10-11, respectively [48]. It implies that H2Y2- and HY3- species are major EDTA anions

adsorbed [107]. Each EDTA4- ion can attach to a metal ion at six different sites since each of four

acetate groups and the two nitrogen atoms have free electron pairs available for coordinate bond

formation as shown in Fig. 3 [118].


doi: 10.5599/jese.2014.0054 309

Fig. 3. Configuration of metal-EDTA complexes

Unless the pH is very high, the EDTA will not be completely deprotonated. In fact, this is the

reason for the high solubility of metal-EDTA complexes. The complexation of metals by EDTA is

dependent on pH. With a metal ion M, it can form a complex MY, a protonated complex MHY, a

hydro complex MY(OH)n and a mixed complex of the form MYX where X is a unidentate ligand.

Field Applications

In most practical applications of electrokinetics, the anodes are iron or aluminum rods and the

cathodes are steel tubes. Sometimes graphite electrodes are also used for both anodes and

cathodes. Lageman et al. [119] reported the results of field applications in the Netherlands. These

studies demonstrated about 60 % of Zn removal at a concentration of 70 g g-1 from sandy clay

soils; 80 % of As removal at a concentration of 90 g g-1 from heavy clayey soils and 75 % of Pb

removal at a concentration of 340 g g-1 from dredged sediment. The energy expenditure ranged

from 60 to 220 kWh m-3 of soil processed. Banerjee et al. [120] applied the electrokinetic

extraction process in conjunction with the pump-and-treat method in a abandoned industrial

hard-chrome plating facility superfund site in Corvallis, Oregon, USA. Their study demonstrated

that chromium removal slightly increased, but they didn’t provide any numerical value of removal

efficiency. They primarily concluded that ion migration plays a significant role in the

decontamination process. In another field study conducted at Stadskanaal, The Netherlands [121],

it is reported that at an energy expenditure of 20 kWh m-3 of soil, Pb concentration reduced to

120 mg kg-1, Cd 150 mg kg-1, and Zn 320 mg kg-1; at 65 kWh m-3 of soil, Pb concentration reduced

to 90 mg kg-1, Cd 50 mg kg-1, and Zn 120 mg kg-1; and at 180 kWh m-3 of soil, Pb and Zn

concentrations reduced to less than 10 mg kg-1and Cd less than 2 mg kg-1. In all cases the initial

concentrations of Pd, Cd and Zn were 210 mg kg-1, 300 mg kg-1, and 480 mg kg-1, respectively.

However, a number of problems not encountered in the laboratory studies arose in the field trails,

e.g., presence of unexpected large objects (> 10 cm) buried in the soil.

Summary and Recommendation

An overview and concept of electrokinetic extraction processes and their potential applications

in geotechnical and geoenvironmental engineering have been reviewed and presented.

Historically, the success of electrokinetics in soil restoration and decontamination in terms of

inorganic contaminants (i.e. heavy metals) has demonstrated its ability to be one of the most cost

effective and viable in-situ remediation processes compared to the conventional remediation

technologies such as soil washing, ligand extraction, vacuum extraction, thermal desorption,

hydraulic fracturing, biotreatment, immobilization by encapsulation, and placement of barrier

-OOCH2C H H CH2COO- N C C N

-OOCH2C H H CH2COO- Mn+


310

systems. Based on the literature review and researches, it is obvious that the field application of

electrokinetic technology to remediate heavy metal contaminated soils /sediments is very limited

and site specific. Additional laboratory studies and more pilot- and full-scale information from field

applications are critical to the further understanding of the technology and to customize the

process in different field conditions.

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doi: 10.5599/jese.2014.0047 315



www.jESE-online.org


Influence of ceramic separator’s characteristics on microbial fuel cell performance

Anil N. Ghadge, Mypati Sreemannarayana, Narcis Duteanu* and Makarand M. Ghangrekar

Department of Civil Engineering, Indian Institute of Technology, Kharagpur -721302, India *University “Politehnica” of Timisoara, Industrial Chemistry and Environmental Engineering, 2 Victoria Sq., 300006 Timisoara, Romania

Corresponding Author: [email protected]; Tel.: +91-3222-283440

Received: January 27, 2013; Revised: February 15, 2014; Published: December 6, 2014

Abstract This study aimed at evaluating the influence of clay properties on the performance of microbial fuel cell made using ceramic separators. Performance of two clayware microbial fuel cells (CMFCs) made from red soil (CMFC-1) typically rich in aluminum and silica and black soil (CMFC-2) with calcium, iron and magnesium predominant was evaluated. These MFCs were operated under batch mode using synthetic wastewater. Maximum sustainable volumetric power density of 1.49 W m-3 and 1.12 W m-3 was generated in CMFC-1 and CMFC-2, respectively. During polarization, the maximum power densities normalized to anode surface area of 51.65 mW m-2 and 31.20 mW m-2 were obtained for CMFC-1 and CMFC-2, respectively. Exchange current densities at cathodes of CMFC-1 and CMFC-2 are 3.38 and 2.05 times more than that of respective anodes, clearly indicating that the cathodes supported much faster reaction than the anode. Results of laboratory analysis support the presence of more number of exchangeable cations in red soil, representing higher proton exchange capacity of CMFC-1 than CMFC-2. Higher power generation was observed for CMFC-1 with separator made of red soil. Hence, separators made of red soil were more suitable for fabrication of MFC to generate higher power.

Keywords Cation exchange capacity; Coulombic efficiency; Charge transfer resistance; Charge transfer coefficient; Exchange current density; Power density; Wastewater treatment

Introduction

Recently considerable attention is being paid on the two major problems of the world, which

are namely maintaining quality of water body and energy crisis. Solution to these problems could

be provided by microbial fuel cell (MFC) to treat organic matter present in wastewater and



J. Electrochem. Sci. Eng. 4(4) (2014) 315-326 CERAMIC SEPARATOR IN MICROBIAL FUEL CELL

316

simultaneously produce bio-electricity [1-5]. Although, considerable progress has been achieved in

the performance of a MFC in the past ten years, one of the main challenge for commercializing

scalable MFCs is the high cost and low mechanical strength of the separator materials used for

fabrication of this device. Tian et al. [6], demonstrated that placing of an anaerobic membrane

filtration process sequentially with an MFC accomplished efficient nutrients removal with low

propensity of membrane fouling. It was reported that the use of poly(tetrafluoro-ethylene) (PTFE)

layered activated charcoal electrode and Zirfon® as separator, improved MFCs performance and

can be used to replace costly polymeric membrane and expensive catalyst in MFCs [7]. Proton

exchange membrane (PEM) such as Nafion [8], nano-composite membrane made of sulfonated

polymer (ether ether ketone) and Montmorillonite Clay [9], nano-composite membranes of Nafion

and montmorillonite clay [10] were used in the MFCs to separate anodic chamber from cathodic

chamber. However, these polymeric membranes or composite membranes are costly; hence, limit

the practical application of the MFCs.

Ceramic membranes are found to be promising materials in MFCs because of their low

production cost and better structural strength, thus, providing an alternative for the costly

polymeric membranes [11-15]. Application of such ceramic membranes in MFC has been practiced

since last ten years and its utility has been demonstrated through different studies. This was the

first attempt where Park and Zeikus [11] developed a porcelain septum separator for single

chambered MFC using 100% Kaolin and found comparable performance with that dual chambered

MFC. A three layered cathode composed of a cellulose acetate film, a ceramic membrane, and a

porous graphite plate to create a single chamber MFC that linked with solar cell to enhance power

generation [12]. Behera and Ghangrekar [13] studied the effect of different thickness of such

ceramic membrane on performance of the dual chambered MFC, and reported better power

output for MFC having smallest thickness of the membrane. Use of terracotta pot for making

single chamber MFC, after coating outer surface with conductive graphite paint, demonstrated

Coulombic efficiency of 21 ± 5 % with power density of 33.13 mW m-2 [14]. More recently, Winfield

et al. [15] compared performance of MFCs made from terracotta and earthenware by considering

wall thickness, porosity and cathode hydration. More porosity of earthenware proved to be the

better material compared to terracotta. However, these studies do not include the effect of soil

pH, conductivity and cation exchange capacity (CEC) on the performance of MFCs.

For effectual use of such ceramic membranes, made from clay minerals, they should have

higher cation exchange capacity. The existence of the pH dependent charge portion of the cation

exchange capacity of soils is widely accepted for many years [16]. Electrical conductivity of the soil

is the measure of salt concentration in the soil solution. Bulk electrical conductivity of soil is

generally assumed to be dominated by the electrical conductivity of the soil solution, with perhaps

a small contribution from surface charges associated with soil solids [17].

In MFCs, the rate of proton consumption at the cathode is often higher than the transfer rate

through the membrane [18,19]. Hence, for enhancing power generation of this device the

separator used should offer higher rate of proton/cation transfer. The transfer of protons from a

protonated species to an uncharged molecule at the surface of the clay mineral is an important

process [20]. The soil used for making ceramic separator in MFC participates in exchange of

cations from anodic chamber to cathodic chamber. Protons released during the oxidation of

organic matter from the anodic chamber are being adsorbed on to the surface of the soil by

replacing the loosely held cations. The layered silicate clay minerals like smectite clays, show

attractive hydrophilic properties and good thermal stability at high temperature [21]. The layered

A. N. Ghadge at al. J. Electrochem. Sci. Eng. 4(4) (2014) 315-326

doi: 10.5599/jese.2014.0047 317

silicates commonly used for proton exchange membrane fuel cell applications are montmorillonite

made of silica tetrahedral and alumina octahedral sheets which has advantageous hygroscopic

properties [22,23]. The cations Ca2+, Mg2+, K+ and Na+ are called the base cations and H+ and Al3+

are called acidic cations. The acidity of the soil is the amount of the total cation exchange capacity

(CEC) occupied by the acidic cations [24]. More than proton, this cation migration also affects the

performance of cathode, hence overall performance of MFC.

Porosity of the soil represents the hydraulic conductivity which depends upon the pore throat

radii of clay materials. Typically clays have very low hydraulic conductivity due to their small throat

radii. For MFC made with such clayware separator, different soil porosities play a vital role in the

seepage of substrate from anodic to cathodic chamber [25]. Apart from loss of fuel, this may lead

to the availability of organic matter at higher concentration on the cathode, supporting

heterotrophic bacterial growth on cathode and thereby reducing cathode potential. Under such

circumstances, the cathode often gives negative potentials (vs. Ag/AgCl), than the positive

potential it is expected to give, while using oxygen as an electron acceptor [26]. Hence, hydraulic

conductivity and cation exchange capacity are the important properties of the materials to be

selected as a separator in MFC.

The objective of this study was to investigate the effect of different soil properties like pH,

conductivity, porosity, cation exchange capacity on the performance of MFCs having ceramic

separators made from two different soils. In addition, the electrode reaction kinetics was

investigated for assessing performance of these MFCs.

Experimental

Construction of Microbial Fuel Cell

The study was carried out using dual-chambered Clayware Microbial Fuel Cells (CMFCs). The

anodic chambers of these CMFCs were made up of baked clayware pot and the wall material of

the pot (about 5 mm thick) itself acted as a separator allowing transfer of protons from anode to

cathode. The pots were made from the red soil (typically rich in aluminum and silica) in CMFC-1

and black soil (rich in calcium, iron and magnesium predominant) in CMFC-2. The anodic chamber

of the CMFC-1 and CMFC-2 had a liquid volume capacity of 550 ml and 700 ml, respectively.

Cathodic chambers in both the MFCs were made up of plastic container having 5 liter capacity.

Although there is difference in anodic chamber volume of both the MFCs, however, cathodic

chamber volume of 5 litre, which was kept same in both the MFCs. It is important to note here

that the rate of proton transfer largely depends on the separator area to anodic chamber volume

ratio (Sa/v). In the present study, this ratio was 83.3 and 86.5 m2/m3, respectively, for CMFC-1

(separator made of red soil) and CMFC-2 (separator made of black soil), which indicates that there

was no significant difference in Sa/v ratio. Carbon Felt (Panex®35, Zoltek Corporation) with 230

cm2 and 261 cm2 projected surface areas were used as cathode in CMFC-1 and CMFC-2,

respectively. Anodes in CMFC-1 and CMFC-2 were made from stainless steel mesh having total

surface area of 268 cm2 and 304 cm2, respectively. An aquarium aerator was inserted at the

bottom of cathodic chamber to supply air continuously with an aquarium air pump (SOBO

Aquarium Pump, China). The connections between two electrodes were made with concealed

copper wire through external resistance of 100 Ω.


318

Inoculation and operation of CMFCs

Anaerobic mixed sludge collected from septic tank was used as an inoculum in the anodic

chamber of the CMFCs. When mixed anaerobic sludge is used as source of inoculum, it contains

both electrogenic as well as non-electrogenic (mostly methanogenic) bacteria. In the anodic

chamber of MFC, it is necessary to dominate electrogenesis to obtain higher Coulombic efficiency.

Methanogens in the MFCs compete for substrate and electrode space with electrogenic bacteria

and reduce the power output. Therefore, the inoculum sludge was given a heat pre-treatment

(heated at 100 °C for 15 min) to suppress methanogens and required amount of sludge was added

to the anodic chamber [27]. Synthetic wastewater containing sodium acetate as a source of carbon

with chemical oxygen demand (COD) of about 3000 mg L-1 was used in this study. The sodium

acetate medium was prepared by adding 3843 mg L-1 CH3COONa, 4500 mg L-1 NaHCO3,

954 mg L-1 NH4Cl, 81 mg L-1 K2HPO4, 27 mg L-1 KH2PO4, 750 mg L-1 CaCl2.2H2O, 192 mg L-1

MgSO4.7H2O and trace metals like Fe, Ni, Mn, Zn, Co, Cu, and Mo as per the composition given by

[27]. The feeding frequency of 5 days was adopted. These CMFCs were operated at temperatures

varying from 33 to 37 °C under batch mode. The pH of tap water used as catholyte remained in the

range of 8.2-8.5; whereas, anolyte pH was in the range of 7.1-7.4.

Analysis and calculations

The pH and conductivity of anolyte and catholyte was measured using pH meter (Cyber Scan pH

620) and TDS meter (Cyber Scan CD 650, Eutech instruments, Singapore), respectively. COD

concentrations were measured according to APHA standard methods [28], using closed reflux

method. The performance of CMFCs was evaluated in terms of voltage (U) and current (I) mea-

sured using a digital multimeter with data acquisition unit (Agilent Technologies, Malaysia) and

converted to power according to P = UI, where P = power, W; I = current, A; and U = voltage, V.

Power density and power per unit volume were calculated by normalizing power to the anode

surface area and net liquid volume of anodic chamber, respectively. The current density id was

calculated using

d

ext d

Ui

A

R (1)

where, Rext is the external resistance (Ω) and Ad (m2) is the surface area of the anode. Polarization

studies were carried out by varying the external resistance from 10000 to 10 Ω using the resis-

tance box (GEC 05 R Decade Resistance Box) and cell voltages (U) were recorded. Internal resis-

tance of the CMFCs was measured from the slope of the line from plot of voltage versus current

[29]. Columbic Efficiency (CE) was determined by integrating the current measured over time, t,

and compared with the theoretical current on the basis of COD removal and calculated as [1]:

0

d

=

t

M I t

CEFbV COD

(2)

where, V is the volume of the anodic chamber of MFC; M = 32, molecular weight of oxygen; F,

Faraday’s constant = 96485 C mol-1; b = 4, the number of electrons exchanged per mole of oxygen;

ΔCOD is the difference in the influent and effluent COD for time t.


doi: 10.5599/jese.2014.0047 319

Analysis of the soil properties

The pH and the conductivity of the soil samples were measured according to the Indian

Standard method of test for soils. Indian standard IS: 2720 (Part 26) – 1987 was used for determi-

nation of pH value [30] and conductivity of the soil was measured according to IS 14767: 2000

[31]. Cation exchange capacity of the soil was measured according to Indian standard, IS: 2720

(Part 24) – 1976 [32]. Chemical constituents for the red and black soils were obtained through the

X-Ray Fluorescence (XRF) analysis. Porosity of the soil was indirectly measured from the

percentage water absorbed by the clayware pot made from respective soils after immersing in

water for 24 hours.

Reaction kinetics at electrodes

Tafel plot, as derived from equation (3) [33], was employed to measure the reaction kinetics for

working electrode (anode and cathode) and Ag/AgCl was used as the reference electrode. The

reference electrode was placed in the working chamber during the measurements.

0

lnF

RT

i

i

(3)

where i0 is exchange current density, i is the electrode current density (mA m-2), is the electron transfer coefficient, R is the ideal gas constant (8.31 J mol-1 K-1), F is the Faraday’s constant (96,485 C mol-1), T is the absolute temperature, K and η is the activation overpotential. Purpose of using

Tafel plot is to calculate the i0 and value. The i0 is a fundamental parameter in the rate of electro-oxidation or electro-reduction of a chemical species at an electrode at equilibrium. The charge transfer resistance (Rct) was calculated from the following equation:

ct

0

RTR

nFi (4)

where, n is the number of electrons.


Physico-chemical properties of the soil used in CMFCs

The soils used for manufacturing the pots showed different pH. The electrical conductivity and

cation exchange capacity of the red soil is higher than that of the black soil, indicating usefulness

of the former in the clayware separator application (Table 1). However, the porosity of the pot

made from black soil was higher than the pot made from red soil. Higher porosity may allow the

anolyte to come to the cathode resulting in not only the physical substrate loss but also it will

allow oxygen to penetrate in anodic chamber, reducing Coulombic efficiency of the system due to

direct oxidation of the substrate.

Table 1. Physical, chemical and electrical properties of red and black soil used for making separators

Sl. No Soil properties Red soil (CMFC-1) Black soil (CMFC-2)

1 pH 7.4 8.5

2 Porosity, % 11.6 17.6

3 Electrical conductivity, mS cm-1 2.403 0.045

4 Cation exchange capacity (CEC), mmol (kg soil)-1 125 20


320

Wastewater treatment

After inoculating the anodic chamber of the CMFCs with heat pretreated anaerobic mixed

consortia, synthetic feed was supplied and wastewater treatment performance of the CMFCs

under different feed cycles was observed. Average COD removal efficiency of 78.9 ± 3.9 % and

89.6 ± 3.2 % was observed in CMFC-1 and CMFC-2, respectively. COD removal efficiency in the

CMFC-2 was higher than the CMFC-1. It was observed that porosity of clayware pot used in

CMFC-2 was 52 % higher (Table 1) than that of CMFC-1, because of which probably it has

permitted more diffusion of oxygen from cathodic chamber to anodic chamber to support aerobic

oxidation of fraction of substrate present in anodic chamber to establish higher COD removal

efficiency. The oxygen diffusion coefficient in the range of 5.38 x 10-6 to 6.67 x 10-6 cm2 s-1 is

reported in early studies for this clayware separator by Behera and Ghangrekar [13]. In addition,

due to more porosity of separator used in CMFC-2, exchange of water molecules due to osmosis

across the membrane might have diluted the anolyte, resulting in higher COD removal rate.

Electricity generation

Performance of CMFCs was evaluated by measuring the open circuit voltage and operating

voltage. The current and voltage gradually increased with time of operation. The maximum voltage

across 100 Ω resistance of 286 mV and 280 mV was observed in CMFC-1 and CMFC-2, respectively.

CMFC-1 generated a maximum sustainable power density (normalized to the anode surface area)

and volumetric power (normalized to the working volume of anodic chamber) of 30.5 mW m-2 and

1.49 W m-3 (Fig. 1), respectively; whereas, CMFC-2 generated power density of 25.7 mW m-2 and

volumetric power of 1.12 W m-3. The power produced by CMFC-1, made from red soil, was 1.33

times higher than the CMFC-2 wherein the separator was made from black soil with lower CEC and

electrical conductivity. It is interesting to note here that in spite of having higher separator area

and more liquid volume (anodic chamber) for CMFC-2, it generated less power compared to

CMFC-1.

Figure 1. Volumetric power density of CMFC-1 and CMFC-2

0 2 4 6 8 10 12 14 16 18 20 22

0.4

0.6

0.8

1.0

1.2

1.4

1.6

Po

we

r d

en

sity

, W m

-3

Time, days MFC-1 MFC-2


doi: 10.5599/jese.2014.0047 321

Effect of chemical properties of soil used for making separator on electricity generation

The CEC of red soil is 6.25 times (Table 1) higher than black soil, indicating more number of

exchange sites are available for the transfer of cations in red soil. Due to availability of more

exchange sites in CMFC-1, better transfer of the protons occurred to improve the power

generation in CMFC-1 compared to CMFC-2. In addition, the XRF data (Table 2) confirms that the

aluminum content of the red soil is more than the black soil which makes the red soil more acidic

than black soil. The pH of the red soil (Table 1) was lower than black soil confirming that the red

soil is more acidic and has high capacity to hold the H+ ions which improved the performance of

CMFC-1 in terms of power generation.

Electrical conductivity is the measure of salt concentration in the soil solution. Soils high in

smectite often exhibit high electrical conductivity due to water associated with the clays. The soil

with high montmorillonite mineral can act as better proton exchange material due to its

hydrophilic nature and the high cation exchange capacity [34]. The conductivity of soil used as

separator in CMFC-1 is almost 53.4 times (Table 1) more compared to CMFC-2, authenticating

utility of the red soil for making separator to harvest more power from the CMFCs.

Table 2. Chemical compounds present in red and black soil

Sl. No

Compound Content, % Sl.

No Compound

Content, %

Red soil Black soil Red soil Black soil

1 Na2O 3.95 0.273 14 Co 0.406 0.441

2 MgO 0.654 3.86 15 Ni 0.004 0.006

3 Al2O3 26.3 21.6 16 Cu 0.274 0.282

4 SiO2 57.5 53.4 17 Zn 0.005 0.023

5 P2O5 1.13 0.204 18 Ga 0.001 0.001

6 SO3 0.258 0.162 19 Rb 0.007 0.008

7 K2O 1.78 0.798 20 Sr 0.004 0.017

8 CaO 0.791 10.4 21 Y 0.004 0.002

9 Fe2O3 4.70 6.75 22 Zr 0.013 0.010

10 Cl 1.45 0.071 23 Nb 0.001 0.0005

11 Ti 0.658 1.45 24 Ba 0.012 0.020

12 Cr 0.01 0.01 25 Ce 0.021 0.023

13 Mn 0.067 0.093 26 Pb 0.003 0.002

Coulombic efficiency

Coulombic efficiency compares the recovery of the coulombs through the external circuit

against theoretical coulombs that is present in the organic matter. CMFC-1 showed average CE of

7.69 ± 1.52 %, whereas in CMFC-2 it was 6.39 ± 1.40 %. Higher CE of CMFC-1 than CMFC-2 might

have been due to the difference in the CEC of red and black soil and also due to more diffusion of

oxygen in case of black soil due to high porosity. In MFCs higher CE is reported with pure inoculum

culture and with synthetic wastewater [35].


322

Polarization and Internal resistance

Polarization curve helps to understand the performance of MFC in terms of power generation

and internal resistance. It represents the cell voltage and power density as a function of the

current density. Figure 2 shows power and polarization curves obtained using variable resistor box

for CMFC-1 and CMFC-2.

Figure 2. Polarization curve for CMFC-1 and CMFC-2

During polarization, the maximum power density observed for CMFC-1 was 51.65 mW m-2

(E = 0.204 V, Rext = 30 Ω) and that of CMFC-2 it was 31.20 mW m-2 (E = 0.217 V, Rext = 50 Ω). This

indicates that the higher CEC of red soil supported better proton transfer from the anode to the

cathode in CMFC-1. Conversely, lower power output observed in CMFC-2 could be attributed to

the lesser CEC of black soil used for making separator (Table 1). It is well documented that the

cation exchange capacity of soil plays vital role in the proton transfer mechanism in soil [36].

Internal resistances of CMFC-1 and CMFC-2 measured from the slope of the plot of voltage vs.

current were 36 Ω and 56.5 Ω, respectively. In the region of low current density (Fig. 2) rapid

voltage drops were observed in both the MFCs, and in the region of high current density, voltage

decreased linearly at lower rate. Lower proton transfer rate and low conductivity of black soil used

for making separator of CMFC-2 increased the internal resistance.

Electrode Potential

Electrode potentials represent the energy level of the electrons at anode and cathode.

Electrons move from area of higher potential energy to area of lower potential energy. As the

anode has a higher potential energy so electrons move from anode to cathode through an external

circuit. During polarization, cathode of CMFC-1 well supported for the reduction reaction up to

0.5 mA current at 1000 Ω external resistance. However, the cathode potential of CMFC-2 dropped

to zero (vs. Ag/AgCl) at 0.25 mA current at 2100 Ω external resistance, showing inefficiency of

cathode for reduction reaction at higher current (Fig. 3).

0 50 100 150 200 250 300 350 400

0

100

200

300

400

500

600

700

CMFC-2 Cell voltage CMFC-1 Cell voltage

CMFC-2 Power density CMFC-1 Power density

Current density, mA m-2

Volt

age,

mV

0

5

10

15

20

25

30

35

40

45

50

55

Pow

er d

ensi

ty, m

W m

-2


doi: 10.5599/jese.2014.0047 323

Figure 3. Change of the cathode and anode potentials during polarization in CMFC-1 and CMFC-2

The open circuit potentials (OCP) for anode observed before polarization (vs. Ag/AgCl) for

CMFC-1 and CMFC-2 were -610 mV and -580 mV, respectively. During polarization, increase in

anode potentials was observed in both the MFCs due to transfer of electrons from anode to

cathode, thus positive overpotential was observed. However, the anode potentials during polariz-

ation in both the MFCs were only slightly increased, indicating better stability of the anodes.

Electrode Kinetics

Tafel plot analyses were carried out to determine the exchange current density (i0), charge

transfer coefficient () and charge transfer resistance (Rct). The values obtained from these

analyses are summarized in Table 3. Based on the Tafel-type linear equation obtained from the

graphs (Figs. 4A, 4B), the slope is F/RT and the y-axis intercept is the logarithm of the exchange

current.

For anodic reaction at 25°C, the slope of Tafel plot is

b = 0.059 / 1 - (5)

and at the same time the slope for cathodic reaction is

b = 0.059 / (6)

The value of i0 represents the rate of exchange current density at equilibrium state when the

reaction overpotential is zero. Higher the exchange current (i0) faster is the reaction rate, resulting

in a lower activation energy barrier of forward reaction [37]. The electrode materials used in both

the CMFCs were same, however different reaction kinetics at electrodes was observed. The

reactions at cathodes were faster than anode. Presence of very high actual surface area of the

carbon felt material resulted in producing a low cathodic overpotential [38]. Comparing the i0

values for the reduction reactions at cathode of CMFC-1 and CMFC-2, the CMFC-1 with separator

made from red soil had better performance. It indicates that the reactions at cathode of CMFC-1

were faster; might be due to the higher transfer of H+ ions and other cations in CMFC-1, enhancing

the reaction rates at cathode. Comparing the anodes of both the CMFCs, reactions at the anode of

CMFC-2 were slightly faster than CMFC-1. Apart from the differences in the CEC, the reactions at

cathode of CMFC-2 were slower than CMFC-1. This could be probably due to the higher porosity of

0 2 4 6 8 10 12-600

-500

-400

-300

-200

-100

0

100

200

300

Ele

ctro

de

po

ten

tial

, mV

Current, mA

CMFC-1 Cathode potential CMFC-1 Anode potential

CMFC-2 Cathode potential CMFC-2 Anode potential


324

the clayware separator used in CMFC-2, due to which the substrate exchange occurred and oxygen

supplied in the cathodic chamber was utilized by the substrate. The exchange current densities of

CMFC-1 and CMFC-2 cathodes were 3.38 and 2.05 times more than that of the respective anodes,

clearly indicating that the reaction at cathode was much faster than anode.

A

B

Figure 4. Tafel plots for A - cathode of CMFC-1 and CMFC-2, and

B - anode of CMFC-1 and CMFC-2

According to the Butler–Volmer model of electrode kinetics, the charge transfer coefficient ()

is used to describe the symmetry between the forward and reverse electron transfer steps and the

magnitude of ranges in value from 0 to 1. Charge transfer coefficient signifies the fraction of the

interfacial potential at an electrode-electrolyte interface that helps in lowering the free energy

barrier for the electrochemical reaction. Lower electron transfer coefficient indicates less

activation energy required for the electron transfer, resulting lower activation loss [37]. Charge

transfer resistance (Rct) represents the capability to resist the transfer of charge from electrode-

electrolyte interface. It is interesting to note that the Rct and (Table 3) of cathodes for both the

MFCs were much lesser than the anode of both the MFCs.

-250 -200 -150 -100 -50 00.0

0.5

1.0

1.5

2.0

Cathode CMFC-1 Cathode CMFC-2

Overpotential, mV

ln (

i / m

A m

-2)

0 20 40 60 80 100 120 140 1602

3

4

5

6

7

8

Overpotential, mV

Anode CMFC-2 Anode CMFC-1

ln (

i / m

A m

-2)


doi: 10.5599/jese.2014.0047 325

Table 3. Tafel analysis of CMFC-1 and CMFC-2

Parameter CMFC-1 CMFC-2

Anode Cathode Anode Cathode

Exchange current density (i0), mA m-2 0.60 2.03 0.74 1.52

Charge transfer coefficient () 0.30 0.032 0.45 0.038

Charge transfer resistance (Rct), Ω m2 10.70 3.16 8.67 4.22

The Rct and values for cathode of CMFC-1 were lower than CMFC-2, which supports that the

clayware membrane made from red soil supported better reaction at the cathode. This is because

of high cation transported from CMFC-1 to the cathode side, increased the rate of electrochemical

transformation with lower electrical energy loss, thus charge transfer coefficient gets reduced.

Lower Rct for anode of CMFC-2 than anode of CMFC-1 indicated that the anode of CMFC-2 was

performing slightly better, as also evident from the exchange current density. However, due to

limitations of the cathodic reactions the overall performance of CMFC-2 was inferior as compared

to CMFC-1.

Conclusions

Properties of the clayware separator such as CEC, pH and electrical conductivity influenced the

performance of MFCs. The power generation of MFC having separator made from red soil was

better than the black soil, due to high CEC, low pH and higher electrical conductivity of the red soil.

Results of Tafel plots showed that lower exchange current density and higher charge transfer

resistance of anodes compared to cathodes, contributed towards more activation loss in both the

MFCs. In spite of similar electrode materials in both CMFCs, variation in electrode kinetics

accentuate effect of properties of separator on the performance of CMFCs. Detailed studies on the

mineral composition of soils are required to enhance the CEC for further improving power

generation of MFC made with such low cost clayware separator. Development of such efficient

and cheaper separator material will help in drastically reducing fabrication cost of MFC for field

implementation.

Acknowledgement: Grants received from Department of Science and Technology, Government of India (File No. DST/TSG/NTS/2010/61) to undertake this work is duly acknowledged.

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