Advanced water treatment electrochemical methods 9780128192283, 0128192283

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Advanced Water Treatment

Electrochemical Methods

Edited by Mika Sillanpää Department of Civil and Environmental Engineering, Florida International University, Miami, FL, United States

Elsevier Radarweg 29, PO Box 211, 1000 AE Amsterdam, Netherlands The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States Copyright © 2020 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein). Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress British Library Cataloguing-in-Publication Data A catalogue record for this book is available from the British Library ISBN: 978-0-12-819227-6 For information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

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Contributors Tuan Pham Anh _ Water & Environment, Amata City Bien Hoa J.S.C, Bien Hoa, Dong Nai, Vietnam Thuy Duong Pham Department of Green Chemistry, Lappeenranta University of Technology, Lappeenranta, Finland Heikki Särkkä Department of Built Environment, Aalto University, Espoo, Finland Marina Shestakova LUT Chemtech, Lappeenranta University of Technology, Lappeenranta, Finland Mika Sillanpää Department of Civil and Environmental Engineering, Florida International University, Miami, FL, United States Mikko Vepsäläinen Mineral Resources, CSIRO, Clayton, VIC, Australia

ix

Chapter

1

Electrocoagulation in the treatment of industrial waters and wastewaters Mikko Vepsäläinen1, Mika Sillanpää2

1

2

Mineral Resources, CSIRO, Clayton, VIC, Australia; Department of Civil and Environmental Engineering, Florida International University, Miami, FL, United States

CHAPTER OUTLINE

List of Publications 2 Abbreviations 3 1. Introduction 4

1.1 Basic concepts and theory of coagulation and flocculation with hydrolyzing metal salts 5 1.1.1 1.1.2 1.1.3 1.1.4 1.1.5

Interface of colloidal particles 5 Stability of colloids in aqueous solutions 8 Destabilization mechanisms of colloids 9 Commonly used metal salt coagulants 10 Applications of metal salt coagulants in industrial raw water and wastewater treatment 12

1.2 Theory of electrocoagulation 1.2.1 1.2.2 1.2.3 1.2.4 1.2.5

16

Main reactions 17 Side reactions 19 Properties of the sludge 21 Treatment parameters 22 Comparison of electrocoagulation and chemical coagulation 27

1.3 Practical considerations of electrocoagulation

29

1.3.1 Constructions of electrocoagulation systems 29 1.3.2 Applications of electrocoagulation 31 1.3.3 Combinations of electrocoagulation and other water treatment technologies 31 1.3.4 Economical and ecological considerations 36

2. Objectives and Structure of the Work

38

2.1 Surface water treatment by electrocoagulation 38 2.2 Wastewater treatment by electrocoagulation 38

3. Materials and Methods

38

3.1 Water samples and chemicals 38 3.2 Water treatment procedure 40 Advanced Water Treatment. https://doi.org/10.1016/B978-0-12-819227-6.00001-2 Copyright © 2020 Elsevier Inc. All rights reserved.

1

2 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

3.3 Analytical methods 41 3.3.1 Chemical analysis 41 3.3.2 Toxicity analysis 42

3.4 Statistical methods

4. Results and Discussion

43

44

4.1 Quality of the statistical models 4.2 Dissolving of electrodes 45

44

4.2.1 Aluminum electrodes 45 4.2.2 Iron electrodes 48

4.3 Change of pH and conductivity 4.4 Surface water treatment 50

48

4.4.1 Organic matter removal 50 4.4.2 Residual metals and turbidity 56

4.5 Wastewater treatment 57 4.5.1 Organic matter removal 58 4.5.2 Sulfide precipitation 61 4.5.3 Resin acids, copper, and toxicity removal 62

5. Conclusions and Recommendations References 66

n

65

LIST OF PUBLICATIONS

This summary is based on the following papers. I M. Vepsäläinen, J. Selin, M. Pulliainen, M. Sillanpää, Combined electrocoagulation and chemical coagulation of paper mill mechanically cleaned water, J. Pulp Pap. Sci. 33 (2007) 233e239. II M. Vepsäläinen, M. Pulliainen, M. Sillanpää, Effect of electrochemical cell structure on natural organic matter (NOM) removal from surface water through electrocoagulation (EC), Sep. Purif. Technol. 99 (2012) 20e27. III M. Vepsäläinen, M. Ghiasvand, J. Selin, J. Pienimaa, E. Repo, M. Pulliainen, M. Sillanpää, Investigations of the effects of temperature and initial sample pH on natural organic matter (NOM) removal with electrocoagulation using response surface method (RSM), Sep. Purif. Technol. 69 (2009) 255e261. IV M. Vepsäläinen, J. Selin, P. Rantala, M. Pulliainen, H. Särkkä, K. Kuhmonen, A. Bhatnagar, M. Sillanpää, Precipitation of dissolved sulphide in pulp and paper mill wastewater by electrocoagulation, Environ. Technol. 32 (2011) 1393e1400. V M. Vepsäläinen, H. Kivisaari, M. Pulliainen, A. Oikari, M. Sillanpää, Removal of toxic pollutants from pulp mill effluents by electrocoagulation, Sep. Purif. Technol. 81 (2011) 141e150.

Abbreviations 3

n

ABBREVIATIONS

AC ANOVA BOD CCF COD DAF DC DHAA DLVO DOC DOE EC EC50 EDL EF FTIR GC-MS HMM HPAC ICP-OES IPA MLR NOM NTU OHP ORP PACl PAF-SiC PASiC PFC PFS PFSiS PICl PLS PDADMAC PSF PtCo PXRD

Alternating current Analysis of variance Biological oxygen demand Central composite face Chemical oxygen demand Dissolved air flotation Direct current Dehydroabietic acid DerjaguineLandaueVerweyeOverbeek (theory) Dissolved organic carbon Design of experiment Electrocoagulation Half maximal effective concentration Electrical double layer Electroflotation Fourier transform infrared spectroscopy Gas chromatographemass spectrometry High molecular mass Highly efficient composite polyaluminum chloride Inductively coupled plasma optical atomic emission spectrometry Isopimaric acid Multiple linear regression Natural organic matter Nephelometric turbidity unit Outer Helmholtz plane Oxidationereduction potential Polymeric aluminum chloride Polyaluminum ferric silicate chloride Polyaluminum silicate chloride Polyferric chloride Polyferric sulfate Polyferric silicate sulfate Polymeric iron chloride Partial least squares Polydiallyl dimethyl ammonium chloride Polysilicate ferric Platinumecobalt scale Powder X-ray diffraction

4 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

RSM SCE SEM-EDS SRB SUVA SVI TOC TSS USD UV/Vis XPS

1.

Response surface model Saturated calomel electrode Scanning electron microscope with energy dispersive spectroscopy Sulfate-reducing bacteria Specific ultraviolet absorbance Sludge volume index Total organic carbon Total suspended solid United States dollar Ultravioletevisible spectroscopy X-ray photoelectron spectroscopy

INTRODUCTION

There is a growth in demand for new water treatment technologies as the world’s population increases and freshwater sources are polluted. Waterborne diseases are still common in developing countries due to the lack of funding or appropriate know-how for water purification. Industry also uses these limited water sources and has to acquiesce to lower quality raw water as a higher proportion of freshwater is required for human consumption. Wastewater treatment technologies used in both municipal and industrial applications have to be further developed to reduce the pollution of receiving waterbodies. Chemical coagulation and flocculation are commonly used as a part of the water purification systems for the removal of pollutants from raw waters and wastewaters. Their main function is to enhance particle separation in the subsequent processes, such as filtration, sedimentation, or flotation. To understand destabilization of particles by coagulants and flocculants, it is crucial to understand the mechanisms which stabilize particles in aqueous solutions. Chemical coagulation and flocculation are used in both industrial and municipal raw water and wastewater treatment systems. They can enhance the removal of several types of pollutants from the water streams. Typical examples of pollutants to be removed are nutrients, toxic heavy metals, and natural organic matter (NOM). The most commonly used coagulants are aluminum or iron salts, such as sulfates and chlorides. These metal salts form various hydrolysis products in the water depending on water chemistry, such as pH and the concentration of anions. Metal cations and hydroxides destabilize colloid pollutants in

1. Introduction 5

water by reducing repulsion forces between the colloids and by entrapping particles in the sludge. Electrocoagulation (EC) has been suggested as an advanced alternative to chemical coagulation in pollutant removal from raw waters and wastewaters. In this technology, metal cations are released into water through dissolving metal electrodes. Simultaneously, beneficial side reactions can remove flocculated material from the water. However, there are also adverse side reactions, such as deposition of salts on the electrode surface, which may cause deterioration of removal efficiency after long operation. As in the case of chemical coagulation with metal salts, aluminum or iron cations and hydroxides are the active compounds in EC. Chemical coagulation and EC have fundamentally similar destabilization mechanisms, and it is therefore important to go through the theory of colloid destabilization with metal salt coagulants, because chemical coagulation has been studied more extensively than EC.

1.1 Basic concepts and theory of coagulation and flocculation with hydrolyzing metal salts Pollutants in raw waters and wastewaters are typically colloidal particles, which are not easily removed with typical filtration, sedimentation, or flotation due to their stability in water. These particles have special properties due to their small size and large total surface area. The properties of the interface of colloidal particles and the stabilization of colloidal particles by hydrolyzing metal salts are discussed in the first section of this work.

1.1.1 Interface of colloidal particles Colloid is a microscopic particle, typically having at least one dimension in the range of 1 nme10 mm, which is dispersed throughout the other substance [1]. This medium where particles are dispersed can be gas, liquid, or solid. The combined surface area of colloids in dispersions is large due to their small size, and therefore surface properties play an important role in their characteristics. Natural waters and typical wastewaters are heterodispersions, having a wide variety of particles with different particle sizes [2]. Stability and destabilization of colloids in solutions is the result of their surface charge. Surface immersed into a solution can attain a charge by

6 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

ionization of surface groups, by ion adsorption, by dissolution of ionic solids, or by isomorphous substitution [1]. Many surfaces contain ionizable functional groups, such as eOH, eCOOH, or eNH2. Surface charge therefore depends on the ionization of these functional groups and consequently on the pH of the solution [3]. Isomorphous substitution occurs when lattice imperfection occurs at the crystal due to the replacement of some atom in the crystal by another ion that has a different amount of electrons, which results in a charged surface. This occurs, for example, in clay particles. Adsorption of ions has an impact on surface charge of the particles. Dissolution of ionic solids can cause a charge on the surface if the dissolution of anions and cations from the solid is unequal. When a charge forms on the surface, it also affects the ions in the surrounding solution. The ions of opposite charge are attracted toward the surface, whereas the ions of the same charge are repelled from the surface. This separation of charges on the particle surface results in the formation of electrical double layer (EDL) presented in Fig. 1.1 [1e4]. The interface of a charged surface has been explained by the models of Helmholtz, Gouy, and Chapman [4]. Stern later combined these models and further developed the model of the EDL. In this model, there is an inner region (Stern layer or Helmholtz layer) and an outer diffusion region (GouyeChapman layer). In inner layer, ions are tightly bound to the surface, whereas in the outer layer, they move about under the influence of diffusion. According to GouyeChapman model, the potential distribution in a flat double layer can be described by Eq. (1.1): d2 j r ¼  dx2 ε

(1.1)

where j is the potential at a point in the diffuse layer versus infinity at the bulk solution, r is the charge density at the same point, and ε is permittivity. In Eq. (1.1), the charge density at the potential j is described by Eq. (1.2). The number of positive and negative ions in the diffuse layer is distributed according to the MaxwelleBoltzmann distribution (Eq. 1.3 for cations and Eq. (1.4) for anions): r ¼ zeðnþ  n Þ

(1.2)


 zej nþ ¼ n0  exp kT

(1.3)

1. Introduction 7

n FIGURE 1.1 Conceptual representation of the electrical double layer. Reprinted from J.C. Crittenden,

R.R. Trussell, D.W. Hand, K.J. Howe, G. Tchobanoglous, Water Treatment e Principles and Design, second ed. (Knovel ebook) John Wiley & Sons, USA, 2005; with permission from John Wiley & Sons.
 zej n ¼ n0  exp  kT

(1.4)

where nþ and n are the respective numbers of positive and negative ions per unit volume at the point where potential is j. n0 is the concentration of ions at the infinity (bulk solution). z is the valency of the ions, e is the charge of electron, k is Boltzmann’s constant, and T is temperature. The potential of the particle surface versus bulk solution is called Nernst potential (j0). Between the inner layer and the outer layer is the outer Helmholtz plane, which has potential, jz, versus bulk solution. However, this potential cannot be directly measured, and therefore a common parameter, which depicts the surface charge of the colloid, is z-potential (jm), which is the

8 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

electrical potential between the plane of shear and the bulk solution. The diffuse part of the double layer is analogous to the plate condenser. Even though in reality the double layer extends to infinity, the DebyeeHückel length, K1, is used to describe the thickness of the double layer.

1.1.2 Stability of colloids in aqueous solutions Colloids are said to be stable in aqueous solutions when their aggregation or sedimentation is so slow that they make virtually stable dispersions. This is important when considering raw water or wastewater treatments because these particles cannot be removed by sedimentation in a reasonable period of time. Stability (or destability) of colloids in water is a balance between the repulsive electrostatic force and attractive, Londonevan der Waals force. The theory of DerjaguineLandaueVerweyeOverbeek estimates these energies of attraction and repulsion [5,6]. The Londonevan der Waals forces of attraction, caused by the permanent or induced dipoles, are important at very short distances. The repulsion force is caused by the overlapping of the EDLs, which leads to higher counterion concentration and hence to repulsion between the particles. The repulsion force is an exponential function of the distance between the particles with a range of the order of the thickness of the EDL, whereas the attraction force decreases as an inverse power of the distance between the particles. There are two minimums in the summation of attraction and repulsion forces between the particles having similar charges. The primary minimum, at a close distance between the particles, is where particles reside after coagulation. There is also a secondary minimum at longer distance, and this is the energy minimum of flocculation. Potential energy at the primary minimum is significantly smaller than at the secondary minimum and hence the particles at the primary minimum are more tightly attached. Between the primary and secondary potential energy minimums, an energy barrier exists which the particles have to overcome with their kinetic energy before coagulation can take place. Detailed calculations of attraction and repulsion forces have been discussed in the articles of Overbeek [7] and Verwey and Overbeek [6]. Ionic strength affects the thickness of double layer. The double layer compresses when ionic strength increases according to Eq. (1.5). In addition to ionic strength, the valence of the ions also affects the thickness of the EDL. Multivalent counterions are concentrated in the double layer more than

1. Introduction 9

monovalent ions according to Eqs. (1.3) and (1.4). According to the SchulzeeHardy rule, the optimum concentration for coagulation for tri-, di-, and monovalent cations is 800:12:1. However, the required ion concentration of multivalent ions is often less than the SchulzeeHardy rule predicts [2]. 1=2  1 2000e2 NA I ¼ 1010 K εε0 kT

(1.5)

where 1/K is Debye length, NA is Avogadro number, I is ionic strength, and ε0 is permittivity in vacuum.

1.1.3 Destabilization mechanisms of colloids As discussed in Section 1.1.2, colloids can be stable in water due to electrostatic repulsions between the particles. In coagulation and flocculation technologies, particles are destabilized with the addition of inorganic or organic chemicals which have an effect on the properties of EDL. Coagulants are chemicals which reduce the repulsive energy between the particles (energy barrier). Therefore, particles can more easily agglomerate at the primary minimum of potential energy. Flocculation occurs when these agglomerated particles attach to each other with a weak bond (the secondary minimum of potential energy). Flocculation aids improve floc properties, such as settleability and filterability. Mechanisms which can destabilize colloidal particles in water are as follows [2,3]:  Compression of EDL. The increase of the concentration of ions in bulk solution compresses EDL, and particles come together more easily as the length of EDL decreases. Optimum destabilization is achieved when z-potential is close to 0 mV. Excess salt concentration does not lead to restabilization of the particles. This mechanism is not employed in water treatment because very high salt concentrations are required for destabilization.  Adsorption destabilization. This occurs when oppositely charged ions or polymers are adsorbed on the surfaces of particles. They reduce surface charge and thus repulsive force between the particles. Destabilization occurs typically at z-potential values close to 0 mV. Too high a coagulant dose can lead to the restabilization of the particles because the charge is reversed.  Interparticle bridging. Polymerized metal coagulants or organic chemicals can form bridges between the particles. Polymer can adsorb on the particle surface by several mechanisms, such as chargeecharge

10 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

interactions, dipole interactions, hydrogen bonding, or van der Waals interaction. When one polymer chain adsorbs on multiple particles, bridging occurs and molecular weight increases. Bridging is a typical mechanism with long-chained, high-molecular weight organic polymers. The z-potential of the particles which are destabilized by bridging mechanism is not typically close to zero. Restabilization can occur when the surface is completely covered by the polymer.  Precipitation and enmeshment mechanism. This destabilization mechanism is typical with the high concentrations of metal salts in near neutral pH. In this pH, metal salts, such as alum or ferric sulfate, form insoluble hydrolysis products and polymerize. Colloidal particles can then be enmeshed into these sweep flocs. The required dosage of coagulant is virtually independent of the type and concentration of colloids in the solution. Adsorption destabilization and interparticle bridging mechanisms are susceptible to restabilization when excess coagulant is added. In the bridging mechanism, restabilization occurs when the adsorption sites of the particles are occupied and no further bridging can occur. In the overcharging phenomenon, the inner part of the EDL carries more countercharge than exists on the surface of the particle [8,9]. In this case, potentials jz and jm are reversed from negative values to positive ones in typical colloids existing in raw water and wastewaters. However, double layer compression and precipitation mechanisms can again destabilize particles when even higher coagulant concentrations are used. In typical coagulation processes of raw water and wastewater treatment plants, there are several mixing stages. At the point of coagulation, fast mixing is required to properly disperse chemicals into the water stream and promote particle collisions. After this short fast mixing stage, water goes into slow mixing stage (flocculation) where induced velocity gradients provide particles opportunities for contact and further particle size growth. However, if the water velocity in the flocculation is too high, it may cause floc breakdown. There are several types of mixers for fast and slow mixing stages and required conditions depend, e.g., on pollutants to be removed and primary mechanism of destabilization. Detailed description can found in Refs. [2,3].

1.1.4 Commonly used metal salt coagulants Aluminum and iron metal salts are used in raw water and wastewater treatment. Both metals can form multivalent ions, Al3þ, Fe2þ, and Fe3þ and various hydrolysis products [10]. Fe(II) is poor coagulant itself and is

1. Introduction 11

typically oxidized to Fe(III) form during the coagulation process to obtain higher efficiency. The most commonly used metal salts are simple aluminum and iron sulfates and chlorides. Metal cations go through a series of hydrolytic reactions depending on the pH of the solution and mononuclear (Fig. 1.2 and Eq. 1.6) and polynuclear hydroxides form in the solution. Neutral amorphous metal hydroxides, Al(OH)3 and Fe(OH)3, are poorly soluble species. Distribution of mononuclear species of metals can be presented as solubility diagrams. Me3þ ðaqÞ þ nH2 O4MeðOHÞ3n þ nHþ ðaqÞ n

n FIGURE 1.2 Concentrations of soluble monomeric hydrolysis products of Fe(III) and Al(III) in

(1.6)

equilibrium with the amorphous hydroxides at zero ionic strength and 25 C. Reprinted from J. Duan, J. Gregory, Coagulation by hydrolysing metal salts, Adv. Colloid Interface Sci. 100e102 (2003) 475e502; with permission from Elsevier.

12 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Polynuclear species, e.g., Al13 [AlO4Al12(OH)7þ24], are effective in destabilization of colloids due to their large size and high positive charge [10,11]. In prehydrolyzed metal coagulants, such as polyaluminium chloride (PACl) and polyferric chloride, these polymerized species are preformed by partially neutralizing metal salt solution to different basicity ratios. These coagulants have less temperature or pH dependence, and their alkalinity consumption is lower. Recently, new types of coagulants have been developed where prehydrolyzed aluminum or iron coagulants are supplemented with silicates or organic polymers [12]. Commonly used and recently studied metal and composite coagulants are presented in Table 1.1. A more detailed description of these coagulants can be found in our review article [12]. In addition to pH, competing anions also have an influence on the composition of metal hydroxides. The effect of anion depends on its electronegativity and tendency to react with metal ion or the positively charged sites of metal hydroxide precipitates [3]. Sulfate ions are weakly basic and strongly coordinated with metal ions. They broaden the pH zone of destabilization to acidic side. Phosphates shift the optimum pH of destabilization to the acidic side. Chlorides and nitrates have only a slight effect in high concentrations.

1.1.5 Applications of metal salt coagulants in industrial raw water and wastewater treatment Metal salt coagulants are used in industrial raw water treatment processes and wastewater treatment plants to remove various pollutants. Typical contaminants in surface waters that can be removed by coagulation are, for example, NOM, mineral substances, and microorganisms. The variety of pollutants found in industrial wastewaters is enormous, and only some examples of applications could be given in this study. NOM is one of the most important pollutants found in surface waters. Basically, NOM is a complex mixture of different organic materials, such as bacteria, viruses, humic acids, fulvic acids, polysaccharides, and proteins. NOM can be divided into hydrophilic and hydrophobic fractions [52,53]. The hydrophilic fraction of NOM is composed mostly of aliphatic carbon and nitrogenous compounds, such as carboxylic acids, carbohydrates, and proteins, whereas the hydrophobic NOM primarily consists of humic (Fig. 1.3) and fulvic acids [54]. Humic substances are highly heterodisperse macromolecules with high molecular weights. High molecular weight hydrophobic NOM fractions are more easily destabilized than low molecular weight hydrophilic fractions [3,52,53,55]. Hydrophilic fractions can be partially removed when sufficiently high coagulant concentrations are used at low pH. In these conditions, metal ions form

1. Introduction 13

Table 1.1 Overview of the Metal Salt and Composite Coagulants Used in Recent Research Studies [12]. Coagulant

Features

Positive

Negative

Reference

Alum, aluminum chloride

Trivalent aluminum ions are released into a solution from the respective salt. They are hydrolyzed and form soluble complexes possessing high positive charges. Coagulation efficiency depends on, e.g., coagulant dose, mixing, pH, temperature, and particle and natural organic matter (NOM) properties. During coagulation, the most effective range of pH is suggested to be 5.0e6.5. Ferric salts hydrolyze similarly as aluminum salts when added to water and form different hydrolysis products. Effectiveness of coagulation depends on the same factors as during alum salt coagulation. The most effective range of pH is suggested to be pH 4.5 e6.0.

Stable, easily handled, readily soluble. Better turbidity removal than with ferric salts in many cases. Can be more effective than ferric in low doses. Higher color removal efficiency.

Relatively high coagulant residuals in the finished water in some cases. Possible link with Alzheimer’s disease. Ferric salts have been noted to be better at removing NOM than aluminum salts in many investigations. High alkalinity consumption. Sulfate and/or chloride in finished water increases corrosivity.

[11,13e26]

Ferric salts have been noted to be better at removing NOM than aluminum salts in many investigations. Especially, the removal of middlesized NOM fractions is noted to be more efficient. Not as sensitive to temperature changes as alum.

[15,16,20,25, 27e29]

Made by partially neutralized (prehydrolyzed) aluminum chloride. Enhanced amounts of high-charged, moderatemolar-mass hydrolysis species, e.g., Al13.

Less temperature and pH-dependent than alum salts. Lower alkalinity consumption. Better NOM removal capacity than alum in many cases. Lower dose requirement and less sludge produced. Lower residual aluminum in treated water.

Ferric-based coagulant acid strength and associated optimized flocculation pH ranges can produce purified water with less buffering capacity and require greater chemical addition for stabilization and corrosion control. High alkalinity consumption. Sulfate and/or chloride in finished water increases corrosivity. The effectiveness of coagulant is significantly affected by coagulant hydrolysis species speciation. Preformed Al species are stable and cannot be further hydrolyzed during coagulation. Might not be so efficient at removing HMM and highly hydrophobic NOM.

Ferric chloride, ferric sulfate

PACl

[11,26, 30e33]

Continued

14 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Table 1.1 Overview of the Metal Salt and Composite Coagulants Used in Recent Research Studies [12]. continued Coagulant

Features

Positive

Negative

Reference

PICl, PFC

Made by partially neutralized (prehydrolyzed) ferric chloride. Enhanced amounts of highcharged, moderatemolar-mass hydrolysis species. Made by partially neutralized (prehydrolyzed) ferric sulfate. Enhanced amounts of highcharged, moderatemolar-mass hydrolysis species. Combination of aluminum and/or ferricbased polymeric flocculants with polysilicate.

Wider pH range, lower sensitivity to temperature, reduced amounts of coagulants, lower residual iron concentration. Less corrosive than ferric chloride. Wider pH range, lower sensitivity to temperature, reduced amounts of coagulants, lower residual iron concentration. Less corrosive than ferric chloride. Enhancement of bridging ability of coagulant. Flocs formed are relatively large. More resistant to pH variations. Wider pH range.

More at development stage. Not so widely used.

[34,35]

Hydrolysis conditions have major impact on speciation of hydrolysis/ polymeric species. More at development stage. Not so widely used.

[36,37]

[31,38e45]

Composition of organic polymer (PDADMAC) and inorganic polymeric coagulant (PACl) with additives such as active silicates. Composition of organic polymer (PDADMAC) and inorganic polymeric coagulant (PFC). High content of Feb hydrolysis species.

Effective in waters with high alkalinity or pH. Wider working pH range.

Although flocs are large, they are nonsettleable, thus creating high turbidity values with overly low dosages. More at development stage (especially ferric-based coagulants). Flocs formed can be small or incompact, thus not favorable to sedimentation, although flotation is much more efficient. More at development stage. Not so widely used.

PFS

PASiC, PAF-SiC, PFSiS, PSF

HPAC

PFC-PDADMAC

Stronger charge neutralization and hence efficient dissolved organic carbon, SUVA, and turbidity removal capacity.

[26,38,46,47]

[35,48e51]

complexes with ionogenic groups of the hydrophilic colloid and electroneutral precipitate forms. Hydrophobic NOM contains more aromatic compounds than hydrophilic fractions and has a higher level of charge due to ionizable groups, such as carboxylic and phenolic groups. Humic substances can also be classified as humus coal, fulvic acids, hymatomelanic acids, and humic acids. Humus coal is an insoluble fraction, whereas others are soluble in

1. Introduction 15

n FIGURE 1.3 Hypothetical molecular structure of humic acid. Reprinted from J. Duan, J. Gregory,

Coagulation by hydrolysing metal salts, Adv. Colloid Interface Sci. 100e102 (2003) 475e502; with permission from Elsevier.

neutral water. Molecular weights of the soluble fractions of humic substances in order of decreasing molecular weight are humic acid > hymatomelanic acid > fulvic acids. Humic acid fraction is more hydrophobic than fulvic acid fraction and is therefore more easily destabilized by coagulation. Industrial wastewaters often contain colloidal suspensions which can be destabilized by the addition of chemical coagulants [56]. Coagulant technologies can be used before biological wastewater treatment to remove toxic substances from the wastewater stream that could affect the survival of microbes at biological wastewater treatment plants [57]. Coagulant chemicals can be added into primary clarifiers to reduce the load on biological processes. With coagulation technology, it is common to achieve high removals of pollutants: 95% total suspended solids, 65% chemical oxygen demand (COD), and 50% biological oxygen demand (BOD). Chemical coagulation can also be used after the biological wastewater treatment plant to remove residual phosphorus from the water stream. Phosphorus causes eutrophication in the receiving wastewaters, and therefore discharges of wastewater treatment plants have to be minimized. Chemical precipitation can be used to remove orthophosphates and particulate phosphorus species [3]. The stoichiometric of Aladded:Premoved has been reported to vary from 1.4 to 4.3. The stoichiometric relationship is valid at high phosphorus concentrations, and higher doses are required when concentration decreases. Typically optimum phosphorus removal occurs within the pH range of 5.5e6.0.

16 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Iron (II) and iron (III) salts are also used in wastewater treatment plants or long pipeline systems to precipitate dissolved sulfides present in the wastewater. Hydrogen sulfide (H2S) has a low odor threshold, and besides common odor problems, hydrogen sulfide also has adverse effects on human health and causes corrosion of the sewers and other equipment [58e62]. Precipitation of dissolved sulfides is normally performed using iron salts, such as ferric chloride or ferrous sulfate [63e65]. Iron salts can be added before sulfide formation or into wastewater already containing sulfides. A higher concentration of iron salt is required if dosing is conducted before sulfide formation, because iron hydroxides and carbonates are formed, and iron cations only partially react with dissolved sulfide ions. Sulfides form black ferrous sulfide (FeS) according to Eq. (1.7) in the presence of Fe(II). Fe(III) can remove sulfides by chemically oxidizing them to elemental sulfur (Eq. 1.8), while the Fe(II) formed during the reaction can subsequently produce FeS (Eq. 1.7). According to Firer et al. [64], the stoichiometry of the reaction with Fe(II) is approximately 1.3 mol of Fe2þ per 1 mol of S2 removed, and with Fe(III) 0.9 mol is required per 1 mol of S2 removed. Fe2þ ðaqÞ þ HS ðaqÞ/FeSðsÞ þ Hþ

(1.7)

2Fe3þ ðaqÞ þ HS ðaqÞ/2Fe2þ ðaqÞ þ S0 ðsÞ þ Hþ

(1.8)

Common issues that arise when iron chemicals are used include the corrosion of equipment and a localized decrease in pH at the feeding point, which can release hydrogen sulfide into the sewer atmosphere. Because of the corrosive nature of the chemicals, special materials, such as titanium, are used in the tanks and piping, increasing capital costs of the equipment [66].

1.2 Theory of electrocoagulation EC has a long history: the first plant was built in London in 1889 for the treatment of sewage [67,68]. Despite some promising results, the success of this technology has been limited. However, there has been renewed scientific, economic, and environmental interest in this technology in recent years due to demand of alternative water treatment technologies. EC understandably has not only several similarities with the chemical coagulation but also significant differences, such as side reactions, which are discussed in this section. In the EC system, there are multiple electrochemical reactions occurring simultaneously at the anodes and cathodes. These mechanisms can be divided into the main mechanisms that cause destabilization of pollutants and side reactions, such as hydrogen formation. The most important reactions are summarized in Fig. 1.4.

1. Introduction 17

n FIGURE 1.4 Schematic representation of typical reactions during the electrocoagulation treatment.

1.2.1 Main reactions Electrodes which produce coagulants into water are made from either iron or aluminum. In addition, there can be inert electrodes, typically cathodes, which are sometimes used as counterelectrodes in the system. Iron and aluminum cations dissolve from the anodes according to Eqs. (1.9) and (1.10).  FeðsÞ / Fenþ ðaqÞ þ ne

AlðsÞ / Al3þ ðaqÞ þ 3e

(1.9) (1.10)

In typical aqueous environments and conditions of the EC process, iron can dissolve in divalent Fe(II) and trivalent Fe(III) forms, whereas aluminum dissolves only in trivalent form Al(III). Fe(II) can further oxidize to Fe(III) (Eq. 1.11) if oxidationereduction potential (ORP) and pH conditions are

18 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

suitable. Oxygen has to be present, and pH has to be neutral or alkaline to achieve a reasonable reaction rate [69]. Thermodynamically stable forms of iron and aluminum in different ORP and pH conditions can be estimated using E-pH diagrams (Fig. 1.5). 4Fe2þ ðaqÞ þ 10H2 O þ O2 ðaqÞ/4FeðOHÞ3 ðsÞ þ 8Hþ

(1.11)

Moreno et al. [67] studied the electrochemical reactions of iron electrodes in EC system. According to their results and the thermodynamical data they presented, the potential of the iron anode in an EC system is in the region where Fe(III) iron is produced. However, some authors suggest that the potential of the cathode is in the region where Fe(III) is reduced to Fe(II) form, and therefore both forms exists in the EC system [70,71]. Iron is produced on the anodes mainly in Fe(II) form [72]. In low pH, the chemical dissolution of iron can be significant, and total iron concentration can be higher than would be theoretically expected. Sasson et al. also studied oxidation of produced Fe(II). According to their results, no significant oxidation occurred at pH 5, oxidation rate was moderate at pH 6, and very rapid oxidation occurred at pH 7e9. The amount of metal cations dissolved during the reactions at the anode can be calculated according to Faraday’s law (Eq. 1.12). m ¼

ItMw zF

(1.12)

where I is the current, t is the operation time (s), Mw is molecular weight of the substance (g/mol), F is Faraday’s constant (96,485 C/mol), z is the number of electrons involved in the reaction (2 for Fe2þ and 3 for Fe3þ and Al3þ), and m is the quantity of metal dissolved (g). Several studies have reported current yields higher than 100% for the dissolving of aluminum

(A)

(B)

n FIGURE 1.5 E-pH diagrams of (A) iron and (B) aluminum at 25 C, 1 bar, and 10e6 M.

1. Introduction 19

electrodes [73]. It seems that aluminum also dissolves from the cathodes. This occurs when pH on the surface of the cathode decreases due to the formation of HO (Eq. 1.13) or by the consumption of hydronium ions/protons (Eq. 1.14). 2H2 O þ 2e /H2 ðgÞ þ 2HO ðaqÞ

(1.13)

2Hþ ðaqÞ þ 2e /H2 ðgÞ

(1.14)

At high pH, aluminum dissolves as aluminate (Eq. 1.15).   2AlðsÞ þ 6H2 O þ 2HO ðaqÞ/2 AlðOHÞ4 ðaqÞ þ 3H2 ðgÞ

(1.15)

Picard et al. [73] studied cathodic dissolution during the EC process. They compared hydrogen evolution from the stainless steel cathodes and aluminum cathodes. Hydrogen produced at the stainless steel cathodes followed Faraday’s law. In tests with the aluminum cathodes, hydrogen production was higher due to the chemical dissolution of aluminum according to Eq. (1.15). Hydrogen formation and aluminum dissolved from the cathodes increased exponentially with the current intensities. When aluminum ion, aluminate, or iron ions are produced on the electrodes, they experience hydrolysis or dehydrolysis reactions in the solution. Green rust is formed when iron electrodes are used [67]. Green rust contains both Fe(II) and Fe(III) hydroxides and anions, such as Cl, CO3 2 , and SO4 2 . Other metal cations, such as Cu(II) and Ni(II), can also substitute Fe(II) in green rust if they exist in the solution [74,75]. In an EC system, green rust and hydrogen are formed according to Eq. (1.16) as follows: 6FeðsÞ þ ð12 þ xÞH2 O/1=2ð12  xÞH2 ðgÞ þ xFeðOHÞ3 $ð6  xÞFeðOHÞ2 ðsÞ (1.16)

Mononuclear hydrolysis products of Fe(III) and Al(III) were presented in Fig. 1.2 as a function of pH. It is probable that similar polynuclear species are formed in EC and chemical coagulation. However, this has not been verified by experimental studies.

1.2.2 Side reactions In addition to dissolving of aluminum and iron production, other electrochemical reactions can also take place in the EC system. They are  hydrogen formation at the cathodes due to Eqs. (1.13)e(1.15)  increase of pH due to the formation of hydroxyl ions or the consumption of hydronium ions/protons (Eqs. 1.13 and 1.14)  reduction of metal ions on the cathodes.

20 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Some articles have also speculated that oxygen is also produced on the anodes [76,77]. However, it seems that this does not take place in typical conditions (electrochemical potential) of EC systems as dissolution of anodes follows Faraday’s law [72,78]. However, at alkaline pH, dissolution of iron anodes is lower than calculated according to Faraday’s law, which indicates that other electrochemical reactions are taking place in these conditions. There is technology called electroflotation (EF), which can be used to produce bubbles that can effectively separate particles from the solution [79e81]. In EF technology, oxygen bubbles are produced on the anodes and hydrogen bubbles on the cathodes, whereas in EC technology only hydrogen bubbles are produced on the cathodes. The efficiency of the flotation in EC and EF technologies depends on the size of the bubbles. Smaller bubbles can provide a larger surface area for particle attachment. EF produces smaller diameter bubbles than commonly used dissolved air flotation (DAF) technology. The mean diameter of the hydrogen and oxygen bubbles generated in EF ranges from 17 to 50 mm, whereas in DAF, the typical mean diameter of the bubbles ranges from 48 to 60 mm [80]. Usually, the diameters of the electrolytically generated bubbles obey log-normal distribution. Electrode material, current density, and pH affect bubble size. The smallest hydrogen bubbles are produced at neutral or acidic pH [80]. Stainless steel plates have been found to produce the smallest bubbles. There has been some controversy in the reported effects of current density on the bubble size. Sarkar et al. [82] studied the effect of current density and electrode on bubble size produced on the cathodes. They discovered that a significant proportion of hydrogen produced on the cathodes can be dissolved in the solution. According to their results, bubble size diameter is a function of hydrogen production rate, bubble nucleation rate, and dissolved gas concentration field. In contrast to chemical coagulation, EC treatment increases the pH of the solution when it is in an acidic, neutral, or slightly alkaline region and decreases when it is highly alkaline. This change of pH during the EC treatment affects the speciation of aluminum and iron hydroxides. At highly acidic pH (pH 2), the alkalinity produced during the EC is not sufficient to increase the pH of the solution, whereas at pH 3 and higher initial pH values, pH rises during the treatment [78]. This is easily understandable, as concentration of hydronium ions increases by factor 10 when pH decreases from pH 3 to pH 2. When initial pH is significantly alkaline (pH > 9), pH probably decreases due to the formation of aluminate [AlðOHÞ4  ], which is an alkalinity consumer [83,84]. It seems that the pH

1. Introduction 21

change rate and final steady-state pH depends on the concentration of anions in the solution. According to Trompette and Vergnes [85], pH increases more in sulfate solution than in chloride solution. As mentioned in Section 1.1.4., sulfates can replace hydroxyl ions in the hydroxide precipitates, and therefore less hydroxyl ions are bound to hydroxides. Because electrochemical reactions occur at the surface of the electrodes, the concentration of reaction products is highest at the electrode surface, and the concentration gradient exists from the surface toward the bulk solution. Therefore, pH decreases at the vicinity of the anodes and vice versa at the cathode surface. This can lead to precipitation of inorganic salts on the electrode surface if their solubility changes as a function of pH, e.g., precipitation of calcium carbonate on the cathodes [86]. If electrochemical potential at the cathodes is in the right region, the direct electrochemical reduction of metal cations may occur at the cathode surface (Eq. 1.17). Menþ ðaqÞ þ ne /Me0 ðsÞ

(1.17)

There are several studies where the removal of metals from wastewaters has been studied, as discussed later in Section 1.3.2. Results indicate that metals are mostly removed by coagulation and coprecipitation. However, the reduction of metal ions may have some effect on removal efficiency even if they are not reduced to the metallic state. Heidmann and Calmano [87] studied the removal of Zn(II), Cu(II), Ni(II), Ag(I), and Cr(VI) from aqueous solution by EC. They proposed that Zn(II), Cu(II), Ni(II), and Ag(I) were removed from the solution by hydrolyzation and coprecipitation, whereas Cr(VI) was reduced first to Cr(III), after which it was precipitated as hydroxide. It is probable that dissolving of aluminum at the cathodes prevents plating of the cathode with other metals. With iron or inert cathodes, some plating could be expected at least with some metal ions, e.g., Ni(II) and Cu(II). It is possible that plating is a slow reaction in typical concentrations and removal occurs mainly by coagulation coprecipitation due to these kinetic reasons.

1.2.3 Properties of the sludge The properties of produced sludge are important because sludge treatment and disposal are one of the major cost factors in water and wastewater treatment, especially when sludge has been produced by chemicals. The dried EC sludge produced by different combination electrodes during arsenic removal has been characterized by FTIR, PXRD, XPS, SEM-EDS, and Mössbauer spectroscopy [71]. Dried sludge formed by AleAl electrodes contained amorphous aluminum hydroxide and/or aluminum oxyhydroxide.

22 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

FeeFe electrodes produced crystalline phases, such as magnetite, and poorly crystalline phases, such as iron oxyhydroxides and lepidocrocite. Sludge produced by combined FeeAl electrodes contained the aforementioned products as well as mansfeldite, diaspore, and iron oxide. Crystallinity of iron species decreased, possibly due to the substitution of iron by aluminum. Emamjomeh and Sivakumar studied the sludge produced during the fluoride removal by EC using X-ray diffraction [88]. Identified products were aluminum fluoride hydroxide complexes and aluminum hydroxide. Shafaei et al. [89] identified amorphous manganese and aluminum species in sludge produced by EC during Mn(II) removal. The settleability of sludge produced by EC during paper mill [90] and textile wastewater treatment has been studied [91]. When paper mill wastewater was treated, the sludge volume index was 0.207e0.310 L/g and 0.081e0.091 L/g for aluminum and iron electrodes, respectively. The sludge formed by iron electrodes was heavier and produced a more compact layer. The authors observed similar results with textile wastewaters. Hydrogen produced during the EC induces flotation of particles and decreases settling velocity. As discussed in Section 1.2.2, the operating parameters of EC affect the properties of hydrogen bubbles. Lai and Lin [92] treated chemical mechanic polishing wastewater by AleFe electrodes and studied settling properties of produced sludge. They concluded that complex models of third- or fourth order described the sludge settling data. The heat value of the sludge produced by EC has been measured using a bomb calorimeter [93]. Heat value was 5.3 MJ/kg, and authors concluded that it could be used as a fuel in the furnaces, and the produced ash could be blended with the cementitious mixtures. It is probable that sludge produced during the treatment of other wastewaters containing high organic matter also has a high heat value. The combination of EC and electrodewatering has been studied for the reduction of water content in sewage sludge [94]. The combination of EC and electrodewatering reduced water content of the sludge to 55%, whereas water content in sludge treated with only pressure filtration was 78%.

1.2.4 Treatment parameters There are various parameters which have an effect on the efficiency of the EC in removing the pollutants from water. Parameters which are known to have an effect are  Material of the electrodes can be iron, aluminum, and/or inert material (typically cathodes). Iron and aluminum ions and hydroxides have different chemistries and applications.

1. Introduction 23

 pH of the solution has an effect on the speciation of metal hydroxides in the solution and also on the z-potential of the colloidal particles. It also affects the dissolution of aluminum cathodes.  Current density is proportional to the amount of electrochemical reactions taking place on the electrode surface.  Treatment time or electric charge added per volume is proportional to the amount of coagulants produced in the EC system and other reactions taking place in the system.  Electrode potential defines which reactions occur on the electrode surface.  Concentration of the pollutants affects the removal efficiency because coagulation does not follow zeroth-order reaction kinetics but rather pseudoesecond- or first-order kinetics.  Concentration of anions, such as sulfate or fluoride, affects the composition of hydroxides because they can replace hydroxide ions in the precipitates.  Temperature affects floc formation, reaction rates, and conductivity. Depending on the pollutant, increasing temperature can have a negative or a positive effect on removal efficiency.  Other parameters, such as hydrodynamical conditions and interelectrode distance, may have effect on efficiency of the treatment and electricity consumption. However, these parameters have not been widely studied or they have a negligible effect on the removal efficiency.

1.2.4.1 Electrode materials Electrode material defines which electrochemical reactions take place in the EC system. Aluminum and iron electrodes have both been used successfully in EC systems. Aluminum dissolves in all cases as Al(III), whereas there is some controversy as to whether iron dissolves as Fe(II) or Fe(III) [67,72]. Most results indicate that iron dissolves as Fe(II), such as [72,95,96], and is oxidized in bulk solution to Fe(III) if there are oxidants, such as oxygen, present in sufficient concentration and pH is alkaline (Fig. 1.5A, see p. 28). Fe(II) is a poor coagulant compared with Fe(III) due to higher solubility of hydroxides and lower positive charge, which explains some poor results obtained with iron electrodes, such as in the study of Bagga et al. [96]. Optimal material selection depends on the pollutants to be removed and the chemical properties of the electrolyte. In general, aluminum seems to be superior compared with iron in most cases when only the efficiency of the treatment is considered. However, it should be noted that aluminum is more expensive than iron.

24 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Inert electrodes, such as metal oxideecoated titanium, are used as cathodes in some constructions. When water has significant amounts of calcium or magnesium ions, the inert cathode material is recommended [76]. There are also some studies where combinations of aluminum and iron electrodes have been used. Linares-Hernández et al. [95] obtained high removal of color with aluminum electrodes, while iron was more effective than aluminum in reducing COD from industrial wastewater. A combination of iron and aluminum removes both color (71%) and COD (69%) with high efficiency. Similar results were obtained when paper mill wastewaters were treated with various aluminum and iron electrode combinations [97]. Aluminum electrodes were most effective in removing color of the wastewater, whereas iron electrodes removed COD and phenol from the wastewater more effectively than aluminum electrodes. A combination of aluminum and iron electrodes removed color, COD, and phenol with high efficiency. Combination electrodes have been studied for arsenic removal from groundwater [71]. Iron electrodes and a combination of iron and aluminum electrodes gave the highest arsenic removal efficiencies. Similar results were obtained for copper, chromium, and nickel removal from metal plating wastewater [98]. FeeAl pair has been most effective in removing indium from water [99].

1.2.4.2 pH of the solution One of the key parameters of EC treatment is the pH of the solution to be treated. It has an effect on the conductivity of the solution, dissolution of the electrodes, speciation of hydroxides, and z-potential of colloidal particles. As discussed in Sections 1.1.3 and 1.1.4, aluminum and iron cations and hydroxides cause destabilization of colloids. Effective coagulant species are formed in acidic, neutral, and slightly alkaline pH. In highly alkaline pH, AlðOHÞ4  and FeðOHÞ4  ions are formed, and these ions have poor coagulation performance. As can be seen in Fig. 1.2 (see p. 20), Fe(III) is effective in a wider pH area than Al(III) and works also in slightly alkaline pH. It is also known that competing anions have an effect on the optimum pH of the coagulation. The effect of water pH on the efficiency of pollutant removal can mostly be explained by the aforementioned mechanisms. However, as discussed in Section 1.2.3, pH increases during the EC treatment, making it a constantly changing parameter; therefore, mechanistic studies of EC systems are difficult to conduct. In pHs lower than 3, the release rate of aluminum during electrolysis with a constant charge per volume was lower than in pHs above 3 [78]. Chemical dissolution of aluminum cathodes occurs because pH increases to a level where aluminate is formed. It is probable that acidic bulk solution inhibits

1. Introduction 25

this reaction because produced hydroxyl ions are consumed by the acid in the solution. In acidic pH, the dissolution of iron electrodes was significant even without electricity, whereas oxidation of Fe(II) to Fe(III) occurs only at pHs above 5 [72]. The dissolution rate decreases at high pH, which is understandable as the corrosion rate of iron decreases in alkaline pH in the presence of oxygen because a passive layer forms on the surface. It has been suggested that initial pH is key parameter when either chemical coagulation or EC is selected for the water treatment [100]. EC is more suitable when higher pH is desired (acidic solutions), whereas chemical coagulation is preferred when pH should decrease before discharge. There are also some pollutants which have specific optimum pHs of treatment, such as phosphorus and metal cations. The effect of initial pH on phosphate removal from wastewater by EC with iron electrodes has been studied [101]. The highest removal efficiency was observed at the lowest tested initial pH (pH 3). Wastewater pH increased to a highly alkaline level (pH 10e12) during the long treatment, probably due to partial replacement of HO by PO4 3 . In high pH, the removal rate was poor due to the formation of FeðOHÞ4  and higher solubility of FePO4. In higher pH, there are more HO ions to compete with PO4 3 , and therefore less FePO4 is formed. Other researchers have had similar results with iron electrodes when they studied the removal of phosphorus from secondary effluent [102]. According to Janpoor et al., phosphorus removal from laundry wastewater with aluminum electrodes was poor when initial pH was lower than 6 or higher than 8 [103]. As with aluminum and iron, other metal cations can also form hydroxides in water. Most nonionic hydroxides have low solubility in water and can be removed by precipitation and coprecipitation with EC systems. Hanay and Hasar [104] studied removal of Cu(II), Mn(II), and Zn(II) by aluminum electrodes. Removal efficiency increased when initial pH of the wastewater increased. Similar results have been obtained for the removal of Co(II) [105], As(V) [106], Cu, Cr, and Ni [98], Cu(II), Pb(II), and Cd(II) [107], Cu(II), Ni(II), Zn(II), and Mn(II) [108], In(III) [99], Mn(II) [89], and Cr(III) [109]. Hg(II) removal was not significantly affected by initial pH in the range of 3e7 [110].

1.2.4.3 Current density and treatment time Current density is directly proportional to the rate of electrochemical reactions taking place on the electrode surface, and it also has an influence on the electrode potential, which defines the reactions taking place on the electrode surface. It seems that on iron and aluminum anodes, dissolution reaction is the primary reaction, and the proportion of other reactions is

26 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

insignificant at the typical current densities and electrode potentials when pH is neutral or acidic [72,78]. At alkaline pH, the dissolution rate of iron anodes can be lower than the value calculated by Faraday’s law, which indicates that there can be other reactions at the anode in these conditions [72]. Coagulant produced by electrolysis can usually be calculated according to Faraday’s law when current and treatment times are known (Eq. 1.12). Coagulant concentration produced by electrolysis on anodes is typically directly proportional to the electric charge added per volume (coulombs per liter). However, the total amount of coagulant dissolved also includes chemical dissolution of the electrodes in low pH and the dissolution of aluminum cathodes. Current density probably has some influence on the chemical dissolution of aluminum cathodes, as it affects the rate of hydroxyl ion production at the cathodes. Mouedhen et al. [78] studied aluminum dissolution from the cathodes with an electrochemical cell constructed of platinized titanium anode and aluminum cathode. They used constant charge per volume (540 C/L) and various current densities. According to their results, as the current density decreases, the amount of aluminum generated increases. These results indicate that even low current density (70%) is typically obtained with optimum parameters. Aluminum, iron, and combination electrodes can be used. In general, iron electrodes give higher organic matter removal, whereas higher color removal is obtained with aluminum electrodes.  Purification of surface waters from NOM, inorganic pollutants, or microbes (Table 1.4). Typically high removal of pollutants (>90%). Aluminum electrodes are more commonly used than iron electrodes in these applications. Tables 1.2e1.4 review relevant and recent publications in these categories. There is also a review article where most of the applications have been discussed [128]. Practically all articles report a higher removal percentage when higher current density and/or treatment times are used. Therefore, this finding was not included in Tables 1.2e1.4. This result is easily understandable as produced coagulant concentration in solution increases as a function of electric charge added as discussed in Section 1.2.4.

1.3.3 Combinations of electrocoagulation and other water treatment technologies Besides the sludge separation technologies, filtration, flotation, and sedimentation, other water treatment technologies have also been combined with EC. EC has been combined with chemical coagulants [173], Fenton oxidation [174], hydrogen peroxide [175,176], ozone [177], photocatalysis [178e180], and biofiltration [102].

32 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Table 1.2 Recent Studies in Which Electrocoagulation (EC) Has Been Used to Remove Metal Pollutants from Water. Pollutants

Matrix

Electrode Material

Studied Parameters

Optimum Conditions and Notes

Arsenic as As(VI) or As(III)

Synthetic solutions, groundwater

Fe, Al, and combination electrodes

Initial pH, initial concentration, current density, treatment time, electrochemical cell designs

[70,71,106, 129e131]

Chromium as Cr(VI) or Cr(III)

Synthetic solutions, wastewaters

Al, Fe, and combination electrodes

Cobalt as Co(II)

Synthetic solutions

Al electrodes

Fe(II) added into tap water

Al electrodes

Initial pH has significant effect on removal efficiency. Higher removal efficiency in neutral or alkaline initial pH. Practically complete removal with long treatment time and/or high current density.

[105]

Iron as Fe(II)

Mercury as Hg(II)

Synthetic solutions, surface water

Al and Fe electrodes

Practically complete removal of mercury. Higher efficiency of treatment with iron electrodes.

[110]

Indium as In(II)

Synthetic solutions

Al, Fe, and combination electrodes

Highest removal efficiency with FeeAl electrodes. Kinetics of indium removal follows pseudoesecondorder mechanism.

[99]

Manganese as Mn(II)

Synthetic solutions

Al electrodes

Improved removal in neutral and alkaline pH. Possible direct reduction at the cathode surface.

[89,104]

Mixed metal wastewaters

Synthetic solutions, wastewaters

Al, Fe, and combination electrodes, stainless steel

Initial pH, current density, electrolyte concentration, electrode material, conductivity, treatment time Current density, initial pH, conductivity, initial concentration, treatment time Current density, initial concentration, treatment time, interelectrode distance Current density, initial pH, treatment time, interelectrode distance, electrode material Initial concentration, supporting electrolytes, electrode material, initial concentration, applied voltage Initial pH, current density, treatment time, anions, conductivity, initial concentration Electrode material, current density, initial pH, conductivity, treatment time, anions, flow rate, initial concentration

Can obtain practically total removal of As. Optimum removal with modified flow or airlift reactor and neutral to alkaline pH. Oxidation of As(III) to As(VI) can take place. Practically complete removal of Cr possible. Reduction of Cr(VI) to Cr(III) can take place during EC.

Generally improved removal at neutral or alkaline pH. Iron or combination electrodes may give improved results compared with aluminum. Reduction of metals can take place.

[70,87,98,104,107, 134e136,138]

References

[70,87,98,109,116, 119,132e136]

[137]

1. Introduction 33

Table 1.3 Recent Studies in which electrocoagulation Has Been Used to Remove Organic Pollutants from Wastewaters. Pollutants

Matrix

Electrode Material

Studied Parameters

Optimum Conditions and Notes

Dyes, textile wastewater

Synthetic solutions, wastewaters

Al, Fe, and stainless steel electrodes

Pulsed current frequency, power supply type, initial pH, current density, treatment time, electrode potential, dye auxiliary chemicals

Slaughterhouse wastewaters, manure

Wastewaters

Al and Fe electrodes

Treatment time, current density, initial pH, supporting electrolyte, electrode material

Tannery wastewaters

Wastewaters

Al and Fe electrodes, stainless steel cathodes

Current density, initial pH, conductivity, initial concentration, treatment time, electrode configuration

Alternating pulsed current gives improved results compared with direct current (passivation prevention). High removal efficiency (>90%) obtained with optimal parameters. Na2CO3 causes adverse effects. Color and chemical oxygen demand (COD) abatement followed pseudoefirst-order kinetics. Over 70% removal of COD, color, and oilgrease obtained. Highest removal efficiency in acid or neutral pH. Al electrodes are more effective in removing COD and color, whereas iron is more effective in removing oil grease. Fe electrodes are more effective for the removal of COD and sulfide than compared with aluminum electrodes. Better treatment efficiency obtained when initial pH was acidic. Typically high removal of COD, biological oxygen demand (BOD), total suspended solid (TSS), sulphide, oil-grease, Cr, Fe, and turbidity (80% e100%) obtained. Kinetics follows pseudo efirst- or second-order kinetics.

References [124,126, 139e142] [70,71,106, 129e131]

[143,144]

[145e149]

Continued

34 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Table 1.3 Recent Studies in which electrocoagulation Has Been Used to Remove Organic Pollutants from Wastewaters. continued Electrode Material

Studied Parameters

Optimum Conditions and Notes

Wastewaters, waste streams

Al and Fe electrodes and combinations

Electrode material, initial pH, current density, treatment time, supporting electrolyte, stirrer speed, electrode distance, temperature

Olive mill wastewaters

Wastewaters

Al and Fe electrodes

Current density, initial pH, treatment time, supporting electrolyte, supporting oxidants, and coagulants

Oily waters, petroleum refinery wastewaters

Wastewaters

Al, Fe, and stainless steel electrodes

Electrode configuration, treatment time, electrode material, initial pH, initial concentration, supporting electrolyte, temperature

Al is more effective. Slightly acidic or neutral initial pH gives higher removals. Increasing rotation speed and electrode distance increases removal of pollutants. Increasing temperature from 20 C to 60 C decreases efficiency. Typically 70% e100% removal of COD, BOD, polyphenols, and color. Combination electrode gives high efficiency. Typically over 70% removal of COD, polyphenols, turbidity, SS, and color. Significantly decreases toxicity of wastewater. Highest removal obtained when initial pH is neutral or slightly acidic. Fe can give higher removal efficiency than Al. Oxidants and other coagulants can improve removal. Typically >90% removal of BOD5, oil and grease, petroleum hydrocarbons, TSS, turbidity, and sulfate obtained. High removal when initial pH is acidic, neutral, or slightly alkaline. Fe electrodes can be more efficient than Al. Removal increases at lower temperatures.

Pollutants

Matrix

Pulp and paper wastewaters and streams

References [90,97, 150e154]

[155e157]

[122,158,159]

Table 1.4 Recent Studies in which electrocoagulation (EC) Has Been Used to Remove Pollutants from Surface Waters or Groundwaters and Nutrients from Wastewaters. Electrode Material

Matrix

Studied Parameters

Surface water, natural organic matter

Synthetic solutions, surface waters

Al and Fe electrodes

Boron

Geothermal waters

Al electrodes

Fluoride

Synthetic solutions

Al electrodes

Anions, current density, initial concentration, initial pH, flow rate, residence time

Microorganisms

Synthetic solution

Al, Fe, and stainless steel electrodes

Nitrate

Synthetic solutions

Al and Fe electrodes

Current density, treatment time, electrode material, initial pH, treatment time, salinity Initial pH, treatment time, current density, initial concentration

Phosphates and phosphorus

Wastewaters, synthetic solutions

Al and Fe electrodes, stainless steel cathodes

Initial pH, current density, applied voltage, supporting electrolytes, initial concentration, treatment time, interelectrode distance Initial pH, current density, temperature, treatment time

Electrode material, treatment time, initial pH, controlled pH, current density, temperature, anions, initial concentration, interelectrode distance, conductivity, supporting electrolyte

Optimum Conditions and Notes

References

Slightly acidic initial pH gives optimum results by preventing formation of gel layer on the electrode surface. NaCl or Na2SO4 are favorable supporting electrolytes. Simultaneous removal of microbes obtained.

[160e165]

High boron removal efficiency (96%) obtained. Increasing temperature increases removal efficiency. Optimal pH is slightly alkaline (pH 8). Fluoride competes with other anions. Sulfate inhibits localized corrosion of the electrodes. Chloride and nitrate prevent the effect of sulfate and enhance corrosion. Neutral pH is optimum. Practically complete removal of fluoride possible. Al electrodes were slightly more effective than Fe or stainless steel. Complete removal of microorganisms is possible. Efficiency of the treatment increases with increasing temperature. Optimal removal at alkaline pH (pH 10 e11). Nitrate removal follows first-order reaction kinetics. Current density did not affect removal when same dose of coagulant was added during EC. Over 90% removal can be obtained. Practically complete removal of phosphate is possible. The adsorption process follows second-order kinetics. Competing anions affect removal efficiency. Very low temperature decreases removal efficiency. Al more efficient than Fe. High efficiency in acidic initial pH.

[114]

[88,112,166]

[165,167,168]

[169,170]

[101 e103,113,171,172]

1. Introduction 35

Pollutants

36 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

The addition of oxidants such as hydrogen peroxide or ozone increases the removal of organic material from wastewaters. Kabdasli et al. [174] studied the removal of organic matter and heavy metals from metal plating effluents using a combined EC-Fenton process. Stainless steel electrodes were used as anodes and cathodes. The combined process increased the removal of organic matter compared with EC only. The optimal dose of hydrogen peroxide was 15e30 mM. Roa-Morales et al. [176] treated wastewater from pasta and biscuit processing with aluminum EC combined with peroxide additions. Hydrogen peroxide additions slightly increased COD removal compared with pure EC. Ozonation of effluent after EC can further decrease the concentration of organic matter in the effluent [177]. Higher removal efficiencies could be obtained when iron electrodes are used instead of aluminum because the presence of Fe(II) catalyzes the decomposition of oxidants to hydroxyl radicals, which can mineralize organic pollutants. Residual organic matter after EC can be also removed with photocatalysis according to our study [180] and Boroski et al. [178,179]. Can et al. [173] combined EC with chemical coagulants PACl and alum. The addition of chemical coagulants decreases pH, whereas EC increases it and with a combination system, the final pH can be stabilized to neutral region. EC combined with PACl gave higher COD removals than combination with alum. PACl has more prepolymerized species, which can be more efficient in pollutant removal, and it also contains chlorides which increase localized corrosion of electrodes and reduce energy consumption. Yu et al. [102] combined EC with biofiltration. Combining EC to biofiltration increased the removal of COD and phosphorus from 69.1% to 9.6% to 76.6%e83.7% and 70.7%e93.0%, respectively.

1.3.4 Economical and ecological considerations Operating cost calculations have been made in a few articles. Calculations typically include the cost of chemicals, electrodes, and energy. It should be noted that the price of materials and energy changes over the course of time, and therefore, operating costs are only rough estimates. Cost calculations do not typically include investment costs, which may be significant including, for example, power supplies, electrochemical cell vessels, and sludge separation systems. The treatment cost of dye-polluted wastewaters has been estimated by Eyvaz et al. [126]. The estimated operating cost was USD 1.3e3.4 per kg TOC removed depending on parameters, such as treatment time. Sridhar et al. [151] carried out an economic analysis of the operating cost of EC treatment of pulp and paper industry bleaching effluents. Operating costs varied

1. Introduction 37

from USD 1.52 per m3 to 1.72 per m3. Ölmez [133] estimated the cost of Cr(VI) removal from wastewaters. The total cost of the EC process was about twofold compared with the conventional process because of higher electricity consumption. Meunier et al. [135] compared the cost of electrochemical and chemical precipitation with calcium hydroxide or sodium hydroxide. EC was up to five times cheaper than chemical precipitation. Emamjomeh et al. [88] estimated operating costs of fluoride removal from water. Operating cost was AUD 0.36e0.61 per m3 when initial fluoride concentration was 5 mg/L. Kobya et al. [181] have published their operating cost calculations on the treatment of waste metal cutting fluids using EC. The operating cost was USD 0.025e0.90 per m3 and USD 0.01e0.79 per m3 with iron and aluminum electrodes, respectively. Espinoza-Quiñones et al. [149] made operational cost calculations for tannery wastewater treatment with EC. They estimated that a 60 min treatment (high removal of pollutants) costs USD 1.64 per m3, which was less than with conventional coagulants. According to Bayramoglu et al. [124], the operating cost of chemical treatment of textile wastewater by chemical coagulation was 3.2 times as high as that of EC. It can be noticed that there are few articles where the costs of chemical coagulation and EC have been compared. It might be difficult to obtain reasonable information on the price of common coagulants. Some chemical coagulants, such as aluminum sulfate and ferric chloride, are bulk chemicals and have a low price, whereas prehydrolyzed metal coagulants are more expensive. The price of chemicals naturally depends on the amount that is required in the process, and therefore, major consumers, such as pulp and paper mills or municipal water treatment plants, pay less for their chemicals than small users. However, iron and aluminum are typically more expensive in their metallic state than as metal salts because production of metals is energy consuming. Coagulation chemicals are produced from the minerals and are not transformed into the metallic state during this process. For example, aluminum sulfate is manufactured through the reaction of aluminum trihydrate and sulfuric acid. Aluminum trihydrate is made from bauxite mineral by purification. In addition, the ecological effects of EC are somewhat unknown. As mentioned earlier, production of metallic aluminum and iron consumes high amounts of energy [182]. It has been estimated that aluminum production consumes 5% of the electricity generated in the United States. Typically, aluminum is recycled, and this process requires significantly less energy than the production of pristine metal. The ecological effect of water purification with EC and chemical coagulants should be compared, taking energy and material consumption into account during the manufacture of metals or chemicals.

38 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

2. OBJECTIVES AND STRUCTURE OF THE WORK 2.1 Surface water treatment by electrocoagulation The objective in Papers IeIII was to study the parameters affecting NOM removal by EC. The effect of current density, treatment time, electric charge added per volume, temperature, initial pH, and electrochemical cell structure on the removal efficiency of NOM was investigated. Relevant water quality parameters were measured before and after treatment to evaluate the performance of the treatment. In Paper I, one objective of the study was to investigate whether chemical coagulation and EC could be used in combination. Statistical design of experiments (DOEs) and models were used in Papers IIeIII to interpret the effect of treatment parameters and their combinations on NOM removal and to optimize parameters of EC.

2.2 Wastewater treatment by electrocoagulation The objective in Papers IV and V was to investigate two specific applications of EC in wastewater treatment; the removal of sulfides in Paper IV and the removal of toxic pollutants in Paper V. In both papers, the main parameters affecting removal efficiency were studied. Wastewaters and effluents used in the studies were collected from pulp and paper mills. Synthetic pollutants were also used in Paper V to study the removal of single pollutants using EC treatment. Statistical methods were used in Paper V to interpret the results and to optimize the treatment parameters.

3. MATERIALS AND METHODS 3.1 Water samples and chemicals Surface water samples (IeIII) were collected from a water treatment plant of a Finnish paper mill. In the mill, water passed through a mechanical sieve that removed coarse material from water, but concentrations of colloidal and dissolved material were not affected. The water was typical Nordic surface water that contained high concentrations of NOM and was highly colored but had low turbidity (Table 1.5). The water was neutral (pH 6.2e6.5), and it had low conductivity (2.8e3.6 mS/m). Wastewaters and effluents were also collected from pulp and paper mills (Table 1.6). Wastewater used in the sulfide experiments (IV) contained high amount of organic matter and sulfurous anions. Before the experiments, wastewater was filtered through a coarse filter and gently stirred at 40 C to promote the growth of sulfate-reducing bacteria (SRB) which converted sulfates to sulfides. Debarking effluent used in the toxicity removal tests (V) was highly colored and contained high concentrations of organic matter.

3. Materials and Methods 39

Table 1.5 Initial Quality of Surface Waters Used in the Studies. Turbidity (NTU) Apparent color (mg/L PtCo) TOC (mg/L) SUVA [L/(m mg)] z-Potential (mV) Residual Al (mg/L) Residual Fe (mg/L) Residual Mn (mg/L)

Paper II

Paper III

0.34 102 18.29 4.52 15.2 0.315 0.884 0.121

0.51 83.7 16.41 3.51 17.1 0.257 0.348

Table 1.6 Initial Parameters of the Wastewaters and Debarking Effluents Used in the Studies. Wastewater (IV) Turbidity (NTU) pH Conductivity (mS/cm) Oxidationereduction potential (mV vs. SCE) Apparent color (mg/L PtCo) TOC (mg/L) Dissolved organic carbon (mg/L) CODcr (mg/L) z-Potential (mV) Sulfate (mg/L) Dissolved oxygen (mg/L) Dissolved sulfide (mg/L) Phosphorus (mg/L) Toxicity (EC50)

6.5e7.1 1.4e3.0 300e(200)

300e450 800e1500

Debarking Effluent (V) 216 5.21 0.86

1910.00 575.00 437.00 18.5

600 0.5e2.0 7.5e9.5 1.3e1.5

Debarking effluents are known to contain tannins and resin acids which are both toxic to microorganisms and also to animals, such as fish [183e189]. Debarking effluent was also filtered through a coarse filter before the experiments. Synthetic resin solutions were used in the toxicity removal tests to analytically study the removal of resin acids and metal pollutants and their toxicity by EC (V). Solutions were prepared using Polish wood rosin (Hercules Corporation, USA) and copper nitrate Cu(NO3)2. The composition of Polish wood rosin is presented in Table 1.7. Additional tests were conducted using

7.6

40 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Table 1.7 Breakdown of Resin Acid Content of the Polish Wood Rosin (Hercules Corporation, USA). Resin Acid

% (w/w)

Pimaric acid Sandaracopimaric acid Isopimaric acid Palustric acid Dehydroabietic acid Abietic acid Neoabietic acid Levopimaric acid

8.0 1.7 3.1 20.7 30.6 22.4 5.3 8.2

pure dehydroabietic acid (DHAA) and isopimaric acid (IPA) (99%, Helix Biotechnologies, Canada). Lake water was used to dilute solutions to obtain a realistic background for EC tests. Water pH was adjusted (IeIII and V) to the desired level using sulfuric acid and sodium hydroxide solutions. Auxiliary electrolytes, sodium chloride (NaCl) (I) and sodium sulfate (Na2SO4) (V), were used in some studies to acquire reasonable conductivity. A chemical coagulant, PACl, was used in tests of combined EC and chemical coagulation (I). All inorganic salts and pH adjustment chemicals used in this study were of analytical grade except PACl, which was typical commercial grade used in the water treatment plant of a paper mill.

3.2 Water treatment procedure Possible auxiliary chemicals were added into the water before electrochemical treatment. Electrochemical treatments were conducted in regular glass beakers. During the treatment, water was stirred with a standard magnetic stirrer. The electrochemical cell was constructed of aluminum (IeIII), iron (IV and V), or combinations of aluminum and inert electrodes (II). Combinations of aluminum and inert electrodes were constructed so that either anode (Cell A) or cathode (Cell C) or both electrodes (Cell T) were made of aluminum. In the Cell A and Cell C, other electrodes were made of inert material which does not dissolve in these conditions. In all tests, electrodes were in a monopolar configuration. Electrode assembly was constructed of 70 mm  120 mm (I) or 50 mm  70 mm (IIeV) plate electrodes. In total, four (I, IV, and V) or six (II and III) electrode plates were used (half anodes and other half cathodes). The distance between the electrode plates was 5 mm (IV) or 10 mm (IeIII and V). Electrochemical

3. Materials and Methods 41

Table 1.8 Electrolysis Parameters of This Study.

Surface water (I eIII) Wastewater (IV) Debarking effluent (V)

Current Density (mA/cm2)

Electric Charge (C/L)

Voltage (V)

0.36e1.79

0e150

3e32

3.6e17.9 7.1e21.4

0e4600 0e2700

2e11 8e25

cells were constructed using plastic nuts and bolts to prevent short-circuiting of electrode plates. Standard laboratory power supplies were used in the experiments. Electrolysis parameters are presented in Table 1.8. In some of the studies, samples were transferred into a programmable stirrer (I) or an overhead stirrer (II and III) to promote flocculation. Sedimentation was used (Ie III and V) to enhance removal of coagulated material from the electrochemically treated water. Filtration through filter paper (I, III, and IV) (Whatman or VWR) or GF/C glass fiber filter (Whatman) (II and V) was used in most of the studies for the separation of coagulated material from the treated water. In few tests, samples were conveyed by pipette from the glass beakers from the center of the beaker after sedimentation or flotation (I). Flocculation, sedimentation, and filtration steps were used to simulate particle separation in full-scale water treatment plants by filtration, sedimentation, or flotation. In addition, a few flotation tests were conducted using the bench-scale DAF system (I). Temperature adjustment (III) during the tests was undertaken at 295 K using heating plate, at 285 K using water flow and at 275 K with an ice bath. According to measurements, temperature was maintained within a 1 K range of the set temperature throughout all stages. Oxidation of sulfide was prevented during the sulfide precipitation tests (IV) by adding nitrogen into the sample bottle. Samples were pumped into the beakers to further ensure that mixing with air was minimized. After electrochemical treatment, the electrodes were washed with 4%e10% HCl solution and distilled water.

3.3 Analytical methods 3.3.1 Chemical analysis Water samples were stored, after filtration, in the cold, except sulfide samples which were measured right after the experiment because sulfides oxidize to sulfates when they come into contact with air. Water pH, conductivity, ORP, and dissolved oxygen were measured using standard laboratory

42 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

devices and sensors. The turbidity of samples was measured with Spectroquant NOVA60 (Merck) (I) or Model 2100P ISO turbidimeter (Hach) (II, III, and V). Apparent color was measured using a photometer 2100AN (I) or DR 2800 (Hach Lange) (II, III, and V) according to standard EN ISO 7887:1994. Absorbance of the samples at 254 nm wavelength was measured with a Lambda 45 UV/Vis spectrophotometer (IIeIII). The organic matter content of surface water and wastewater samples was investigated by analyzing permanganate index (I), COD (IV), TOC (II, III, and V) or dissolved organic carbon (DOC) (IV and V). The permanganate index was measured as stated in standard EN ISO 8467:1995 using a 2100AN (Hach) photometer. COD was measured according to standard SFS 5504:1998. TOC was measured using a TOC analyzer (Shimadzu), and DOC was measured with the same analyzer after the samples were filtered through a 0.45 mm pore size membrane. Absorbance at 254 nm wavelength was measured to study the composition of residual NOM. Specific UV absorbance (SUVA) is absorbance at 254 nm divided by DOC concentration. When long-chained humic acids are present in water, SUVA values are high [3]. When SUVA value is 4e5 L/(m mg), the water is composed of hydrophobic material which is high in aromatic structures. When SUVA is less than 3 L/(m mg), the water is composed mainly of hydrophilic material. SUVA values were measured during the temperature and cell construction tests (II and III). The concentration of metals was measured from acidified samples (0.2 mL 37% HNO3 to 10 mL sample) using iCAP 6000 series ICP-OES (Thermo Electron) (IIeV) or photometric method (I). Residual metal concentrations were measured after filtration, whereas metals dissolved from the electrodes were measured from the samples which were taken during or directly after the treatment. The z-potential of the samples was measured using a Zetasizer Nano ZS analyzer (Malvern). Resin acids were measured according to the method described by Soimasuo et al. [190] using the HP 6890 GC-MS system (HewlettePackard) with HP-5 capillary column. Sulfate, sulfide, and phosphorus concentrations of the samples were measured using photometric methods (Lange). Dissolved sulfide and phosphorus concentrations were measured after passing the samples through a 0.45 mm pore size membranes.

3.3.2 Toxicity analysis Bacterial toxicity was measured using a Microtox-Flash assay based on the bioluminescence of the marine bacterium Vibrio fischeri. Toxic substances affect the metabolism of the cells and break their structure, thus reducing

3. Materials and Methods 43

bioluminescence. Bacterial toxicity tests were conducted according to the BioTox method (Aboatox, BO1243-500), standard draft ISO CD 2006b, and the method developed by Lappalainen et al. [191] using a Sirius luminometer. 2% NaCl solution was used as negative control, whereas K2Cr2O7 was used as a positive reference chemical. In addition to bacterial toxicity, toxicity of samples to algae was also measured. Analysis was carried out according to standard ISO 8692:2004. The sample size was reduced to fit 48-well microplates (Greiner bio-one GmbH). The average algal cell concentration before exposure to wastewaters was measured using a microscope (Olympus BX40) and cytometer (Assistant). Fluorescence after 0 h, 24 h, 48 h, and 72 h exposure was measured using Fluoroskan Ascent fluorometer (Labsystems). Negative control solutions contained nutrients, whereas K2Cr2O7 was used as the positive reference substance.

3.4 Statistical methods Statistical DOE and statistical models were used to study the effect of treatment parameters (factors) and their interaction terms on pollutant removal. Either full factorial design containing all possible variations (II and III) or central composite face (V) design was used. The factors range and levels are presented in Tables 1.9e1.11. Full factorial design (Table 1.9) was conducted separately for all three electrolysis design cells when the effect of cell construction on NOM removal was studied (II). Statistical models were made of the results using MODDE software (Umetrics) with either partial least squares regression (PLS) (II and V) or multiple linear regression (MLR) (III). Both are commonly used statistical methods which are used to interpret the effect of variables (factors) to responses. In MLR and PLS, the coefficients of the model are computed to minimize the sum of squares of residuals, i.e., the variance not explained by the model. MLR assumes responses to be independent, whereas in PLS the covariances of responses are taken into account. These methods are described in detail elsewhere, such as [192]. An equation describing the relationship between the response variable and predictor variables (factors) is obtained. The Table 1.9 Factor Range and Levels Used in Electrolysis Cell Construction Studies (II). Factor Range and Levels Variable X1 X2

Electric charge (C/L) Initial pH

L1

0

1

48 3

96 4

144 5

44 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

Table 1.10 Factor Range and Levels Used in Temperature Studies (III). Factor Range and Levels Variable X1 X2 X3

Electric charge (DL) Initial pH Temperature (K)

L1

0

1

48 4 275.15

96 5 285.15

144 6 295.15

Table 1.11 Factor Range and Levels Used in Debarking Effluent Studies (V). Factor Range and Levels Variable X1 X2 X3

Current (A) Time (s) Initial pH

L1

0

1

1 60 4

2 210 6

3 360 8

general form of the regression equation including single terms, square terms, and interaction terms is presented in Eq. (1.18): Y ¼ b0 þ

Xk Xi pH 4 > pH 5. This order was the same at all tested temperatures. The effect of initial pH on the dissolution rates of aluminum anodes and cathodes was studied with different cell constructions (II). Dissolution of Cell A was close to the theoretical values calculated according to Faraday’s law at the tested initial pHs. According to this, the proportion of side reactions on anode, such as the production of oxygen or chlorine, was low or negligible in the tested conditions. Mouedhen et al. [78] obtained similar results in their tests with aluminum electrodes.

48 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

The results of Cell C indicated that there was short time lag at acidic initial pH (pH 3 and pH 4) before the chemical dissolution of cathodes initiated. Mouedhen et al. [78] reported that in pHs lower than 3, the released amount of aluminum during the electrolysis with constant charge per volume was significantly lower than in pHs above 3. Acidic pH seemed to inhibit or decrease the dissolution rate of cathodes depending on the pH. The highest aluminum concentrations per added electric charge were measured for Cell T, but it should be noted that aluminum concentrations of Cell T were not the sum of concentrations measured from samples treated with Cell A and Cell C. There was some variation in dissolved aluminum concentrations in temperature tests (III) and cell construction tests (II), even when initial pH and temperature were the same in both tests (initial pH 5 and temperature 295.15 K). Variation in initial water quality in these tests, such as different amount of salts in water, might have caused this effect. It is known that chlorides and sulfates have a significant effect on the dissolution rate of electrodes during EC [85].

4.2.2 Iron electrodes Dissolution of iron electrodes was studied in various current densities during the sulfide treatment (IV). According to the results, current density did not have a significant effect on the dissolution rate in the studied range (3.6e17.9 mA/cm2) (Fig. 1.9). Iron dissolved at the rate of 0.27 mgFe/C, whereas the theoretical value for dissolution of Fe(II) is 0.29 mgFe/C. It can be concluded that iron dissolved in Fe(II) form and dissolution followed Faraday’s law. This was consistent with the results of other researchers [72,95,96]. This result was significant because it is known that Fe(II) is a poor coagulant and should be oxidized to Fe(III) form before it can be used to remove organic matter [3]. However, Fe(II) can form precipitates with sulfides according to Eq. (1.7). The proportion of other electrochemical reactions on anodes was low in the tested conditions. Water pH does not have significant effect on the dissolution of iron electrodes in near neutral pHs [72].

4.3 Change of pH and conductivity Change of pH during the EC treatment is a well-known side effect of EC [76,127,194]. The effects of initial pH, temperature, electric charge, and electrolysis cell construction on pH change were studied (II and III). When initial pH was highly acidic (pH 3), the alkalinity produced by electrochemical reaction was not sufficient to neutralize pH and final pH was 3.3e4.4 when the electric charge was 48e144 C/L. When initial pH was 4e6, the final pH was neutralized to pH 5.6e7.8. According to the results,

4. Results and Discussion 49

2

3.6 mA/cm 2 10.7 mA/cm 2 14.3 mA/cm 2 17.9 mA/cm

110 100 90 80

95% confidence level 2 Linear fit (R =0.99)

Fe (mg/L)

70 60 50 40 30 20 10 0 0

50

100

150

200

250

300

350

400

Electric charge (C/L) n FIGURE 1.9 Dissolving of iron electrodes in various current densities (IV).

the pH increment was higher when initial pH was acidic (except in very acidic initial pH 3), and therefore, it could be concluded that less HO ions were bound to the sludge produced in water having low initial pH. This may have an effect on sludge properties. Tests with different cell constructions indicate that final pH also depended on the cell construction (II). This effect was due to different Al/HO ratios produced by different cell constructions. In initial pH 5 and electric charge 144 C/L, the final pHs of waters treated with Cell T, Cell C, and Cell A were 6.83, 7.47, and 7.76, respectively. Respective aluminum concentrations were 18.42 mg/L, 16.26 mg/L, and 12.04 mg/L. The amount of HO ions produced by electrolysis followed Faraday’s law. Aluminum, on the other hand, dissolved from both electrodes but chemical dissolution of cathodes did not follow Faraday’s law. A higher concentration of free HO- ions led to higher final pH. The effect was similar in other tested initial pHs. Decreasing temperature slightly increased the final pH due to this same effect; aluminum cathodes dissolved at a lower rate at lower temperatures, and therefore, there were more free HO ions in the solution. Water pH also increased when iron electrodes were used (IV and V), except at initial pH 8 when short treatment times were used. When DOC and COD removals were studied with high electric charges (4500 C/L), the final pH of the wastewater increased as high as 10.4. It was expected that pH increases

50 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

faster and higher with iron electrodes than with aluminum electrodes because less coagulant was produced per produced HO ions, because iron did not dissolve on cathodes. Produced Fe(II) bound less HO ions compared with Al(III), and therefore pH increased to a higher level because sludge contained less HO ions. Al(III) forms AlðOHÞ4  species at high pH, which stabilizes pH change. The transformation of Fe(II) to Fe(III) at high pH and oxidizing conditions decreased pH because Fe(III) bound more HO ions. Conductivities of samples depended on the concentration of residual metal coagulant and final pH. When initial pH was acidic, conductivity was naturally higher than in neutral initial pH because more ions were added into the solution. In temperature tests, water conductivity without EC treatments was 2.8 mS/m, and after small pH correction (initial pH 6) and treatment with 144 C/L at 295.15 K the final conductivity was 3.0 mS/m. Water conductivity was 3.5 mS/m and 3.3 mS/m when waters having initial pH 4 and 5 were treated with the same parameters, respectively. It could be concluded that optimized EC treatment had a negligible effect on water conductivity. This differs from metal salt coagulants, which significantly increase the concentration of salts, such as chlorides or sulfates, in water.

4.4 Surface water treatment 4.4.1 Organic matter removal The effects of current density, electric charge, initial pH, temperature, and electrolysis cell construction on NOM removal from surface water samples were studied (IeIII). In addition, the removal of NOM by combined chemical coagulation and EC was investigated (I). When current density increased, the same amount of organic matter was removed in a shorter treatment time as when the production rate of coagulant increased (I). However, when organic matter removal was observed as a function of electric charge, low current density was more effective. According to Mouedhen et al. [78], the dissolution rate of cathodes does not increase as a function of current density in contrast to the dissolution of anodes. Even low current density initiated chemical dissolution of cathodes, and therefore, more aluminum was produced per electric charge in low current density, which in turn meant that low current density was more energy efficient. DOC or TOC removal increased when added electric charge per volume increased. Initial pH had a significant effect on the removal of organic matter. At initial pH 3, there was no significant effect whether aluminum originated from anodes or cathodes (II). The TOC removal rate was

4. Results and Discussion 51

18

Cell T Cell A Cell C Linear fit (R2=0.94) 95% confidence level

16

TOC (mg/L)

14 12 10 8 6 4 0

5

10

15

20

Aluminum (mg/L) n FIGURE 1.10 TOC concentrations of the samples treated with different cell constructions at initial pH

3 (II).

0.83 mgTOC per mgAl added into the solution. TOC concentration depended linearly on the aluminum concentration added into the solution. The statistical model showed somewhat false results at this initial pH, especially for Cell C (Fig. 1.10). At low pH, the mechanism of NOM removal was mainly charge neutralization and compression of double layer, because hydroxides were not stable in this environment [3,10]. It is noteworthy that the TOC removal rate as a function of aluminum concentration did not decrease, even at low TOC concentrations at this initial pH. Therefore, it is possible that even higher TOC removal could be obtained if even higher aluminum concentrations would be used. When initial pH was 4e6, the removal efficiency increased when initial pH decreased, as seen in Fig. 1.11 (III). Significantly higher aluminum concentrations were required per removed DOC at initial pH 6 than at initial pH 4 or 5. At initial pH 4e5, even the lowest obtained aluminum concentration removed most of the organic matter from water. When pH increased, aluminum species lost charge because they bound more hydroxyl ions and therefore more aluminum was required for charge neutralization. The primary mechanism of NOM removal at near neutral pHs was, therefore, precipitation and enmeshment, which required a higher concentration of aluminum compared with charge neutralization [2,3]. Similar results were obtained when the effect of cell constructions was studied at initial pH 4 and 5 (II). When initial pH was 6, the NOM concentration of water did

52 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

14 13

pHi 4

12

pHi 5 pHi 6

DOC (mg/L)

11

Linear fit (pHi 4, R2=0.33)

10

Linear fit (pHi 5, R2=0.68) Logarithmic fit (pHi 6, R2=0.77)

9 8 7 6 5 4 5

10

15

20

25

30

35

40

45

Aluminum (mg/L) n FIGURE 1.11 Dissolved organic carbon concentration of samples treated at different initial pHs (pHi).

Temperature during the treatment was 275e295 K (III).

not decrease linearly as a function of aluminum concentration, rather it followed logarithmic or asymptotic fit. Similar results have been reported in literature by other researchers [160,161,163]. According to their results, the efficiency of the NOM removal increased when initial pH decreased. At near neutral initial pH 6.2, there was a short time lag before any organic matter was removed by EC (I). According to this, a threshold aluminum concentration exists at neutral pH which has to be passed before enmeshment mechanism initiates. TOC removal stabilized at approximately 4 mg/L or 5 mg/L when initial pH was 4 or 5, respectively (II and III). It seemed that residual NOM was not easily removed by EC at these pHs. Water temperature and cell construction affected through aluminum concentration, but besides that they did not have a major effect on NOM removal. For example, at 295 K, more NOM was removed than at 275 K or 285 K (Fig. 1.12) as more aluminum was dissolved into the solution. Chemical coagulant additions reduced the required electric charge per volume for optimum organic matter removal (I). The apparent color of surface water was due to NOM, and both organic matter and apparent color were removed simultaneously during the treatment (IeIII). However, the color intensity of water also varied with pH. Increasing pH increases the color intensity [3]. The lowest measured color value was 3 mg/L PtCo (96.4% removal) (III) when water having initial

4. Results and Discussion 53

n FIGURE 1.12 Contour plots of dissolved organic carbon concentration (mg/L) in (A) 275 K, (B) 285 K, and (C) 295 K according to statistical model (III).

pH 4 was treated with 144 C/L at 295.15 K. Aluminum originating from anodes and cathodes had similar color removal efficiency as can be seen in Fig. 1.13. At initial pH 3, color was removed linearly with increasing aluminum concentration, whereas at initial pHs 4 and 5 the color removal followed asymptotic or logarithmic fit. More color was removed per electric charge when low current density was used, because more aluminum was produced per electric charge in low current density (I). In all tested current densities, cell configurations, temperatures, and initial pHs, high color removal could be obtained when aluminum dosage was sufficient. The initial SUVA value of surface water was 4.52 L/(m mg) (II) or 3.51 L/ (m mg) (III). According to this, the water contained mainly hydrophobic

130

Apparent color (mg/L PtCo)

120 110

Cell T (pH 3)

Cell A (pH 3)

Cell C (pH 3)

Cell T (pH 4)

Cell A (pH 4)

Cell C (pH 4)

Cell T (pH 5)

Cell A (pH 5)

Cell C (pH 5)

100 Linear fit (pH 3, R =0.91)

90

Asymptotic fit (pH 4, R =0.89)

80

Logarithmic fit (pH 5, R =0.90)

70 60 50 40 30 20 10 0 0

5

10

15

20

Aluminum (mg/L) n FIGURE 1.13 Apparent color of samples treated in different initial pHs (pHi) and cell constructions (II).

54 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

humic material. The SUVA value decreased during the treatment simultaneously with DOC or TOC and color. Temperature or cell constructions did not have a significant effect on SUVA values besides their effect on the dissolving of electrodes. The SUVA value decreased to 1.5e2.5 L/(m mg) during the treatment, except at two points at initial pH 6 and one point at initial pH 4 where SUVA values were higher (III). The effects of aluminum concentration and initial pH on SUVA values were also studied with different cell constructions (Fig. 1.14). SUVA values exhibited similar behavior to apparent color values. At initial pH 3, the SUVA value decreased linearly with aluminum concentration, whereas at initial pH 4 it followed asymptotic fit. At initial pH 5, there was more variation in the results. However, according to the results, EC removed mainly hydrophobic material, and residual NOM fraction was composed mostly of hydrophilic nonhumic material which was low in aromatic chemical structures. When these results were compared with other studies, it could be concluded that EC and chemical coagulation are similar in terms of aluminum doses per TOC or DOC removed, optimum pH of the treatment and fractions remaining in water after the coagulation [11,19,195]. Chow et al. [19] studied DOC removal with alum from four different surface waters. They obtained optimum DOC removal at pH 5, which was between the optimum initial pH and final pH of the EC treatment, according to this study. Alum treatment decreased the SUVA value of waters, and recalcitrant DOC was

6.0 5.5

Cell T (pH 3)

Cell A (pH 3)

Cell C (pH 3)

Cell T (pH 4)

Cell A (pH 4)

Cell C (pH 4)

Cell T (pH 5)

Cell A (pH 5)

Cell C (pH 5)

Linear fit (pH 3, R =0.86)

5.0

Asymptotic fit (pH 4, R =0.93)

SUVA [L/(m mg)]

4.5

Linear fit (pH 5, R =0.66)

4.0 3.5 3.0 2.5 2.0 1.5 1.0 0

5

10

15

20

Aluminum (mg/L) n FIGURE 1.14 Specific UV absorbance values of samples treated in different initial pHs (pHi) and cell

constructions (II).

4. Results and Discussion 55

found to have smaller molecular weight and lower SUVA value compared with the DOC of the raw water. Shi et al. [195] studied the coagulation of humic acid by preformed and nonpreformed aluminum species. According to their results, decreasing pH reduces the coagulant dosage required for complete UV254nm removal. Large molecular and hydrophobic humic material was more effectively removed with nonpreformed aluminum species, which are also released during the EC treatment. The z-potential of the untreated water samples was 15.2 mV (II) or 17.1 mV (III). The results indicated that temperature affected z-potential, due to its effect on the dissolving speed of the aluminum electrodes, as reflected by the contour plots of the model (Fig. 1.15). The z-potential of the residual organic matter decreased when more aluminum was added to the solution. Water pH also had a significant effect on the z-potential of pollutants and metal hydroxides as discussed in Section 1.1.1. However, the model quality terms were low for the z-potential at the temperature studies. This was probably due to other noncontrolled parameters which affected z-potential values. Indeed, there was correlation between residual aluminum concentrations and z-potential values. At low initial pHs 4 and 5, the z-potential increased as a function of residual aluminum concentration, whereas at initial pH 6 it decreased slightly. These results also indicated that aluminum species had a more positive charge when they were formed in acidic initial pH than in neutral initial pH, even when the final pH of water was similar. The effect of initial pH on the z-potential was also studied with different cell constructions (II). At initial pH 3, the aluminum had a higher charge (Al3þ) due to lack of hydrolysis, and the z-potential of residual aluminum increased linearly with increasing aluminum concentration (Fig. 1.16). At this pH,

n FIGURE 1.15 Contour plots of z-potential (mV) in (A) 275 K, (B) 285 K, and (B) 295 K according to statistical model (III).

56 CHAPTER 1 Electrocoagulation in the treatment of industrial waters and wastewaters

20 15 10

Cell T (pH 3)

Cell A (pH 3)

Cell C (pH 3)

Cell T (pH 4)

Cell A (pH 4)

Cell C (pH 4)

Cell T (pH 5)

Cell A (pH 5)

Cell C (pH 5)

Linear fit (pH 3, R =0.96)

ζ-potential (mV)

5 0 -5 -10 -15 -20 -25 -30 -35 -40 0

5

10

15

20

Aluminum (mg/L) n FIGURE 1.16 z-potential as a function of aluminum concentration at different initial pHs (pHi) and

cell constructions.

small precipitates were formed that were not easily removed with sedimentation, flotation, or filtration, and therefore the residual metal concentration in water was higher than at initial pH 4 or 5. When the aluminum concentration was over 12 mg/L, the z-potential exceeded 0 mV, which may cause restabilization of colloids. z-potentials of the samples at initial pHs 4 and 5 were similar after filtration. At these pHs, aluminum existed predominantly as hydroxides, and easily settleable or filterable flocs were formed, and therefore, residual aluminum did not have a significant effect on the z-potential of the samples. It was interesting that residual organic matter had more negative z-potential values at initial pH 4 and 5 than that measured from the untreated samples. According to this, some highly charged fraction of NOM was not removed by EC.

4.4.2 Residual metals and turbidity Residual metals may cause problems, such as deposition and process chemical decomposition, in processes where water is consumed. There are also concerns that residual aluminum in drinking water may cause Alzheimer’s disease [21]. The effects of current density, temperature, initial pH, and cell construction on concentrations of residual metals were studied. The water had low initial turbidity, and therefore turbidity was mostly associated with residual aluminum concentration as aluminum hydroxides form milky gel precipitates in water.

4. Results and Discussion 57

Residual aluminum concentration depended on water pH and aluminum added to the solution. When aluminum concentrations were low, there was more residual aluminum in water, probably because floc size was small and passed more easily through the filters. At highly acidic initial pH 3, the residual aluminum concentrations were high (2.5e8.3 mg/L) because final pH remained acidic, and more soluble aluminum and small flocs were present than at higher initial pHs (II). When initial pH was 4.0e6.2 and added aluminum concentrations were high, the residual aluminum concentrations were low (IeIII). Negligible aluminum concentrations were measured in these conditions (90%

[67]

40% COD

[68]

40% COD

[68]

100% >70% COD

[69]

96% 82.5% COD

[70]

100%

[71]

98.8% 43.1% TOC

[72]

95.5% 74.5% TOC

[73]

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 97

Table 2.2 Performance of Different Electrode Materials in Direct and Indirect Electrochemical Oxidation of Organic Compounds. continued Anode Type

Targeted Pollutant

Working Parameters

Pyridine (100 mg/L)

Na2SO4 (10 g/L) I ¼ 30 mA/cm2 pH 3 tel ¼ 3 h NaClO4(10 mM) I ¼ 10 mA/cm2 tel ¼ 1.5 h NaClO4(10 mM) I ¼ 10 mA/cm2 tel ¼ 1.5 h Na2SO4 (5 wt.%) I ¼ 20 mA/cm2 tel ¼ 5 h Na2SO4 (0.2 M) I ¼ 50 mA/cm2 tel ¼ 2 h NaClO4(10 mM) I ¼ 10 mA/cm2 tel ¼ 1.5 h H2SO4 (0.5 M) I ¼ 30 mA/cm2 tel ¼ 10 h T ¼ 23 C Eel ¼ 5 V tel ¼ 2 min Na2SO4 (0.1 M) I ¼ 5 A/dm2 pH 3 tel ¼ 4 h Na2SO4 (0.25 M) I ¼ 10 mA/cm2 tel ¼ 6 h I ¼ 0.3 A tel ¼ 6 h Na2SO4 (0.2 M) I ¼ 20 mA/cm2 pH 7 tel ¼ 0.5 h Na2SO4 (0.2 M) I ¼ 20 mA/cm2 pH 7 tel ¼ 0.5 h Na2SO4 (0.2 M) I ¼ 20 mA/cm2 pH 7 tel ¼ 0.5 h

Perfluorooctanoic acid (100 mg/L) Ti/SnO2eSb/PbO2

Perfluorooctanoic acid (100 mg/L)

Ti/TiOxHy/SbeSnO2

Aniline (500 mg/L)

PbO2eZrO2

Methylene blue (30 mg/L)

Ti/SnO2eSb/MnO2

Perfluorooctanoic acid (100 mg/L)

Sn0.86eSb0.03eMn0.10e Pt0.01eoxide/Ti

4-Chlorophenol (8 mM)

Stainless Steel/SnO2e CeO2 Ti/RuO2eIrO2eSnO2e TiO2

Dye wastewater (color, 1565 PCU; COD, 188.94 mg/L) Phthalic anhydride (2 g/L)

BDD

Phenol (100 mg/L)

Sanitary landfill leachate (COD 6.2 g/L) Dimethyl phthalate (0.03 mM)

Diethyl phthalate (0.03 mM)

Diheptyl phthalate (0.03 mM)

Removal Efficiency

References

98%

[74]

90.3%

[75]

91.1%

[75]

85% 71% COD

[76]

100% 72.7% COD

[77]

37.1%

[75]

100%

[78]

83% color 48.6% COD 88% COD

[79]

100% 95.4% TOC

[73]

40% COD

[68]

100% 50% COD

[64]

100% 50% COD

[64]

50% 10% COD

[64]

[80]

98 CHAPTER 2 Ultrasound-assisted electrochemical treatment

2.2 Sonochemical destruction methods of organic pollutants 2.2.1 The theory and mechanism Ultrasound is a sound wave having a frequency above the audible sound typically higher than 20 kHz and up to 106 kHz. Ultrasound in a medium is transmitted via series of compression and rarefaction cycles, which lead to formation of high and low pressure zones in the medium, respectively [81]. In rarefaction cycle, when high power of the cycle exceeds the attraction forces of the liquid molecules, cavitation bubbles filled with saturated liquid vapor are formed [82]. Bubbles undergo a continuous growth and compression during subsequent cycles in a process called rectified diffusion (Fig. 2.2). Stable cavitation bubbles will reach an equilibrium size and continue to vibrate producing acoustic microflows. Transient cavitation bubbles will continue to grow in size and then collapse violently generating tremendous pressures up to 400 MPa and temperatures of gases inside the bubbles (2000e5000K) and in the liquid surrounding the cavity (1500e2000K) [83]. Immediately after collapse, spherical shock waves are transmitted in the liquid rapidly attenuating. Static equilibrium of bubbles in water is described by Eq. (3) [84]. pg þ pv ¼ phs þ

2g r

(Eq. 3)

where pg and pv are pressure of gas and vapor inside the bubble, respectively, phs is the hydrostatic pressure, g is the surface tension, and r is the bubble radius.

n FIGURE 2.2 The mechanism of cavitation bubble formation and collapse.

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 99

A sudden drop in hydrostatic pressure in the liquid below a certain value causes the formation of cavitation bubbles as a result of rupture of fluid continuity. Pure liquids without bubble inclusions have a high tensile strength. For example, to form a cavity in water with a radius of 10
10 m at 293K, it is necessary to apply a negative pressure p ¼ 2g/r ¼ 2  0.073/ 10
10 ¼ 1.46 GPa. However, it is a very difficult task to obtain a pure liquid without heterogeneity. The presence of gas bubbles and gas nuclei trapped in suspended solid particles create “weak spots” in the liquid, which consequently reduce tensile strength of the liquid. As a result, significantly less pressure (about 103e104 times) is required to initiate cavitation [85]. The cavitation bubble growth depends on ultrasonic frequency. Higher ultrasonic frequencies promote faster bubble collapse thus producing smaller cavitation bubbles [83]. External conditions and fluid properties influence the cavitation development and intensity (the amount of energy passed per unit time per unit area perpendicular to the ultrasound beam, J/s/m2 ¼ W/m2). Degassing of the liquid reduces cavitation and increases the intensity of the shock waves created by cavitation bubble. The higher the heat capacity and solubility and the lower the thermal conductivity of the ambient gas, the better the cavitation induced [86,87]. While increasing the medium temperature, the pressure inside the bubble increases and the fluid viscosity decreases. These conditions promote more ease cavitation occurring at lower powers and less violent bubble collapse [88]. The minimum ultrasonic intensity required for initiation of cavitation is called cavitation threshold, which depends on many factors such as ultrasonic frequency, liquid properties, ambient temperature, etc. [89]. The asymmetric bubble collapse (Fig. 2.3) on/or close to a solid surface provokes the generation of liquid microjets directed toward the surface with speeds up to few hundred meters per second [83,90]. The presence of solid particles or other foreign liquids in the area of liquid microjet action leads to their active degradation due to high pressures at bubble collapse. The physical effect of ultrasound on compound degradation is also based on

n FIGURE 2.3 Asymmetric cavitation bubble collapse near a solid surface producing a liquid microjet.

100 CHAPTER 2 Ultrasound-assisted electrochemical treatment

formation of liquid microjets in addition to shock waves, acoustic streaming, and shear forces. Acoustic cavitation in liquids initiates various physical and chemical phenomena such as sonoluminescence, sonochemical reactions, solid surfaces erosion, mixing and homogenization of immiscible liquids, etc. Because imploding bubbles can produce excited molecules in high-energy states, these molecules emit visible light when returning to ground state, a phenomenon known as sonoluminescence [91]. Sonochemical reactions depend on many parameters including ultrasonic frequency and power, pH, temperature, materials types, etc. It was shown in many studies that at higher frequency, higher amount of OH radicals are generated, which are the main reactants in chemical transformations [92,93]. At the same time, collapse that is more violent is produced at lower frequencies providing higher temperatures and pressures [94]. This is explained by long acoustic cycles and as a result larger bubble sizes. Violent cavitation bubble collapse is also formed at higher ultrasonic intensities [95]. It should be noted that, unlike most chemical reactions, sonochemical reactions decrease with increasing ambient temperature. This is because the higher the ambient temperature, the greater will be the extra vapor intake inside the cavity, which cushions the implosion of the cavity and lowers the temperature of its collapse [83]. A summary of different parameters on cavitation process is listed in Table 2.3.



Table 2.3 The Main Parameters that Influence the Cavitation. Parameter Frequency

Dissolved gases

Intensity

Effect of Different Parameters of Cavitation Process Higher frequency leads to shorter acoustic cycles, smaller cavitation bubble size, shorter bubble collapse time, and higher OH formation. Lower frequencies provide more violent collapse. Reduced tensile strength of the liquid initiates the cavitation process at lower pressures. The more the gas nuclei in the liquid, the lower the intensity of generated shock waves. Monatomic gases generate more energy on collapse than multiatomic gases, which is related to the heat conductivity of gases. There is a minimum level of intensity, which is required to initiate cavitation. Higher intensity generates more bubbles. However, too many bubbles may prevent the distribution of ultrasonic power into entire solution; therefore, an optimal level of intensity should be found.

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 101

Table 2.3 The Main Parameters that Influence the Cavitation. continued Parameter Temperature

Pressure

Effect of Different Parameters of Cavitation Process Higher ambient temperature reduces viscosity of the liquid and surface tension, which results in cavitation threshold decrease and higher bubbles formation; however, bubble collapse is less violent. Too many bubbles may reduce the distribution of ultrasonic power into entire solution. Temperatures above the boiling point of the media significantly reduce the effect of sonochemical reactions. Higher ambient pressure reduces the number of bubble collapses at the same intensity; however, collapses are more violent.

In addition to the described parameters, transducer placement within the reactor and reactor geometry influences the distribution of ultrasonic field and, as a result, efficiency of sonochemical reactions. However, the modern ultrasonic systems are equipped with amplitude unit control, which allow automatic adjustment of amplitude to a constant value, and a unit for automatic overlapping ultrasonic waves, which is achieved by sweeping the frequency in the range of þ1 kHz, which eliminates the formation of dead zones and hot spots [96,97]. There are three theories explaining cavitation bubble collapse. These are (i) electrical, where electrical discharge is produced during asymmetric bubble implosion [98]; (ii) plasma theory, implying microplasma formation inside the cavitation bubble [99]; and (iii) hot spot theory, which is the most used due to its simplicity and the fact that other theoretical approaches could not describe various findings [95]. In the hot spot theory, there are three regions where chemical reactions take place (Fig. 2.4). A hot spot hydrophobic gaseous nucleus (1), which is created due to the compression of gases, and liquid vapors inside the cavity initiates high local temperatures inside the cavity and in interfacial region (2). Because temperature of hot spot is enormously high, the heat disperses rapidly and warms up the bulk solution (3). It was suggested that degradation of compounds inside the cavity (hot spot) including solvent decomposition occurs through the thermal decomposition mechanism. Pyrolysis and hydroxylation are the main pathways of compounds decomposition in interfacial region. Decomposition of compounds in the bulk solutions occurs by hydroxylation and reactions with hydroxyl radicals [100].

102 CHAPTER 2 Ultrasound-assisted electrochemical treatment

n FIGURE 2.4 The hot spot model of cavitation.

Energy released during the collapse of the bubble is enough to excite, ionize, and dissociate water molecules, gases, and substances. At this stage, any gas present in the liquid is an active component participating in the transfer of excitation energy, recharging, and other processes. The oxidation of organic species presenting in those three areas occurs through reaction with OH radicals produced during water thermolysis and H2O2 obtained from the recombination reaction of OH radicals [101,102]. H2O / OH þ H

(2.30)

H, þ O2 /HO2 ,

(2.31)

2 OH / H2O2 

(2.32)

2HO2 , / H2 O2 þ O2 H þ O2 / OH þ O 



(2.33)



The action of reactive gases as, for example, O2 and H2 inside the cavitation bubble is ambivalent. H2 and O2 are involved in radical transformation reactions (2.31 and 2.34) 

OH þ H2 / H þ H2O

(2.34)

O2 / 2O

(2.35)

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 103

Additionally, inert and noble gases participate in transfer of electron excitation energy to water molecules. The presence of gases, salts, or nanoparticles in treated solution can enhance the degradation rates of pollutants due to availability of additional centers for the cavitation or the formation of additional radicals. Persulfate and periodate ions can be considered as an example of high reactive secondary radicals formation [103e105]. S2 O8 2
þ ÞÞÞ/2SO4 ð
,Þ

(2.36)

IO4
þ ÞÞÞ/IO3 , þ Oð,
Þ

(2.37)

O,
þ Hþ /, OH

(2.38)

, OH þ IO4
/OH
þ IO4 ,

(2.39)

The addition of carbon tetrachloride was found to enhance sonodegradation of organic pollutant through the degradation of organic compounds through the formation of chlorine-containing radicals and recombination with H, which prevents OH recombination with H [106e108]. CCl4 þ ))) / CCl3 þ Cl

(2.40)

CCl3 þ H / HCCl3

(2.41)

HCCl3 / HCCl2 þ Cl

(2.42)

HCCl2 þ H / H2CCl2

(2.43)

H2CCl2 þ ))) / H2CCl þ Cl

(2.44)

Cl þ Cl / Cl2

(2.45)

Cl2 þ H2O / HClO þ HCl

(2.46)







2.2.2 Sonochemical degradation of organic compounds The first reports on the effect of ultrasound on the course of chemical reactions were published by Wood R.W. and A.L. Loomis [109]. In 1929, A.L. Loomis and E.N. Harvey showed the effect of ultrasound on deactivation of luminous bacteria [110]. In 1950, Miller was one of the first suggesting formations of reactive species from the disruption of water molecules during acoustic cavitation [111]. A few years later, Henglein and Weissler showed that hydrogen peroxide was generated during the sonication of water

104 CHAPTER 2 Ultrasound-assisted electrochemical treatment

[112,113]. In 1950, Weissler also found that the addition of CCl4 enhances the degradation of KI, generating free chlorine under at CCl4 sonication [114]. The first works confirming the direct generation of OH and H radicals and their effect on ultrasonic degradation of organic compounds were conducted in 1970se1980s [115,116]. For the first time, the mechanism of thymine degradation at 450 kHz ultrasonication was described through radical formation, and degradation products were identified [115]. Creation of inexpensive ultrasonic generators in the 1980s promoted further studies on organics remediation using ultrasound [83]. Despite the considerable amount of research conducted on the sonochemical degradation of organic compounds, comparison of different studies provides a random error due to the different transducers geometry and spatial set-up used in the studies. For the last 30e40 years of intensive development of ultrasonic technologies, sonochemical remediation has been found effective in a wide range of applications as an individual process and assisting other water treatment technologies such as EO, photodegradation, chemical oxidation, adsorption, membrane filtration, biological treatment, coagulation, etc. The individual ultrasound remediation techniques use the energy of ultrasound as a main reactant for pollutant degradation. Degradation of organic compounds such as pesticides, dyes, pharmaceuticals, herbicides surfactants, solvents, and industrial chemicals is the most studied field in ultrasonic wastewater treatment [117,118]. However, dyes are the leaders among other pollutants investigated in sonochemical degradation. The recent study of the degradation of rhodamine B, acid orange 7, and malachite green dyes at 300 kHz (0.085 W/mL output power) sonication provided 85%, 80%, and 95% dyes degradation efficiency of 5 mg/L dyes, respectively, after 1 h at 25 C-treated solution temperature [119]. The mechanism of pollutant degradation was the reaction with OH radicals. Degradation experiments were conducted in a cylindrical glass reactor with a transducer incorporated into the bottom of the reaction vessel. The increase of liquid temperature up to 55 C significantly approved dye degradation, which was explained by the fact that OH radical formation and number of cavitation bubbles increase with the temperature rise. The degradation efficiency of magenta dye equal to 80% was obtained in ultrasonic bath after 3 h sonication at pH 7.2 of 30 mg/L dye solution [120]. The ultrasonic frequency used in the study was 50 kHz with output power of 0.085 W/ mL. Addition of NaCl to the treated solutions slightly enhances the degradation efficiency of the dye till 85% while applying the same working parameters.

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 105

It was found that the thermal decomposition mechanism is responsible for the degradation of hydrophobic compounds such as trichloroethylene, tetrachloride, chlorobenzene, etc., which occurs mostly inside the cavitation bubble [121,122]. Degradation of hydrophilic compounds occurs in water phase mostly through their reactions with formed radicals generated by gaseous vapors thermolysis [92,123]. It was shown that phenol decomposition occurs through the reaction with hydroxyl radicals resulting in generation of phenoxy radical, which react with HO once again forming nonradical degradation products [124]. HO þ PhOH / PhO þ H2O

(2.47)

PhO þ HO / nonradical products

(2.48)

The degradation of phenol (1100 mol/L initial concentration) using 28 kHz ultrasonic bath followed by the zero-order kinetics model and efficiency of degradation after 3 h of sonication did not exceed 25% [124]. Most of the studies applied ultrasound of low to medium frequency which was found effective, for example, for dyes decomposition [86,94,125]; however, the most used frequencies are between 20 and 100 kHz, which is explained by the availability of low ultrasonic frequency devices on the market. Only a few studies used ultrasound frequency higher than 1000 kHz [126e129]. Thus, it was shown that irradiation of solution containing 10 mol/L MB dye in ultrasonic bath type reactor with 1640 kHz frequency (0.08 W/mL actual power) for 1 h provided 25% of dye degradation efficiency [130]. To the best of our knowledge, 1640 kHz was the highest frequency tested in MB degradation. However, while comparing the sonication of MB solution at 1640 kHz with applied frequencies of 22.8, 127, 490, and 970 kHz, the highest degradation of about 50% was achieved using 490 kHz. Therefore, once again the necessity of careful optimization of ultrasonic working parameters was shown. At the same time, the higher ultrasonic powers applied showed the better degradation rates of MB at every studied frequency. The positive effect of the pulsed mode, especially over longer period of sonication, was also confirmed in some studies [131,132]. Sonication of high-molecular weight polysaccharide extracted from Phellinus linteus mycelia at 20 kHz and 10e30 W/mL output power in a pulsed mode (2 s ond2 s off) provided 64%e93% molecular weight removal with higher decrease at higher output powers. The initial molecular weight of 3.1$108 g/mol in 5% w/v polysaccharide solution dropped to about

106 CHAPTER 2 Ultrasound-assisted electrochemical treatment

0.2 and 1.1 g/mol for 10 and 30 W/mL, respectively, after 3 h of sonication [133]. It was found that the use of multiple frequencies of (20 þ 30 þ 50) kHz allowed enhancing the degradation of FA by nearly 50% comparing with single sonication of 100 ppm FA solution at 30 kHz [134]. Multiple and single frequency experiments were carried out in a hexagonal flow cell with each set of transducers incorporated into each side wall of the reaction vessel. In the same study, it was shown that higher volumes of treated solutions with the same pollutant concentration of 100 ppm provided higher degradation efficiency of FA of 9% using 1700 mL solution against 4% while sonicating 750 mL of the treated solution. This was explained by the fact of standing wave formation at higher volumes, which lead to higher ultrasonic intensity generated at higher volumes. Another study conducted on the degradation of rhodamine B dye using simultaneous irradiation of solution with double frequency mode (25 þ 40 kHz) showed the degradation efficiency be equal to a sum of efficiencies of singular degradation processes at 25 and 40 kHz [135]. Reaction was conducted in a cylindrical cell with three set of transducers for each frequency fixed at opposite walls of the reactor. Degradation of FA at 20, 200, and 607 kHz frequencies showed 6 and 8 times higher degradation of the pollutant at higher frequencies of 200 and 607 kHz (transducers incorporated at the bottom of the reactor), respectively, comparing with 20 kHz (ultrasonic probe system) sonication [136]. The mechanism of sonochemical degradation of FA showed to be complicated with formation of different intermediates such as CO2, H2O, H2C2O4, and CO [136]. However, the main degradation occurs through FA decarboxylation shown in reaction 2.49e2.51 [136]. H2O þ )))) / H þ OH

(2.49)

HCOOH þ OH / HCOO þ H2O

(2.50)

HCOO þ OH / CO2 þ H2O

(2.51)

Mechanical stirring (200 rpm) of treated solution could improve the FA degradation by a factor of three from 2% to 6% compared with sonication without stirring when using 590 kHz ultrasonic frequency with output power intensity of 0.13 W/mL in degradation of 1000 mg/L FA in a cup-horn reactor with a transducer fitted at the bottom of the vessel [137]. The further increase of rotation speed decreased the efficiency of the process thus showing a need for stirring condition optimization. Addition of 4% NaCl

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 107

to the treated solution enhanced the efficiency of FA degradation, which was explained by creation of additional centers of cavitation and concentration of pollutant in the hot spot of cavitation bubbles. Higher amounts of salt decreased the efficiency of the process that was attributed to the fact of surface tension and vapor pressure changes preventing the formation of cavitation effect. The more detailed information on different organic compounds degradation by sonication is represented in Table 2.4.

Table 2.4 Efficiency of Ultrasonication in Degradation of Different Organic Compounds in Water. Working Parameters

Pollutant

Initial Concentration/ Comments

Power, W/mL

Frequency, kHz

Time, min

Removal Efficiency

Bisphenol A

0.228 mg/L

0.2

30

26.9 mg/L pH3; O2 2 mg/L 50 mg/L 10e20 mg/L

0.27 1

90 180

12

20

60

2.4 mg/L 110 mg/L O2 8.88 mg/L

0.6e1.8 0.201

20 500

60 100

61% 93% 100% 47% 30% 83% depolymerization 42%e58% 97%

[138]

Bisphenol A Bisphenol A

28 580 300 130

0.2

60

0.1 0.065 126 4.8

60 120 360 120

0.67

36

30

1 2.4

200 22.5

20 6

50% 90% 95% 88% 30% 5%TOC 13%TOC 8% 14% 95.6% intrinsic viscosity 79% 7.5%

[144]

40 mg/L 45 mM 450 mg/L (TOC) 100 mg/L 1000 mg/L 27.93 mL/g intrinsic viscosity 10 mg/L 1.4% 60 mL/g intrinsic viscosity 2 mg/L 10 mg/L 150 rpm

20 577 861 1145 20 283 20 22.7

0.17

35

200

63% 91%

[151]

Dextran 1000 (MW 1.2$106) Diazinon Dichlorvos Diclofenac

Diclofenac Diethyl phthalate Dinitrotoluene Formic acid Guar gum Linuron Low-density polyethylene Malachite green

References

[139] [140] [141] [142] [143]

[145] [146] [147] [134] [148] [149] [150]

Continued

108 CHAPTER 2 Ultrasound-assisted electrochemical treatment

Table 2.4 Efficiency of Ultrasonication in Degradation of Different Organic Compounds in Water. continued

Pollutant

Initial Concentration/ Comments

Methomyl Methylene blue Methyl orange Methyl parathion Methyl parathion Oxalic acid

25 mg/L 50 mg/L 10 mg/L 20 mg/L 50 mg/L 210 mg/L

Reactive brilliant red Ke2BP Sulfamethazine 2,4,6Trichlorophenol 17a-Ethinylestradiol 2,4-Dichlorophenol 2,3,7,8-TeCDD

Working Parameters Power, W/mL

Frequency, kHz

Time, min

Removal Efficiency

20 42 20 20 40 40

75 60 90 60 80 60

20

40

30% 10% 8% 8% 16% 3% 13% 12%

[152] [153] [154] [155] [156] [157]

20 mg/L pH3

1 0.5 0.36 0.072 1 0.35 0.56 4

9 mM 100 mg/L

0.2e0.5 1.8

800 20

60 150

41%e95% 25%

[158] [159]

296 mg/L

0.2

30

0.4

67% 98% 100%

[138]

50 mg/L pH 2e11; O2 20 ng/L pH 2

28 580 489

1.5

20

100%

[161]

60e 120 60

References

[106]

[160]

2,3,7,8-TeCDD, 2,3,7,8-Tetrachlorodibenzo-p-dioxin.

2.3 Sonoelectrochemical destruction methods EO/US degradation of organic compounds is a relatively new developing technique, which shows high mineralization rates in degradation of organic compounds. EO/US degradation methods are based on the synergetic effect of separate mechanism of EO and US degradation, which were described in Sections 2.1 and 2.2. The effects of sonication involving shock waves and shear forces promote almost all electrochemical processes. Acoustic streaming and liquid microjets activate an electrode surface by removing passivating impurities from the electrode surface, enhancing electrolytic current mode, and facilitating mass transfer of electroactive species [11]. Degassing effect of ultrasound can reduce polarization of electrodes by removing the generated gases from the electrodes surface. The extreme conditions generated at cavitation bubble collapse provide additional mechanisms of pollutant degradation through generation of hydroxyl radicals and thermal decomposition. However, free radicals produced during EO/US process can be also consumed by recombination reactions. When summarizing the

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 109

advantages and disadvantages of EO and US treatments, it can be concluded that combined EO/US treatment is suitable for the degradation of highly toxic compounds regardless of their concentration and turbidity at ambient conditions. Moreover, electrical energy is used as the main reactant in the process, and no additional chemicals are required. The synergetic effect of combined EO and US processes can be calculated using Eq. (4) [162]. S ¼

kUS=EO kEO þ kUS

(Eq. 4)

where S is the synergetic index and kEO, kUS, and kEO/US are rate constants in EO, US, and EO/US processes, respectively. One of the main disadvantages EO/US is potential enhanced electrode corrosion at combined sonication and electrolysis, which instead of increase can decrease the degradation process efficiency. Higher electrolysis currents lead to higher gas evolution such as oxygen and toxic chlorine gas (generated from the presence of Cl
ions in the electrolyte solution). The produced gases can initiate corrosion of the electrodes. At the same time, ultrasound is known for its cleaning effect due to the formation of liquid microjets and shock waves during the cavitation bubble collapse, which can assist the initiated corrosion of electrodes at higher currents. In this regard, the working parameters such as applied current or voltage in electrochemical experiments and actual power, reactor geometry in sonochemical experiments, which were effective in singular processes, need a careful optimization (often decrease for voltage and US powers) to be applied to sonoelectrochemical processes. A spatial setup of electrodes and transducer in the reaction vessel significantly influences sound distribution and a result has an effect on pollutant degradation efficiencies and electrode deterioration. The effect of ultrasound on decrease of decomposition voltage can be explained by the fact of reducing anodic reactions overpotential and increasing cathodic reac tions overpotentials. This was shown in the example of AgðS2 O3 Þ23
Ag redox couple where shift of anodic potentials was observed at 20 and 500 kHz sonication of different intensity [163]. Cathodic polarization depression was obtained also in chromium deposition process under continuous acoustic irradiation [164]. The two-sided effect of ultrasound, which can both promote and suppress the corrosion of metals and metal alloys, was reported in a number of works [165e167]. Some of the first uses of ultrasound in electrochemistry were for the removal of any passivation layer from electrode surfaces, for the homogenization of electrolytes, and for the improvement of structural properties in films deposited during electroplating [164,168,169]. The degradation of

110 CHAPTER 2 Ultrasound-assisted electrochemical treatment

phenol at applied current of 6.8 mA/cm2 and ultrasonic irradiation at 20 and 500 kHz was the pioneering work [170]. Ti/Pt grid was used as an anode in the study and Ni form was cathode. EO/US showed the enhanced degradation rates comparing to US alone, and higher frequency provided faster degradation, which were equal to 95% after 10 min of treatment comparing with 75% within the same time when using 20 kHz sonication. Moreover, 20 kHz sonication combined with EO contributed to the formation of toxic decomposition product such as quinones, which were absent in 500 kHz/EO degradation with acrylic and chloroacrylic acids as the main decomposition products. The intensive studies related to the use of sonoelectrochemical methods in pollutant degradation date back to the turn of the 21st century [4] where ultrasound assisted the electrochemical degradation of diuron herbicide, Procion Blue dye, and N,N-dimethyl-p-nitrosoaniline [171e173]. It was found that procion blue can be directly oxidized in US field on BDD anodes at potentials below 2.5 V versus SCE in acidic conditions because contamination of electrodes was enhanced in alkaline solutions [171]. US-assisted EO was conducted to decompose diuron herbicide using glassy carbon plate anode and applying 35 and 20 kHz ultrasonic irradiation [173]. Glassy carbon anode was subjected to polarization at potentials higher than 1.4 V; therefore, the electrolysis was conducted at lower potential of 1.16 V. After 8 h of EO/US process, the degradation of diuron was equal to 72%. When conducting electrolysis in 35 kHz ultrasonic bath at 1.3 V applied voltage, the degradation efficiency did not exceed 43%. That was explained by the fact that cleaning efficiency of the US horn is higher than in US bath, thus preventing the passivation of the electrodes.

2.3.1 Reactor types used in EO/US degradation All EO/US methods can be set up in three major types depending on ultrasonic reactor type. The first type is an ultrasonic bath type with working electrodes immersed into the reactor. The distribution of acoustic field in bath type reactors is considered to be even [4]. The second group consists of working electrodes and ultrasonic horn placed in the reactor volume. Both setups are used in this study and will be discussed in Chapter 4. The third setup uses a combination of working electrode and sonotrode in one device, which is shown in Fig. 2.5 [174]. The first study on EO/US reduction of benzaldehyde and benzoquinone in different solvents using the working Ti cathode combined with 20 kHz sonotrode was conducted in 1996 [174]. It was shown that US increased the current intensity at constant voltage applied and consequently enhanced

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 111

n FIGURE 2.5 Schematic representation of sonoelectrochemical reactor with working electrode acting as

a sonotrode.

the reduction rates of studied compounds. EO/US process conducted within 2 h provided 84% of benzoquinone reduction, which was seven times higher when using electrolysis alone. Ultrasonic enhancement of reduction rates left mechanism of reduction unchanged, but changed the behavior of electroactive species. One of the drawbacks of combined ultrasonic emitter and working electrode use is the necessity for a precise potentiostatic control, which can allow the electron transfer control on such electrodes [175]. A special attention should be paid to the distance between ultrasound source in the case of horn-type sonotrodes and electrodes because it influences the intensity of mass transfer, etching of electrode surfaces, polarization intensity, etc. [176]. The approximate liquid height (hpeak) in sonochemical reactor where the maximum sonochemical efficiency is produced can be calculated using Eq. (5) [177]: hpeak ¼

23400
22:9 f

(Eq. 5)

where f is the ultrasonic frequency. Another approach to calculate an approximate distance from the transducer to the vicinity of reactor where the maximum peak of US pressure occurs is described by Eq. (6) [178]: N ¼

D2 f 4c

(Eq. 6)

112 CHAPTER 2 Ultrasound-assisted electrochemical treatment

where N is near field distance, D is transducer’s diameter, f is frequency, and c is sound velocity in water. It is worth to notice that lower ultrasonic frequencies promote higher cleaning and mechanical effect of ultrasound in enhancement of degradation processes with a domination of turbulent acoustic streaming [179]. Higher frequencies have major contribution to the degradation through free radical formation [180]. Cyclic voltammograms recorded in ultrasonic field proved the significant enhancement of anodic currents and decrease in diffusion layer thicknesses especially at close ultrasonic emitter-working electrode distances [179]. The fact of increased anodic current can be attributed to the enhanced mass transfer due to reduction of diffusion layer thickness caused by cavitation in the case of higher frequencies [181]. Enhanced mass transfer can assist the EO efficiency of pollutants especially at their low concentrations.

2.3.2 Dyes degradation Over the last few years, ultrasonically assisted electrochemical methods were tested for the degradation of a range of different compounds such as pesticides, dyes, pharmaceuticals, etc. [169,182e185]. Table 2.5 contains a summary of extended works conducted in the field of EO/US degradation of organic pollutants. Degradation of dyes remains the most studied in EO/ US treatment processes. Table 2.5 Efficiency of EO/US in Degradation of Different Organic Compounds in Water. Working Parameters

Pollutant/Initial Concentration

Electrochemical

Sonochemical

Amaranth dye 100 mg/L Cl reactive black 8 100 mg/L Pentachlorophenol 2 mg/L

BDD anode I ¼ 35 mA/cm2 Ti/RuO2eIrO2 anode I ¼ 31.7 A/cm2 Pt anode Eel ¼ 30 V

Perchloroethylene 60 mg/L Phenol 0.5 mM Sandolan Yellow 50 mg/L

PbO2 anode I ¼ 3.5 mA/cm2 Stainless steel anode Eel ¼ 30 V Pt anode I ¼ 60 mA

20 kHz 523 W/cm2 20 kHz 100 W/L 22.5 W/L 35 kHz 170 kHz 300 kHz 500 kHz 700 kHz 20 kHz 0.27 W/mL 850 kHz 170 W 40 kHz 0.011 W/mL

Removal Efficiency

Time/ Conditions

References

95.1% TOC

1.5 h

[186]

32.4% COD

1.5 hpH 5.4

[187]

75% 65% 50% 45% 40%

1h

[188]

100% FC 56% 98%

6h

[189]

1h

[190]

75% decolor

80 min

[182]

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 113

Table 2.5 Efficiency of EO/US in Degradation of Different Organic Compounds in Water. continued Working Parameters

Pollutant/Initial Concentration

Electrochemical

Sonochemical

Removal Efficiency

Time/ Conditions

Reactive Blue 19 dye 50 mg/L Diuron 18e26 mg/L

PbO2 anode Eel ¼ 10 V BDD anode I ¼ 60 mA/cm2

80 kHz 0.3 W/mL 20 kHz 750 W

90% decolor 56% TOC 100% 43% TOC

[184]

Rhodamine B 5 mg/L Lissamine Green B 20 mg/L Acid Black 24 75 mg/L Methyl orange 20 mg/L Reactive Black 5; 65 mg/L Trupocor Red 50 mg/L Methyl paraben 100 mg/L

Pt anode Eel ¼ 4 V Graphite anode Eel ¼ 5 V

22 kHz 400 W 20 kHz 1.05 W/mL

91% decolor

2h pH 8 8h 10 C pH 12 6 min pH 6.5 2h

BDD 21.6 mA/cm2

20 kHz 523 W/cm2

[193]

Methylene blue 200 mg/L

Ti/TiO2eIrO2eRuO2 anode I ¼ 11 mA/cm2 Pt/Ti anode I ¼ 4 mA/cm2 Eel ¼ 10 V DCN anode Silicon carbide Stainless steel Ti/Pt electrodes I ¼ 300 A/m2

45 kHz 0.42 W/mL

92% TOC

2h pH 5.7 T ¼ 25 C 1h

24 kHz 0.032 W/mL 850 kHz 0.425 W/mL

96%

6h

[183]

67% 61% 81%

15 min

[195]

500 kHz 0.75 W/mL

47% TOC

1.5 A h

[196]

Trichloroacetic acid 0.005 M Triclosan 10 mg/L

2,4-Dihydroxybenzoic acid 300 mg/L

95% decolor 60% COD 50.2% decolor 19% COD 84% decolor 27% COD 73% decolor 22% COD 78% decolor 24% COD 50% TOC

Amaranth dye degradation on BDD anode in alkaline media provided 98% dye mineralization after 1.5 h when coupling US at 20 kHz (523 W/cm2) and electrolysis at 50 mA/cm2 [197]. The degradation followed the pseudoe first-order kinetic model with rate constant equal to 0.035 min
1. The lower applied current of 35 mA/cm2 showed a slightly lower TOC removal, which reached 90% after 1.5 h. The initial concentration of dye in treated solution

References

[191]

[185] [192]

[194]

114 CHAPTER 2 Ultrasound-assisted electrochemical treatment

was equal to 100 mg/L, and K2SO4 (0.05 mol/L) was used as an electrolyte. The proposed dye degradation mechanism involves the rupture of azo bond by cathodic reduction on the first stage. The next step is the reaction with OH radicals, consistent elimination of sulfonic acid and NH groups and 2 cleavage of carbon bonds in aromatic ring, decarboxylation, hydroxylation, and the formation of CO2 and H2O final products. The main intermediates of the amaranth decomposition are naphthalenediol, tautomers, phthalic acid, phthalic anhydride, benzoic acid, phenol fumaric, and aliphatic acids. Novel Ti/SnO2eSb2O3/PTFE-La-Ce-b-PbO2 anodes were tested in EO/US degradation of 200 mg/L cationic gold yellow X-GL dye in acidic media. The maximum removal efficiency of the dye was about 99.95%, and COD removal was 74% under 50 kHz (36.71 W/cm2) US irradiation and electrolysis with applied current density of 71 mA/cm2 after 2 h treatment [198]. The same working parameters of EO/US process tested with different anodes modifications (Ti/SnO2eSb2O3/PTFE-b-PbO2, Ti/SnO2eSb2O3/PTFE-Lab-PbO2, Ti/SnO2eSb2O3/PTFE-Ce-b-PbO2) provided about 10% lower removal efficiency toward the dye and COD. The synergetic effect of sonication and electrolysis in EO/US process was confirmed in the study. Reactive red 195 was completely decolorized at MMO electrodes when assisting EO with 20 kHz (0.04 W/mL output power) US [199]. For comparison, EO degradation provided 90% decolorization and only 3% of color removal was obtained in US degradation after 2 h of treatment. The initial dye concentration was 100 mg/L, and 2 g/L NaCl was used as supportive electrolyte. The COD removal was below 30%. Better hydrodynamics and mass transfer explained the improvement in dye decolorization in EO/US compared with EO and US processes. Among different electrolytes tested in the degradation of the dye, KCl and NaCl showed the best effect on reactive red 195 degradation. When using Na2SO4 and Na2CO3 as electrolytes, color removal did not exceed 50%, and COD removal was below 10%. The better results obtained in KCl and NaCl electrolyte solution is explained by in situ generation of chlorine and hypochlorite, which are strong oxidizing agents. The higher powers of ultrasonic irradiation were found to be less efficient in dye decolorization. This fact is explained by degassing of formed chlorine, thus preventing the formation of hypochlorite (2.46).

2.3.3 Removal of phenolic compound Degradation of phenolic compounds is the second most studied after dyes. Ultrasound was found to increase significantly the current efficiency of EO. It was shown that combination of EO with 20 mA/cm2 current density and 33 kHz US (50 W output power) for phenol degradation promoted increase

2. State-of-the-Art Research Developments in Sonoelectrochemical Oxidation of Organic Compounds 115

in current efficiency by 100% on BDD electrode and by 49% on Pt anode [54]. Moreover, US enhanced the degradation of phenol in EO/US process by 301% on BDD electrode and by 51% on Pt one. The difference in degradation efficiencies on BDD and Pt electrodes was explained by mechanism of the pollutant decomposition and adsorption. Thus, in the case of BDD electrode, US provided 79% decrease of phenol adsorption on the surface of the electrode versus 56% adsorption decrease obtained on Pt anode. BDD provided higher degradation rates of phenol and less variety of intermediate products generated at pollutant decomposition. When applying ultrasonic irradiation, significant enhancement in diffusion coefficients by 375% and 42% was observed for both BDD and Pt anodes, respectively. Nitroaromatics are recalcitrant compounds, which are hardly degradable by conventional wastewater treatment methods. In this regard, AOPs including EO/US method could be potentially effective for their degradation. It was shown that US enhanced the electrochemical degradation of 1,3dinitrobenzene (DNB) and 2,4-dinitrotoluene (DNT) [200]. The cathodic reduction of DNB and DNT to less toxic dihydroxylamines using combined Ti ultrasonic horn and cathode reached 60% and 50%, respectively, after 30 min EO/US with 50 mA applied current. Even though the parameters for electrolysis were not optimized, however, the enhancement in DNB and DNT oxidation was threefold and twofold higher in EO/US, respectively, compared with single electrochemical degradation. A novel alternated pulsed ultrasound and EO process provided 94.1% of p-nitrophenol removal after 2 h of degradation in a dual-pulse EO/US mode in acidic media (pH 3) [201]. The increase of the pH of treated solution to 11 reduced the degradation efficiency to 35%. The pulse time of potentiostatic EO at 10 V was set to 50 ms. The pulse time of sonication at 22 kHz (48.8 W output power) was equal to 100 ms. After 2 h of degradation, continuous EO/US process could provide only 89% efficiency versus 58.9% and 2.4% obtained at degradation of 100 mg/L p-nitrophenol in pulsed EO and pulsed US processes, respectively. The degradation of the pollutant fitted to the pseudoefirst-order kinetic model for every degradation process. Ti/SbeSnO2 DSA was used as a working electrode.

2.3.4 Removal of pharmaceuticals One of the recent studies on degradation of refractory pharmaceutical cephalosporin wastewater showed that low frequency ultrasound (45 kHz) can enhance the catalytic activity of nanocoated Ti/RuO2eIrO2 anode by promotion of OH radicals formed at the electrode surface to the bulk solution without radicals adsorption step on the electrode. Moreover, facilitated

116 CHAPTER 2 Ultrasound-assisted electrochemical treatment

chlorine production was obtained in the process [202]. The initial COD of wastewater was equal to 17,630 mg/L and TOC to 7340 mg/L. Stainless steel cathode was used in the study. While applying 8 mA/cm2, 45 kHz frequency with output power of 100 W for 20 min, complete decolorization of wastewater was obtained with COD removal efficiency equal to 80% and TOC reduction to around 67% at pH 5. Comparing the degradation of cephalosporin wastewater to single electrolysis and sonication processes, COD reduction in electrochemical treatment did not exceed 40% and was below 3% in sonication alone. While conducting the treatment at the same conditions for 1 h, 94% of COD removal was achieved. Sonoelectrochemical degradation of commonly used ibuprofen on Pt anodes at constant voltage of 30 V and 1000 kHz (100 W/L) ultrasonic irradiation for 1 h provided 89.3% drug removal in alkaline media [203]. The initial concentration of ibuprofen used in the treatment was equal to 2 mg/L; stainless steel electrode was used as a cathode. The degradation followed the pseudoe first-order kinetics model. Low ultrasonic frequencies provided low ibuprofen degradation rates, which reduced from 0.027 min
1 when sonicating at 1000 kHz to 0.022, 0.018, and 0.014 min
1 at 500, 300, and 35 kHz, respectively. Moreover, the highest energy efficiency of 0.062 mmol/kWh was observed for 1000 kHz frequency. A general trend of higher degradation efficiencies at higher ultrasonic powers was observed. Removal of antiinflammatory drug diclofenac exceeded 90% at pH 5.8 when using EO/US process for 5 min [162]. The initial concentration of the drug was 50 mg/L, and reaction followed the pseudoefirst-order kinetics with the rate constant of 0.505 min
1. Kinetic rate of the drug decomposition reduced at lower applied voltage. The potentiostatic electrolysis (7.2 V) on BDD electrodes immersed on 12.1 cm was combined with 850 kHz (94.1 W/L) sonication. Smaller applied voltage of 2.8 V at the same ultrasonic conditions provided only 54% diclofenac removal. The influence of immersion depth on the pollutant removal was insignificant, while the effect of electrodes distance was prominent at higher voltages. The smaller distance (1 cm) between the electrodes provided higher degradation rates (93%) compared with the degradation efficiency of 75% at higher distance (3 cm) when using 7.2 V voltage. The removal was dependent on the media pH and decreased from 96.8% at pH 4%e85.1% at pH 9.9.

3.

OBJECTIVES OF THE WORK

The aim of this research was to develop novel environmentally and costeffective electrodes working as electrocatalysts and test their activity in the electrocatalytic oxidation of organic compounds such as dyes and

4. Experimental Work 117

organic acids. Moreover, the goal of the research was to enhance the efficiency of electrocatalytic degradation processes by assisting it with ultrasound to eliminate the main drawbacks of a single EO such as electrodes polarization and passivation. The specific aims were 1. To prepare novel Ti/Ta2O5eSnO2 electrodes, which would not contain toxic or expensive Pt-group metals in their composition, and optimize the electrode bulk composition and number of active oxide layers for coating (Paper I). 2. To characterize the obtained electrodes structurally and electrochemically. The goal of structural characterization was to study the electrode crystal structure (Papers I and III), their surface topography, and the bulk composition of the obtained oxide films (Paper I). Electrochemical investigation was conducted to find out the electrode electrocatalytic properties toward MB oxidation and applicability for water oxidation. 3. To optimize the annealing temperature of Ti/Ta2O5eSnO2 electrodes and compare the surface area properties of electrodes annealed at different temperatures with the surface area properties and commercially available Au electrode (Paper III). 4. To check the electrode performance toward MB oxidation under continuous agitation fields (Paper III) and reproducibility of electrode activity after applying different posttreatment procedures (Paper I). 5. To test the electrode performance in electrochemical MB dye and FA oxidation experiments and find the best media conditions for the electrolysis process (Papers II, III, and IV). 6. To estimate the effect of a wide range of ultrasonic frequencies and powers on MB degradation efficiency (Paper II). 7. To improve the efficiency of MB and FA degradation by assisting EO of model pollutants with ultrasound and study the synergetic effect of electrolysis and sonication in sonoelectrocatalytic degradation experiments (Papers II and IV). 8. To investigate the energy efficiency of electrochemical (Papers II and III), sonochemical, and sonoelectrochemical (Papers II an IV) degradation experiments using model organic pollutants.

4. EXPERIMENTAL WORK 4.1 Ti/Ta2O5eSnO2 electrode preparation Preparation of novel Ti/Ta2O5eSnO2 electrodes is described in details in Papers I, II, and III. Briefly, electrodes were first prepared by drop-casting

118 CHAPTER 2 Ultrasound-assisted electrochemical treatment

technique and thermal decomposition using Ti substrates and organic solventebased metal chloride precursor. All chemicals required for precursor and pretreatment solution preparation were of analytical grade and used without further purification. The electrode composition was modified by changing the concentration of tantalum and tin in the precursor solution and keeping the total metal ion concentration constant at 0.04 M. The atomic content of Ta in the prepared electrodes varied from 0% to 50%. The amount of active layers deposited on Ti substrate was ranged from 1 to 10. Annealing of electrodes was conducted for 10 h at temperatures of 450, 550, and 650 C in muffle furnace.

4.2 Physicochemical and electrochemical characterization of the electrodes The surface microstructure and composition of electrodes was determined by scanning electron microscopy (SEM) coupled with energy dispersive X-ray spectroscopy for elemental analysis of microareas (Papers I and III). X-ray diffraction spectrometry (XRD) was used to analyze the crystal structure of the electrodes (Papers I and III). The electrochemical characterization of electrodes was conducted by Potentiostat/Galvanostats PGSTAT12 (Papers I and III) and PGSTAT 30 (Paper I) using cyclic voltammetry (CV) technique with scan rates between 5 and 100 mV/s in a conventional three electrode cell. Solution of 0.1 mM MB and 0.1 M Na2SO4 was used for the detection of electrocatalytic activity of Ti/Ta2O5eSnO2 electrodes at different pH media, which was controlled by H2SO4 or NaOH. Solution of 0.1 M Na2SO4 was used as reference solution. To estimate the active surface of prepared electrodes and compare it with common available electrodes, CV measurements were conducted in 0.5 mM potassium hexacyanoferrate (II) trihydrate and 0.1 M Na2SO4 solution using Ti/Ta2O5eSnO2 electrodes (Papers I and III) and gold disk electrode (Paper III) as working electrodes. In Paper I, the reproducibility of electrocatalytic activity of Ti/Ta2O5eSnO2 electrodes with Ta chosen nominal content of 25 and 37.5 at.% was checked by applying different posttreatment procedures such as 10 voltammetric cleaning cycles between 0.2 and 2.5 V in 0.1 M Nas2SO4, annealing for 1 h at 550 C and cleaning in ultrasonic bath for 30 min. To prove the ability of mechanical agitation of the solution to improve and maintain a constant mass transport of electroactive species near to the electrode surface, CV measurements in magnetic stirring and ultrasonic irradiation field were carried out (Paper III).

4. Experimental Work 119

4.3 Experimental setup in degradation experiments Fig. 2.6 shows a general experimental setup used in sonication, electrolysis, and sonoelectrochemical degradation experiments. Power amplifier, ultrasonic sinus wave generator, and transducer were used for sonication part, and power supply was used in electrochemical experimental part. Electrolysis was conducted using two electrodes setup with Ti/Ta2O5eSnO2 electrode as an anode and Ti plate as a cathode. While carrying out sonication or electrolysis alone, either sonochemical or electrochemical part worked. In the case of sonochemical experiment, electrodes were withdrawn from the reaction vessel. Combined EO/US experiments were conducted with a simultaneous work of sonication and electrolysis parts. In all experiments, a glass beaker played a role of reaction chamber. The majority of

n FIGURE 2.6 Experimental setup (A) in Paper II using 20 kHz ultrasonic probe (Sonics and Materials Inc., VCX 600, 600 W of max output power, 12.5 mm tip)

and Thulby PL320 DC power supply; (B) in Paper IV using ultrasonic bath (Meinhardt, Ultraschalltechnik, 250 W of maximum output power) with transducer type E/805/TM (75 mm diameter) fitted at the bottom of cylindrical jacketed glass reactor [97] and programmable power supply GW Instek, PSP-405; (C) in Paper II using 40 kHz ultrasonic bath (Langford Electronics, Sonomatic 375TT, 300 W of output power), 850 kHz bath (Meinhardt, K80-5, 140 W of max output power, 69.6 mm transducer diameter fitted at the bottom of cylindrical reactor), and multifrequency bath operated at 380, 1000, and 1176 kHz (Meinhardt, M11, 250 W of max out power, 52 mm transducer diameter fitted at the bottom of cylindrical reactor) and Thulby PL320 DC power supply.

120 CHAPTER 2 Ultrasound-assisted electrochemical treatment

sonication experiments were conducted using indirect ultrasonic irradiation. In Paper II, the decolorization of dye was controlled online using a peristaltic pump for solution circulation through a UV-Vis spectrophotometer (Fig. 2.6C). In Papers II and IV, when conducting indirect sonication of solutions containing synthetic pollutant, the beaker was placed into a jacketed ultrasonic bath with water circulation for cooling (Fig. 2.6B and C). Direct sonication was conducted in Paper II using low frequency 20 kHz ultrasonic probe which was placed in the beaker with working solution (Fig. 2.6A). Working parameters of degradation experiments are listed in Table 2.6. The actual ultrasonic power generated within the working solution was determined by calorimetry measurements.

4.4 Analysis of liquid samples The decolorization efficiency of MB was controlled by UV-Vis spectrophotometer (Lambda 45, UV-2450, PerkinElmer Instruments in Papers II and III) through the monitoring light absorbance by MB at 664 nm in neutral and acidic media and at 591 nm in basic one. COD was controlled by Dr. Hach Lange 2800 (Paper II) and 200 (Paper III) systems. Mineralization efficiency of pollutant was controlled by measuring total organic carbon (TOC) and nonpurgeable organic carbon (NPOC) by TOC analyzer (Sartec Limited DC190, Sartec Group in Paper II and TOC-Vcpn, Shimadzu in Papers III and IV). Shimadzu LC-20 high-performance liquid chromatograph using Shodex RSpak KC-811 column and UV-Vis detector SPD20AV measured FA concentration in the working solutions. Table 2.6 Experimental Parameters for Pollutants Degradation.

Paper Number

Type of Degradation Experiment

Pollutant and Pollutant Concentration

Volume of Working Solution, mL

Applied Current Density, mA/cm2

Ultrasonic Frequency, kHz

Supporting Electrolyte and Its Concentration

II

EO/US

MB 0.025 mM MB 0.025 mM

50

9.1

380, 850

Na2SO4, 0.1 M

50

e

Na2SO4, 0.1 M

MB 0.025 mM MB 0.1 mM FA, 250 mg/L FA, 250 mg/L FA, 250 mg/L

50

9.1

20, 40, 380, 850, 1000, 1176 e

30

9.1

e

Na2SO4, 0.1 M

40 40 40

4.5; 9.1; 13.6 e e

e 381 381, 863, 992, 1176

NaCl, 3 g/L NaCl, 3 g/L NaCl, 3 g/L

US

EO III

EO

IV

EO US EO/US

Na2SO4, 0.1 M

5. Results and Discussion 121

4.5 Energy efficiency control and kinetics The energy consumption (EC, kWh/m3) per volume of working solution and current efficiency (CE, %) in electrolyses processes were calculated using Eqs. (7) and (8): 
EC kWh m3 ¼ IVt=Vs

(Eq. 7)

CEð%Þ ¼ ððDCODÞFVs =8ItÞ  100

(Eq.8)

where I is the average applied current (A), V is the average cell voltage (V), t is the electrolysis time (h in the case of EC) or time of the COD decay (s in the case of CE), Vs is the solution volume (dm3), DCOD is the COD reduction (g/dm3) at time t, F is the Faraday constant (96,487 C/mol), and the constant 8 is the oxygen equivalent mass (q/equiv). The EC for the sonication process was estimated from the actual ultrasonic powers determined by calorimetry measurements and sonication time. The EC for sonoelectrochemical process was determined by summarizing the EC of electrolyses and sonication processes. Kinetics rate constants derived from the plot of natural logarithm versus degradation time were best fitted to the first-order kinetic model represented by Eq. (9): lnCt ¼ lnC0
kt

(Eq. 9)

where C0 is the initial concentration of pollutant, Ct is the concentration of pollutant at time t, and k is the first-order rate constant which was estimated from the slope by plotting.

5. RESULTS AND DISCUSSION 5.1 XRD analyses XRD analysis was used to identify the crystal structure of the prepared thin films of Ti/Ta2O5eSnO2 electrodes in Papers I and III. All patterns of prepared electrodes for the exception of electrodes annealed at 650 C contained intense peaks of hexagonal Ti substrate corresponding to the (0 0 2), (1 0 1), (1 0 2), (1 1 0), (1 0 3), (1 1 2), and (2 0 1) crystal orientations of titanium. XRD patterns on Fig. 2.7A contained additional peaks associated with the (0 0 4) and (1 0 4) planes which can be explained by a wider range of 2q scale applied in the studies. XRD patterns of Ti/Ta2O5eSnO2 electrodes annealed at 650 C contained only the one clear pick of hexagonal Ti appearing at 2q ¼ 40.141 degrees (Fig. 2.7B) which could be explained by high crystallinity of thin films formed at 650 C and dense Ta2O5eSnO2 coating

122 CHAPTER 2 Ultrasound-assisted electrochemical treatment

n FIGURE 2.7 X-ray diffractograms for Ti/Ta2O5eSnO2 thin film electrodes of different composition of Ta (A) and different annealing temperature (B).

obtained. The presence of low intensity tetragonal rutile (TiO2) peaks corresponding to the (101), (111), and (211) main atomic planes in the electrode films annealed at 550 C except those with Ta content of 50% and 7.5% (8 layers) can be explained by the oxygen solubility in the metal lattice of the titanium substrate. Tantalum XRD patterns showed the reflections corresponding to the (0 01), (1 11 0), (2 8 0), (1 11 1), (0 0 2), and (3 10 1) atomic planes (Fig. 2.7A) and (0 0 1), (1 1 0), (2 0 0), (1 1 1), (2 0 1), (0 0 2), (3 1 0), (2 0 2), (3 1 2), and (4 0 2) planes (Fig. 2.7B) typical of orthorhombic b-Ta2O5. Electrodes containing eight active layers of Ta2O5eSnO2 films and annealed at 550 C contained peaks corresponding to the (0 0 1), (1 0 0), and (1 0 1) planes (Fig. 2.7A) and (0 0 3), (2 0 0), (2 0 3), (0 0 6), and (2 2 0) planes (Fig. 2.7B), which are typical of hexagonal d-Ta2O5. Tin was represented by tetragonal rutile SnO2 with main peaks corresponding to the (1 1 0), (1 0 1), and (2 1 1) atomic planes (Fig. 2.7A and B) and minor peaks corresponding to the (1 1 1), (3 1 0), (3 0 1), (2 0 2), (3 2 1), (4 0 0), and (3 3 0) reflections. Even though electrodes annealed at 650 C revealed the highest crystallinity, annealing temperatures higher than 550 C had insignificant effect of electrodes electrocatalytic efficiency toward MB degradation (see below). Ti/Ta2O5eSnO2 electrodes annealed at 450 C as well as Ti/Ta2O5, Ti/Ta2O5eSnO2 (x ¼ 7.5 at.%, four layers) and Ti/Ta2O5eSnO2 (x ¼ 50.0 at.%) electrodes annealed at 550 C revealed the low intensity of

5. Results and Discussion 123

metal oxide peaks and a high intensity hexagonal Ti peaks, which can be explained by a low crystallinity of the film formed at this temperature. Ta2O5 peaks were absent in the diffractograms of Ti/Ta2O5 and Ti/Ta2O5eSnO2 (x ¼ 7.5 at.%, 4 layers) electrodes. Ti/Ta2O5eSnO2 (450 C annealing) and Ti/Ta2O5eSnO2 (x ¼ 50.0 at.%) electrodes did not contain SnO2 peaks.

5.2 SEM and EDX analyses Fig. 2.8 shows SEM images of Ti/Ta2O5, Ti/SnO2, and Ti/Ta2O5eSnO2 of different annealing temperature and nominal composition. EDX element mapping of the Ti/Ta2O5eSnO2 electrodes annealed at 450, 550, and 650 C with Ta content of 7.5 at.% represented on Fig. 2.8 (NeO, QeR, TeU). In general, the surface structure of electrodes prepared at different annealing temperatures and with different nominal composition is slightly cracked which is typical of many MMO electrodes [204e206]. The surface of Ti/Ta2O5 electrode (Fig. 2.8A) was porous in contrast to the surface of Ti/SnO2 electrode (Fig. 2.8B), which was clogged. The addition of Ta to the SnO2 coating partly recovers the roughness of electrode surface due to the formation of Ta2O5 oxides (Fig. 2.8CeG). Among annealing temperatures of 450, 550, and 650 C, the surface of the electrode prepared at 550 C (Fig. 2.8P) is rougher with cavities and pin holes on it. This allows assuming a potentially larger active surface area available for mass transfer at electrolysis. The influence of surface roughness and porosity on diffusion to the electrode surface will be discussed below. The element distribution over the surface of Ti substrate was studied using EDX mapping. As it can be seen from Fig. 2.8O, R, and U Sn distributions over Ti substrate are nearly homogeneous with a greater content on electrodes annealed at 550 C and smaller accumulation on the surface of electrodes annealed at 450 and 650 C. Distribution of Ta over Ti substrate and SnO2 films is less uniform (Fig. 2.8N, Q, and T) with formed Ta2O5 particles varying in a size from a few nanometers (dots on the mapping images) to a few micrometers (bright blue agglomerates on the mapping images). Paper I contains more detailed information on experimental bulk composition and Sn/Ta atomic ratios of the Ti/Ta2O5eSnO2 electrodes, which were determined by EDX microanalysis. Shortly, the electrodes containing eight active Ta2O5eSnO2 catalytic oxide layers showed the best agreement between nominal and prepared Sn/Ta atomic ratios. The surface texture of the Ti/Ta2O5eSnO2 electrodes annealed at 550 C and containing different number of active metal oxide layers is investigated in the Paper I. The dependence of porosity on the number of active layers was insignificant.

124 CHAPTER 2 Ultrasound-assisted electrochemical treatment

(A)

(B)

(C)

(D)

(E)

(G)

(M)

(N)

(O)

(P)

(Q)

(R)

(S)

(T)

(U)

n FIGURE 2.8 Scanning electron microscopy images of (A) Ti/Ta2O5, (B) Ti/SnO2, and Ti/Ta2O5eSnO2 electrodes of different nominal composition for 550  C

annealing temperature: (C) 50.0 at.%, (D) 25.0 at.%, (E) 12.5 at.%, and (F) 7.5 at.%; different annealing temperatures of 450  C (GeI), 550  C (JeL), 650  C (MeO), and element mapping for the electrode with x ¼ 7.5 at.%: (H, K, and N). Ta mapping and (I, L, and O) Sn mapping.

5. Results and Discussion 125

5.3 Cyclic voltammetry 5.3.1 Characterization of electrodes CV was chosen to characterize electrodes annealed at different temperatures. To compare the electrodes’ relative surface area, the cyclic voltammograms were taken in a simple ferrocyanideeferricyanide redox couple solution. Fig. 2.9 shows CVs run in aqueous solution of 0.5 mM K4Fe(CN)6 and 0.1 M Na2SO4 at different potential scan rates for Ti/Ta2O5eSnO2 electrode annealed at 550 C. The process was classified as quasireversible with well-formed oxidation and reduction peaks. The separation between the peaks (Table 2.7) was varying from 98 mV at the scan rate of 5 mV/s to 129 mV at the sweep rate of 100 mV/s which is greater than 60 mV/s for an ideal reversible process. This could be explained by a high resistance of prepared electrodes. The diffusion coefficient was estimated from the linear dependence of the anodic peak current Ip on the square root of the sweep rate using the RandleseSevcik equation. The value of diffusion coefficient was equal to 9.2$10
6 cm2/s, which is in good correspondence with literature data [207,208]. Table 2.7 demonstrates the peak current densities for the oxidation of 0.5 mM K4Fe(CN)6 obtained at Ti/Ta2O5eSnO2 electrodes of different annealing temperatures, Pt deposited on vitreous carbon electrode and commercial gold disc electrode. The CVs were recorded at a potential scan rate of 5 mV/s.

n FIGURE 2.9 Cyclic voltammetry Ti/Ta2O5eSnO2 electrode annealed at 550 C in 0.5 mM K4Fe(CN)6$3H2O and 0.1 M Na2SO4 at different scan rates n.

126 CHAPTER 2 Ultrasound-assisted electrochemical treatment

Table 2.7 Peak Currents for the Oxidation of 0.5 mM K4Fe(CN)6 in 0.1 M Na2SO4 for Ti/Ta2O5eSnO2 Electrodes, Gold Disk Electrode, and Pt Deposited on Vitreous Carbon at a Scan Rate of 5 mV/s. Electrode Ti/Ta2O5eSnO2

Au disc electrode

Annealing Temperature

Peak Current Density/mA/cm2

450 550 650 e

41.2 35.4 32.8 22.1

As it is seen from Table 2.7, Ti/Ta2O5eSnO2 electrodes annealed at different temperatures had anodic oxidation peak of about the same order of magnitude with the highest value observed for the electrode annealed at 450 C and the lowest one for the electrode annealed at 650 C. These values were compared with the values obtained from commercial Au disc electrode and Pt deposited on vitreous carbon. In both cases, oxidation peak currents from Ti/Ta2O5eSnO2 electrodes were 1.5e2 times higher. This proves the resulting surface roughness of prepared electrodes shown on Fig. 2.8 and active working area of about 1.5e2 times higher than that for Au disc and Pt on vitreous carbon electrodes. The decrease of peak current density with the annealing temperature increase could be explained by formation of poorly conductive oxide layer between the coating and Ti substrate which was observed by other workers (Paper III) [209].

5.3.2 Water and MB oxidation Fig. 2.10 shows the results on optimization of Ta content in Ti/Ta2O5eSnO2 electrodes with Ta(x)eSn(100ex) nominal composition of metal ions in the precursor solution. CV measurements run on the electrodes with x atomic content varied from 0% to 100% in a blank solution of 0.1 M Na2SO4 are shown on Fig. 2.10A. As it was expected, Ti/SnO2 and Ti/Ta2O5 electrodes could not produce a significant anodic current up to 2.5 V which is explained by a wide band gap of n-type semiconductors. While increasing the amount of Ta in the precursor solution from 0 to 2.5 at.%, the anodic current density corresponding to the OER increased from 34.5 at Ti/SnO2 electrode to 115 mA/cm2. When raising Ta content to 5 at.% anodic current density increased by the factor of 72 to 2.5$103 mA/cm2 showing the highest performance of this electrode toward water oxidation. Further increase of Ta content led to lowering of anodic current densities of OER until it dropped to the level of Ti/SnO2 and Ti/Ta2O5 electrodes at x ¼ 50 at.%. This could be

5. Results and Discussion 127

n FIGURE 2.10 Cyclic voltammograms of Ti/SnO2, Ti/Ta2O5, and Ti/Ta2O5eSnO2 electrodes of eight active oxide layers in (A) 0.1 M Na2SO4 and (B) 0.1 M Na2SO4 and 0.1 mM MB; v ¼ 50 mV/s.

explained by improvement of carrier density and lowering the resistivity of SnO2 films owing to Ta doping effect. To determine a possible applicability of Ti/Ta2O5eSnO2 electrodes for organic compound electrocatalytic oxidation, CVs were run in 0.1 M Na2SO4 and 0.1 mM MB aqueous solutions. As it can be seen from Fig. 2.10B, all voltammograms for the exception of those obtained on Ti/SnO2 and Ti/Ta2O5 electrodes contained a new anodic oxidation peak with a peak maximum at 1.1e1.2 V versus SCE. The peak was attributed to MB oxidation, and Ti/Ta2O5eSnO2 electrodes were proved to have electrocatalytic activity toward MB oxidation. To find the highest

128 CHAPTER 2 Ultrasound-assisted electrochemical treatment

electrocatalytic activity of the electrodes, their nominal composition was modified by addition of Ta. The addition of Ta in the amount of 5 at.% gave a significant increase in an anodic current of MB oxidation. However, the maximum of electrocatalytic activity was reached at Ta content of 7.5 at.%. Because Ta2O5eSnO2 films with a nominal composition of x ¼ 7.5 at.% were constituted only by hexagonal d-Ta2O5, it was suggested that hexagonal d-Ta2O5 enhanced the electrocatalytic activity of the films (see (0 0 1), (1 0 0), and (1 0 1) atomic planes on Fig. 2). Further rise of Ta content in oxide films decreased the electrocatalytic activity of the electrodes and shifted the onset potential of OER to more positive values. That could be explained by the enhanced adsorption of MB on Ta2O5-active centers, which blocks the catalytic activity of the electrodes with Ta content higher than 7.5 at.%. Relatively reversible anodic current at potential of 0.2e0.5 V which was absent on CVs obtained in the blank solution was attributed to pseudo/capacitive process of MB adsorption/desorption. The second anodic oxidation peak on voltammogram of Ti/Ta2O5eSnO2 electrode with x ¼ 7.5 at.% at potential of 1.6 V was assigned to the oxidation of potentially formed oxidation products caused by nascent oxygen generation at films surface. To estimate the effect of the Ta2O5eSnO2 film thickness on the electrocatalytic properties, several Ti/Ta2O5eSnO2 electrodes were prepared with a different number of active oxide layers and tested in the blank solution of 0.1 M Na2SO4 and working solution of 0.1 M Na2SO4 and 0.1 mM MB. The results on CVs run are presented on Fig. 2.11. As the number of layers increases, the onset potential for the OER shifts to more negative values, and the currents generated by OER increase. Fig. 2.11A shows that for one and two active layers, the potentials needed to generate, for example, a 100 mA/cm2 current density are the most positive ones (2.6e2.7 V), while increasing the number of active layers reduces the required potential from 2.6 to 2.8 V to less than 2 V for 6e10 layers. For current density equal to 250 mA/cm2, the overpotential of OER decreased by 110 mV for 10-layer electrodes compared with the 1-layer electrode. The activity of the electrodes also increased with the number of layers at constant potential, which could be explained by the influence of underlying Ti substrate clearly detected during SEM and XRD measurements (Paper I). The maximum current reached 1430 mA/cm2 for 10 active layers of Ta-doped SnO2 electrode at 2.5 V. On the other hand, its electrocatalytic activity for MB oxidation (Fig. 2.11B) was similar to that of the eightlayer electrode. However, the peak profiles of the MB oxidation process are different for the 8- and 10-layer electrodes, indicating slight differences

5. Results and Discussion 129

n FIGURE 2.11 Cyclic voltammetry of Ti/Ta2O5eSnO2 electrodes of nominal composition x ¼ 7.5 at.% and different number of active layers in (A) 0.1 M Na2SO4 and (B) 0.1 M Na2SO4 and 0.1 mM MB; v ¼ 50 mV/s. Inset Fig. 2.4B: different voltammetric cycles for Ti/Ta2O5eSnO2 electrodes with eight active layers in 0.1 M Na2SO4 and 0.1 mM MB (scan 1e3) and third scan in 0.1 M Na2SO4 (blank); v ¼ 50 mV/s.

in their electrocatalytic activity. The current generated at potentials more positive than 1.6 V was mainly used for water oxidation. While running CVs with this electrode in the working solution, the height of the anodic peaks decreased with each scan (see inset in Fig. 2.11B). It can be explained by poor mass transfer at the surface of the electrode derived from the low concentration of the dye. During the first scan, the MB coverage on the electrode surface was high, and thus the oxidation current was also high. However, with each subsequent scan, the coverage of MB decreased together with the anodic oxidation peaks. Moreover, the CV profiles corresponding to the OER recorded in the working solution were shifted toward positive potentials with respect to those of the blank solution. Therefore,

130 CHAPTER 2 Ultrasound-assisted electrochemical treatment

poor mass transfer was also aggravated by fouling the electrode with degradation products of MB leading to the formation of organic films on the electrode surface, as is frequently the case for MMO [210,211]. To prove the best performance of Ti/Ta2O5eSnO2 electrodes for MB oxidation, the annealing temperatures and media for conducting electrolysis experiments were optimized. Fig. 2.12A shows the CV results of

n FIGURE 2.12 Cyclic voltammetry of Ti/Ta2O5eSnO2 electrodes prepared at different temperatures (Fig. 2.4A) and run in different media (Fig. 2.4B) in working solution of 0.1 mM MB and 0.1 M Na2SO4 (Fig. 2.4B, red dashed line) adjusted to pH ¼ 2 (Fig. 2.4B, green line) and pH ¼ 12 (Inset Fig. 2.4B) with H2SO4 or NaOH, respectively, for Ti/Ta2O5eSnO2 electrode annealed at 550 C. n ¼ 50 mV/s.

5. Results and Discussion 131

prepared Ti/Ta2O5eSnO2 electrodes annealed at different temperatures and recorded from a working solution of 0.1 mM MB and 0.1 M Na2SO4. The voltammograms of electrodes annealed at 550 and 650 C contain an anodic current peak of MB oxidation with a maximum at potential of 1.1 V. Anodic current of 60 mA/cm2 at potential of 1.1 V is 3 and 8 times higher for the electrode annealed at 550 C than for electrodes annealed at 650 and 450 C, respectively. The higher capacitance current observed at potentials between 0.2 and 0.5 V from the electrode annealed at 550 C also confirms a greater microroughness. Therefore, it was confirmed that annealing temperature of 550 C for tantalum/tin oxide coating shows promise as a nonprecious metal electrode for the electrocatalytic oxidation of MB. To optimize an electrochemical behavior of Ti/Ta2O5eSnO2 electrodes for MB oxidation at different pH, CVs were conducted in the working solutions of 0.1 M Na2SO4 and MB at pH 2, 6.5, and 12 (Fig. 2.12B). Anodic current density peaks of MB oxidation with the values of 60 and 30 mA/cm2 were recorded on voltammograms at pH 6.5 and at pH 2, respectively. However, no anodic current of MB oxidation was observed at pH value of 12 (Inset Fig. 2.12B). The absence of MB oxidation peak is explained by a significant shift of the OER onset potential to more negative values from about 1.8 to 1.4 V and overlapping the MB oxidation peak. Once more, it was confirmed that alkali media is therefore more favorable for OER rather than MB oxidation. To check the reproducibility of these results and the ability of electrodes to recover their electrocatalytic activity, some posttreatment procedures were applied to the Ta-doped SnO2 electrodes of nominal composition x ¼ 25 at.% and 37.5 at.%. The chosen electrodes were (i) submitted to 10 voltammetric cycles between 0.2 and 2.5 V in the blank solution, (ii) annealed at 550 C for 1 h, and (iii) cleaned in an ultrasonic bath (FinnSonic M08, 200 W, 40 kHz) for 30 min. CVs were recorded after each treatment both in the blank solution and in the working solution. All the treatment applied to the electrodes demonstrated either a partial or a complete recovery of the electrocatalytic activity toward MB oxidation. Annealing at 550 C was found the most effective way to recover the electrode electroactivity toward water oxidation. Electrochemical cleaning had either insignificant or no effect on the recovery of electrode activity for water electrolysis and oxidation of MB. Ultrasonic treatment appeared to be a good way to clean the electrode due to the increase of electrode films electroactivity toward both water oxidation and MB oxidation. For more information, please refer to Paper I.

132 CHAPTER 2 Ultrasound-assisted electrochemical treatment

To avoid depletion of mass transfer near to the electrode surfaces, a continuous agitation of working solution should be maintained. Typically, this is achieved on an industrial scale by a number of techniques, for example, rapid stirring, pumping, or using turbulence promoters [212]. Ultrasound can also be applied to enhance mass transport of electroactive species to the electrode surface. Moreover, the sonication can initiate the generation of hydroxyl radicals increasing the oxidation rates of organic pollutants. Fig. 2.13 shows the first and sixth scans from the CV measurements using a magnetic stirrer (Fig. 2.13A) and with a 30 kHz ultrasonic field applied (Fig. 2.13B). The cyclic voltammogram exposed to continuous stirring and ultrasonic irradiation showed no change in the current density response. This means that MB is replenished as fast as it is consumed at the electrode surface. Therefore, applying ultrasonic irradiation to electrolysis processes can provide good electrolyte movement and accelerate the MB oxidation of MB through hydroxyl radical generation.

n FIGURE 2.13 Different cyclic voltammetric scans of Ti/Ta2O5eSnO2 electrodes made in the working solution of 0.1 mM MB and 0.1 M Na2SO4 in acidic media (pH ¼ 2) under the influence of ultrasonic field (Fig. 2.9A) and with magnetic stirrer (Fig. 2.9B). n ¼ 50 mV/s.

5. Results and Discussion 133

5.4 Electrochemical degradation 5.4.1 MB degradation Electrocatalytic activity of Ti/Ta2O5eSnO2 electrodes was tested in electrochemical degradation experiments using 0.1 mM MB in 0.1 M Na2SO4 as a model pollutant. A series of electrolysis experiments was conducted in different media at pH 2, 6.5, and 12 using magnetic stirring of 1000 rpm (Fig. 2.14). According to Fig. 2.14, decolorization efficiencies of 95% were obtained in original media of the working solution (pH 6.5) followed by 80% achieved in acidic media and 78% in basic conditions after 2 h of degradation. The lower degradation efficiency achieved in basic conditions was explained by low overpotential toward OER and was confirmed by CV measurements. The complete decolorization of MB solution regardless of the medium acidity was obtained after 6 h of oxidation process. The COD decrease obtained at pH 2, 6.5, and 12 was equal to 70%, 85%, and 26%, respectively, after 2 h of oxidation (Paper III). Decolorization and COD removal efficiencies obtained after 2 h of MB electrolysis at Ti/Ta2O5eSnO2 electrodes was comparable to MB degradation efficiencies obtained in other studies at Pb/ MnO2, PbO2eZrO2, and TiO2-NTs/Co-PbO2 electrodes [65,77,213]. As it is seen, all mentioned electrodes contain potentially toxic Pb in the structure, thus emphasizing the advantages of Ti/Ta2O5eSnO2 electrodes to be environmentally friendly. NPOC reduction data obtained during electrolysis showed that MB mineralization reached 71%, 74%, and 76% at pH of 12, 6.5, and 2, respectively, after 8 h of degradation for the electrodes annealed at 550 C (Paper III).

n FIGURE 2.14 Color removal efficiency after 8 h of electrolysis of 0.1 M Na2SO4 and 0.1 mM MB working solution.

134 CHAPTER 2 Ultrasound-assisted electrochemical treatment

5.4.2 FA degradation To estimate the effect of applied current on FA removal efficiency, galvanostatic electrolysis was conducted at applied currents of 10, 20, and 30 mA (4.5, 9.1, and 13.6 mA/cm2, respectively). Fig. 2.15 shows the dependence of FA removal efficiency and its kinetic behavior (Inset Fig. 2.15) from the electrolysis time with currents applied. The degradation followed the first-order kinetic model. The first-order kinetic rate constants were estimated from the slope by plotting the logarithm of FA concentration versus time for the degradation experiments. The slowest kinetics of FA degradation was obtained at 10 mA applied current (0.0079 min
1) with the maximum FA removal efficiency of 64% after 2 h of electrolysis. When increasing current densities to 9.1 and 13.6 mA/cm2, the kinetic rate constants increased more than twofold to 0.0185 and 0.0207 min
1 showing nearly the same FA removal efficiencies of 92% and 92.8%, respectively, after 2 h of degradation. NPOC removal reached 58.6%, 74.6%, and 77.7% after 2 h of electrochemical degradation experiments using 4.5, 9.1, and 13.6 mA/cm2 current densities, respectively. This can be explained by mass transfer limitations in the process and higher electrode polarization at higher current densities [214]. For this reason it was considered that 9.1 mA/cm2 was the best current density to employ because 13.6 mA/cm2 provided only a slight

n FIGURE 2.15 Effect of the current applied on formic acid removal efficiency and variation of the

kinetic behavior with the applied current.

5. Results and Discussion 135

improvement in rate constant and a negligible increase in removal efficiency. It was decided to conduct further studies of the sonoelectrochemical experiments at a constant current density of 9.1 mA/cm2. The direct mechanism of FA degradation was suggested taking into account the chemical structure of the acid and relatively high degradation efficiency and was similar to that introduced elsewhere for BDD electrode [215]: HCOOH þ M(OH) / CO2 þ 2Hþ þ 2e


(2.52)

5.5 Sonochemical degradation 5.5.1 MB degradation Sonochemical MB degradation was conducted in a range of different frequencies (20, 40, 380, 850, 1000, and 1176 kHz). The values of the real power entering into the reaction vessel were calculated using calorimetrically technique and can be found in Paper II. To estimate the effect of ultrasound on MB decolorization, a detailed examination of dependency of color removal efficiency on different frequencies and powers was conducted. Fig. 2.16 shows the decolorization removal obtained during the sonication of the working solution at different ultrasonic frequencies. The ultrasonic

n FIGURE 2.16 Effect of sonication at different frequencies on decolorization of 0.025 mM MB and

0.1 M Na2SO4 solution.

136 CHAPTER 2 Ultrasound-assisted electrochemical treatment

frequency of 1000 (0.8 W) kHz did not have any effect on the dye degradation. The frequency of 1176 kHz (0.4 W) had a slightly better efficiency giving about 6% of MB color removal. The low efficiency at 1000 and 1176 kHz frequencies is explained by low actual ultrasonic powers generated by the equipment, thus producing lower cavitation effect. The complete decolorization of solution was achieved using 380 (2.7 W) and 850 kHz (9.3 W) after 3 and 1.5 h of sonication, respectively. When applying 20 kHz (8.8 W) and 40 kHz (4.9 W) ultrasonic irradiation for 3 h, the color removal reached only 58% and 19%, respectively. Despite the higher powers employed at 20 and 40 kHz comparable with those produced at 380 and 850 kHz baths, the higher frequencies provided better degradation rates. This was explained by the higher production of hydroxyl radicals at higher frequencies as a result of a shorter cavitation bubble lifetime and increased bubble collapse per a unit of time [44e46]. While making a general comparison of a direct (ultrasonic probe is immersed into solution) and indirect (ultrasonic energy is transmitted to the working solution through an intermediate media; water in our case) sonication, direct system transmits more ultrasonic energy directly to the working solution. It is possible to proceed smaller volumes with higher ultrasonic intensities in direct setups. In the case of indirect sonication, significant amount of energy is reflected from the intermediate liquid/reactor vessel wall interface. However, the main advantages of indirect setup are applicability to sterile solutions and the elimination of sample lost and solution foaming. Effect of different ultrasonic frequencies on dye decolorization was further analyzed using about the same ultrasonic powers for the treated solution irradiation (Paper II). The results did not show any clear dependence of decolorization rates on ultrasonic frequencies. In this regard, a careful optimization of working parameters is required in sonication processes, which was confirmed by a number of studies [216,217]. The effect of different ultrasonic powers on MB decolorization was studied using ultrasonic irradiation at 850 kHz (Fig. 2.17), which has the highest efficiency toward MB degradation. Fig. 2.17 shows a general trend of higher degradation rates at higher powers, which can be attributed to an increased number of cavitation bubbles produced at higher powers leading to enhanced hydroxyl radicals’ formation [203].

5.5.2 FA degradation Ultrasonic irradiation of 250 mg/L FA in 3 g/L NaCl at 1176, 992, 862, and 381 kHz provided a negligible effect on pollutant degradation with a maximum degradation efficiency of 2.5% obtained after 2 h sonication at 381 kHz (Fig. 2.18). Taking into account resistivity of FA to treatment

5. Results and Discussion 137

n FIGURE 2.17 Effect of different ultrasonic powers at constant frequency on decolorization efficiency of

0.025 mM MB and 0.1 M Na2SO4 solution.

n FIGURE 2.18 Effect of ultrasonic frequencies on removal of 250 mg/L FA in 3 g/L NaCl in

sonochemical degradation experiments.

processes and low oxidation efficiency of MB in the previous studies, it was decided to change electrolyte from Na2SO4 to NaCl because chloride ions facilitate the indirect oxidation of pollutants by means of electrogeneration of strong oxidizing agent such as hypochlorite. The degradation followed the first-order kinetic model with kinetic rate constants within the range from 0.00,004 to 0.0001 min
1 (Paper IV). When comparing the sonochemical degradation of FA with electrochemical treatment, electrolysis provided 90% removal of FA in the same time.

138 CHAPTER 2 Ultrasound-assisted electrochemical treatment

Based on literature data, it was reported that 22.7 kHz horn and 20 kHz ultrasonic bath can provide 14% and 6% degradation of 250 mg/L FA [134]. When using direct sonication (590 kHz) for degradation of 1000 mg/L FA without stirring, the maximum degradation efficiency of 3% can be achieved in 1.5 h [137]. When applying 200 rpm mechanical stirring, the degradation efficiency increased more than twice up to 6%. The increase of degradation rate at stirring can be explained by improved compound distribution in the reactor, which enhances degradation rates. The low degradation efficiencies obtained in this work is probably attributed to the low powers used, indirect sonication, and absence of stirring in the reactor. When conducting indirect sonication, only around 37% of energy is transmitted across water/pyrex glass reactor, and about 63% is lost due to reflection interface (based on acoustic impedance data).

5.6 Sonoelectrochemical degradation 5.6.1 MB degradation Fig. 2.19 shows the dependence of MB decolorization on the time of degradation in EO, US, and EO/US experiments. Ultrasonic frequencies of 380 (0.054 W/mL) and 850 kHz (0.186 W/mL) were used in combined EO/US process due to their high efficiency obtained in sonication. As can be seen

n FIGURE 2.19 Comparison of electrochemical, sonochemical, and sonoelectrochemical decolorization of

0.025 mM MB and 0.1 M Na2SO4 solution.

5. Results and Discussion 139

in Fig. 2.19, electrolysis and sonication applying 380 kHz irradiation provided almost complete decolorization of the working solution in 3 h of degradation. Combined electrolysis and sonication provided complete decolorization of MB in 45 min compared with 90 and 180 min obtained in independent sonication and electrolysis, respectively, while using 850 kHz frequency and 20 mA current (see vertical lines in the figure). When combining 380 kHz sonication and 20 mA electrolysis, the complete decolorization rate increased by a factor of three and was achieved in 1 versus 3 h in separate processes of sonication and electrolysis (Inset Fig. 2.19). COD and TOC reduction values confirmed the obtained improvement of MB degradation rates (Paper II). The COD removal efficiency of combined sonication at 850 kHz and electrolysis at 20 mA achieved 85.4% after 2 h of degradation compared with 40.4% and 78.5% obtained after separate sonication and electrolysis, respectively. To investigate a contribution of EO/ US process to mineralization efficiency, TOC data were investigated. The TOC removal efficiency achieved 38.4% in 2 h of sonoelectrolysis versus 3.6% and 26% obtained in sonication and electrolysis, respectively, within the same degradation time.

5.6.2 FA degradation To confirm the choice of optimal current, EO/US degradation of FA was conducted using 10, 20, and 30 mA applied currents (4.5, 9.1, and 13.6 mA/cm2 current densities) and 381 kHz (0.02 W/cm3) frequency (Fig. 2.20). The kinetic rate constant increased twice from 0.0185 min
1 in 9.1 mA/cm2 electrolysis to 0.0374 min
1 in combined use of sonication and electrolysis and almost 200-fold comparing with sonication alone (Paper IV). The kinetic rate constants obtained in this study for EO and EO/US processes of FA degradation were among the highest values when comparing with different AOPs [134,137,218]. The higher kinetic rate constant of 955 min
1 was observed while using photocatalytic degradation over Fe/ TiO2 catalyst [219]. However, the initial concentration of FA used in photocatalytic degradation was equal to 2.3 mg/L and considered to be negligible comparing with this study. Photocatalytic degradation of FA of higher concentration (103 mg/L) provided only 6.29$10
4 min
1 kinetic rate constant [220]. The time required for 90% FA degradation reduced from 2 h in electrolysis to 1.1 h in sonoelectrolysis. A combination of 4.5 and 13.6 mA/cm2 electrolysis 381 kHz sonication had no significant effect on kinetic rate constant improvement compared with separate EO. EO/US decomposition process slightly improved the removal efficiency of FA from 64% to 78%, 92% to 99%, and 93% to 97% for 4.5, 9.1, and

140 CHAPTER 2 Ultrasound-assisted electrochemical treatment

n FIGURE 2.20 Effect of applied current densities on sonoelectrochemical degradation of 250 mg/L FA

in 3 g/L NaCl at a constant ultrasonic frequency of 381 kHz (0.02 W/cm3 power).

13.6 mA/cm2 applied current densities, respectively, compared with single EO after 2 h of treatment. It should be noticed that combined 13.6 mA/ cm2 electrolysis and 381 kHz sonication provided lower degradation of FA comparing with EO/US process, which could be explained by higher corrosion of the electrodes. Increased currents enhance the gas evolution such as oxygen and toxic chlorine gas (generated from the presence of Cl
ions in the electrolyte solution), which can initiate the corrosion of electrodes. The known cleaning effect of ultrasound due to the formation of liquid microjets and shock waves during the cavitation bubble collapse can assist the initiated corrosion of electrodes at higher currents. Moreover, it was reported that ultrasound can both promote and suppress the corrosion of metals and metal alloys [165e167]. Comparable degradation rates (over 99%) of FA were obtained in photocatalytic and wet oxidation method combined with the use of catalyst within 1 and 8.3 h, respectively [220,221]. The same behavior was obtained while comparing mineralization efficiencies of EO/US treatment, which were equal to 75%, 95%, and 92% after 2 h of degradation process at 381 kHz (0.02 W/cm3) irradiation and at 4.5, 9.1, and 13.6 mA/cm2 applied currents, respectively. In this regard, the careful optimization of working parameters should be done for EO, US, and EO/ US processes.

5. Results and Discussion 141

Among tested currents in EO/US degradation experiments, 9.1 mA/cm2 was found optimal, which was confirmed by electrolysis experiment (Fig. 2.15). In this regard, the further studies on EO/US degradation of FA were carried out using 9.1 mA/cm2 current. Fig. 2.21 shows the effect of ultrasonic irradiation at 381 (0.02 and 0.007 W/cm3), 863 (0.0067 W/cm3), 992 (0.003 W/cm3), and 1176 kHz (0.003 W/cm3) frequencies on the EO/US degradation of FA. The highest degradation efficiency (99%) and kinetic rate constant (0.0374 min
1) were obtained after 2 h of sonoelectrolysis while irradiating the working solution with 381 kHz, which also provided the highest power of 0.02 W/cm3. The lowest degradation efficiency of 91% and kinetic rate constant of 0.0184 min
1 was obtained after 2 h of sonication at 992 kHz combined with 9.1 mA/cm2 electrolysis. The same low degradation efficiency of 0.8% was also obtained in a separate sonication at 992 kHz irradiation. The removal of FA for 2 h of EO/US experiments at frequencies of 381 (0.007 W/cm3), 863 (0.0067 W/cm3), and 1176 kHz (0.003 W/cm3) was approximately at the same range of 96%e97%. As it seen from Fig. 2.21, the efficiency of FA degradation after 2 h of sonoelectrochemical experiments is comparable with the degradation efficiency obtained in singular electrochemical decomposition. Therefore, it is difficult to compare these processes and estimate the effect of US on the

n FIGURE 2.21 Effect of different ultrasonic frequencies on formic acid degradation in

sonoelectrochemical experiments.

142 CHAPTER 2 Ultrasound-assisted electrochemical treatment

degradation of FA. In this regard, comparison of synergetic indexes of combined sonochemical and electrochemical degradation rates calculated by Eq. (4) could allow better understanding of degradation process. The data on kinetic rate constants and synergetic indexes are listed in Table 2.8. Table 2.8 shows a clear synergy between the sonochemical and electrochemical degradations of FA in sonoelectrochemical process for all tested ultrasonic frequencies except 992 and 1176 kHz. The highest S index equal to 2 was obtained for combined sonication at 381 kHz (0.02 W/cm3) and electrolysis at 9.1 mA/cm2 followed by S index of 1.63 obtained while sonicating the working solution at the same frequency of 381 kHz and actual power of 0.007 W/cm3 with applied 9.1 mA/cm2 current density. The value of synergetic index significantly above 1 for 381 and 863 kHz indicated the enhancement of electrochemical decomposition of FA by US. This can be caused by the improved mass transfer near to the electrode surface or activation of electrode’s electrocatalytic properties and cleaning of the electrode surface by ultrasound. When combining sonication at 992 and 1176 kHz with electrolysis at 9.1 mA/cm2 applied current density, no synergetic effect

Table 2.8 Rate Constants and Synergetic Index Obtained for Sonoelectrochemical Decomposition of FA. Working Parameters The Type of Degradation Experiment Sonication

Electrolysis

Combined electrolysis and sonication

Ultrasonic Frequency, kHz (Actual Power W/cm3)

Applied Current Density, I, mA/cm2

The First-Order Kinetic Rate constant, k, minL1

Synergetic Index, S

381 (0.02) 381 (0.007) 863 (0.0067) 992 (0.003) 1176 (0.003) e e e 381 (0.007) 381 (0.02) 381 (0.02) 381 (0.02) 863 (0.0067) 992 (0.003) 1176 (0.003)

e e e e e 4.5 9.1 13.6 9.1 9.1 4.5 13.6 9.1 9.1 9.1

0.0002 0.0002 0.0001 0.00004 0.00005 0.0079 0.0185 0.0207 0.0304 0.0374 0.0097 0.0273 0.0234 0.0184 0.0201

e e e e e e e e 1.63 2 1.2 1.31 1.26 0.99 1.08

5. Results and Discussion 143

was observed. It can be speculated that generated power at these frequencies is too low to generate sufficient cavitation effect. A smaller amount of cavitation bubbles is formed at lower powers, which means fewer cavitation bubble collapses in the unit of time and lower physical effect of US. The best mineralization efficiency was obtained in EO/US processes (9.1 mA/cm2) at the highest frequency of 1176 kHz (97%) followed by 381 kHz (0.02 W/cm3) and 863 kHz having similar mineralization efficiency of 94.7%. Overall, the NPOC reduction data in EO/US degradation experiments with 20 mA constant current can be placed in series with increasing mineralization at 381 kHz (0.007 W/cm3) 4) [117]. On the other hand, nanoparticles of iron species could be excellent catalysts for organic decontamination by ultrasonication. Nanoparticles of superparamagnetic iron oxide were significantly more effective than micro-sized zerovalent iron or reactive divalent iron (Fenton’s reagent) [118]; and magnetic nanoparticles of zerovalent iron proved to be an effective low-cost alternative catalyst [101] in diclofenac removal by high-frequency (861 kHz) ultrasound. The positive effects were attributed to the synergy of massive surface area of nanoparticles, enhanced mass transfer rate, and enhanced cavitation events [101,118]. Most of the studies on sono-Fenton for dye wastewater treatment were conducted with low-frequency ultrasound (20e60 kHz), and high-frequency ultrasound of 1700 kHz was used only in one study [119]. In all these studies, the combination of ultrasound and Fenton’s reagent (H2O2 and iron catalyst) performed much better than either Fenton oxidation or ultrasonication alone as ultrasound irradiation accelerating the production of hydroxyl radicals in the Fenton’s reaction [119e125]. It was found that the optimum pH for these systems is 3, and the increase of ultrasonic input power and Fenton’s reagent concentration favors the increase of decolorization rate. Decolorization rate was also higher when dissolved oxygen was present as compared with nitrogen and argon [125]. Sonocatalytic degradation of Acid Red B (ARB) dye using Fe-doped zeolite Y catalysts with the assistance of low-frequency (20 kHz) ultrasonication was found to be accelerated by the reaction between Fe (II) and Fe (III) ions and H2O2 generated in situ by water sonolysis. While the dye degradation rates by ultrasonication alone and with the presence of undoped zeolite Y were less than 10% and up to 12%, respectively, the degradation rates were remarkably increased up to 86% and 85% when the Fe (II)- and Fe (III)-doped zeolite Y catalysts presented, without any other additional

2. Literature Review 245

oxidants, respectively. Thus, the presence of ultrasound eliminated the need for external H2O2 to enhance the Fenton-like process [126].

2.1.2.2 Sonophotolysis and sonophotocatalysis Since OH radicals are formed during TiO2 photocatalysis, combination of the TiO2 photocatalyst and ultrasound can take advantage of the sonoluminescence phenomenon to enhance the generation of these hydroxyl radicals. Using ultrasound (40 and 39 kHz) as an irradiation source to induce TiO2 performing catalytic activity for dyes degradation (azodyes, Methylene Blue) confirmed that the degradation increased markedly in the presence of TiO2 [127e129]. Sonocatalytic (42 kHz) degradation of various organic dyes by powder and nanotubes TiO2 indicated that powder TiO2 was more favorable for the treatment of anionic dyes while nanotubes TiO2 was more effective for cationic dyes [130]. It might be attributed to the higher surface charge and higher surface area of the negatively charged TiO2 nanotubes that made it easier for them to adsorb cationic organic dyes. On the other hand, the less charged TiO2 powder was more easy to absorb the anionic organic dyes with high molecule weight and large hydrophobic middle section [130]. The combination of 283 kHz ultrasonication (sonolysis) and ultraviolet irradiation (UV) (photolysis) showed synergistic effect for diethyl phthalate mineralization [131]. The H2O2-assisted homogenous sonophotolysis of wastewater from food industry significantly enhanced the mineralization (82%) as compared with each individual process (less than 41%) [132]. Recently, sonophotocatalysis, the combination of ultrasound, UV, and catalysts, has appeared to be a promising alternative water treatment method. The synergistic effects of UV and TiO2 catalyst and ultrasonication (25 and 200 kHz) were observed in the degradation of phenol and aldehydes, respectively [133,134]. The shock waves generated by ultrasound promote photocatalytic degradation of organic pollutants through many pathways such as increasing the specific surface area of TiO2 by reducing its particle size, improving mass transfer to the TiO2 surface, increasing dispersion of TiO2, and interrupting recombination of excited electrons and holes on the catalyst. On the other hand, TiO2 catalyzes the decomposition of H2O2 generated by ultrasonication, thus producing hydroxyl radicals, and TiO2 particles act as nuclei of cavities, thus promoting the production of active species and pyrolysis [134]. Fenton and Fenton-like’s reagents were also efficiently employed for the sonophotocatalytic degradation of fenitrothion (insecticide) at 20 kHz, linuron (herbicide) at 200 kHz, and p-nitrophenol at 25 kHz [135e137]. It was found that optimum loading of H2O2 exists; thus, beyond that the removal

246 CHAPTER 4 Ultrasonic and electrokinetic remediation

efficiency can be reduced. It was interesting to note that while normal Fenton reaction has optimum pH at 3, the simple handling and effective sonophotocatalysis with ferrioxalate can be applied in a wide pH range [135]. Furthermore, some visible light-responsive catalysts have been synthesized as an attempt to develop a photocatalytic system that can operate efficiently not only under UV but also under solar light irradiation. Shifting the optical response to the visible range can expand the photocatalytic efficiency as UV light contributes only a small fraction (5%) of the solar spectrum. The sonophotocatalytic processeapplied visible light-responsive catalysts, Bi2O3/ TiZrO4 (at 20 kHz ultrasonication), and ZnFe2O4/TiO2-granular activated carbon (at 37 kHz ultrasonication) showed initial positive results in the degradation of 4-chlorophenol and phenol [138,139]. Concerning the synergistic effect of sonophotocatalysis, comparative experiments among sonolytic (US alone), photocatalytic (UV þ TiO2 alone), and sonophotocatalytic processes (combining simultaneously US þ UV þ TiO2) were conducted for Orange-G and Methyl Orange degradation. Although there were many other factors to consider, it is interesting to note that there was obvious synergistic effect of the sonophotocatalytic treatment at low (20 kHz) ultrasonic frequency (degradation rate of 90% compared with 8% of the US alone and 68% of the UV þ TiO2 alone processes) [140]. On the other hand, the sonophotocatalytic treatment of Orange G at high frequency (213 kHz) showed no synergistic enhancement but simply an additive effect of combining sonolysis and photocatalysis (degradation rate of 85% compared with 35% of the US alone and 59% of the UV þ TiO2 alone processes) [141]. Comparison among photocatalytic, sonocatalytic, and sonophotocatalytic treatments of Rhodamine B dye using natural sunlight and 35 kHz ultrasound as irradiation source and pristine ZnO nanoparticles and ZnO/carbon nanotubes composites as catalysts confirmed the synergistic effect of the ZnO/carbon nanotube sonophotocatalysis [142]. The synergistic effect was also achieved in the sonophotolytic advance oxidation (US þ UV þ Fe3þ) system for Reactive Black 5 dye treatment. However, the system could be affected differently in the presence of different organic ligands. While oxalate, citrate, tartrate, and succinate could enhance the dye degradation, NTA and EDTA exhibited strong inhibitions [143]. Very recently, the sonophoto-Fenton system, combining ultrasonication (850 kHz), UV irradiation, and Fenton’s reagent for azo dye Orange II decolorization confirmed its synergistic effect, compared with either individual method. The decolorization rate was increased with the addition of Fenton’s reagent at the optimum molar ratio Fe2þ:H2O2 of 1:50. The decolorization rate obtained in the sono-Fenton system using heterogenous Fenton’s

2. Literature Review 247

catalyst (Fe-containing ZSM-5 zeolite/H2O2) was lower than in the system using traditional homogenous Fenton’s reagent (FeSO4/H2O2), due to more difficulty of the reaction between Fe2þ and H2O2 [144].

2.1.2.3 Sonoelectrochemical remediation The integration of sonolysis and electrolysis, though still underdevelopment, has emerged as a potentially promising remediation method. Electrochemical process is promoted by the effects of ultrasonication such as shock waves, acoustic streaming, and microjets, which activate electrode surface, enhancing electrolytic current mode and facilitating mass transfer of reaction solution. Both low-frequency (22 and 24 kHz) and high-frequency (850 kHz) ultrasounds successfully improved electrochemical decontamination of organic pollutants (nitrobenzene, trichloroacetic acid, and phenol) [145,146]. Significant reduction of energy consumption could be achieved through specifically designed sonoelectrochemical reactor [145] and innovative operating mode using dual-pulse system, which synchronized alternatively pulsed ultrasonic waves and electric waves [70]. The sonoelectrochemical system was found to be an effective and potential alternative for dye decolorization in recent researches [99,147,148]. Studying effects of different ultrasonic frequencies (40e60e80 kHz) showed that the highest degradation rate was achieved at 80 kHz, but by increasing frequency from 80 to 100 kHz, no significant effect was observed [99]. Studying effects of different supporting electrolytes showed that KCl and NaCl were much more effective than Na2SO4 and Na2CO3. Moreover, studying effects of various ultrasonic power showed that increase in ultrasonic power has decreased decolorization rate due to the degassing effect of the intermediate chlorine gas by ultrasound (chlorine gas generated from the electrode surface would escape rather than dissolved in the liquid to form hypochorite, a strong oxidant for decolorization) [148]. Furthermore, the integrated sonoelectrochemical-Fenton processes using low-frequency ultrasound (20 and 24 kHz) again confirmed their positive synergistic effect in dye removal [149,150]. Furthermore, powerful synergistic effects were observed in the combination of 20 and 120 kHz sonoelectrochemical processes and Fenton’s reagent (Fe2þ/H2O2) for hydrophilic chloroorganic pollutants and nitrotoluenes removals. While external H2O2 addition was used in one study [151], H2O2 in situ electrogenerated by reduction of oxygen dissolved at the cathode (from oxidation of water at the anode) was used in another study [152]. H2O2 can be generated in situ, which is one of the advantages of electroFenton oxidation compared with traditional Fenton oxidation. In addition, Fe2þ can be regenerated from the electroreduction of ferric salts, thus it

248 CHAPTER 4 Ultrasonic and electrokinetic remediation

can be recycled in situ and coupled with H2O2 to continuously produce the Fenton’s reagent [149,152]. It was noted that, despite degassing phenomenon by ultrasonication (which caused less amount of oxygen dissolved in water at the same temperature), the yield of electrogenerated H2O2 under sonoelectrolysis was higher than that of electrolysis due to the significantly enhanced mass transfer rate of oxygen toward the cathode also by ultrasonication [152]. Sonoelectrochemical degradation of the chloroorganic compound 2,4-D by catalysts Pd and Pd/Fe resulted in complete mineralization of the substrate with the greatly shortened reaction time in comparison to traditional electrocatalytic processes. The bimetallic Pd/Fe catalyst performed even faster (in 5 min) than the pure Pd catalyst (in 10 min) and was more cost-efficient [153].

2.1.2.4 Other sonocatalytic/sono-assisted oxidation Of various heterogeneous catalysts tested (Pt, Pd, Ru, CuO.ZnO/Al2O3) on the removal of sodium dodecylbenzene sulfonate (SDBS) by low-frequency (20 kHz) ultrasonication, CuO.ZnO/Al2O3 was found to be the most effective in terms of both SDBS decomposition and total oxidation rates as well as H2O2 formation [82]. Cadmium selenide graphene and zerovalent aluminum were efficiently used as catalysts for dye degradation enhancement under 20 kHz ultrasonication [154,155]. In both cases, the improvement was due to the sonocatalytic effects of accelerating the formation of hydroxyl radicals. In the case of zerovalent aluminum, it was found that the acidic environment was more favorable with the optimum pH 2.5 [155]. Investigating the catalytic effects of TiO2 nanoparticles, SiO2 nanoparticles, and Al2O3 microparticles on the ultrasonication of monolinuron herbicide at 20 kHz showed that TiO2 gave a greater efficiency (about twice that with ultrasound alone) and SiO2 was a little more efficient (about 1.5 times), but no real effect on the ultrasonication efficiency was found for Al2O3. It was explained that the nanoparticles provide nucleation sites for cavitation bubbles formation at their surfaces, leading to an increase in the number of bubbles when the liquid is irradiated by ultrasound, thus enhancing sonochemical reaction yield. The additional positive effect of TiO2 was suggested due to the long-lived active {TiOH}þ species (formed by reaction of OH from the cavitation collapse at the surface of TiO ) that can react with 2 adsorbed monolinuron or can be decomposed into OH, which can either react with monolinuron in the solution or be recombined into H2O2. Furthermore, ultrasonications of monolinuron with and without TiO2 catalyst were again conducted at high frequency of 800 kHz. Interestingly, no

2. Literature Review 249

improvement in the ultrasound efficiency at 800 kHz was found in the presence of TiO2. It was supposed that these active species are not formed at 800 kHz, probably because at this frequency, particles are not used as nuclei for the generation of cavitation bubbles [156]. Comparing catalytic activities among three different composites CeO2/ TiO2, SnO2/TiO2, ZrO2/TiO2, and pure TiO2 under ultrasonic irradiation (40 kHz) of the organic dye ARB, it was found that the sonocatalytic degradation rates varied significantly in this decreasing order: CeO2/ with TiO2 > SnO2/TiO2 > TiO2 > ZrO2/TiO2 > SnO2 > CeO2 > ZrO2, the corresponding removal ratios of ARB being 91.32%, 67.41%, 65.26%, 41.67%, 28.34%, 26.75%, and 23.33%, respectively [157]. The CeO2/ TiO2 composite demonstrated much higher sonocatalytic activity than pure TiO2 powder. Furthermore, to improve the catalytic activity of TiO2, composites of doping agents and TiO2 such as Fe3þ/TiO2 [158,159], Fe3þ-C60/TiO2 [160], Ce3þ/TiO2 [159], La3þ/TiO2 [161], InVO4/TiO2 [162], CdS/TiO2 [163], CdSeC60/TiO2 [164], multiewalled carbon nanotubes-CdS/TiO2 [165], Er3þ:YAlO3/TiO2 [166,167], and Er3þ: Y3Al5O12/TiO2 [168] were synthesized and proved their effectiveness of enhancing sonophotocatalytic dyes degradation in many studies. Results from studying the effects of some additives (Mn3O4, Cuþ2, Fe0, KIO3) and some radical scavengers (Na2CO3, perfluorohexane [C6F14] and t-butyl alcohol [C4H10O]) on the 35 kHz ultrasonication of olive mill wastewater showed that total phenol (88%), total aromatic amines (79%), and toxicity were removed efficiently and cost-effectively by ultrasonication alone at 60 C within 150 min. Moreover, the addition of Fe0, Fe3O4, KIO3, Cuþ2, and perfluorohexane separately enhanced the ultrasonication decontamination of the olive mill wastewater. It was suggested that total phenols are mainly eliminated by OH radicals outside the cavitation bubble during ultrasonication, while total aromatic amines are mainly degraded by high temperature of pyrolysis in ultrasonic cavities [169]. Recently, sono-activated persulfate process has emerged as a highly promising advanced oxidation technique for wastewater treatment. Persulfate (S2O82) is a strong oxidant and relatively stable during storage and handling. The combination of ultrasound and persulfate proved significant synergistic effects in the degradation of dinitrotoluenes, trichloroethane, and ammonium perfluorooctanoate [170e172]. This was attributed to the ability of ultrasound in activating persulfate to generate sulfate radical (SO4L), which is considered even more efficient and powerful oxidant than OH. In another study, the combination of oxone (2KHSO5.KHSO4.K2SO4, a source of the strong oxidant that can generate

250 CHAPTER 4 Ultrasonic and electrokinetic remediation

sulfate radicals), cobalt catalyst, and ultrasound (Oxone þ Co2þþUS) for amoxicillin degradation displayed the best performance among other setups tested as shown in the order of removal efficiency Oxone þ Co2þþUS (85%)> Oxone þ US (63%)> Oxone þ Co2þ (51%)> Oxone (22%) [173]. Ultrasound has been used to initiate catalytic reaction (sonocatalysis) for decolorization of dye solutions in several recent studies. While pH seems to not demonstrate significant effect in the process of ultrasonic irradiation alone [60], it can affect differently in the presence of different catalysts. Using immobilized cobalt ions (Co), persulfate (PS), and 35 kHz ultrasound (US) in different setups for the treatment of various dyes showed that for all the dye solutions, the decolorization efficiencies of these systems were in the order of PS þ Co þ US > PS þ US > PS þ Co > PS. This confirmed the feasibility of the combined cobalt activated persulfate and ultrasonication system, due to the enhanced formation of highly reactive sulfate radicals (SO4L) and hydroxyl radicals through persulfate (S2O82) activated by cobalt ions and ultrasound. Studying the effect of pH with the range from pH 2 to 11, it was found that pH has no practical influence in the system [174]. Ultrasonic (300 kHz) degradation of malachite green was intensified by the addition of bromide ions. The positive effect of bromide ions was explained due to the presence of dibromine radical anions (Br2L) generated by reaction of bromide ions with hydroxyl radicals produced from water sonolysis. Although less reactive than hydroxyl radicals, dibromine radicals undergo radical recombination at a lesser extent and thus would be more available than hydroxyl radicals for substrate degradation [175]. Ultrasound has also successfully assisted ozone in organic decontamination of several studies, especially for dye wastewater treatment [176e178]. Although ozonation is known as an effective advanced oxidation process for textile effluent treatment, the mass transfer of ozone from the gas phase to the liquid phase is often the limiting factor. Thus, ultrasound has been used in some studies for the enhancement of the ozone oxidation process. The degradation of reactive dyes using ultrasound (300 kHz, 520 kHz), ozone, and combination of both methods showed that the effectiveness of ultrasound was lower than ozone for both bleaching (decolorization) and mineralization; however, the efficiency of ozone treatment is significantly enhanced by simultaneous irradiation of the treated solution with ultrasound. The synergistic effect was attributed to the mechanical effects of ultrasound to enhance mass transfer of ozone to accelerate its direct reactions with the dyes and the generation of excess radical species from the sonolysis of water,

2. Literature Review 251

the normal chemical degradation of ozone, and the thermal decomposition of ozone in the acoustic cavitation bubbles [176,177]. Moreover, low-frequency (20 kHz) ultrasoundeassisted ozone oxidation proved to be a rapid, efficient, and low-energy consumption process to decolorize the high concentration malachite green (triphenylmethane dye) wastewater [178]. Ozonation combined with ultrasonication in the degradation of p-aminophenol (PAP) resulted in a synergistic increase of the overall rate. Although ozonation (72% and 90% PAP removal at 10 and 30 min, respectively) was more effective than ultrasonication (3% and 4% at 10 and 30 min), the efficiency of the combination of ozone and ultrasound (88% and 99% at 10 and 30 min) exceeded even the sum of those using ozone and ultrasound alone. It was explained that the synergy observed in combined treatment was mainly due to the effects of sonolysis in enhancing the decomposition of ozone in collapsing bubbles to yield additional free radicals [179].

2.1.2.5 Ultrasound-assisted adsorption As adsorption is a common and effective process for the removal of dyestuff from textile wastewater, the ultrasound-assisted adsorption was investigated, using high-frequency (850 kHz) ultrasound and activated carbon as adsorbent for treatment of the reactive dye Yellow HE4R. While ultrasonication alone could only achieve decolorization (80.62%) and was ineffective in COD removal, in the presence of activated carbon, 99.9% decolorization as well as 85.22% COD removal were achieved. Thus, the combined process performed much better due to the physical and chemical effect of ultrasonication on the activated carbon [180]. Another type of adsorbent, the new porous carbon-based exfoliated graphite, together with low-frequency (28 kHz) ultrasonication was applied for decolorization of azo dye scarlet 4BS and acid black 210. The combined processes also indicated enhancement in removal ratio and reduction of the treatment time compared with either individual method [181,182]. The decolorization efficiency was better in the combined exfoliated graphite and ultrasound process (98%) than in the combined activated carbon and ultrasound (58%) [182]. Furthermore, the coupling of ultrasound-assisted exfoliated graphite adsorption with photocatalysis (TiO2/exfoliated graphite-40 kHz US þ UV) and ultrasound-assisted exfoliated graphite adsorption with H2O2 (exfoliated graphite-28 kHz US þ H2O2) proved very positive synergistic effects in decolorization of some azodyes [183,184].

252 CHAPTER 4 Ultrasonic and electrokinetic remediation

Ultrasonication was used as a pretreatment process to enhance the adsorption capacity of sepiolite, a natural clay adsorbent, for Methylene Blue removal. The physical effect of ultrasonication (20 kHz, 5 h) caused significant increase in the specific surface area of sepiolite, leading to improve its uptake capacity of Methylene Blue [185]. Very recently, an experimental method using gold nanoparticles loaded on activated carbon in the ultrasound-assisted (40 kHz) adsorption process for Methylene Blue removal was effectively developed [186].

2.1.2.6 Ultrasound-assisted biological treatment Ultrasonication can be applied before biological treatment step to assist the process by stimulating enzyme activity [187] or breaking the pollutant molecules into simpler ones, thus facilitating the followed biodegradation by microorganisms [188]. Ultrasound can also be used as an auxiliary process for biochemical treatment of dyes. The decomposition of azodyes by laccase (enzyme), ultrasound (850 kHz), and combination of both treatments was investigated [189]. The laccase treatment showed high decolorization rates but could not degrade all the azodyes to the same extent, and the enzyme could be deactivated in some unfavorable conditions. On the other hand, ultrasound treatment could decolorize all tested dyes after 3 h at a high energy input, and prolonged sonication leads to nontoxic ionic species. The combination of laccase and ultrasound performed the higher degradation rates among these operations. The findings indicated the possibility of saving time and energy by applying a simultaneous combination of laccase and low energy input ultrasound treatments for decolorization of azodyes [189]. In another study, continuous low-frequency (20 kHz) ultrasonication of 5 h was used as an initial step, followed by 8 h microbial (Rhodotorula mucilaginosa) treatment for decolorization of Reactive Red 2, Reactive Blue four, and Basic Yellow 2 dyes. This hybrid method proved fairly effective with the removal rates up to 93%, 88%, and 40% for Reactive Red 2, Reactive Blue four, and Basic Yellow 2 dyes, respectively [190].

2.1.2.7 Ultrasound-assisted coagulation Ultrasound proved to significantly enhance the removal of toxic algae by coagulation through breaking down gas vesicles in algae cells [191,192]. In both studies, ultrasonications were conducted within only short duration times of 5e15 s. It was observed that variation in ultrasonic frequency did not have remarkable effect, while increasing ultrasonic power beyond certain value resulted in negative outcome as higher power makes it difficult for the aggregate algae to form larger clusters [191]. Therefore, proper power

2. Literature Review 253

supply should be considered to obtain the most effective, energy-efficient, and economical performance. In another study, short-term pretreatment with ultrasound, followed by alum coagulation, gave a marked increase in permeate flux of a biological microfiltration effluent. It was attributed to ultrasound in fouling mitigation as it can break up suspended solids. However, as the result, smaller particles can also accelerate the clogging of the membrane pores, thus prolonged ultrasonication feed pretreatment may counteract the reduction of irreversible fouling [193]. Furthermore, low-frequency ultrasound (20e22 kHz) also effectively intensified electrocoagulation treatment of effluents containing surfactants, oil, and heavy metals [89,194]. Main benefits that ultrasound brings during electrocoagulation treatment are (i) saving electrical energy due to the free radicals and outgassing effect in the cavitation raising the electrical conductivity, thus maintaining a constant current at lower voltages, (ii) intensifying electroflotation due to the release of gas at the electrodes [89], and (iii) increasing the sorption capacity of the coagulant (magnetite) through increasing surface area by reducing their particle size [194].

2.1.2.8 Ultrasound-assisted membrane filtration Employment of ultrasound to enhance ultrafiltration and membrane cleaning has emerged as an effective and promising approach in wastewater treatment. Ultrasound assisted filtration is less dependent on the feed properties [195]. The application of ultrasound in both mechanical and chemical cleaning for fouled membranes resulted in much higher flux recovery, particularly at low-frequency and high power setups [196]. Lower frequency ultrasound seems to have higher cleaning efficiencies than higher frequency ultrasound [197,198]. Ultrasonic membrane cleaning was based on the liquid jets generated by ultrasonic cavitation and the mechanical effects of ultrasound that break the fouled layer at the membrane surface and thus increased the flux [195]. Intermittent ultrasound irradiation was found more desirable than continuous irradiation mode, not only for its effective flux enhancement but also for lowering energy consumption and prolonging the lifetime of the membranes used [195,196,199].

2.1.2.9 Ultrasound-assisted disinfection High-power, low-frequency (20e40 kHz) ultrasound has proved to be effective in water disinfection, due to the shear forces generated by the collapse of acoustic cavitation bubbles, which are able to break up the cells and suspended particles [200,201]. However, disinfection by ultrasonication alone requires very high energy, therefore ultrasonication should not be considered

254 CHAPTER 4 Ultrasonic and electrokinetic remediation

as a sole alternative to conventional disinfection for economic aspects but rather be coupled together with other techniques [42,202,203]. Low-frequency 20 kHz ultrasonication combined with chlorine (chlorination) or chlorine dioxide significantly improved the biocidal action [42,204]. On the other hand, it is interesting to note that low-power, highfrequency (1.5e2 MHz) ultrasound employed in the Ashland’s patented sonoxide ultrasonic water treatment technology eliminates the need for chemical addition in controlling of bacterial, algae, and biofilms throughout an entire chemical-free industrial system [205]. While ultraviolet germicidal (UV-C) irradiation has been widely known and used, the UV disinfection is often hindered by suspended solids as they scatter UV light and provide shielding for bacteria. Novel disinfection techniques that combine low-frequency pretreatment ultrasonication (20e40 kHz) and a subsequent ultraviolet irradiation have shown to be both effective and cost-efficient [202,206,207]. The synergistic disinfection effect of the system was attributed to the capability of ultrasound in reducing the mean size of suspended solids, thus leading to increasing germicidal effect of UV-C. Results from the study of low-frequency (40 kHz) ultrasound coupled with electrolysis to disinfect saline solution showed that ultrasonication amplified the effect of electrolysis through (i) enhancing the mixing of bacterial suspensions in the vicinity of the electrode surface where the hypochlorite is being generated; (ii) the mechanical breaking effect of cavitation on the bacterial cell wall, making them more susceptible to attack by hypochlorite; and (iii) the cleaning effect of ultrasound on the electrode surface, preventing fouling build up, thus maintaining more efficient electrolysis [208]. Recently, the first study to demonstrate 36 kHz sonoelectrocatalytic disinfection using TiO2 fabric as an anode for effective inactivation of Escherichia coli bacteria was conducted, showing excellent synergistic performance [209].

2.1.2.10 Ultrasound-assisted radioactive wastewater treatment Although ultrasonic surface cleaning is a well-established technology used in many industrial processes, there are only very few reports on ultrasound used in radioactive surface cleaning [210]. Kumar et al. (2014) are the pioneers to develop the ultrasonic cleaning for external surface of plutonium-bearing components in the nuclear fuel fabrication. The radioactive particulates were contained inside the liquid medium (for later filtration), reducing the chance for any airborne activity to release. The

2. Literature Review 255

technique was conducted without any damage to the thin-walled fuel tubes and was shown to be quite effective, achieving the cleaning efficiency of more than 99% with 38 kHz ultrasonication in 30 min [210]. High decontamination efficiency of more than 97% was also achieved in another radioactive wastewater treatment process (40 kHz, 100 min) with specific ultrasonic reactor design that applied the ultrasonic standing wave effect for gathering the suspension radioactive oxide particles into the pressure node plane [211].

2.1.3 Ultrasound in sludge stabilization Recently, ultrasound application has emerged as a promising technique for the pretreatment of waste-activated sludge for subsequent sludge treatment, mostly in anaerobic digestion and also in aerobic digestion [212,213]. Comparing with other pretreatment methods, ultrasonication exhibits a great potential of not being hazardous to environment and economically competitive [214]. The use of ultrasound to enhance sludge digestion could be achieved at full scale and effectively result in less retention time, thickened sludge, improved biodegradability, improved solids destruction, substantial increases in methane gas production, and better residual solids dewatering [214e218]. It was suggested that the positive effects of ultrasonication were attributed to mechanisms such as hydromechanical shear forces, oxidizing effects of free radicals, thermal decomposition of volatile hydrophobic substances in the sludge, and increase of temperature during ultrasonication [219]. Low-frequency ultrasound was more effective than high frequency for sludge treatment, indicating that the mechanical effects, instead of free radicals, were primarily responsible for the enhancement [217,220]. The most popular frequency used for sludge stabilization is 20 kHz or in the range of 20e40 kHz. However, high-frequency (200 kHz) ultrasound also showed positive results in sludge disintegration and, moreover, in sludge decontamination (surfactant removal) [221]. High power density ultrasound could improve sludge disintegration, cell lysis, and inactivation [220]. The higher the sonication power employed, the higher rate of sludge disintegration, but increasing power beyond certain value showed no significant effect on excess sludge reduction [222,223]. Ultrasonication could disintegrate sludge solids, and with longer duration time, bacteria cells could be destroyed [224,225]. It was found that low density and long duration ultrasonication was more efficient than high density and short duration ultrasonication at the same energy input for sludge disintegration [222,226].

256 CHAPTER 4 Ultrasonic and electrokinetic remediation

Low-power ultrasound treatment could improve sludge dewaterability, depending on the sludge disintegration level. When sludge disintegration level was too low, sludge dewaterability did not change significantly. When sludge disintegration level was high, many fine particles were produced, leading to decrease in sludge dewaterability. With a proper sludge disintegration level, an improvement of sludge dewaterability could be achieved as after disintegration and the sludge fragments could be reflocculated to tighter flocs with the help of conditioning agents [226]. Ultrasonic pretreatment followed by anaerobic digestion can improve both sludge digestion and dewatering [227]. The dewatered sludge had less viscosity and elasticity than untreated sludge [228,229]. Synergistic effects were found in combined alkaline (NaOH) and ultrasonication pretreatment for enhancement of aerobic as well as anaerobic digestion of waste-activated sludge [223,230]. The efficiency of these combinations was in descending order: simultaneous treatment (NaOH and ultrasonication [US] at the same time) > NaOH treatment followed by US > US followed by NaOH treatment [230]. It was found that lower specific energy input in ultrasound pretreatment yielded higher synergistic effect [223]. Significant improvement of sludge disintegration was also observed in the combined sonothermal pretreatment. In addition, methane production and total COD reduction were also enhanced by the combined ultrasonication and thermalization. However, because of their high energy consumption, this method was assessed as unfeasible for practical application [231]. Furthermore, the optimization experiments on ultrasonication pretreatment of activated sludge under pressures resulted in the optimum pressure of 2 bars regardless of temperature conditions. As best energy efficiency would correspond to short ultrasonication at the optimum pressure under adiabatic condition, the under pressure ultrasonication pretreatment promises a significant potential for energy saving in sludge treatment [232].

2.1.4 Ultrasound in sediment and soil remediation Ultrasound can not only degrade the chemicals itself but also increase porosity and percolation rate of soil, accelerating desorption and facilitating removal of entrapped contaminants. Therefore, ultrasound has been investigated for sediment and soil remediation from variety types of contaminants from heavy metals to organic compounds [233].

2. Literature Review 257

2.1.4.1 Ultrasound-assisted heavy metals removal Ultrasound has been used to aid precious metals recovery by the cleaning action that removes an unwanted clay coating from raw ore, accelerates leaching of minerals from the ore, and improves filtration rates [234]. The mechanism of metal removal is based on the mechanical effects of ultrasound, which cause particle size reduction and detachment [235]. More than a decade ago, low-frequency (20e22 kHz) ultrasound was investigated for enhancement of several soil washing processes such as passing water across the contaminated soil on an ultrasonically shaken tray [234], coupling ultrasonication with vacuum pressure extraction in an integrated multistep technology [235], or ultrasonication of diluted soil [236]. In recent studies, low-frequency power ultrasound was efficiently used to aid the chemical leaching of heavy metals by acid [237,238], acid and Fenton’s reagent [239], thiourea [240] from electroplating, sewage sludge, and kaolin, respectively. Ultrasound improved the soil remediation in all the studied cases.

2.1.4.2 Ultrasound-assisted organic desorption Power ultrasound (low frequency) has been used in many soil (sludge)washing studies for desorption of nonaqueous phase liquid such as vegetable oil [241] and petroleum hydrocarbons (diesel [242e246], crude oil, bitumen, etc. [102,247,248]). The flushing processes were sometimes combined with surfactants [242,246], air floating [247], or mechanical mixing [244]. Ultrasound-induced desorption is based on the mechanisms of physical breakage of bonds by hot spots, particle surface impingement, the fragmentation of long-chain hydrocarbons by microjets, and microstreaming generated by acoustic vortices in the media pores [102,242]. Experimental results indicated that ultrasonication can enhance oil removal considerably, and the level of improvement depends on factors such as ultrasonic power, water washing flow rate, and soil type. The finer the particle size, the higher the surface area and the capillary force, leading to reduction in the removal efficiency [245,246,248]. Removal increased slightly with power intensity, but only up to a certain level [241,242,245]. The solid concentration of the slurry also played an important role [243]. Removal efficiency could be improved with a multistage ultrasonic treatment process [242]. Furthermore, as carbon particles provide strong sorption sites for hydrophobic organic contaminants and thus reducing their freely dissolved concentrations, a powdered activated carbon amendment assisted with ultrasonication was developed and proved to be more effective than mechanical mixing in decreasing the bioavailability of phenanthrene and pyrene [249].

258 CHAPTER 4 Ultrasonic and electrokinetic remediation

2.1.4.3 Ultrasound-assisted organic destruction Power ultrasound of 20 kHz was found to be able to destroy major contaminants, herbicides, and pesticides such as atrazine, simazine, dichlorodiphenyl-trichloroethane (DDT), lindane, 2,4,5-T, TBT, endosulfan, PCB, PAHs, and petroleum hydrocarbons [34,250e252] in soils and sediments. It was observed that radical formation was negligible, but pyrolysis dominated in cavitation reactions in slurries [250]. Several advantages of high power ultrasonic technology compared with conventional methods were indicated including high destruction rates, no dangerous breakdown products, and low energy demand leading to low cost. Moreover, the technology can be made rather compact and transportable, allowing on-site treatment [34]. On the other hand, the application of low-power high-frequency (1.6 MHz) ultrasound was showed to be effective in degrading DDT dispersed in water and sand slurry. However, in practice, due to intensity limitations of lowvolume coverage and high attenuation of energy, currently available highfrequency ultrasound equipment is not ideal for heavy duty application such as soil remediation [253].

2.1.4.4 Ultrasound-assisted advanced oxidation of organic pollutants A combination of ultrasound (47 kHz) and Fenton-advanced oxidation (sono-Fenton process) for soil remediation from volatile organic pollutants was initially developed with toluene and xylenes as model contaminants and with different Fenton-like catalysts (iron sulfate, iron chloride, and copper sulfate) as well as different concentrations of H2O2. It was found that, when H2O2 was present, the addition of a Fenton catalyst increased the efficiency of the process, especially iron sulfate for toluene and copper sulfate for xylenes removals. It was also observed that increasing H2O2 concentration enhanced the removal of all the contaminants. The total efficiency of the process was noticeably improved when applying ultrasonication [32]. Another sono-Fentonelike process using naturally occurring mineral iron as catalyst with different ultrasonic powers (100, 200, 400 W) was investigated for soil remediation from naphthalene. The results indicated that mineral iron was able to catalyze the degradation of naphthalene in the presence of ultrasound at various concentrations of H2O2. At the optimum condition, the maximum of 97% reduction in naphthalene concentration in soil after 2 h of treatment could be achieved. For practical use, it was suggested that to prolong the lifetime of the sonotrode, improvement in the reactor design is necessary [254].

2. Literature Review 259

2.1.5 Ultrasound in waste treatment and recycling Recently, interests on ultrasonic applications have been even extended into waste treatment and recycling. High-intensity (1e2 kW), low-frequency (20 kHz) ultrasound has been found effective in facilitating the deinking of recoverable office waste and Indigo prints within 10e20 min with the removal efficiency that can reach up to 100% [255,256]. On the other hand, sonochemical reactors have successfully synthesized biodiesel from waste cooking oil due to the enhanced emulsification effect and mass transfer rate by low-frequency ultrasound (20 kHz, 200 W). After the same operation time of 40 min, the sonochemical reactors showed clearly better conversion efficiency (89.5%) than the conventional stirring method (57.5%) [257]. While there have been several studies on the employment of ultrasound as pretreatment before biological process in wastewater treatment and sludge stabilization, the study by Cesaro and Belgiorno (2013) is the first attempt to use ultrasound as pretreatment to enhance further anaerobic digestion of solid organic waste. The efficiencies of sonolysis (20 kHz) and ozonation in improving anaerobic biodegradability of the organic fraction of municipal solid waste for enhancing biogas production and energy recovery were investigated. It was found that both pretreatments significantly improved the solubilization of the solid organic waste. However, ultrasonication appeared to be more competitive than ozonation as it was more effective, less costly, and was without formation of undesired by-products [258].

2.1.6 Ultrasound in air pollution control The application of ultrasound in air pollution control is based on acoustic agglomeration, a process in which ultrasonic waves induce relative motion and collisions among pollutants suspended in the air, causing them to adhere together to form larger assembled particles [35,37]. This acoustic agglomeration facilitates the precipitation of small air pollutant particles for easy removal. It was noted that while low-frequency ultrasonic agglomeration is more cost and energy efficient, high-frequency ultrasonic agglomeration is more retention efficient, particularly for very small particles in submicron range [259]. The presence of humidity could enhance the acoustic agglomeration of submicron particles much smaller than 1 mm in diesel exhausts [260].

2.1.7 Ultrasound in environmental analysis The application of ultrasound in solid (food, sediment, soil) sample pretreatment before analysis for determination of contaminants has attracted remarkable interests recently. The number of studies on the topic has grown significantly; within just over last 5 years, more than a 100 publications have been published and reviewed in 4 recent articles [264e267]. Compared with

260 CHAPTER 4 Ultrasonic and electrokinetic remediation

other conventional methods, main advantage of the ultrasound-assisted extraction is the dramatic reduction of the preparation time. For example, it can take only 30 min instead of 24 h required by conventional shaking method [261]. Moreover, the use of ultrasound as assistance in solvent extraction for environmental analysis also brings many other benefits such as high recovery rate (which can be up to near 100% [262]), significant reduction of solvent consumption [263], high purity of the extracts obtained, no cross-contamination, simple setup, and ease of operation [264e267]. On the other hand, in recent decades, microwave heating has been used in analytical and organic laboratory practices as a very effective and nonpolluting method as compared with conventional digestion, which often takes at least several hours of prolonged heating and stirring in strong acid solution. Then, the simultaneous microwave and ultrasound irradiation has been recognized as a new technique for atmospheric pressure digestion of solid and liquid samples in chemical analysis. The coupling microwave ultrasound gave significant improvement such as reduction of digestion time, reduction of the quantity of reagents, and reduction of contamination. In addition, the process could be totally automatic and more safe. The combination of these two types of irradiation in physical processes such as digestion, dissolution, and extraction appears very promising [268].

2.2 Electrokinetic remediation of organic contamination Electrokinetics is defined as the formation of fluid flow and transportation of charged particles in soilewater systems (porous media) under the influence of an electric field. A low-intensity direct current applied through the soil can cause the contaminants drive from the soil to one of the electrodes where it would be collected or treated. EK is also known as electroremediation, electroreclamation, or electrochemical soil remediation [269], a soilflushing technique enhanced by electric field, particularly effective in finegrained soil of low permeability [270].

2.2.1 Principles of electrokinetics The principle of EK, the transportation, and removal of contaminants in soil solution are based on three mechanisms: electroosmosis, electromigration, and electrophoresis. Electromigration is the transport of ions and ionic species to the electrode of opposite charge, electroosmosis is the movement of soil moisture or groundwater from the anode to the cathode, and electrophoresis is the transport of charged particles or colloids under the influence of an electric field; contaminants bound to mobile particulate matter can be transported in this way [271].

2. Literature Review 261

Electromigration is faster than electroosmosis. But electromigration of metals through the soil requires pH control at the cathodes. Specialized electrodes should also be utilized to prevent the corrosion by acid generation at the anode and that requires higher maintenance and operation costs. However, using electroosmosis for organics removal reduces the need for specialized electrodes as most organic remediation is not affected by pH. While electromigration has been used for the removal of metal contaminants, electroosmosis has often been used for the removal of organic contaminants (nonionic). Because soil particles often have a negative surface charge, when immersed in an electrolyte, the particles attract cations, creating a positively charged boundary layer (referred to as the charged double layer) next to the surface of the soil particles. Application of a voltage difference across a section of soil causes movement of cations and associated water within the double layer toward cathode. Then, the remaining interparticle pore fluid also moves in the same direction as the double-layer fluid due to viscous drag interactions [272]. Electroosmosis provides uniform pore water movement in most types of soil [273]. Electroosmotic flow rate is primarily a function of applied voltage. The entire soil mass between the electrodes is basically treated equally, and the whole bulk fluid moves at the same rate as the cation double layer, thus electroosmosis is so effective in clayey soils. Electroosmosis can be utilized to remediate contaminated soils in situ by flushing out the pore fluid and contaminants (or to deliver nutrients, surfactants, etc.). The transport of larger charged molecules and particles (cationic or anionic surfactant micelles, microorganisms) under the influence of an electric field is called electrophoresis [272]. Contaminants bound to mobile particulate matter can be transported in this manner. Besides these transport processes, many other reactions may occur in a direct current (DC) field such as desiccation due to heat generation, gas generation due to electrolysis of water, decomposition or precipitation of salts and minerals, ion exchange, development of pH gradients, sorption processes, and electrochemical transformations [274]. The electricity applied to the soil directly results in heating of the soil. The soil warming not only increases the mobilization of volatile organics but also increases the electroosmotic permeability by lowering the viscosity of the pore water. Application of direct electric current in water induces electrolysis. This electrolysis of water is the dominant electron transfer reaction at electrodes during the electrokinetic process [271].

Oxidation at anode: 2H2O e 4e / 4Hþ þ O2[ Reduction at cathode: 4H2O þ 4e / 4OH þ 2H2[

262 CHAPTER 4 Ultrasonic and electrokinetic remediation

Oxidation of water at the anode generates an acid front, accumulated with Hþ ions, while reduction at the cathode produces a base front with excess OH ions. As a result, a pH gradient is always generated between anodes (usually with pH as low as 2) and cathodes (usually with pH up to 12), in the electrokinetic process. Electrokinetic technology has been shown to be efficient in the removal of partially polar organic species, such as phenol and nonpolar ones such as PAHs, atrazine, chlorinated solvent perchloroethylene (PCE), and BTEX compounds (benzene, toluene, ethylene, and xylene) [273,275e279]. With nonpolar organic species, the applicability of EK was effective by the assistance of different solubilizing agents.

2.2.1.1 Factors affecting electrokinetics Soil types: Bench-scale and pilot-scale tests indicated that the technology can be successful in clayey to fine sandy soils. However, contaminant transport rates and the efficiency of the process depend heavily on soil type and mineral composition [280]. Soil pH: Changes in soil pH can change the chemical states of contaminants as well as the magnitude and direction of electroosmotic flow, affecting the transport of contaminants in soil pore fluid. When the soil pH is lower than the point of zero charge (pzc, a zero zeta potential of soil particle surface), the direction of electroosmotic flow is reversed, from the cathode toward the anode [281,282]. Degree of saturation: Depending on the target transport mechanism, the degree of saturation affects both electroosmosis and ionic migration. Soils of high water content and low ionic strength provide the most favorable conditions for transport of contaminants by electroosmotic and ionic migration [280]. Voltage and current levels: Electric current intensities used in most reported studies are in the order of a few A/m2. High current levels generate more acid and increase the total ionic concentration that will decrease the overall electroosmotic flow. Selection of the most appropriate current density and voltage gradients depends on the soil electrochemical properties, especially electric conductivity. Soils with higher electric conductivities require more charge and higher currents than lower conductivity soils. A voltage gradient in the order of 1 V/cm can be used as an initial estimate. Increasing the current densities (or voltage gradients) will increase transport rates under ionic migration but will also increase energy expenditure and cost of the process. An optimum current density or voltage could be appropriately selected based on the soil properties, electrode spacing, and time requirements of the process [280].

2. Literature Review 263

Electrode materials and arrangement: Decontamination by electrokinetics is also influenced by the chosen electrode materials and arrangement [283]. Suitable electrode materials were often selected based on their performances in terms of current density, chemical stability, and corrosion resistance. Moreover, the study on effects of electrode arrangement showed that electrodes in direct contact with the soil presented less electrical resistance and more decontamination than the electrodes separated from the soil by physical barriers [283].

2.2.1.2 Electrokinetics impacts on soil health Application of electrokinetics alters both the physicochemical characteristics of the soil and the exposed microbial community. While soil pH change by electrokinetics may reduce microbial number and diversity, the low direct electric current can increase biodegradation of hydrocarbons and stimulate soil enzyme activities, making electrokinetics and bioremediation combination possible [284e286]. However, sometimes, electrokinetics may increase negative effects of toxic contaminants or electrolytes on microbial community under specific pH change [286,287]. Therefore, soil parameters, electric current, electrode, electrolyte, and their interactions should be carefully considered for a healthy soil environment or a successful combination of electrokinetics and bioremediation. Main advantages and limitations of electrokinetic soil remediation [1,271,276]: Advantages:  Useful for site remediation under conditions that normally limit in situ approaches such as fine-grained sediments or low permeability soils.  Versatile and can be used to enhance other treatment methods such as bioremediation, chemical oxidation, or soil vapor extraction.  Less impact on existing landscaping, buildings, or structures than ex situ technologies or soil washing.  Not require heavy equipment, excavation, or installation of large plants, thus reducing cost.  Cost efficient compared with costs for other in situ and ex situ methods. Limitations:  Limited by solubility of contaminants in aqueous phase and desorption of contaminants from the soil matrix. The process is also not efficient when the target ion concentration is low and nontarget ion concentration is high (not selective).

264 CHAPTER 4 Ultrasonic and electrokinetic remediation

 Acidic conditions and electrolytic decay can corrode some anode materials. Inert electrodes, such as carbon, graphite, or platinum, must be used so that no residue will be introduced into the treated soil mass.  When higher voltage is applied to the soil, the efficiency of the process decreases because of the increased temperature.  Removal efficiency is significantly reduced if soil contains large rocks or gravel. Electrokinetic technology has been applied successfully in the remediation of inorganic contaminant and has been shown to be highly efficient in the removal of partially polar organic species, such as acetic acid and phenol. However, because of their low solubilities and slow desorption rates, hydrophobic organic compounds are difficult to remove from subsurface environments with traditional electrokinetic technology. Therefore, the remediation process of these compounds using the electrokinetic method has been combined with other treatment processes. Besides combination with other processes, electrokinetics also have been developed in different designs (such as vertical design, circulation of electrolytes) or with various operation modes (rotation of electrode matrix, nonuniform, periodic modes) to enhance its performance. Fig. 4.4 summarizes enhanced electrokinetic processes for organic contaminated soil remediation, which will be briefly described in the following sections.

2.2.2 Electro-Bioremediation Bioremediation has potential to restore contaminated soil, but it is time consuming. The method can be accelerated by using electrokinetics. Electrokinetic bioremediation technology is developed to activate the growth of microorganisms that are capable of transforming organic contaminants in soil by promoting the transport of nutrients [276]. Electrokinetics can produce uniform transport and mixing of additives for bioremediation [288]. Injecting nitrate and ammonium to organic, tropical, clayey soil was feasible; however, the injection of phosphorous did not prove to be successful [289]. Disseminating bacteria by electroosmosis in three different soil types (garden soil, fine sand, and clay) was compared with those controls without electricity, showing clearly that electroosmosis stimulates bacterial spreading even in low-permeability soil such as clay, although the migration velocity was lower than in other soils tested [290]. The high clay content of the soil limited oxygen and nutrient supply to the microbial community because of slow diffusion and low hydraulic conductivity [291]. Thus, EK is especially indicated for clayey soils, to increase the

2. Literature Review 265

Electrokinetic Remediation (EK)

Heavy metal removal (electromigration)

Organic decontamination (electroosmosis) EK enhanced by design and operation modes

EK combined with other processes

EK + Bioremediation EK + Flushing Agent Enhancement

EK + Fenton/Oxidation

Induced Polarization Electro-ChemicalGeo-Oxidation Circulationenhanced EK

Upward EK

EK + Permeable Reactive Barriers

Periodic electric potential application

EK + Ultrasound

Non-uniform EK & rotational mode

n FIGURE 4.4 Enhanced electrokinetic processes for organic decontamination.

diffusion rate of nutrients and microorganisms. It was observed that the creosote degradation proceeds 10-fold faster in soil treated with an electric field than in the control cells without current or microbial activity [291]. Electrokinetics can enhance bioremediation in overcoming the limitation of cold climate areas, where biodegradation is often a slow process, because electricity increased the soil temperature to a level suitable for active microbial degradation [292]. It was found that high initial contaminant concentrations showed significantly higher relative degradation than low initial concentrations [292].

266 CHAPTER 4 Ultrasonic and electrokinetic remediation

Electroreclamation in the form of an electrokinetic fence can be applied as an in situ method to fence off and remediate polluted sites or groundwater plumes [293]. ElectroBioFence is a long and narrow zone that intensifies biodegradation due to increased temperature and availability of nutrients. Nutrients are often electrically charged compounds and thus can be dispersed through the soil electrokinetically. The organic pollutants are degraded by enhanced microbiological activity, either within or downstream of the zone. Electrokinetic biofences can be independent of the subsurface soil composition and applicable to relatively great depths (>10m) and under buildings without disturbance of the groundwater flow regime. Because it is an in situ treatment, there is no need to pump huge amount of groundwater. Godschalk and his company, Holland Milieutechniek (2005), have performed successfully an electrokinetic biofence remediation project for the site of a chemical laundry, polluted with volatile chlorinated hydrocarbons, at Wildervank, Netherlands. After running the ElectroBiofence for nearly 2 years, the contaminants were being dechlorinated and the chloride index was decreasing [294]. One interesting thing was that the power supplied for the biofence was from solar energy panels. Electrokinetic bioremediation of soil mixed contaminated by both organic and heavy metal pollutants is often challenging because the toxicity of heavy metals can be harmful for microbial community. A new electrokinetic bioremediation was developed for mixed Pb and oil decontamination from polluted soil, with the addition of EDTA to enhance heavy metal removal and the addition of Tween 80 to enhance oil removal. This enhanced electrokinetic bioremediation resulted in greatly reduced Pb toxicity, leading to improved microbial degradation of oil. It was found that regular electrolyte refreshment was favorable for maintaining high electric current and under optimum operation condition, more than 80% removal of both Pb and total petroleum hydrocarbons (TPHs) was achieved [295].

2.2.3 Electrokinetics with flushing agent enhancement The contaminant transport during electrokinetic treatment of soil contaminated with nonpolar contaminants occurs primarily by electroosmosis. Therefore, the process is not effective unless the contaminants are soluble in pore fluid or are converted to soluble form. Hydrophobic compounds adsorb strongly to the soil and hence the electrokinetic treatment of soil contaminated with such compounds will be feasible only if the contaminants are desorbed from the soil and made soluble in the pore fluid. Thus, surfactants, cosolvents, chelants, and complexing agents are often employed to increase desorption and solubilization. As they may react differently with different contaminants, compatibility of flushing solutions

2. Literature Review 267

and contaminants should be considered. In general, nonionic surfactants are often chosen because of their higher solubilization capacity, lower cost, and higher biodegradability as compared with cationic and anionic ones [296]. Many studies have investigated the applications of surfactants and cosolvents in EK to enhance the mobility of low polarity or nonpolar contaminants through low permeability soil [275,297e304]. As electroosmosis is the main electrokinetic movement of low polarity or nonpolar contaminants, optimization of this phenomenon is necessary for effective decontamination. It was noted that controlling pH at 7 in the anode chamber and increased electric charge could accelerate electroosmotic flow through soil, leading to increased mobilization and removal of low polarity contaminants such as PAHs [303,304]. On the other hand, contrary to the observation at the laboratory scale, the recent pilot-scale research for surfactant (sodium dodecyl sulfate)enhanced electroremediation of natural soil contaminated with phenanthrene showed that gravity and evaporation fluxes were more relevant than electrokinetic fluxes, and the desorption of phenanthrene promoted by electric heating seemed to be a significant removal mechanism. Although the treatment time was as long as 4 months, the average removal rate achieved was only 25%. The pilot-scale experiment indicated that decontamination of the soil was feasible but could require long operation time and high-energy consumption [305]. Because of their nontoxicity, biodegradability, and low affinity of sorption to a solid phase at a wide range of pH values, cyclodextrins are considered advantageous over regular surfactants. The hydrophobic cavity of the cyclodextrin serves as the binding site for the low polarity contaminant, and the hydrophilic shell with its charged group facilitates movement through the soil toward the electrode well under electrokinetic conditions, enabling this method for removal capability of both polar and low polarity contaminants [298]. Cyclodextrineelectrokinetic process was proven to be a costeffective and environmental friendly method for remediation of soil contaminated by low polarity organics, naphthalene and 2,4-dinitrotoluene, and petroleum hydrocarbons [298,306]. Hydroxypropyl cyclodextrin (HPCD) was also used as a flushing solution during the electro-Fenton process for the removal of pentachlorophenol (PCP) from soil [307]. It was reported that HPCD increased the efficiency of pollutant degradation, and this beneficial effect of HPCD on PCP degradation rate was explained by the formation of a ternary pollutantecyclodextrineiron complex capable of directing the hydroxyl radicals toward reaction with the pollutant [307].

268 CHAPTER 4 Ultrasonic and electrokinetic remediation

Four flushing agents, including two surfactants (5% Igepal CA-720 and 3% Tween 80), a cosolvent (20% n-butylamine), and a cyclodextrin (10% HPCD), were used in a feasibility study to enhance EK of manufactured gas plant soils contaminated by both PAHs and heavy metals [301]. It was found that all four flushing agents were capable to enhance PAHs removal, with Igepal CA-720 surfactant yielding the highest removal efficiency due to partial solubilization of PAHs. While maximum electroosmotic flow was observed in the cosolvent-enhanced system, followed by the HPCD-enhanced system and comparatively low flow observed in the surfactant-enhanced system, PAHs solubilization was more effective in the surfactant- and HPCD-enhanced systems than in the cosolvent system. Because of high soil pH due to high soil buffering capacity, heavy metals remained strongly adsorbed, precipitated, and were not removed in all the tested systems [301]. Later on, similar attempts to use flushing agenteenhanced electrokinetics for both PAHs and heavy metals mixed contamination remediation had been conducted in several studies [270,308e312]. It was confirmed that solubilization, electroosmotic flow, and concentration of the flushing agents were the critical factors that contribute to the removal of PAHs [309]. Catholyte conditioning maintained the soil pH as acidic, leading to enhanced desorption of metals but making the soil positively charged, resulting in decreased and reverse electroosmotic flow which in turn inhibited metal removal through electromigration and PAH removal through electroosmotic flow [270]. Surfactants and complexing agents were commonly used to increase desorption and solubility of organic (PAHs) and metal pollutants, respectively [311], but the ecotoxicity of chelating agent such as EDTA should be considered [310]. The selection of processing fluid and pH control are the key variables for effective removal; hence, there must be a preevaluation of flushing agents and their compatibilities to the soil types and the pollutants [311]. Both the simultaneous addition of complexing agents and surfactants [311] and the sequential application of a chelating agent (such as citric acid), followed by a surfactant-enhanced electrokinetics [312], were shown to be effective in the mixed contamination remediations. The surfactant-aided electrokinetics coupled with carbon nanotube barriers was studied for the removal of 1,2-dichlorobenzene from spiked soil [313]. Because of its high specific surface area, high reaction ability, and high electron transfer capacity, carbon nanotube barrier is expected to highly remove the soluble contaminants by adsorption. Although both methods helped in the soil remediation, results showed that the removal of 1,2 dichlorobenzene in the coupling system was mainly contributed by surface sorption on carbon nanotube barriers rather than by electrokinetic process [313].

2. Literature Review 269

Moreover, an innovative two-stage process combining soil EK and liquid electrochemical oxidation was developed for benzo(a)pyrene removal from spiked kaolin. In the first step, ethanol and Brij 35 were tested as flushing solutions. While no presence of the contaminant in either chamber was observed in the case of ethanol as flushing solution, 17% of initial benzo(a)pyrene was detected in the cathode chamber without pH control and up to 76% of initial benzo(a)pyrene was detected in the cathode chamber when pH was set at 7.0 in the anode, in the case of Brij 35 as flushing solution. In the second step, mobilized benzo(a)pyrene in the liquid collected from the electrokinetic process was further oxidized by electrochemical treatment, reaching 73% contaminant degradation after 16 h treatment [296].

2.2.4 Electro-Fenton (EK-Fenton) and other oxidationenhanced electrokinetics Fenton and other oxidation processes were successfully integrated with electrokinetics to treat low permeability soil contaminated with organics [314e317]. Among various advanced oxidation processes, the Fenton process has received much interest in the destruction of biorefractory organic pollutants in various media because of its strong oxidizing capability [318,319]. The EK-Fenton phenomenon is characterized by a complex process, which includes the effects of electroosmosis, electromigration, electrolysis reactions, and the mineral-catalyzed Fenton-like reaction. Fenton’s reagent can either be generated in the system or applied externally [320e322]. The EK and Fenton processes were successfully combined for in situ treatment of various organic compounds (e.g., TCE, phenol, 4chlorophenol, and diesel fuel) in different soil types, catalyst types, and electrode materials. Biodegradation was even further incorporated in the EK-Fenton process for treating PCP-contaminated soils. It was found that 100% PCP destruction could be obtained within a reasonable treatment time by combining these three technologies. Results of cost analysis have shown that the EK-Fenton process is very low in operating cost [314]. The study on the applicability of the EK-Fenton process for the remediation of low permeability soil contaminated with phenanthrene reported that the phenanthrene degradation yield was proportional to the transfer rate of the acid front and H2O2 stability. Therefore, to effectively treat sorbed contaminant on soils during the EK-Fenton process, an injection of acid is necessary in such an extent that it does not decrease the electroosmotic flow rate [323].

270 CHAPTER 4 Ultrasonic and electrokinetic remediation

The relationship between the chlorine content of contaminant molecule and degradation rate was studied using various chlorophenols including PCP, 4chlorophenol (4-CP), 2,4-DCP, and 2,4,6-trichlorophenol (2,4,6-TCP), and the degradation sequence was obtained as 2,4-DCP>2,4,6-TCP > PCP>4CP [324]. Results from the study of EK-Fenton processes for the removal of hexachlorobenzene (HCB) from kaolin under different conditions indicated that EK-Fenton was effective in HCB remediation from kaolin; about 64% HCB removal achieved in the test using beta cyclodextrin as flushing agent [325,326]; and 76% HCB removal achieved in the test using high concentration of hydrogen peroxide (30% H2O2) with inherent iron in the kaolin and without cyclodextrin [327]. It was observed that in the absence of cyclodextrin, the oxidation was faster, but there was an accumulation of contaminants in the anode part of the system [327]. Moreover, it was also noted that the positions of electrodes and the way Fenton’s reagent was added into the system had a significant influence on the remediation efficiency [326,328]. EK-Fenton remediation of dredge marine sediment contaminated with petroleum hydrocarbons (TPH) and metals were carried out, testing EDTA and Tween 80 as flushing solutions. After 30 days of treatment, the EKe FentoneEDTA showed the better performance with the highest removal of TPH and metals [329]. A new oxidation-enhanced EK was developed, applying different oxidants (H2O2, NaClO, KMnO4, and Na2S2O8) and catholyte pH control (3.5 and 10) for the remediation of soil contaminated by both heavy metal copper and organic compound pyrene. The results showed that low pH favored the migration of copper, and KMnO4 was the best among the tested oxidants for pyrene degradation; but unfortunately, KMnO4 prevented the migration of copper by forming copper oxide. It was found that the use of Na2S2O8 and catholyte pH at 3.5 were the optimum operation conditions for remediation of the mixed contaminated soil by copper and pyrene, with removal rates achieved of 50% and 94%, respectively [330]. Recently, combination of the surfactant Igepal CA-720 and the oxidant persulfate in EK of PCB-contaminated soil was proven to be effective [331]. It was observed that PCBs was partially solubilized and transported toward anode by Igepal CA-720, while persulfate moved from cathode to anode by electromigration; and the optimum dosage of Igepal CA-720 and persulfate was 2% and 20%, respectively, for PCBs extraction and oxidation from the contaminated soil. It was noted that the addition of zerovalent iron as activator for persulfate turned out not to be a good choice as the zerovalent iron consumed most of the persulfate and limited its transport into the cell [331].

2. Literature Review 271

2.2.5 Electrokinetics and permeable reactive barriers To inhibit the extension of contaminants in the groundwater, “sorptive barriers” or “permeable reactive barriers” (PRB) may be installed across the flow path, capturing the contaminated plume and destroying the contaminants without soil or water excavation [332]. This innovative technology has been tested to treat both inorganic and organic pollutants.

2.2.5.1 Zero-valent metal permeable reactive barrier Zerovalent metals (ZVMs) can be used in reactive barriers as reducing agents for dechlorination of many chlorinated hydrocarbons. Combination of electrokinetics and ZVM for the remediation of PCE-contaminated soils showed that the best PCE removal efficiency can reach 99% after 10-day treatment, and the zerovalent zinc processes gave better PCE degradation than zerovalent iron. The EK-ZVM technique can maintain the neutral pH status with the appropriate operational parameters of sodium carbonate 0.01 M as anolyte, 1.0 V/cm voltage gradient, and ZVM installation close to the anode [333]. To investigate the potential use of atomizing slag as an inexpensive PRB material coupled with the EK processes to remediate contaminated ground of low permeability soils, the lab-scale EK and EK with PRB experiments were conducted. The results showed that the TCE concentrations of effluent solution through the PRB material were much lower than those of EK remediation without atomizing slag, due to the reaction between TCE and the reactive material that caused dechlorination of TCE. The removal efficiencies for TCE and Cd both achieved approximately 90%. Thus, the coupled technology of EK with a PRB system could be an effective, in situ remediation, and applying atomizing slag as the PRB reactive material seems very promising for the dechlorination of TCE [334]. As PCP transports back and forth and accumulates between anode and cathode, it is difficult to remove PCP by EK alone. Therefore, a PRB filled with reactive Pd/Fe was installed between anode and cathode to improve the efficiency of EK remediation for PCP. The mechanism of PCP removal was proposed, involving the transport of PCP by EK into the PRB compartment, the dechlorination of PCP to phenol by Pd/Fe in the PRB, and the subsequent moving out of phenol by electroosmosis. When positioning PRB right at the middle of the reactor and controlling PRB pH, the EK-PRB system was proven to be effective for PCP decontamination from soil [335]. The combination of Triton X-100 surfactant-enhanced electrokinetics and microscale Pd/Fe PRB proved to be efficient and promising for the remediation of HCB-contaminated soil. The results suggested that Triton X-100 was an effective enhancement agent for HCB removal, both in the

272 CHAPTER 4 Ultrasonic and electrokinetic remediation

electroosmotic movement and in the Pd/Fe degradation. The combined EKPRB greatly enhanced the removal of HCB compared with the EK alone (60% vs. 13%), indicating that in general, the degradation by the reactive Pd/Fe particles in the PRB rather than the movement by electroosmosis played a predominant role in HCB removal [336].

2.2.5.2 Lasagna process Lasagna is a novel integrated electrokinetic technology that creates permeable zones in close proximity sectioned through the contaminated soil region and turns them into “treatment zones” by introducing appropriate materials (sorbents, catalytic agents, microbes, oxidants, buffers, etc.). Electroosmosis is utilized to transport contaminants from the soil into “treatment zones,” where contaminants are removed from the pore water by sorption, immobilization, or degradation [337]. Thus, Lasagna technology reduces the distance the mobilized contaminant would have to travel before being removed/degraded and at the same time reduces the time required to remediate a site [272]. Lasagna remediates soils and soil pore water contaminated with soluble organic compounds and is especially suited to sites with low permeability soils where electroosmosis can move water faster and more uniformly than hydraulic methods, with very low power consumption. However, Lasagna does not work well in sandy soils where there is not enough hydraulic resistance to create significant electroosmotic gradients. Highly electrically conductive soils such as coastal or saline soils are also not practical because of the high current draw and excessive soil heating. Both vertical and horizontal configurations have been conceptualized, but fieldwork to date is more advanced for the vertical configuration [338]. TCEcontaminated soil was treated effectively using the Lasagna process. The levels of contamination before treatment were as high as 1500 mg/kg, and posttreatment levels were around 1 mg/kg [338]. The removal of the model organic compound p-nitrophenol from the soil was also very efficient, 98% in the pilot unit [337]. While activated carbon is the most common adsorbent used in Lasagna processes, a low-cost, easily available bamboo charcoal was employed, aiming to enhance the in situ electroremediation of 2,4-DCP and Cd from sandy loam at different polarity reversal intervals of 12 and 24 h. It was found that higher removal efficiency and lower energy consumption was achieved with the polarity reversal interval of 24 h, though in both operation modes, the periodic polarity reversal helped to keep soil moisture at a suitable level (favorable for electroremediation of contaminated soil with high permeability like sandy loam) and soil pH stable (around 7.2e7.4, close to the initial pH).

2. Literature Review 273

On the other hand, bamboo charcoal was proven to be an excellent adsorbent for both 2,4-DCP and Cd, making it a potential substitute for activated carbon. Moreover, its high water-holding capacity and high cation exchange capacity may also help to maintain soil moisture content and pH [339].

2.2.5.3 Electrochemical redox barriers One potential application of electrokinetic processes is the development of reactive electrochemical barriers for cleanup of mobile contamination plumes in groundwater [288]. Unlike the previous case of PRBs, the system does not require any reactive materials. The mechanism of this process is based on electrochemical redox reactions originated by water electrolysis under DC currents. The water electrolysis causes reduced environment and generation of hydrogen at the cathode and oxidized environment and generation of oxygen at the anode. The system has enhanced the reductive dehalogenation of chlorinated solvents. Significant advantages of electrochemical redox barriers over other types of barriers are as follow: continuous source of electrons, controllable rate of redox reactions, providing both reducing and oxidizing (or sequential) conditions, no need for additional reactive chemicals, electrode material is not consumed and may be reused, and flexibility in reversing the polarity of the electrodes [288].

2.2.6 Induced polarization ElectroChemical GeoOxidation Electro-Petroleum, Inc. in Wayne, PA, USA, and Electrochemical Processes, LLC. in Stuttgart, Germany, have developed the electrochemical remediation technologies (ECRTs) as an innovative, cost-effective, and rapid method for treating organic contamination in soil [340,341]. ECRTs can destroy organics in situ in the vadose zone and groundwater aquifers by using the ElectroChemical GeoOxidation (ECGO) process, which is considered the next generation in electrokinetics. ECGO is a geophysical process based on the phenomena of induced polarization (IP). ECGO utilizes a low voltage, low amperage alternating current (AC)/DC current passed between one or more electrodes pairs driven into the ground (soil, sediment, sludge, or groundwater) to rapidly address a wide range of both organic and inorganic compounds. When a DC current is imposed in the earth with a superimposed AC via in situ electrodes, the soil particles become polarized and develop electrical properties similar to a capacitor, discharging electricity. The energy given off induces redox reactions, which decompose organic contaminants. Empirical evidence indicates that reaction rates are reversibly proportional to grain size [273]. The soil particle surface area and the soil to water ratio are key parameters in determining the effectiveness of the technology.

274 CHAPTER 4 Ultrasonic and electrokinetic remediation

ECGO is also a patented in situ technology available from ManTech International Corporation that remediates soil and water contaminated with organic and inorganic compounds [276]. The process utilizes induced electric currents to create oxidationereduction reactions that lead to the mineralization of organic constituents (or the immobilization of inorganic constituents) present in a volume of soil and groundwater between the electrode locations. ECGO relies on the IP of naturally occurring conducting surfaces in soil and rock particles. These conducting surfaces are composed of elements such as iron, magnesium, titanium, and elemental carbon. Heavy metal impurities that are also naturally occurring further contribute to the process by acting as catalysts for the redox reactions. Depending on the site conditions, accessibility, and targeted constituents, the ECGO process may take 60e120 days [276].

2.2.7 Circulation-enhanced electrokinetics One major disadvantage of EK is the soil acidification during EK operation, which may dramatically destroy the soil constituents, affecting the soil zetapotential that causes decreasing electroosmotic flow. Therefore, the EK process with circulation system was designed to neutralize the pH of working solution and soil for avoiding the above problems [342e344]. The electrodes installed in the reservoir without attachment on soils can decrease the pH deviation of the soil matrix [344]. A siphon pipe and a pump were used to automatically neutralize the acid produced at the anode with the base produced at the cathode during EK operation. A hexane extractor was used to trap the organic compounds released from the contaminated soils [342]. Experimental results showed that this circulation-enhanced electrokinetics produced roughly stable pH (around 6) and conductivity of working solution, current density, and electroosmotic flow rate. All selected chlorinated organic compounds were effectively removed from the soil with removal efficiency ranging from 85% to 98% after 2 weeks [342].

2.2.8 Upward electrokinetic soil remediation An upward electrokinetic soil remediation (UESR) process is designed with vertical nonuniform electric field generated between an anode embedded in soil and a cathode placed on the soil surface. Unlike conventional electrokinetic treatment that uses boreholes or trenches for horizontal migration of contaminants, the UESR process uses vertical nonuniform electric field causing upward transportation of contaminants to the top surface of the treated soil. The main advantages of the UESR technology are minimization of site disturbance as well as reduction of treatment costs because a clean up of contaminated site takes place on the soil surface [345]. Experiments

2. Literature Review 275

demonstrated the feasibility of simultaneous removal of heavy metals (Cu, Pb) and organics (p-xylene, phenanthrene) from kaolin with this UESR process with removal efficiencies of phenanthrene, p-xylene, Cu, and Pb being 67%, 93%, 62% and 35%, respectively, after 6 days of treatment [345].

2.2.9 Periodic electric potential application It was postulated that applying the electric potential in a periodic mode, or disconnecting the voltage periodically, would increase micellar solubilization and enhance electrokinetic remedial efficiency. Then, experiments were conducted to assess the effects of employing a periodic voltage application during an electrokinetically enhanced in situ flushing process for PAH-contaminated clay soils. Four different bench-scale tests were conducted, using the nonionic surfactant Igepal CA 720 as flushing solution, with the voltage gradient applied continuously or periodically (5 day on and 2 day off), under relatively low voltage (1.0 V/cm) and high anode buffering (0.1 M NaOH) as well as high voltage (2.0 V/cm) and low anode buffering (0.01 M NaOH) conditions [346]. The results of these experiments confirmed the positive effect of the periodic voltage application with considerable high contaminant removal, which was attributed to a pulsed electroosmotic flow, a pulsed surfactant molecular movement produced, which generated a flushing action increasing solubilization and physically mobilizing the PAHs contaminants. Moreover, compared with continuous mode, the periodic mode tests sustained a relatively high average current value for a longer duration [346].

2.2.10 Nonuniform electrokinetics and rotational mode Nonuniform electrokinetics system with an electrode matrix, which operated with periodic polarity reversal or rotational operation mode to accelerate bioremediation by mixing organic pollutants and bacteria in soil, was studied with a phenol-contaminated sandy loam. Compared with the unidirectional and bidirectional operation, the results showed that the rotational operation could effectively stimulate the biodegradation of phenol in the soil if adopting appropriate time intervals of polarity reversal and electrode matrixes. With a reversal interval of 3 h and a square-shaped electrode matrix, a maximum phenol removal of 58% was achieved in 10 days and the bioremediation rate was increased about five times as compared with that with no electric field applied [347]. In another study, the EK with sandy loam as the model soil and 2,4-DCP as the model organic pollutant were investigated at two nonuniform operation modes (bidirectional and rotational) in a hexagonal matrix. Periodically, the electric field reverses its direction at bidirectional mode and revolves at a given angle at rotational mode. The results showed that the nonuniform

276 CHAPTER 4 Ultrasonic and electrokinetic remediation

electric field could effectively stimulate the desorption and movement of 2,4-DCP toward anode. At the bidirectional mode, an average 2,4-DCP removal of 73.4% was achieved in 15 days, and the in situ biodegradation of 2,4-DCP was increased by about three times as compared with that uncoupled electric field, whereas 34.8% of 2,4-DCP was removed on average in the same period at the rotational mode. While the bidirectional mode was more effective and cost efficient, the rotational mode provided better remediation’s uniformity in soil [348].

2.2.11 Ultrasound-assisted electrokinetic remediation While EK was applied to remove mainly heavy metals and ultrasound was applied to decontaminate mainly organic substances, these two techniques was combined for the removal of heavy metal Pb and phenanthrene in natural clay, confirming the synergistic effects on the pollutant migration as well as decontamination [31]. The decontamination efficiency in the combined electrokinetic and 30 kHz, 200 W ultrasonic system (91% for Pb and 90% for phenanthrene) was higher than in the simple EK alone (88% for Pb and 85% for phenanthrene). It was observed that ultrasound caused decreasing fluid viscosity, increasing flow rate, porosity, and permeability as well as mobilizing sorbed contaminants [31]. Based on the literature review, it was found that very few studies (probably only one [31], as far as we know) that combined both electrokinetics and ultrasound in the remediation of POPs contaminated clayey soils. As the integration of electrokinetics and ultrasound may bring positive synergistic effects, more exploration is needed in this research field.

3.

OBJECTIVES

The aim of the study was to investigate the effectiveness of ultrasonic treatment, individually and as an enhancement for electrokinetic treatment in the remediation of persistent organic contamination from low permeability clayey soils. The specific objectives of this study were as follow:  To investigate the effects of ultrasonic treatment with various experimental conditions on the decontamination of POPs from different types of clayey matrices (paper I, II).  To investigate the coupling effects of electrokinetic and ultrasonic remediation on the decontamination of POPs from kaolin, in different experimental setups (paper III, IV, V).  To investigate the comparative enhancements of EK using 2hydroxylpropyl-b-cyclodextrin surfactant and ultrasound for the decontamination of POPs from kaolin (paper VI).

4. Materials and Methods 277

4. MATERIALS AND METHODS 4.1 Materials 4.1.1 Model persistent organic pollutants and clay The representative POPs chosen in the study were HCB, a typical polychlorinated hydrocarbon, and three PAHs, phenanthrene (PHE), fluoranthene (FLU), and chrysene (CHR). HCB is considered as 1 of the 12 worst offenders (known as the “dirty dozen”) in the Stockholm Convention on POPs, and the three PAHs are listed in the US EPA’s 16 priority pollutant PAHs [2,349,350]. HCB and fluoranthene were purchased from Sigma-Aldrich, phenanthrene and hexane solvent from Merck, and chrysene from Acrös Organics. All chemicals were of analytical grade and used without further purification (Table 4.2). Model low permeability clayey soil used in the experiments was white kaolin purchased from VWR International. Kaolin was often used as the model clay due to its high content of clay and negligible content of organic matter [354]. Main characteristics of the kaolin were measured and summarized in Table 4.3. For pH and electric conductivity measurement, 10 g kaolin was mixed with 20 mL distilled water, using the pH meter (pH 730 inoLab, WTW series) and the portable EC meter (waterproof EC Testr low, Eutech Instruments Pte, Ltd.), respectively. The dry bulk density of kaolin was measured by weighing the given volume (100 cm3) of kaolin. Cation exchange capacity was estimated by determination of exchangeable

Table 4.2 Properties of the Model Persistent Organic Pollutants [350e353]. Properties

Hexachlorobenzene (HCB)

Phenanthrene (PHE)

Fluoranthene (FLU)

Chrysene (CHR)

Chemical formula

C6Cl6

C14H10

C16H10

C18H12

284.78 204 231 322 0.005 (insoluble)

178.23 1.18 101 332 1.15 (insoluble)

202.26 1.25 110.8 375 0.265 (insoluble)

228.28 1.27 254 448 0.002 (insoluble)

Chemical structure

Molecular mass (g/mol) Density (g/cm3) Melting point ( C) Boiling point ( C) Water solubility at 25  C (mg/L)

278 CHAPTER 4 Ultrasonic and electrokinetic remediation

Table 4.3 Main Characteristics of Kaolin. pH Dry bulk density (g cm3) Moisture (%) Electrical conductivity (mS cm1) Cation exchange capacity (cmol kg1) Organic content (%) Particle size distribution: % sand (>0.05 mm) % silt (0.05e0.002 mm) % clay ( EK EK þ US > EK (in tests with plexiglass cylinders)

POPs removal

US does not affect pH of soils; pH patterns along kaolin profile are quite similar among EK and EK þ US EK þ US > EK > US For average PHE removal: EK þ US > EK þ SF For average HCB removal: EK þ SF > EK þ US

0.02e0.03 A), which had higher initial current (0.04e0.06 A) and tended to decrease along the time and then became constant (0.01e0.03 A). As current mainly results from the electromigration of ions through pore fluid, the higher currents observed in EK þ US tests may be attributed to the ultrasonic effects, which made the slurries more porous and permeable. The decrease of current values overtime may be explained through the fact that as mobile ions constantly electromigrated toward the electrodes, they were neutralized by reacting with the soil, with other species in the slurries or with the opposite charged electrode [346]. Because of evaporation (and electroosmosis toward cathode), the slurry’s moisture tended to decrease along the time (particularly in the anode parts), and this could also affect the currents. In the tests with plexiglass cylinders, the anolyte level decreased overtime due to the electroosmotic flow from anode to cathode. Therefore, anolyte was added daily to maintain its level. The cumulative amount of this anolyte addition was recorded for calculating the electroosmotic flow. It was observed that the electroosmotic flows of EK þ US tests were higher than of EK tests. This implied that ultrasound increased the liquid outflow due to sonication effects that increased the matrix porosity and permeability [31].

5.2.2 Distribution of pH There was no significant difference of pH distribution along the kaolin profiles between EK and EK þ US tests. This again confirmed the previous finding that ultrasonication did not affect the pH of the kaolin slurry, and thus pH distribution pattern was affected solely by the electrokinetic process. In both EK and EK þ US tests, the section near anode had the lowest pH and the section near cathode had the highest pH because the electrolysis of water created acid front at anode and base front at cathode. Because of

5. Results and Discussion 287

electromigration and electroosmosis, the low pH solution (high Hþ concentration) generated at anode moved toward cathode [346]. As the electromigration of Hþ was concurrent with the electroosmotic flow and the mobility of ion Hþ is about 1.76 times higher than that of OH [280], the low pH solution migrated faster and dominated the kaolin profiles. Moreover, as the initial pH of kaolin was around 4.7, most of the kaolin profiles had the pH values in the range of 4e6.

5.2.3 Persistent organic pollutants removal Certain amounts of POPs were removed from kaolin in all the treatment tests through complex mechanisms of electrokinetic and ultrasonic processes. Ultrasound increased kaolin porosity and permeability as well as desorption of the low soluble POPs, thus enhanced electroosmotic mobilization of these contaminants. Moreover, ultrasonication could induce high fluid-solid sheer stresses [234], thermal decomposition, and hydroxyl radical oxidation [33] that contributed to the sonolysis of these POPs in the slurries. Residues of POPs tended to remain the highest at the central part of the kaolin profile and decreased at the two electrode ends, and this could be seen clearer in EK tests than in EK þ US tests because of the physical water-kaolin mixing effect of ultrasound. Oxygen produced at anode could be attributed to organic oxidation in these parts. During experiments, kaolin accumulated in anode side due to electrophoresis (as kaolin particle’s surface charge is negative), while water accumulated in cathode side (due to electroosmotic flow). Therefore, slurries in the cathode parts were more dilute, and the contaminants there might be more easily desorbed and removed from kaolin. On the other hand, in both EK and EK þ US tests, the two PAHs, PHE and FLU, tended to be removed more in the anode side while HCB tended to be removed more in the cathode side. It could be explained that the two PAHs were more easily destroyed by oxidation than HCB, and HCB can be desorbed, mobilized, and removed from the dilute slurries. Among all of the POPs tested, HCB was the most difficult to treat because of its very stable chemical structure and low water solubility. The initial POPs concentration was an important input parameter as low initial contaminant concentration tests gave better remediation efficiencies than higher initial contaminant concentration tests. In all of the experimental setups, the EK þ US tests gave the highest POPs removal efficiencies and the US alone tests gave the lowest POPs removal efficiencies. In general, with the same conditions, the removal efficiencies from EK þ US tests were 2%e17% higher than the EK alone tests, depending on various setups and on the contaminants. In the case of series tests for

288 CHAPTER 4 Ultrasonic and electrokinetic remediation

CHR remediation, the EK þ US test in pan with iron anode provided the highest average CHR removal. In all of the cases, ultrasound assistance provided better uniformity of the remediation because of its mixing effect. Comparative tests between EK þ SF and EK þ US showed that EK þ US test demonstrated better PHE removal, while EK þ SF test performed slightly better HCB removal. This again implied that the remediation mechanism of PHE involved with oxidation degradation, while HCB removal was more attributed to the enhanced desorption and solubilization.

6.

CONCLUSIONS

This study investigated the effectiveness of ultrasonic treatment, individually and as an enhancement for electrokinetic treatment in the remediation of persistent organic contamination from low permeability clayey soils. HCB, PHE, FLU, and CHR were chosen as the representative POPs and kaolin as the model low permeability clayey soil. From the experimental observation, key findings are summarized in the following points:  Ultrasonication could reduce the POPs contamination in low permeability soils, but the remediation efficiency was low, particularly in short duration time. For complete removal, ultrasonication should not be considered as a single treatment process but rather be integrated with other appropriate techniques.  The efficient treatment of soil by ultrasonication required a certain amount of water. Practically, the reasonable water:soil ratio of the soil slurry could be in the range of 1e2:1.  Because of the heating effect observed during ultrasonication, soil slurry could be dried out, and thus ultrasonic treatment should not be operated over very long time. The intermittent ultrasonication could overcome this heating effect for long-term treatment.  Ultrasonication did not affect pH of the slurries.  The combined electrokinetic and ultrasonic treatment proved positive synergistic effect in POPs removal than each single process alone, though the level of enhancement was not high, only from 2% to 17%.  Ultrasonication sustained higher current, higher electroosmotic flow, and more uniform remediation in the combined EK þ US tests, as compared with the EK alone tests.  Initial contaminant concentration is an important input parameter that can seriously affect the remediation effectiveness. Removal efficiency decreased with increasing initial contaminant concentration.  Among the POPs, HCB was the most difficult to treat because of its high chemical stability and low water solubility.

References 289

Although the lab-scale study showed some potential of ultrasonic and electrokinetic treatment in POPs decontamination from low permeability medium, the technique is far from large-scale practical application. Therefore, for practical scale-up systems, further studies are needed, considering technical design, economic factor, and limitations of physical impacts such as heating effect and noise during ultrasonication. On the other hand, as the EK þ US tests used the model kaolin, more research on real contaminated clay soils should be conducted to examine their interaction and effectiveness on real-life situation. In addition, besides HCB, PHE, FLU, and CHR, further studies on many other POPs contamination remediation are also needed.

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Chapter

5

Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Heikki Särkkä1, Mika Sillanpää2

1

2

Department of Built Environment, Aalto University, Espoo, Finland; Department of Civil and Environmental Engineering, Florida International University, Miami, FL, United States

CHAPTER OUTLINE

List of Publications 312 List of Symbols 313 Abbreviations 313 1. Introduction 314 1.1 Pulp and paper mill circulating waters and wastewaters

314

1.1.1 General 314 1.1.2 Microorganisms in the paper mill environment 314

1.2 Wastewater treatment in the pulp and paper mills 1.2.1 General 315 1.2.2 Primary and secondary treatment 1.2.3 Tertiary treatment 317

315

316

1.3 Electrochemical oxidation in water and wastewater treatment 1.3.1 1.3.2 1.3.3 1.3.4 1.3.5

319

General 319 Theory of electrooxidation 321 Electrodes 323 Treatment of different wastewaters 326 Disinfection of wastewater and drinking water 330

2. Objectives of the Study 332 3. Materials and Methods 333 3.1 Reactors and electrodes 3.2 Chemicals 335

333

3.2.1 Wastewaters used for the experiments 3.2.2 Biocides 335

3.3 Bacterial strains 3.4 Analyses 336

335

336

3.4.1 Bacteria 336 Advanced Water Treatment. https://doi.org/10.1016/B978-0-12-819227-6.00005-X Copyright © 2020 Elsevier Inc. All rights reserved.

311

312 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

3.4.2 Oxidants 337 3.4.3 Cyclic voltammograms 337 3.4.4 Other analyses 337

4. Results and Discussion

338

4.1 Cyclic voltammograms 338 4.2 Electrochemical inactivation of bacteria

340

4.2.1 Aerobic bacteria in synthetic paper mill water 340 4.2.2 Anaerobic bacteria in paper mill wastewater 346

4.3 Electrochemical oxidation of sulfide 347 4.4 Electrochemical oxidation of organics in pulp and paper mill bleaching effluent 349 4.5 Electrochemical degradation of methyl orange dye 349

5. Conclusions and Further Research References 352

n

350

LIST OF PUBLICATIONS

This summary is based on the following papers. I. H. Särkkä, M. Vepsäläinen, M. Pulliainen, M. Sillanpää, Inactivation of Deinococcus geothermalis bacteria in synthetic paper machine water by electrochemical oxidation, J. Pulp Paper Sci. 33 (2007) 95e99. II. H. Särkkä, M. Vepsäläinen, M. Pulliainen, M. Sillanpää, Electrochemical inactivation of paper mill bacteria with mixed metal oxide electrode. J. Hazard. Mater. 156 (2008) 208e213. III. H. Särkkä, K. Kuhmonen, M. Vepsäläinen, M. Pulliainen, J. Selin, P. Rantala, E. Kukkamäki, M. Sillanpää, Electrochemical oxidation of sulphides in paper mill wastewater by using mixed oxide anodes. Environ. Technol. 30 (2009) 885e892. IV. H. Särkkä, M. Kolari, M. Pulliainen, M. Sillanpää, Potential generation of oxidizing radicals in synthetic paper mill water by electrochemical treatment combined with biocides. Curr. Org. Chem. 16 (2012) 2054e2059. V. M. Zhou, H. Särkkä, M. Sillanpää, A comparative experimental study on methyl orange degradation by electrochemical oxidation on BDD and MMO electrodes. Sep. Purif. Technol. 78 (2011) 290e297. VI. K. Eskelinen, H. Särkkä, T.A. Kurniawan, M.E.T. Sillanpää, Removal of recalcitrant contaminants from bleaching effluents in pulp and paper mills using ultrasonic irradiation and Fenton-like oxidation, electrochemical treatment, and/or chemical precipitation: a comparative study. Desalination 255 (2010) 179e187.

Abbreviations 313

n

LIST OF SYMBOLS

HO2* I M *OH R t

n

Perhydroxyl radical Current Electrode surface Hydroxyl radical Organic pollutant Time

ABBREVIATIONS

AOP AOX APC BDD BOD BSTFA CEH CFU COD CV DAF DO DOC DSA EC EF EO GC GCE HFCVD MB MF MMO MO MS MTBE NF NOM PAM PCA RO ROS

Advanced oxidation processes Adsorbable organic halides Aerobic plate count Boron-doped diamond Biological oxygen demand Bis(trimethylsilyl)trifluoroacetamide Chlorination, extraction, and hypochlorite bleaching Colony forming unit Chemical oxygen demand Cyclic voltammogram Dissolved air flotation Dissolved oxygen Dissolved organic carbon Dimensionally stable anodes Electrocoagulation Electroflotation Electrooxidation Gas chromatograph General current efficiency Hot filament chemical vapor deposition Methylene blue Microfiltration Mixed metal oxide Methyl orange Mass selective Methyl tert-butyl ether Nanofiltration Natural organic matter Polyacrylamide Plate count agar Reverse osmosis Reactive oxygen species

314 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

RSM SCE SEM SHE SPEF SPW SS TCF TMCS TOC UF UV

Response surface methodology Saturated calomel electrode Scanning electron microscope Standard hydrogen electrode Solar photoelectro-Fenton Synthetic paper mill water Stainless steel Totally chlorine free Trimethylchlorosilane Total organic carbon Ultrafiltration Ultraviolet

1. INTRODUCTION 1.1 Pulp and paper mill circulating waters and wastewaters 1.1.1 General The pulp and paper industry, like the primary metal and chemical industries, uses a great deal of raw water [1]. Water is needed at paper mills mainly for the pulp washing procedure and paper forming process in the wet end phase. Modern mills need 15e50 m3 of water to produce a ton of paper or cardboard [1,2] and consumption varies depending on the manufacturing process and country. Pulp and paper mill circulating waters and wastewaters are a very complex mixture of different organic and inorganic compounds. Wastewaters from pulp making mainly contain lignin, resin acids, fatty acids, and dissolved wood extractives [3e5]. After pulp bleaching, inorganic chlorine and AOX compounds may also be present, depending on the bleaching procedure [4]. The wastewater generated from the paper-making process contains particulate waste, organic compounds (e.g., fatty and resin acids), anions, inorganic dyes, and biocides [6e9]. These substances can be highly toxic in natural waters causing fish death and negatively affecting the whole ecosystem [3].

1.1.2 Microorganisms in the paper mill environment A recent trend is the circulation of process waters, and purified wastewater may even be used again in the paper-making process. Recycling of these waters is important to utilize as much fibers as possible in paper sheete making process [6]. This can reduce freshwater consumption but introduce other challenges, such as the neutralization of cationic retention chemicals, corrosion problems, and the biofouling of pipelines by microbes. Biofouling is usually prevented by using chemicals (biocides), which are often relatively toxic to handle, and the risk of negative effects in the receiving waters is

1. Introduction 315

increased after wastewater treatment [6]. Some biocides can also retain in fibers and accumulate in the final paper product. Because of the nutrient-rich, moist, and warm paper mill environment, many microbes exist in paper mill circulating waters. Microorganisms are constantly introduced into paper machines through raw materials, water, fibers, and paper-making chemicals [10e12]. Microbial growth, if not controlled, can result in slime formation which can seriously disturb the paper-making process and have a negative impact on the quality of the paper [11]. Free-living bacteria are not a major problem (some degrade fiber, for example), but when these species attach to the pipelines creating floccules or slime, the process is easily disturbed, causing breaks in the paper web [12]. Attached bacteria growing on a surface as a biofilm are not easy to remove by physical means such as washing. Mechanical cleaning is also a common method, though switching off the paper machine for cleaning the pipelines causes financial losses. Thus, the paper industry uses biocides to control excessive bacterial growth [6,12e14]. Biocides are usually classified as either oxidizing or nonoxidizing, and chlorine, hydrogen peroxide, and peracetic acid are examples of oxidizing species used in mills. Paper mill microbe communities vary a lot depending on the mill and papermaking process used. Direct microscopic examination of slime deposits shows microbes with different shapes and sizes [11]. Filamentous organisms can be fungi or bacteria which are capable of accumulating large amounts of colorful slime. Many species have been isolated in these environments, such as Deinococcus, Pseudoxanthomonas, Meiothermus, and Bacillus [12,14,15]. Väisänen et al. [12] isolated totally 390 strains of aerobic bacteria from printing paper machines. Deinococcus geothermalis was one of the species that could adhere to stainless steel surface only within 1 day. It is known for forming firm colored biofilms and for its persistence against cleaning and chemical treatments [13]. Some of these species can attach firmly to the surface of the pipelines and sometimes further species can attach to these, speeding up the biofouling process [16,17]. In favorable conditions, microbes can also produce odorous and toxic compounds such as H2S or cause localized corrosion under the biofilm deposits. Many strains also degrade paper-making raw materials, such as fibers and chemicals.

1.2 Wastewater treatment in the pulp and paper mills 1.2.1 General The wastewater produced during paper-making process has to be purified before piping into natural waters or recycling back into process. The quality

316 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Table 5.1 Characteristics of Wastewater at Various Pulp and Paper Processes. Parameters Process

pH

TS (mg/L)

SS (mg/L)

BOD5 (mg/L)

COD (mg/L)

Color (PteCo)

Reference

Kraft mill Bleached pulp mill Pulp and paper Paper making Paper mill Paper machine

8.2 7.5 7.8 7.8 8.7 4.5

8260 e 4200 1844 2415 e

3620 1133 1400 760 935 503

e 1566 1050 561 425 170

4112 2572 4870 953 845 723

4667.5 4033 Dark brown Black Dark brown 243

[18] [19] [20] [21] [22] [19]

Adapted from D. Pokhrel, T. Viraraghavan, Treatment of pulp and paper mill wastewater e a review, Sci. Total Environ. 333 (2004) 37e58.

of the wastewater varies significantly depending on the processes and chemicals used for pulp- and paper-making operations. Some wastewater quality parameters for various pulp and paper processes are presented in Table 5.1 [3]. It can be concluded that the variation between wastewater parameters is significant in different processes, increasing the challenges for wastewater treatment processes at mills. The main treatment process used at pulp and paper mills is primary clarification followed by biological treatment (secondary treatment). A tertiary process is needed, e.g., for removing recalcitrant organic compounds or color before water is circulated again into the process or to rivers or lakes (Fig. 5.1) [1]. Primary clarification can be achieved by either settlement or flotation, secondary treatment by aerobic or anaerobic biological treatment, and tertiary treatment, for example, by membrane processes, advanced oxidation processes (AOPs), or coagulation/ flocculation.

1.2.2 Primary and secondary treatment Sedimentation units usually achieve high removal rates for suspended solids, but little organic material (BOD and COD) is eliminated [1]. Flotation is also used for the removal of suspended solids [3]. Recently, electrochemical techniques such as electroflotation (EF) have been applied in industrial wastewater effluents as a primary or tertiary treatment [23,24]. Ben Mansour et al. [24] reported that the purification efficiency of suspended solids by combined coagulation/EF process exceeded 95%.

1. Introduction 317

Screen

Weir

From mill

Chemicals

Nutrients Primary treatment – Settlement and flotation

Secondary biotreatment

Return activated sludge

Secondary clarifier

Sludge and excess biomass for disposal

To river or tertiary treatment

n FIGURE 5.1 Generalized schematic diagram of the plant for the treatment of paper mill effluent. Adapted from G. Thompson, J. Swain, M. Kay, C.F. Forster, The treatment of pulp and paper mill effluent: a review, Bioresour. Technol. 77 (2001) 275e286.

Aerobic treatment is still the most popular secondary treatment technique for the removal of soluble biodegradable organic pollutants from pulp and paper mill effluents. There are numerous aerobic biological treatment systems available, but the most common is the activated sludge process, which can achieve high removal efficiencies for BOD and COD [1,3]. This process has certain disadvantages, however, such as the production of sludges with very variable settlement properties, sensitivity to shock loading and toxicity, and limited capacity to remove poorly biodegradable toxic substances. An anaerobic process has potential advantages over aerobic treatment, such as lower sludge production and chemical consumption [1]. It also produces methane which can be used for energy production. It also has its limitations, including potential hydrogen sulfide production in pulp and paper mill effluents due to the high sulfur content of wastewaters.

1.2.3 Tertiary treatment Tertiary treatment of pulp and paper mill wastewaters is often needed to reach the target limit values set by the authorities for the further removal of residual COD, toxicity, color, and microorganisms. Membrane-based

318 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

techniques, AOPs, and physicochemical processes are often used for this purpose [1,3]. Among the advanced treatment processes, membrane technology offers an alternative [25]. Membrane techniques include microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO). Pizzichini et al. [26] pilot-scale tested these techniques to remove salts, COD, and total organic carbon (TOC) from the paper mill wastewater. MF followed by RO filtration could recycle more than 80% of the original wastewater back into paper-making process. Beril Gönder et al. [25] found out that biologically treated pulp and paper mill wastewater could be used again as process water using two-step NF. UF removed metals from totally chlorineefree wastewater, and the process was enhanced by adding water-soluble polymeric ligands [27]. Chemical coagulation using alum, ferric chloride, ferric sulfate, and lime has been used extensively in the treatment of different wastewaters [28]. Interest in the use of synthetic polyelectrolytes as flocculants for pulp mill wastewater treatment has grown recently [29]. The main advantage of polymeric flocculants is their ability to produce large, dense, compact, and stronger floccules with good settling characteristics compared with those obtained by coagulation alone [28]. In particular, polyacrylamide flocculants have been used intensively in pulp and paper mill effluent treatment and proved their economic feasibility [28,29]. Coagulation/flocculation treatment can also be combined successfully with other techniques, such as heterogeneous photocatalysis [30]. Adsorption of organic pollutants from pulp and paper mill effluents is an alternative technique. Zhang and Chuang [31] compared the performance of styrene divinylbenzene copolymer and activated carbon for the acidic bleach plant effluent treatment. It was observed that resin is more effective than activated carbon in color removal, and that it is possible to regenerate resin by washing with sodium hydroxide solution. AOPs for wastewater treatment have received significant attention in recent years. AOPs are based on the generation of very reactive nonselective oxidizing species such as the hydroxyl radicals [32]. These radicals can be formed by combining the following oxidizing agents: ozone (O3), hydrogen peroxide (H2O2), ultraviolet (UV) radiation, ferrous and ferric salts (Fe2þ and Fe3þ), and catalysts such as TiO2. Catalkaya and Kargi [32,33] utilized these techniques in combination or individually to purify pulp mill effluents from color, TOC, and AOX compounds. The results showed that TiO2-assisted photocatalysis (UV/TiO2) achieved the highest TOC and toxicity removals [32], and in another study [33], Fenton’s reagent utilizing

1. Introduction 319

H2O2/Fe2þ resulted in the highest color, TOC, and AOX removals under acidic conditions compared with the other AOPs tested. Solar photocatalysis [34] with Fenton reagent and TiO2 and solar photo-Fenton [35] reactions have also been shown to be effective in the removal of refractory organic compounds. Catalyzed ozonation has demonstrated promising results compared with conventional ozonation treatment. Homogeneous and heterogeneous catalysts have been used for the treatment of pulp and paper mill effluents [36,37]. Activated carbon together with ozonation treatment could increase the BOD5/COD ratio from 0.11% to 0.28%, and 87% of color was removed from chlorination, extraction, and hypochlorite bleaching effluent [36]. Fontanier et al. [37] used TOCCATA catalyst in treating three different wastewaters from pulp and paper mills. It was shown that organic matter was removed through the steady conversion of organic carbon to carbon dioxide. Ozonation has also been applied to NF concentrate to increase its biodegradability [38]. Electrochemical techniques have been applied recently in pulp and paper mill wastewater treatment. Vepsäläinen et al. [39] investigated natural organic matter (NOM) removal by electrocoagulation (EC) together with chemical coagulation. The results indicated that the combined method was efficient in removing NOM even with small electric charges per liter. EC has also yielded promising results in sulfide and toxic pollutant removal from pulp and paper mill effluents [40,41]. Although several pulp and paper mill wastewater treatment techniques are available, many of these are still associated with high investment and running costs or lack of efficiency to remove refractory organic pollutants from effluents. It is important to maximize purification efficiencies and develop new, simple in situ techniques for treating pulp and paper mill effluents and circulating waters. Strategies that promote lower energy use, reduce the amount of solid waste produced, and increase efficient energy recovery are economically sustainable [42].

1.3 Electrochemical oxidation in water and wastewater treatment 1.3.1 General Electrochemical techniques have been applied extensively to treat various wastewaters, disinfect drinking water, or enhance polluted soils [43e52]. These include, e.g., EC, EF, electrooxidation (EO), and electrokinetic treatment. Conventional water purification techniques, such as chemical

320 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

coagulation, biological treatment, or UV oxidation, are not effective against some toxic and refractory organic pollutants. Electrochemical techniques can offer a more efficient means of treating these pollutants. Electrochemical techniques are innovative, inexpensive, and effective methods for purifying wastewaters from many industrial processes before discharge into water systems or circulation back into processes [44]. They could also be called “green technology” methods because little or no chemicals are needed to facilitate water treatment. Electrochemical techniques can also be applied in sludge treatment. Tuan et al. [46] wrote a review about applying electro-dewatering to sewage sludge treatment. It was found that electro-dewatering has several potential benefits, such as lower energy and transportation costs and enhanced solideliquid separation of sludge. Yet high operation costs still hinder the use of electro-dewatering in largescale applications. Electrokinetic Fenton process is a promising technology for the remediation of low permeable soil [47,48]. It is effective for highly biorefractory contaminants, such as hexachlorobenzene. It is important to optimize parameters such as pH level at the cathode region or contact between the contaminant and the oxidant during the treatment, however [47,48]. Ultrasonically enhanced electrokinetics can increase pollutant removal from soils [49,50]. It has several advantages, such as lack of dangerous breakdown products and compact and transportable on-site treatment. Yet it still has technical limitations such as scaling up and physical effects such as noise. EC is currently enjoying both increased popularity and major technical improvement [44]. It is a rather simple and robust technique and can be applied in many environments. However, it involves several chemical and physical phenomena which should be well understood before effective treatment. When “sacrificial” iron and aluminum anodes are releasing ions into water, EF occurs simultaneously by hydrogen and oxygen bubbles released at the electrodes, improving purification efficiency. Disadvantages of EC are the periodical need for replacement of anodes, electrode passivation, and the lack of any systematic approach to reactor design and operation [45]. An important factor in EF treatment is the size of bubbles formed at the electrodes [43]. Smaller bubbles are more efficient in pollutant removal because of the larger surface area available for particle attachment. EF is a promising technique especially in oily wastewater treatment and it shows advantages over either DAF or settling.

1. Introduction 321

EO treatment has received a great deal of interest in wastewater treatment and drinking water disinfection in recent years [43]. The technique is rather simple. Oxidants are produced during the treatment in situ either at the electrodes or indirectly by chemical compounds in the treated water. Unfortunately, the lack of efficient and stable yet economical electrode material has hindered the large-scale application of this technique to date.

1.3.2 Theory of electrooxidation Thermodynamically, the electrochemical degradation of any soluble organic compound in water should be achieved at low potentials, before the thermodynamic potential of water oxidation to molecular oxygen (1.23 V/SHE under standard conditions) as indicated in Formula 5.1 [53]:

2H2O / O2 þ 4Hþ þ 4e

(5.1)

In acidic media, water can be discharged on the electrode producing highly oxidative absorbed hydroxyl radicals (Reaction 5.2):

H2O þ M / M(*OH) þ Hþ þ e

(5.2)

where M means electrode surface. These radicals are physisorbed on the anode surface where the organic pollutant R can be oxidized as follows (Reaction 5.3): RðaqÞ þ Mð*OHÞn=2 /M þ Oxidation products þ

n þ n  H þ e 2 2

(5.3)

where n is the number of electrons involved in the oxidation reaction of R. The reaction of organics with electrogenerated hydroxyl radicals (3) is in competition with the side reaction of the anodic discharge of these radicals to oxygen (Reaction 5.4): Mð *OHÞ / M þ

1 O2 þ Hþ þ e 2

(5.4)

Anodic activity by electrodes depends on their value of overpotential for oxygen evolution [43]. Platinum electrodes have much lower potential value for oxygen evolution reactions (1.3 V vs. SHE) than, for example, SnO2 electrodes (1.9 V vs. SHE) or boron-doped diamond electrodes (Ti/BDD, 2.7 V vs. SHE). This signifies that anodic oxidation by hydroxyl radicals can take place on a Ti/BDD electrode surface at a significantly higher current density with a minimal oxygen evolution side reaction.

322 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Electrochemical oxidation may also occur by an indirect process where oxidants such as chlorine, hypochlorous acid, and hypochlorite [54e60] or hydrogen peroxide/ozone [61e64] are formed at electrodes by following Reactions (5.5e5.11):

2Cl / Cl2 þ 2e

(5.5)

Cl2 þ H2O / HOCl þ Hþ þ Cl

(5.6)

HOCl / Hþ þ OCl

(5.7)

H2O / *OH þ Hþ þ e

(5.8)

2*OH / H2O2

(5.9)

H2O2 / O2 þ 2Hþ þ 2e

(5.10)

O2 þ *O/ O3

(5.11)

At acidic pH, chlorine is present in the solution in the form of hypochlorous acid, which has a higher oxidation potential (1.49 V) than hypochlorite (0.94 V). Under alkaline conditions, hypochlorite is a dominant species [55]. Under higher pH conditions, more chlorate or perchlorate is also formed instead of chlorine/hypochlorite, which decreases oxidation efficiency. Higher initial chloride concentration in the electrolyte solution naturally encourages more chlorine/hypochlorite production [54e56]. A similar effect can be observed by lowering the temperature of the electrolyte solution, but this depends on the electrode material used [56]. Higher current density also increases chlorine production [55,56]. Electrochemical production of reactive oxygen species (ROS) such as ozone and hydrogen peroxide has proved effective in water disinfection without the mediation of active chlorine products [61]. In an electro-Fenton oxidation process, H2O2 is continuously generated in acidic solutions by the reduction of O2 at the graphite cathodes or carbon felt [62]. When Fe2þ is introduced as the catalyst, a Fenton reaction takes place in the solution, generating hydroxyl radicals to decompose pollutants or disinfect bacteria. Chu et al. [63] recognized effective degradation of 4-nitrophenol by

1. Introduction 323

employing a dual-cathode system to generate H2O2 and Fe2þ simultaneously at two cathodes to encourage an electro-Fenton reaction in the solution together with anodic oxidation by hydroxyl radicals at the anode. Guinea et al. [64] achieved almost total mineralization of enrofloxacin solutions by solar photoelectro-Fenton treatment using an O2-diffusion cathode. Some examples of indirect EO studies are shown in Table 5.2. Efficient organic pollutant removal can be achieved also by generating Fe2þ in situ by an EC process where an iron anode is dissolved into a solution, causing a Fenton reaction when hydrogen peroxide is added to the solution [65,66]. Martins et al. [65] could remove 95% of nonylphenol polyethoxylate in 5 min (aqueous solution) and in 10 min (wastewater), respectively.

1.3.3 Electrodes Several electrodes have been used for water treatment by electrochemical oxidation. Anodes used for water and wastewater treatment include lead and lead dioxide [67e70], dimensionally stable anode (DSA) electrodes [71e74], graphite [75e77], and BDD electrodes [70,78e85]. Lead and lead dioxide have been used as anodes because of their stability, low cost, and high oxygen evolution potential which delays O2 evolution in favor of Cl2 evolution [67]. Hamza et al. [68] completely mineralized of 1,3,5-trimethoxybenzene in acid media at a Ta/PbO2 anode. They discovered that all oxidation products were finally oxidized to CO2 by the intermediary of carboxylic acids. Awad and Abo Galwa [69] found out that the electrocatalytic activity of a lead dioxide electrode depends on the conductive electrolyte. They concluded that in the presence of H2SO4 electrolyte, electrode poisoning occurred, because an adherent film was formed on the anode surface. The dissolution of toxic Pb2þ ions also hinders the use of lead and lead dioxide as anodes [43]. DSAs are catalytic oxide electrodes which can effectively generate active hydroxyl radicals and active chloride species [72]. They also have a relatively high overpotential for oxygen evolution. Efficient degradation of paper mill wastewater was achieved using three-dimensional electrodes (Ti/ Co/SnO2eSb2O5) combined with activated carbon treatment [72]. This was mainly due to the fact that the conversion rate within an electrochemical reactor can be increased substantially due to its large specific surface area in comparison with conventional two-dimensional electrodes. So-called “nonactive” electrodes such as SnO2 form hydroxyl radicals on their surface more easily, which can result in the complete oxidation of the organic molecules to CO2 [74]. With “active electrodes,” such as RuO2 and IrO2, only selective oxidation of the organic species in the solution occurs.

Removal Efficiency

Matrix

Pollutant

Anode Material

Synthetic dye wastewater Synthetic dye wastewater Synthetic water

Color Color

Dimensionally stable anode (DSA) DSA

100%

e

IrO2eTa2O5

e

Synthetic water

e

IrO2/RuO2

e

Synthetic dye wastewater

Color

100%

Synthetic dye wastewater Olive oil wastewater

Color

Ti anode covered by Ta, Pt, and Ir thin film Nb/D and Pt/Ti

Synthetic water

Escherichia coli

Synthetic water

Rotenone (COD removal) 4-Nitrophenol (TOC removal) Fluoroquinolone enrofloxacin (TOC removal)

Synthetic water Synthetic water

Phenol, color

Ti anode covered by Ta, Pt, and Ir thin film Boron-doped diamond (BDD) Pt net Ti/SnO2eSb2O5e IrO2 Pt sheet and BDD thin film

100%

Up to 90% Almost 100%

Under detection limit >97% 74.5% Almost 100%

Current Density (mA/cm2)

Reference

Chlorine/ hypochlorite Chlorine/ hypochlorite Chlorine/ hypochlorite Chlorine/ hypochlorite Chlorine/ hypochlorite

36.1

[54]

14.44e36.10

[55]

15

[56]

30

[57]

5, 10, 14, and 20 A

[58]

Chlorine/ hypochlorite Chlorine/ hypochlorite

12e18 V (anode potential) 5, 7, and 9 V (cell potential)

[59]

Ozone, hydrogen peroxide Hydrogen peroxide (EF) Hydrogen peroxide (EF) Hydrogen peroxide (SPEF)

1.5e13.3

[61]

10e60

[62]

0.80 and 0.10 V/SCE 33

[63]

Main Oxidant

[60]

[64]

324 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Table 5.2 Indirect Electrooxidation of Pollutants.

1. Introduction 325

Effective removal of COD (>96%) was achieved when the electrochemical degradation process was catalyzed by transition metals (Co and Cu) [76] or molybdenum- and phosphate (MoeP)-modified kaolin with graphite as the anode and cathode [77]. Pollutants were adsorbed on the surface of the kaolin where they were oxidized by hydroxyl radicals produced at the graphite cathode by the reaction of hydrogen peroxide and transition metals [76]. This process is similar to the electro-Fenton process. Recently, the potential of conducting diamond films for water treatment has been recognized. They have an inert surface with low adsorption properties, remarkable corrosion stability even in strong acidic media, and an extremely wide potential window in aqueous and nonaqueous media [79,81]. They also have the highest oxygen evolution overpotential value [43,79] meaning that more hydroxyl radicals are formed on the anode surface during treatment. BDD electrodes can also degrade refractory organic pollutants completely, and the nature of the pollutant does not affect the efficiency of the process significantly [80]. It is also known that besides hydroxyl radical formation on the electrode surface, diamond electrodes also increase mediated oxidation by other electrochemically formed compounds such as persulfate, perphosphate, percarbonate, or hypochlorite depending on the electrolyte used. The low pressure conversion of carbon to diamond crystals has made it possible to grow a thin layer of diamond film on suitable substrates such as silicon, niobium, tungsten, molybdenum, and titanium [82]. Hot filament chemical vapor deposition (HFCVD) technique has been applied primarily to fabricate active and stable BDD electrodes, mainly using titanium as a substrate material [82,84,85]. Migliorini et al. [85] used an additional H2 gas flux passing through a bubbler containing a solution of B2O3 dissolved in CH3OH with a B/C ratio of 30,000 ppm during the HFCVD coating of diamond films. Two different fluxes were used to produce heavily BDD films. Fig. 5.2 shows the top view of scanning electron microscopic images of the deposited diamond films. The images present well-facetted microcrystalline diamond surfaces for both coatings and a significant increase in the smallest diamond grain population for E2 because as the boron content increases, the diamond grain size decreases. The main advantages and disadvantages of different electrodes in EO treatment are presented in Table 5.3.

326 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

n FIGURE 5.2 Scanning electron microscopic images of E1 and E2 boron-doped diamond films [85].

Table 5.3 Comparison of Electrodes in EO Treatment. Electrode

Advantages

Disadvantages

Ti Pt

Stable Inert, low oxygen evolution overpotential

Passive, expensive Expensive

PbO2

Good current efficiency, cheap, effective in oxidizing pollutants, high oxygen evolution overpotential, easy to prepare Supports indirect oxidation, good current efficiency, high oxygen evolution overpotential, lower cost, higher availability Inert in tough conditions, high oxygen evolution overpotential and electrochemical stability, good current efficiency, high corrosion stability, good conductivity

Corrosive, toxic Pb2þ ions could be released

DSA electrodes

Boron-doped diamond

Compared Other Electrodes Poor efficiency in anodic oxidation of organic compounds

Short lifetime, lack of electrochemical stability

Very expensive

Higher activity

1.3.4 Treatment of different wastewaters EO treatment has been applied in various wastewaters. The treatment of highly refractory dyes has yielded especially promising results [86e89]. Ma et al. [86] found out that 96.47% of COD could be removed from

1. Introduction 327

methylene blue (MB) containing wastewater assisted by Fe2O3-modified kaolin with graphite plate electrodes. The performances of the TiePt/ b-PbO2 and BDD electrodes were investigated in Reactive Orange 16 dye treatment [87]. Total decolorization was achieved by both electrodes, but BDD electrode was more effective with lower energy consumption. A study by Tsantaki et al. [88] showed that complete decolorization of textile dyehouse effluents was achieved by BDD electrodes in 180 min. Song et al. [89] observed that the decoloring efficiency of the azo dye C.I. Reactive Red 195 increased, whereas the mineralization efficiency decreased with increasing concentrations of NaCl, signifying that oxidized active chlorine at the anode favors the oxidation cleavage of the azo bond. Ramirez et al. [90] degraded methyl orange (MO) azo dye in a recirculation flow plant system. BDD electrodes gave an optimum decolorization efficiency of about 94% with a flow rate of 12 L/min, and response surface methodology was used to describe the EO treatment behavior. Coupling ozone and EO treatment significantly improved COD removal from industrial wastewater [91]. The coupled process was efficient at a relatively low current density so the synergistic effect was easily recognized. Xu et al. [92] and Park et al. [93] presented the innovative approach of combining membrane filtration techniques such as NF and MF with EO treatment. It was observed that the concentration polarization and membrane fouling were effectively restrained by electroosmosis, electrophoresis, and EO treatment [92]. During electrochemical degradation of municipal wastewater, the simultaneous production of hydrogen fuel was observed [93]. MF after EO treatment achieved significant TOC and turbidity removals, with a clear reduction in membrane fouling. EO treatment together with biological oxidation was investigated in individual, combined, and integrated methods [94]. It was observed that the combined process performed rather better than the individual, but took longer. It was concluded that combined processes can be improved by optimizing process parameters and experimental design. A study by Gonçalves et al. [95] presented the positive performance of a two-step process consisting of anaerobic digestion followed by EO in olive mill wastewater treatment. A novel catalyst prepared by Chen et al. [96] showed excellent catalytic activity in the electrochemical treatment of 1-naphthylamine wastewater. Kinetic experiments with EO resulted in the complete removal of iohexol and showed great agreement between the experimental results and the kinetic model [97].

328 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

EO treatment of RO concentrate revealed that in the presence of high chloride ions, concentrations of persistent halogenated by-products will be formed [98]. Another study by Bagastyo et al. [99] showed that effective dissolved organic carbon (DOC) removal could be achieved without chlorinemediated oxidation and by-products. Toxic by-products hinder the use of the EO technique in wastewater treatment, and other water purification techniques are needed for removing by-products before water will be recycled back into the process or released into natural waters. Some comparison in the treatment of different wastewaters and pollutants has been presented in Table 5.4. EO treatment achieved 100% removal of color from olive mill wastewaters by DSA electrodes [100,101]. In both cases, indirect oxidation by active chlorine took place. In addition, bulk electrolysis of wastewater showed that degradation proceeded through partial oxidation reactions to intermediates that are eventually mineralized to carbon dioxide and water [101]. The effect of experimental parameters on oxidation results has been investigated in several papers. Commonly, increasing the pH improves pollutant degradation efficiency [106,108,111]. Contrarily, the opposite has also been observed [112]. Higher current density [105,108,111,112], temperature

Table 5.4 EO Treatment of Different Pollutants. Current Densities Used (mA/cm2)

Reference

Matrix

Pollutant

Electrodes

Removal Efficiency

Synthetic dye wastewater Synthetic dye wastewater

COD

Graphite

96.47%

69.23

[86]

Color

100%

10e70

[87]

Synthetic dye wastewater Synthetic dye wastewater Olive mill wastewater Olive mill wastewater Synthetic wastewater

Color

TiePt/b-PbO2, boron-doped diamond (BDD) BDD

100%

4e50

[88]

Color

Ti/SnO2eSb/PbO2

100%

5e40

[89]

COD, color

Ti/TiRuO2

100%

60

[100]

Color, phenols

Ti/IrO2

100%

50

[101]

Paracetamol

BDD

>98% of TOC decay

33e150

[102]

1. Introduction 329

Table 5.4 EO Treatment of Different Pollutants. continued Removal Efficiency

Current Densities Used (mA/cm2)

Reference

Matrix

Pollutant

Electrodes

Tannery wastewater

COD, ammonia, Cr, sulfides

Satisfactory with all anodes

20 and 40

[103]

Pulp bleaching effluent Paper mill effluent Dye wastewater Synthetic wastewater Coking wastewater Synthetic wastewater Domestic wastewater Synthetic wastewater Synthetic wastewater Landfill leachate Synthetic wastewater Synthetic wastewater Synthetic wastewater Synthetic wastewater Synthetic wastewater Synthetic wastewater Citric acid wastewater Synthetic wastewater Synthetic wastewater Synthetic wastewater

Pentachlorophenol

Ti/PteIr, Ti/PbO2, Ti/PdOeCo3O4, Ti/RhOxeTiO2 Graphite

100%

6

[75]

Organic material: COD

Lead

>96%

2.2e11

[67]

Anthraquinone dye Phenol

BDD Ti/SnO2eSb, Ti/RuO2, Pt BDD

100% 100%

30 20

[83] [104]

Almost 100%

20e60

[105]

Almost 100%

[106]

77%e85%

e1.5 to 1.5 V (anodic potential) 10

Ketoprofen

Ti-based oxide electrode Ta/Ir, Ru/Ir, Pt/Ir, SnO2, PbO2 BDD and Pt

100%

4.4, 8.9, and 13.3

[108]

1,4-Dioxane

BDD

>95%

5, 15, and 25

[109]

Ammonium Surfactants (TOC)

BDD BDD

100% 82%

15e90 4e20

[110] [111]

4,6-Dinitro-o-cresol

BDD

100%

33e150

[112]

Triclosan

BDD

>99%

6e15

[113]

Progesterone

BDD

15e100

[114]

Sulfamethoxazole

BDD

15e100

[115]

Chlorpyrifos

BDD

Almost complete To below 0.1 mg/L 100%

15 and 30

[116]

Organic pollutants

Ti/RuO2eIrO2

Almost 100%

[117]

Chloroxylenol

Pt, BDD

100% (with BDD)

9V (cell potential) 33, 100, 150

Diclofenac

Pt, BDD

100%

150

[119]

Atrazine

BDD

Up to 94%

100

[120]

Organic pollutants (TOC) 2,4-Dichlorophenol Sulfide

[107]

[118]

330 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

[105,109,112], and initial pollutant concentration also increase the removal rates [111,112]. Based on the results in Table 5.4, it can be concluded that BDD electrodes are highly efficient for the removal of different organic pollutants. Degradation current efficiencies can vary significantly, however, and it is important to achieve the required removal efficiency by adjusting the applied current density together with the removal time and energy consumption of the process [109]. One novel approach is to use solar photo-assisted EO treatment for organic pollutant degradation [121,122]. It allows more rapid mineralization of carboxylic acids [121], and photovoltaic solar EO is a self-sustaining wastewater treatment process [122]. The main benefits can be summarized as follows: there is no need for energy storage systems and sun energy can be directly supplied to the treatment process [122]. Comparison of different AOPs (such as EO, ozonation, and Fenton oxidation) shows that all technologies can reduce the organic content of wastewaters but with different performances [123,124]. EO was the most efficient in mineralization of enrofloxacin but not as effective as ozonation in COD removal [123]. Removal of organic material also seems to depend greatly on the addition of an electrolyte salt in EO treatment [124]. The differences in purification efficiencies can be explained mainly in terms of the contribution of hydroxyl radicals and other oxidation mechanisms involved in each technology.

1.3.5 Disinfection of wastewater and drinking water Electrochemical oxidation has also been used in disinfection of drinking water and various wastewaters. Jeong et al. [125] applied EO to disinfect Escherichia coli in drinking water. They found that the main mechanism was inactivation of bacteria by hydroxyl radicals produced by water discharge, also involving direct oxidation at the electrode surface. ROS can also cause effective inactivation [61,126e128]. It was observed that different electrodes have different abilities to produce radicals, e.g., BDD electrodes prefer to form hydroxyl radicals and DSA electrodes activate chlorine, depending on the electrolyte solution used [127]. In addition, inactivation of E. coli at BDD and Pt electrodes was mainly achieved by the reaction of hydroxyl radical and the direct electron transfer reaction, respectively. Ma et al. [128] noticed that a hemin/graphite electrode was preferable to produce ROS compounds, such as H2O2 and *OH, at the cathode surface when applying low potentials without any addition of chloride. A sterilizing rate as high as 99.9% could be obtained after 60 min of

1. Introduction 331

inactivation. Fang et al. [129] and Drees et al. [130] also observed effective bacteriophage MS2 inactivation in drinking water. Yet the inactivation rate for bacteriophage MS2 was much lower than for E. coli demonstrating that bacteria are more sensitive to electrochemical inactivation than bacteriophages [130]. Electrochemically generated oxidants were a major cause of inactivation within the electrochemical cells. EO has also been applied in Legionella bacteria disinfection of germinated brown rice circulating waters and cooling tower waters [131]. Disinfection was attributed to the synergistic effects of the oxide anode, the electric field, and the radicals formed during the treatment. This observation strongly suggests that electrochemical oxidation could be applicable to the disinfection of waters from other sources. The technique has shown its potential also in the treatment of municipal wastewaters [132] and for disinfection in seawater desalination systems where biofouling of the desalination plant membranes can be prevented without using chlorination [133]. Total removal of coliform bacteria was achieved with very different raw wastewaters and the main disinfectants produced depended on the applied current density, the concentration of chlorides, and the concentration of nonoxidized nitrogen in the electrolyte [132]. Although EO has shown promising results in wastewater purification from organic pollutants and disinfection of drinking water, it has its drawbacks. For example, some disinfection by-products, such as perchlorates or bromates, can be produced during the treatment. Oh et al. [134] observed that this phenomenon occurred during the desalination treatment of seawater, and bromate concentrations were some orders of magnitude higher than the USEPA regulation. It was strongly indicated that the application of electrochemical treatment to seawater desalination cannot be recommended without the control of bromate by-product formation. It is also possible to find nitrite, ammonia, and monochloramine residues in drinking water disinfection depending on the treatment conditions and original water quality parameters [135]. EO has been applied also to ballast water treatment in the disinfection of Artemia salina [136] and for algae removal using Chlorella vulgaris as a model organism [137]. A current density of 135 mA/cm2 and a treatment time of around 1 min could achieve 100% mortality of A. salina, the main oxidant being chlorine together with direct oxidation at the anode surface [136]. Total inactivation of C. vulgaris was achieved by EO when 100 mg/L of chloride was present in the solution, so the main mechanism killing the algae was long-life oxidants electrogenerated at the anode surface [137].

332 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Table 5.5 EO Treatment in Disinfection of Different Waters. Removal Efficiency

Current Densities Used (mA/cm2)

Reference

90% 5 log reduction

0e100 24 and 216

[125] [126]

2.4 log reduction (BDD) Almost 100%

17e167

[127]

0.6 V versus SCE (anodic potential) 0.20 A

[128]

Even 4 log reduction (P. aeruginosa)

5e350 mA

[130]

Ti/RuO2

100%

1.0 and 1.5 kV (cell potential)

[131]

E. coli Bacillus sp. E. coli

BDD BDD

100% 100% 100%

1.3e13 110 2.1

[132] [133] [138]

Coliform bacteria Coliform bacteria E. coli K-12 E. coli K-12

Ti/RuO2/TiO2

99.9%

0e30

[139]

BDD

4 log reduction

2.5e120

[140]

BDD

100%

5e40

[141]

BDD

Almost 100%

20

[142]

Matrix

Microbe

Anode

Synthetic water Tap water

Escherichia coli E. coli

Synthetic water

E. coli

Synthetic water

E. coli

Pt Boron-doped diamond (BDD) BDD, Ti/RuO2, Ti/ IrO2, Ti/PteIrO2, Pt Graphite felt

Synthetic water

Bacteriophage MS2

Synthetic water

E. coli, Pseudomonas aeruginosa, bacteriophages MS2 and PRD1 Legionella

Germinated brown rice circulating water and cooling tower water Municipal wastewater Seawater Tertiary treated wastewater Saline secondary effluent Biologically treated wastewater Synthetic water Synthetic Water

Ti pellet with a thin layer of IrO2eSb2O5eSnO2 Pt-tipped copper wire

8 log reduction

[129]

Some studies of electrochemical disinfection of different microbes are presented in Table 5.5.

2.

OBJECTIVES OF THE STUDY

The overall objective of this study was to investigate the inactivation of biofilm-forming bacteria present in a paper mill environment by EO. The main goal was to discover electrochemical behavior and the efficiency of

3. Materials and Methods 333

different electrode materials during the treatment. Another focus of the research was the applicability of EO to sulfide and organic material removal in real paper mill wastewaters. The enhancement of treatment together with biocides (such as hydrogen peroxide) was a further research interest. The specific aims of the study were as follows: 1. Investigating the electrochemical behavior of some electrode materials in paper mill water and their inactivation efficiency against primary biofilm-forming bacteria D. geothermalis, Meiothermus silvanus, and Pseudoxanthomonas taiwanensis and discovering the main inactivation mechanisms. Studying the influence of parameters, such as current density and initial pH or chloride concentration of paper mill water on the inactivation efficiencies (Papers I and II). 2. Studying the efficiency of the EO treatment in sulfide removal from paper mill wastewaters and inactivation of sulfide-forming anaerobic bacteria (Paper III). 3. Studying the electrochemical behavior of electrodes combined with biocides in paper mill water and discovering the oxidation and radical formation mechanisms. Finding the synergistic effects when combining biocides with polarization treatment (Paper IV). 4. Investigating EO in the degradation of MO dye in synthetic wastewater and studying effect of key operative parameters on degradation efficiency. Discovering the oxidation mechanism (Paper V). 5. Discovering the purification efficiency of EO compared with some other common physicochemical treatment methods used in pulp and paper mill wastewater treatment (Paper VI).

3. MATERIALS AND METHODS 3.1 Reactors and electrodes The electrochemical treatment system used for EO of D. geothermalis bacteria is shown in Fig. 5.1 of Paper I. For the experiments in Papers II, III, IV, and V, a sterile beaker was used as a reaction chamber. Volumes, electrode materials, and current densities used for the experiments in this study are presented in Table 5.6. SCE was selected as the reference electrode. Fig. 5.3 shows the experimental apparatus of the electrochemical system in Paper VI. The reactor was equipped with circulation. Two types of electrodes were tested as anodes and cathodes; BDD electrodes and mixed metal oxide (MMO) electrodes. The surface area of the BDD anode was 644 cm2, and the current density used was 4.7 mA/cm2. The surface area of the MMO

334 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Table 5.6 Experimental Parameters. Reaction Chamber and Volume

Cathode Material

Current Densities Used (mA/cm2)

Pollutant

Anode Material PbO2, mixed metal oxide (MMO), and boron-doped diamond (BDD) MMO

Pt wire

25e75

Stainless steel rod

5e65

MMO

MMO

14.3e42.9

MMO

e

Stainless steel plate BDD, MMO

30 and 50

Paper I

75 mL

Deinococcus geothermalis E50051

Paper II

250 mL

Paper III

500 mL

Paper IV

500 mL

D. geothermalis E50051, Pseudoxanthomonas taiwanensis JN1109, Meiothermus silvanus B-R2A5-50.4 Sulfide, anaerobic bacteria e

Paper V

200 mL

Methyl orange dye

MMO, stainless steel 2343 MMO, BDD

Paper VI

3L

COD, resin acids

BDD, MMO

n FIGURE 5.3 Experimental system for pulp and paper mill effluent treatment in Paper VI.

4.7 and 9.5

3. Materials and Methods 335

anode was 315 cm2 and the current density used was 9.5 mA/cm2. Stream velocity was adjusted to 1 L/min, and the volume of the water sample was 3 liters. The recycling time was 60 min, and the samples were taken at variable time intervals.

3.2 Chemicals 3.2.1 Wastewaters used for the experiments Synthetic paper mill waters (SPW) were used for the experiments in Papers I, II, and IV. The composition of the SPW is presented and was developed by Peltola et al. [13]. More detailed compositions are represented in published Papers I, II, and IV. In Paper III, real paper mill wastewater was used (Table 5.7). MO was used as a pollutant for the experiments in Paper V. Finally, in Paper VI, wastewater samples (Table 5.8) were obtained from three Finnish pulp and paper mills which produce peroxide bleached mechanical pulp. These mills are hence called mill A, mill B, and mill C. Samples were collected after the primary clarifier and biological process.

3.2.2 Biocides The following biocides were tested in combination with polarization in Paper IV: hydrogen peroxide p.a. (Merck); peracetic acid, 15% equilibrium solution (Kemira Chemicals, Oulu, Finland); formic acid, 85% (Kemira Chemicals, Oulu, Finland); sodium percarbonate, ECOX (Kemira Chemicals, Helsingborg, Sweden); Omacide IPBC 100, 3-iodo-2propynyl-n-butylcarbamate (Arch Chemicals, UK); and Fennosan GL10, 50% glutaraldehyde; Fennosan M9, 9% methylene bisthiocyanate (MBT); Fennocide BIT20, 20% benzisothiazolinone (BIT); and Fennodispo 315, a naphthalene sulfonate containing anionic dispersant, all from Kemira Chemicals, Vaasa, Finland.

Table 5.7 Paper Mill Wastewater Characteristics. Parameter pH Conductivity (mS/cm) Chloride (mg/L) Sulfate (mg/L) COD (mg O2/L) DOC (mg/L) Redox potential (mV) Dissolved oxygen (mg/L) Sulfide (mg/L)

6.5e7.0 1500e1700 60e115 660 1500 270e350 200e300 0.5e2 4e32

336 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

Table 5.8 Wastewater Characteristics of Finnish Pulp and Paper Mills. Characteristics

pH COD, mgO2/L DOC, mg/L Color Turbidity, FAU Lignin, mg/L Cl, mg/L Abietic acid, mg/L Oleic acid, mg/L b-Sitosterol, mg/L Conductivity, mS/cm

Mill A

Mill B

Mill C

After Primary Clarifier

After Biological Process

After Primary Clarifier

After Biological Process

After Primary Clarifier

After Biological Process

6.7 1550 488 826 74 255 38 12.4 7.7 2.4 1.2

5.9 45 11 e e 16 e nd nd nd e

6.4 1029 390 660 101 257 177 1.1 14.3 1.7 2.0

6.8 343 144 e e 213 e nd 5.1 nd e

7.1 1440 297 1230 78 296 64 6.0 5.3 0.1 e

7.7 140 55 e e 189 e nd nd nd e

Remarks: nd, not detected; e, not analyzed.

3.3 Bacterial strains The bacterial strains used in this study were D. geothermalis E50051, Pseudoxanthomonas taiwanensis JN1109, and Meiothermus silvanus B-R2A550.4. They were selected on the basis of being common primary paper mill bacterial strains [143e145] and were received from the Hambi Collection (Department of Applied Chemistry and Microbiology, Faculty of Agriculture and Forestry, University of Helsinki, Finland) and stored in glycerol freezer stocks (22 C). Fresh bacteria were taken for each of the experiments. All bacteria were inoculated into Petri dishes (R2A-agar) and incubated for 3 days (45 C). After this bacteria were inoculated into five test tubes which had 5 mL of R2 stock solution (composition [per liter]: yeast extract 0.5g, beef extract 0.25g, meet peptone 0.25g, tryptone 0.5g, starch [soluble] 1.0g, K2HPO4 0.3g, MgSO4*7 H2O (dried) 0.05g, Napyruvate 0.3g). Test tubes were incubated for 24 h (45 C, 70 rpm) until cultivated bacteria was ready to use.

3.4 Analyses 3.4.1 Bacteria Amounts of each bacteria species were determined as total aerobic plate counts according to the standard method [146] before and after the EO treatments (Papers I and II). The bacterium was cultured on PCA agar and

3. Materials and Methods 337

incubated at 30 C for 72  3 h. After the incubation period, bacterial colonies were counted, and the results calculated as CFU/mL. In Paper III, the method was similar but for measuring bacteria, an anaerobic atmosphere was created in anaerobic jars (Oxoid Ltd., Hampshire, England) using anaerocults.

3.4.2 Oxidants Measurement of total oxidants (chlorine/hypochlorite, ozone, and hydrogen peroxide) was performed according to the standard method [147].

3.4.3 Cyclic voltammograms In Papers I and IV, the electrochemical behavior of the electrodes was investigated by cyclic voltammetry. In Paper I, cyclic voltammograms (CVs) were performed with stirred solutions at scan rates of 25 mV/min by potentiostat (delivered by Savcor Forest Oy). In Paper IV, a Princeton ParStat 2273-Potentiostat/Galvanostat was used for CVs. Anodic CVs were run by range 0e1.6 V versus SCE (up ¼ from 0 to 1.6 V and down ¼ from 1.6 to 0 V) and cathodic curves by range 01.9 V versus SCE (up ¼ from 0 to 1.9 V and down ¼ 1.9e0 V) at a scan rate of 25 mV/s.

3.4.4 Other analyses The conductivities of the synthetic and real wastewater samples were measured by conductivity meter (VWR EC300, VWR International) and pH values and redox potentials by pH meter using different probes (VWR pH100, VWR International). Dissolved oxygen (DO) measurements were done by DO meter (VWR DO200, VWR International). Sulfide, sulfate, and chloride concentrations were measured directly without filtration by HACH Lange photometer (DR 2800 VIS Spectrophotometer) using Lange Cuvette tests. DOC values of the samples were measured by TOC analyzer (Shimadzu, Model TOC-5000A) and COD values by the standard method [148]. DOC samples were filtrated before analysis through 0.45 mm membranes. Wood extractives were analyzed adapting a method from Örså and Holmbom [149]. Four milliliters of the sample were measured in a screw capped test tube and the pH was adjusted to 3.5 with 0.05 M or 0.5 M H2SO4, depending on initial pH of the sample. Bromocresol green was used as an indicator. Two milliliters of methyl tert-butyl ether (MTBE), containing 20 mL/mL of heneicosanoic acid and botulinum, were also added. The sample was mixed vigorously for 2 min and centrifuged at 4500 rpm for 5 min. A clear organic layer was carefully pipetted off. The extraction was repeated twice with 2 milliliters of MTBE (without ISTD) and mixed for 1 min. The

338 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

combined MTBE layers were evaporated in a nitrogen stream. The residue was silylated adding 80 mL of bis(trimethylsilyl)trifluoroacetamide and 40 mL of trimethylchlorosilane. The silylation was carried out in an oven at 70 C for 20 min. A Hewlett Packard 6890 gas chromatograph (GC) coupled with a 5973 Mass Selective Detector GC-MS equipped with a standard capillary column (25  0.32 mm I.D.  0.17 mm film thickness) containing polydimethylsiloxane was used for measuring resin acids. The initial oven temperature was 120 C, and the temperature was programmed to 190 C at a rate of 10 C/min from which it was raised to 340 C at a rate of 3 C/min. The final temperature was held for 10 min. The inlet temperature was 260 C. Helium was used as a carrier gas. All the samples were added by splitless injection with a sample volume of 1 mL. Identification and quantification were performed using heneicosanoic acid as the internal standard for resin acids. Dye concentration was determined on a UV-visible spectrophotometer at the maximum visible wavelength of 465 nm (PerkinElmer Lambda 45). Wastewater mineralization was monitored by the removal of TOC (Shimadzu, Model TOC-5000A).

4. RESULTS AND DISCUSSION 4.1 Cyclic voltammograms To find out the oxygen evolution overpotentials, CVs were recorded for three different electrodes (Paper I). Fig. 5.2 of Paper I presents CVs recorded in the SPW. According to the CVs, BDD has the highest oxygen evolution overpotential (2.5 V vs. SCE) of the electrodes, suggesting that instead of molecular oxygen, hydroxyl radicals are formed on the surface of the anode. In addition, PbO2 has a higher oxygen evolution overpotential than the MMO electrode which belongs to DSA electrodes. This has been reported also by Chen [43]. In Paper I, CV’s are presented also for three electrodes without chloride salts in SPW. BDD seems to have still the highest oxygen evolution overpotential, but the difference between BDD and PbO2 is not significant. This means that PbO2 can form more hydroxyl radicals on its surface in this SPW than in ordinary SPW-containing chloride salts. To achieve one of the main objectives of this study, electrochemical properties of electrodes in paper mill water were investigated. The main idea was to observe how well electrochemical treatment can enhance the radical formation reactions in paper mill waters where biocides are already present.

4. Results and Discussion 339

If more oxidative radicals can be formed, more efficient biofilm prevention will be also achieved. Combinations of CVs for the MMO and stainless steel SS 2343 electrodes with H2O2 concentrations of 50 mg/L are shown in Figs. 5.7e5.8 (Data from Paper IV). CVs without H2O2 additions showed good reproducibility through the runs. It can be seen that the added biocide has a clear effect on the surface of the SS 2343 electrode by increasing the currents (between 0.7e1.4 V vs. SCE) even before the oxygen evolution reactions occur. This was also seen on the cathodic side when hydrogen peroxide caused a current increase during the experiment on the SS 2343 electrode at 1.0 V versus SCE (reduction of H2O2) which was also observed by Patra and Munichandraiah [150]. Thus, different radical reactions may occur with hydrogen peroxide. On the MMO electrode, the current also increased between 0.2e1.1 V versus SCE. Probably, the radicals were formed at the beginning of the run until formation stopped after 1.1 V versus SCE. A clear current increase was also observed on the cathodic side after starting the run on the MMO electrode, indicating radical formation also on this side. On the cathodic side, curves returning back to initial stage show that currents are smaller on both SS 2343 and MMO electrodes. This means that hydrogen peroxide has been degraded during the run to water. More reactive behavior of SS 2343 electrode must be due to its surface structure. On the cathodic side, it can be also seen that hydrogen peroxide has degraded almost completely because it has a similar curve to the SS 2343 electrode without biocide (Fig. 5.8). It is a clear proof that on the cathodic side, hydrogen peroxide has a capacity to form different radicals in combination with electrical treatment until degraded to water. The proposed radical production mechanism could be as follows (Reactions 5.12 and 5.13):

H2O2 / *OH þ *OH

(5.12)

*OH þ H2O2 / HO2* þ H2O

(5.13)

Patra and Munichandraiah [150] suggested that following reactions will occur in direct reduction of H2O2 in a slightly acidic medium (pH ¼ 5.8) (Reactions 5.14e5.18):

H2O2(bulk) / H2O2(surface)

(5.14)

H2O2(surface) þ e / OHad þ OH

(5.15)

340 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

OH þ Hþ / H2O

(5.16)

OHad þ e / OH

(5.17)

OHad þ Hþ þ e / H2O

(5.18)

Reaction 5.14 followed by electron transfer steps (Reactions 5.15 and 5.17) is expected to result in the CV peak current [150] which we observed at 1.0 V versus SCE (Fig. 5.8). It is also known that hydrogen peroxide is relatively stable at pH < 9 [151] which proves that H2O2 was not degraded by itself. Many other biocides were also tested in this study, but they did not give any response (oxidation or reduction peaks) for radical formation with electrical treatment (Paper IV). Peltola et al. [13] showed that removal of D. geothermalis biofilm was enhanced by cathodically weighted pulsed polarization in the presence of oxidizing biocides. ROS compounds were successfully generated during the experiments for biofilm removal from stainless steel surfaces which supports also results of this study.

4.2 Electrochemical inactivation of bacteria 4.2.1 Aerobic bacteria in synthetic paper mill water Fig. 5.4 shows the electrochemical inactivation of D. geothermalis in SPW as the current density varied from 5 to 65 mA/cm2 (Paper II). It can be seen

5 mA/cm2

0 15 mA/cm2

-1 Log (N/N0)

25 mA/cm2

-2 35 mA/cm2

-3 50 mA/cm2

-4 65 mA/cm2

-5 -6 0

2

4

6

8 10 12 Time (min)

14

16

18

n FIGURE 5.4 Inactivation of Deinococcus geothermalis in SPW using a mixed metal oxide electrode and

different current densities during galvanostatic electrolysis (pH ¼ 7). Data from Paper II.

4. Results and Discussion 341

that the effective inactivation (>2 log) of bacteria was reached when current density was higher than 25 mA/cm2, and the time taken was at least 3 min. The inactivation also increased with higher current density which has also been observed in other studies [152,153]. Differences in inactivation in the range of 25e65 mA/cm2 were minor, signifying that above the threshold value (25 mA/cm2) most of the electric energy used in the experiments was spent to form oxygen. In addition, it can be seen from Fig. 5.5 that different oxidants were formed during the EO. With higher current density, it is possible to form more oxidants. The amounts of oxidants produced are in accordance with the inactivation rate. In 3 min, it was possible to reach a sufficient inactivation level, which means that the concentration of oxidants needed is about 3 mg/L. The amount of oxidant was smaller at a current density of 65 mA/cm2 than at 50 mA/cm2. This can be explained by the higher oxygen evolution reaction and hydroxyl radical production at this current density.

Concentration of oxidants (mg/l)

Fig. 5.6 indicates that the pH of SPW did not have a significant influence on the inactivation efficiency of D. geothermalis. Almost equal amounts of total oxidants were formed during the experiments (Fig. 5.7). However, less oxidants were produced at lower pHs. Similar inactivation efficiency can be explained by higher oxidation potential of hypochlorous acid formed (Reaction 5.6) at acidic pH. Hypochlorous acid has higher oxidation potential than hypochlorite present at neutral or alkaline pH (Reaction 5.7) [55]. It was also observed that generation of chlorine is more or less same under the fixed current density [54]. However, in this study also other oxidants than chlorine/hypochlorite could be formed by higher pH.

14

5 mA/cm2

12

15 mA/cm2

10

25 mA/cm2

8 35 mA/cm2

6 50 mA/cm2

4

65 mA/cm2

2 0 0

2

4

6

8

10

12

14

16

18

Time (min)

n FIGURE 5.5 Amounts of oxidants electrochemically generated on a mixed metal oxide electrode

during galvanostatic electrolysis using different current densities at pH ¼ 7. Data from Paper II.

Log (N/N0)

342 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

0

pH 5

-1

pH 6

-2

pH 7

-3

pH 8

-4

pH 9

-5 -6 0

2

4

6 8 10 Time (min)

12

14

16

n FIGURE 5.6 Inactivation of Deinococcus geothermalis in SPW using a mixed metal oxide electrode at

Concentration of oxidants (mg/l)

different pH values during galvanostatic electrolysis (current density 50 mA/cm2). Data from Paper II.

14 12

pH 5

10

pH 6

8

pH 7

6

pH 8

4

pH 9

2 0 0

2

4

6

8

10

12

14

16

18

Time (min)

n FIGURE 5.7 Amounts of electrochemically generated oxidants on mixed metal oxide electrode during

the galvanostatic electrolysis using different pH values (current density 50 mA/cm2). Data from Paper II.

The effect of different chloride concentrations on inactivation efficiency is shown in Fig. 5.8. It is evident that electrochemically generated chlorine/hypochlorite has a significant influence on the inactivation of D. geothermalis (indirect oxidation). This can also be seen from Fig. 5.9. A six log inactivation could be reached in 5 min when the chloride concentration was 130 mg/ L in SPW. However, most likely anodic oxidation by direct electron transfer reaction occurred simultaneously, as well as oxidation by ROS generated from water discharge [125]. It has been also shown that oxidation may occur by peroxodisulfates generated from sulfate ions [141] in SPW, but they were not measured in this study. Chemical composition of SPW is very heterogeneous mixture including many different anions which can be oxidized to more reactive form to inactivate bacteria.

4. Results and Discussion 343

0

Log (N/N0)

-1 chloride 0 mg/l

-2

chloride 65 mg/l

-3

chloride 130 mg/l

-4 -5 -6 0

2

4

6

8

10

12

14

16

Time (min)

n FIGURE 5.8 Inactivation of Deinococcus geothermalis using a mixed metal oxide electrode and different

Concentration of oxidants (mg/l)

initial chloride concentrations during galvanostatic electrolysis (pH ¼ 7, current density 50 mA/cm2). Data from Paper II.

16 14 12 chloride 0 mg/l

10 8

chloride 65 mg/l

6

chloride 130 mg/l

4 2 0 0

2

4

6

8

10

12

14

16

18

Time (min) n FIGURE 5.9 Electrochemically generated oxidants on a mixed metal oxide electrode during

galvanostatic electrolysis using different initial chloride concentrations (pH ¼ 7, current density 50 mA/ cm2). Data from Paper II.

Because it was obvious that chlorine/hypochlorite played a significant role in inactivation, the effect of residual chlorine on D. geothermalis was investigated. Electrical treatment was switched on for 1 min and then switched off. A bacteria sample was taken immediately after switching on and then every minute. Fig. 5.10 shows that with a chloride concentration of 130 mg/L, it was possible to achieve a reasonable inactivation level in 5 min. Fig. 5.10 shows also inactivation was faster when the initial chloride concentration in SPW was higher. D. geothermalis has been shown to be an efficient primary biofilm former in paper machine water [12,16,17], and it forms thick biofilms on which secondary biofilm bacteria can further attach [16]. It is known to be highly resistant toward radiation and desiccation so results in our study show that

344 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

0

Log (N/N0)

-1 -2

chloride 65 mg/l

-3

chloride 130 mg/l

-4 -5 -6 0

1

2

3

4

5

6

Time (min) n FIGURE 5.10 Inactivation of Deinococcus geothermalis using a mixed metal oxide electrode and

different initial chloride concentrations during galvanostatic electrolysis, effect of residual chlorine/ hypochlorite (pH ¼ 7, current density 50 mA/cm2). Data from Paper II.

EO treatment can produce effective oxidants to inactivate it. In addition, when Deinococcus will be inactivated, it also decreases capability of other potential biofilm formers to attach pipelines because they cannot adhere on the surfaces without Deinococcus. Three different paper mill bacteria species (D. geothermalis, P. taiwanensis, and M. silvanus) were compared to discover possible differences in inactivation efficiency. These species have been recognized as pertinent, primary biofilm formers in the wet end of paper machines and they can cause colorful spots to final product [154]. Experiments were conducted as before (each species was treated separately) and responses to all oxidants (SPW) and oxidants without chlorine/hypochlorite (SPW without chloride salts) were measured. The residual effect of chlorine/hypochlorite on the inactivation efficiency of these three bacteria species was also investigated. Fig. 5.11 shows how electrochemical inactivation affected these three bacteria species. The inactivation order was M. silvanus > P. taiwanensis > D. geothermalis. M. silvanus could be inactivated quite effectively (1.5 log) in 1 min. The resistance of other bacteria to oxidants was somewhat higher. The same experiments were conducted without chloride salts in SPW (chloride concentration ¼ 0 mg/L) to compare their influence. Fig. 5.12 indicates that the inactivation order was different to that with chlorine/hypochlorite. Thus, bacteria have different abilities to withstand different oxidants. It was also observed that D. geothermalis was more sensitive to other oxidants than chlorine/hypochlorite. On the other hand, P. taiwanensis was more resistant to these oxidants than to chlorine/hypochlorite. In addition,

4. Results and Discussion 345

0

Log (N/N0)

-1

Deinococcus geothermalis

-2 -3

Pseudoxanthomonas taiwanensis

-4

Meiothermus silvanus

-5 -6 0

2

4

6

8 10 12 Time (min)

14

16

18

n FIGURE 5.11 Inactivation of Deinococcus geothermalis, Pseudoxanthomonas taiwanensis, and

Meiothermus silvanus using a mixed metal oxide electrode during galvanostatic electrolysis (chloride concentration 65 mg/L, pH ¼ 7, current density 50 mA/cm2). Data from Paper II. 0

Log (N/N0)

-1

Deinococcus geothermalis

-2 -3

Pseudoxanthomonas taiwanensis

-4

Meiothermus silvanus

-5 -6 0

2

4

6

8

10

12

14

16

18

Time (min)

n FIGURE 5.12 Inactivation of Deinococcus geothermalis, Pseudoxanthomonas taiwanensis, and

Meiothermus silvanus using a mixed metal oxide electrode during galvanostatic electrolysis without chloride salts (chloride concentration 0 mg/L, pH ¼ 7, current density 50 mA/cm2). Data from Paper II.

M. silvanus was inactivated with a slower response. Li et al. [142] observed that oxidants with high oxidationereduction potential, such as hydroxyl radical, will damage cell structure more due to strong oxidation ability. For weaker oxidants, such as chlorine, reactions with the cell wall are quite limited, and then there is little cell surface deformation. Oxidation with the enzymes in the cell plasma might be the lethal reason. As a conclusion, sensitivity of different bacteria species to different oxidants varies a lot. Fig. 5.13 shows the influence of residual chlorine/hypochlorite on the inactivation efficiency of bacteria species. As expected, the inactivation order was the same as in Fig. 5.11. An efficient inactivation result was achieved with P. taiwanensis and M. silvanus bacteria in 3 min. D. geothermalis

346 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

0

Log (N/N0)

-1

Deinococcus geothermalis

-2 Pseudoxanthomonas taiwanensis

-3

Meiothermus silvanus

-4 -5 -6 0

1

2

3 4 Time (min)

5

6

n FIGURE 5.13 Inactivation of Deinococcus geothermalis, Pseudoxanthomonas taiwanensis, and

Meiothermus silvanus using a mixed metal oxide electrode during galvanostatic electrolysis, effect of residual chlorine/hypochlorite (chloride concentration 65 mg/L, pH ¼ 7, current density 50 mA/cm2). Data from Paper II.

bacteria were more persistent against residual chlorine/hypochlorite. Thus, the effective oxidation noted during the experiments was mainly due to indirect electrochemical oxidation through chlorine/hypochlorite produced on the anode. This mechanism has also been reported in many other studies [127,129,132,140,155e157]. The main achievement of the current study is to show that residual disinfection efficiency by chlorine/hypochlorite can keep the circulating waters clean enough to avoid biofilm formation on the pipeline surfaces. We did not measure in this study hydroxyl radicals which have also important role in disinfection [125,152]. In general, inactivation mechanisms in EO are very complex and depend on many factors, such as electrolyte composition, bacteria species, electrodes, and operating conditions during the treatment. For example, in disinfecting germinated brown rice circulating water and cooling tower water containing Legionella bacteria, it was concluded that electrochemical disinfection was due to synergistic effect of the oxide anode, pulsed electric field, and the hydroxyl radicals formed during the electrochemical treatment [131]. Li et al. [139] did similar observation in electrochemical disinfection of saline wastewater effluent.

4.2.2 Anaerobic bacteria in paper mill wastewater The reduction of sulfate to sulfide by anaerobic bacteria is a serious problem for pulp and paper mills, so it was worth of investigating the inactivation efficiency of EO against them (Paper III). Real paper mill wastewater was used for the experiments. Fig. 5.14 shows that inactivation was less effective with chloride concentration (in this case 62 mg/L) originally

4. Results and Discussion 347

Amount of bacteria (%)

120 100 80

chloride 62 mg/l

60

chloride 164 mg/l

40

chloride 281 mg/l

20 0 0

1

2

3 4 Time (min)

5

6

n FIGURE 5.14 Inactivation efficiency of anaerobic bacteria during experiments with different initial

chloride concentrations in wastewater (current density 42.9 mA/cm2). Data from Paper III.

present in the wastewater. Thus, it was justifiable to increase the initial chloride concentration of the wastewater by adding NaCl to the wastewater. With chloride concentrations of 164 and 281 mg/L, it was possible to inactivate anaerobic bacteria effectively in 5 min, so electrochemically produced chlorine species accelerated inactivation. Yet many studies have shown that electrochemical disinfection can be effective against bacteria and viruses without the generation of chlorine species [125,128,130,141,152,158]. Jeong et al. [125] observed that in chloride-free phosphate buffer solution E. coli inactivation occurred in two distinct stages. The first step was direct anodic oxidation at anode’s surface and another step was oxidation by hydroxyl radicals generated from water discharge. It is also evident that oxygen evolution reaction killed part of the bacteria in this study. Li et al. [141] also showed in their study that oxidants produced in the electrolysis of SO2 4 (such as S2O2 8 ) improved the disinfection process. This phenomenon cannot be ignored in this study either because paper mill wastewater had high concentration of sulfate as well as SPW. Chlorine containing organic and inorganic by-product formation is also expected [134] during electrochemical treatment, but investigation into this was beyond the scope of the present study. However, applying lower current densities could help to avoid formation of hazardous compounds [132].

4.3 Electrochemical oxidation of sulfide Fig. 5.15 shows the change in sulfide concentration in paper mill wastewater by EO. It was clearly seen that the oxidation of sulfide occurred in all cases. It was possible to achieve an almost 100% reduction with all the different initial sulfide concentrations. Yet at a lower initial concentration (4.0 mg/ L), the adequate level of sulfide in wastewater was achieved with a smaller electric charge.

Sulphide remaining (%)

348 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

120 100

Sulfide 4.0 mg/l

80

Sulfide 7.3 mg/l

60

Sulfide 15.6 mg/l

40

Sulfide 28.4 mg/l

20 0 0

250 500 750 1000 1250 1500 1750 2000 Electric charge (C/L)

n FIGURE 5.15 Electrochemical oxidation of sulfide in pulp and paper mill wastewater using a mixed

metal oxide electrode and different initial sulfide concentrations (current density 42.9 mA/cm2). Data from Paper III.

The effect of current density on sulfide removal efficiency is shown in Fig. 5.16. The initial sulfide concentration was ca. 20 mg/L in all the experiments. The best electric charge efficiency was achieved with the lowest current density. Other current densities were not as effective, and it was noted that the energy consumption of the effective treatment was higher. Other experiments of this study showed that pH did not change much during the experiments and the redox potential as well as DO values raised which proves that oxidation of sulfide was mainly occurred by electrochemically generated oxygen.

Sulphide remaining (%)

In earlier studies, EO has been used for the oxidation of sulfide-containing waters [159,160]. It was found that sulfides were oxidized to elemental sulfur and sulfate during electrochemical treatment [159].

120 100

14.3 mA/cm2

80

21.5 mA/cm2

60

35.7 mA/cm2

40

42.9 mA/cm2

20 0 0

250

500

750 1000 1250 1500 1750 2000

Electric charge (C/L)

n FIGURE 5.16 Electrochemical oxidation of sulfide in pulp and paper mill wastewater using a mixed

metal oxide electrode and different current densities with initial sulfide concentration of ca. 20 mg/L. Data from Paper III.

4. Results and Discussion 349

4.4 Electrochemical oxidation of organics in pulp and paper mill bleaching effluent In paper VI, the MMO and BDD electrodes were employed in electrochemical oxidation of the pulp and paper mill bleaching effluents. The method was rather effective in resin acid degradation in the treatment of the mill B effluent by MMO electrode. At a constant pH of 7.0 and 60 min of treatment time, the removal of abietic acid, b-sitosterol, and oleic acid were 51%, 83%, and 76%, respectively. About 28% of COD could be removed. Compared with other techniques used in the study (such as chemical precipitation), removal rates were rather low. However, in this study, continuous circulation of the wastewater was used through the experiments. Comparable study has shown better purification results for COD [161], but it was done in batch mode. This study showed also that MMO electrode was more effective in degradation efficiency of organic material than BDD electrode most likely due to active chlorine produced during the treatment. Jeong et al. [127] also observed that active chlorine was produced more on DSA electrodes than on BDD electrode. Yet electrochemical oxidation seems capable of removing wood extractives which are hazardous in a biological wastewater treatment process. If the method is used as a pretreatment, potential inhibition of the biological treatment can be eliminated due to reduced toxicity in the primary effluent.

4.5 Electrochemical degradation of methyl orange dye Textile industry dye was selected as a refractory organic compound for further analysis to see effectiveness of MMO and BDD electrodes. MO was selected as the model azoic dye because it is persistent and highly soluble in water; therefore, its removal is also a subject of major importance in environmental protection. It was necessary to gain insight on the electrochemical mineralization of MO in relatively high initial concentrations on BDD and MMO electrodes. It was observed that MO dye degradation was more effective on a BDD electrode than on an MMO electrode (Paper V). For instance, at a current density of 50 mA/cm2, the color was almost completely removed on BDD after 90 min of treatment, while for the MMO electrode, the removal ratio was less than 15%. It was noticed that the BDD electrode has higher onset potential for oxygen evolution than the MMO electrode which indicates that the former has a much higher current efficiency.

350 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

The degradation on two electrodes presented different trends on the operative parameters. High current density enhanced the decolorization on both electrodes, but the promotion was not as significant on MMO as on BDD, leading to a sharp increase in specific energy consumption. The decolorization of MO was more successful under acidic conditions for both electrodes, but the pH dependence was not as obvious on BDD as on MMO. The presence of NaCl favored the indirect oxidation of active chlorine on the MMO electrode, which greatly improved the decolorization rate. The presence of NaCl promoted the decolorization also on BDD electrode, as also combustion rate of dye on both electrodes. The formation of active chlorines seemed to be more efficient on the MMO electrode. Oxygen evolution potential on BDD is much higher than on MMO electrode, which prevents the side reactions and greatly improves the current efficiency for hydroxyl radical formation [43,162]. High initial concentration enhanced the general current efficiency, though the COD and TOC removal efficiency was reduced. In general, EO treatment of MO dye was more effective by BDD electrode in tested conditions. BDD performed better in relative wide concentration ranges, and effect of pH was not so important for decolorization efficiency of dye. Moreover, BDD demonstrated more economical way for dye mineralization from the technical point of view, even absence of NaCl.

5.

CONCLUSIONS AND FURTHER RESEARCH

Electrochemical oxidation was an effective method for inactivation of different biofilm and sulfide-forming bacteria species in pulp and paper mill circulating waters and wastewaters. Paper mill bacteria (D. geothermalis, P. taiwanensis, and M. silvanus) were inactivated effectively (>2 log) at the MMO electrode with a current density of 50 mA/cm2 and contact time of 3 min, and the oxidation was mainly due to indirect electrochemical oxidation (electrochemical formation of chlorine/hypochlorite). Increasing current density and initial chloride concentration of SPW speeded up inactivation. The initial pH value of the SPW did not have a significant impact on the inactivation rate. Bacteria species varied in response to different oxidants. Optimizing the operative parameters is important in finding the best current efficiency for inactivation. Electrochemical oxidation showed promising performance in oxidizing the sulfide present in paper mill wastewater. Inactivation of the anaerobic bacteria present in the wastewater was also observed. This supports strongly

5. Conclusions and Further Research 351

use of technique in oxidizing sulfide-containing wastewaters or preventing sulfate reduction by anaerobic bacteria. Based on the CV runs, it was observed that hydrogen peroxide could be degraded to radicals with cathodic potentials used in this study in the SPW. The stainless steel electrode was more reactive than the MMO electrode on cathodic treatment. More reactive behavior of SS 2343 electrode must be due to its surface structure which would offer great opportunity for biofilm prevention on stainless steel pipelines at paper mills. EO treatment of organics in bleaching effluent gave rather good results for resin acid degradation, but COD removal was not as effective most likely due to continuous flow system used in the study. Electrochemical oxidation seemed capable of removing wood extractives which are hazardous in a biological wastewater treatment process. If the method is used as a pretreatment, potential inhibition of the biological treatment can be eliminated due to reduced toxicity in the primary effluent. EO treatment of MO dye (color removal) was more effective by BDD electrode in tested conditions than by MMO electrode. It was observed that the BDD electrode had higher onset potential for oxygen evolution than the MMO electrode which indicated that the former had a much higher current efficiency for dye degradation. The efficiency of the EO process largely depended on cell configuration, electrode material, electrolyte composition, the microorganism or pollutant, and other experimental parameters, such as current density or the temperature of the treated water. Effective EO treatment of primary biofilm forming bacteria species will offer an alternative to biocides. In addition, combined treatment with hydrogen peroxide will produce powerful oxidants (radicals) which are effective but still environmentally friendly in paper mill environment. In future studies, it would be important to measure different radical reactions at the surface of the anode and by-products formed after the treatment. Finding an effective and stable yet economical electrode material would speed up the use of EO technique in wastewater treatment. It would also be important to develop some novel electric power source for the system, such as solar energy. Combining treatment with current tertiary techniques would enhance the purification results. In addition, economical calculations of EO technique in treatment of different pulp and paper mill circulating waters and wastewaters would be needed.

352 CHAPTER 5 Electrooxidation treatment of pulp and paper mill circulating waters and wastewaters

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Index ‘Note: Page numbers followed by “f ” indicate figures and “t” indicate tables.’

A Abietane-type acids, 62 AC. See Alternating current (AC) Acetic acid (CH3COOH), 196 Acid Red B dye (ARB dye), 244e245 Acid treatment, 178 Acoustic cavitation in liquids, 100 Acrylamide (AM), 176 Activated carbon, 272e273 Activated sludge process, 165 Active electrodes, 323 Additives, 249 Adsorbed$OH radicals, 88 Adsorbent, 232e233 Adsorption, 251 destabilization, 9e10 ultrasound-assisted adsorption, 251e252 Advanced oxidation processes (AOPs), 82, 233, 316, 318e319 Aeration, 232e233 Aerobic bacteria in synthetic paper mill water, 340e346 treatment, 317 Air pollution control, ultrasound in, 259 Alternating current (AC), 30e31 Alum, 36 Aluminate, 20e21 Aluminum, 19, 23, 37 electrodes, 45e48 metal salts, 10e11 salts, 175 sulfate, 37 trihydrate, 37 AM. See Acrylamide (AM) Amaranth dye degradation, 113e114 p-Aminophenol (PAP), 251 Ammonium acetate (CH3COONH4), 196 Amplitude unit control, 101 Anaerobic bacteria in paper mill wastewater, 346e347 digestion process, 171 Analysis of variance (ANOVA), 43e44 Analytical methods, 41e43, 195e196 chemical analysis, 41e42

pseudoetotal metal digestion, 195e196 revised BCR sequential extraction, 196 statistics, 197 toxicity analysis, 42e43 Anion concentration, 23, 26e27 ANOVA. See Analysis of variance (ANOVA) AOPs. See Advanced oxidation processes (AOPs) Apparent current efficiency, 90 ARB dye. See Acid Red B dye (ARB dye) Artemia salina, 331 Asymmetric bubble collapse, 99e100, 99f ATAD. See Autothermal aerobic digestion process (ATAD) Automatic overlapping ultrasonic waves, 101 Autothermal aerobic digestion process (ATAD), 171 Auxiliary electrolytes, 40

B Bacillus, 315 Bacteria, 336e337 electrochemical inactivation of, 340e347 Bacterial strains, 336 Bacterial toxicity, 42e43 Bacteriophage MS2, 330e331 Bamboo charcoal, 272e273 BCR. See Community Bureau of Reference (BCR) Belt filter presses (BFPs), 179e181 Bench-scale tests, 275 Benzene, toluene, ethylene, and xylene compounds (BTEX compounds), 262 Benzisothiazolinone (BIT), 335 Benzoquinone reduction, 110e111 BFPs. See Belt filter presses (BFPs) Biocides, 315, 335 Biofouling, 314e315 Biological oxygen demand (BOD), 15, 316e317 Biological stabilization, 165e166 Biological treatment, 231e232

Biological wastewater treatment, 82 Bioremediation, 231e232, 264 Biostabilization, 171 BioTox method, 42e43 BIT. See Benzisothiazolinone (BIT) BOD. See Biological oxygen demand (BOD) BOD5/COD ratio, 319 Boron doping, 95 Boron-doped diamond (BDD) BDD-based anode, 92, 95 electrodes, 330, 333e335 Bridging model, 176f BTEX compounds. See Benzene, toluene, ethylene, and xylene compounds (BTEX compounds) Bubble size diameter, 20 Bulking sludge, 170

C Cadmium selenide graphene, 248 Calcium hydroxide (Ca(OH)2), 191 Capillary suction time (CST), 168 Capillary water, 172 Capping, 232e233 Carbamate pesticides, 241e242 Carbon dioxide (CO2), 82, 209 Carbon tetrachloride, 103 Carbon-based anodes, 93 Carbon-based electrodes, 91e92 Catalyzed ozonation, 319 Cathode hindering of contaminant migration, 212 Cationic polymers, 175e176 Cavitation, 236, 237f bubbles, 98e99, 143 collapse, 101 process, 100, 100te101t threshold, 99 Cell A model, 44e45, 48 Cell C model, 44e45, 48 Cell T model, 44e45, 48 Centrifuge process, 181e182 CF. See Cyclophosphamide (CF) Charge neutralization model, 175f Charge neutralization model, 174e175 Chelating agent, 268

363

364 Index

Chemical oxygen demand (COD), 15, 90, 316e317 Chemical sludge conditioning, 174, 177e178 sludge flocculation, 174e177 Chemicals, 335 analysis, 41e42 bacterial strains, 336 biocides, 335 coagulants, 36e37 coagulation, 4e5, 27e29, 318 extraction, 233 oxidation techniques, 177 precipitation, 15 wastewaters used for experiments, 335 Chlorella vulgaris, 331 Chloride, 29 Chlorinated POPs, 233 Chlorine, 47, 322 Chlorobenzene, 105 4-Chlorophenol (4-CP), 270 CHR. See Chrysene (CHR) Chrysene (CHR), 277 Circulation-enhanced EK, 274 Clay, 264e265, 277e278 Coagulants, 9, 26 chemicals, 15 technologies, 15 Coagulation, ultrasound-assisted, 252e253 Cobalt ions (Co), 250 COD. See Chemical oxygen demand (COD) Colloid stability in aqueous solutions, 8e9 Colloidal particles, interface of, 5e8 Community Bureau of Reference (BCR), 196 Competing anions, 27 Compression of EDL, 9 Conductivity, 48e50 Copper (Cu), 62e65 Copper nitrate (Cu(NO3)2), 39e40 4-CP. See 4-Chlorophenol (4-CP) CST. See Capillary suction time (CST) Current density, 23, 25e26 Cyclic voltammetry (CV), 118, 125e132 characterization of electrodes, 125e126 water and MB oxidation, 126e132 Cyclic voltammograms (CVs), 112, 337e340

Cyclodextrineelectrokinetic process, 267 Cyclophosphamide (CF), 95 Cylindrical electro-dewatering cell, 187

D DADMAC. See Diallyldimethylammonium chloride (DADMAC) DAF. See Dissolved air flotation (DAF) DC. See Direct current (DC) 2,4-DCP. See 2,4-Dichlorophenol (2,4DCP) DC electro-dewatering, 194e195 Debarking effluents, 38e39, 39t DebyeeHückel length, 6e8 Decolorization efficiency of MB, 120 Dehydroabietic acid (DHAA), 39e40 Deinococcus, 315, 343e344 Deinococcus geothermalis, 315, 333, 336, 343e346 Dendrites, 178e179 Design of experiments (DOEs), 38 Destabilization mechanisms of colloids, 9e10 Dewatering process, 172, 179e180 DHAA. See Dehydroabietic acid (DHAA) Diagnostic ultrasound, 234e235 Diallyldimethylammonium chloride (DADMAC), 176 Diamond doping, 95 2,4-Dichlorophenol (2,4-DCP), 244 Diclofenac, 116 Diffuse layer, 174e175 Digested sludge, 171 Dimensionally stable anodes (DSAs), 94 electrodes, 323 1,3-Dinitrobenzene (DNB), 115 2,4-Dinitrotoluene (DNT), 115 Direct current (DC), 30e31, 182, 261 Direct EO, 88 Disinfection of wastewater and drinking water, 330e332 Dissociative adsorption of water, 88 Dissolution reaction, 25e26 Dissolved air flotation (DAF), 20 Dissolved organic carbon (DOC), 42, 328 Dissolved oxygen (DO), 337 DNB. See 1,3-Dinitrobenzene (DNB) DNT. See 2,4-Dinitrotoluene (DNT)

DO. See Dissolved oxygen (DO) DOC. See Dissolved organic carbon (DOC) DOEs. See Design of experiments (DOEs) Dried sludge, 21e22 Drinking water, disinfection of, 330e332 Dry solids (DSs), 165 DSAs. See Dimensionally stable anodes (DSAs) DSs. See Dry solids (DSs) Dual-frequency ultrasonication, 242e243 Dual-pulse system, 247 Duration time, 284e285 Dyes, 83e84, 104 degradation, 112e114, 245

E Earthworms, 231e232 EC. See Electrocoagulation (EC) ECGO. See ElectroChemical GeoOxidation (ECGO) Economic analysis of electro-dewatering process, 190 ECRTs. See Electrochemical remediation technologies (ECRTs) EDL. See Electrical double layer (EDL) EDTA. See Ethylenediaminetetraacetic acid (EDTA) EDX analyses, 123 EF. See Electroflotation (EF) Effluents, 38e39 EK. See Electrokinetics (EK) EK-Fenton process. See Electro-Fenton process (EK-Fenton process) Electrical discharge, 101 Electrical double layer (EDL), 6, 7f Electrical field effect on electro-dewatering, 199e201 Electro-bioremediation, 266e269 Electro-dewatering process, 167. See also Sewage sludge electrodewatering current density during, 197e199 development, 186e187 electrical field effect, 199e201 electrokinetic phenomena, 182e186 migration of organic and inorganic compounds during heavy metals migration, 212e214

Index 365

macrometal migration, 209e212 organic matter migration, 206e207 soluble ions migration, 207e209 polymer and freeze/thaw conditioning effect, 203e206 setup and operation, 187e188 of sewage sludge, 188e190 sludge type effect on electro-dewatering (II), 201e203 Electro-Fenton process (EK-Fenton process), 87, 269e270 ElectroBioFence, 266 Electrochemical inactivation of bacteria, 340e347 aerobic bacteria in synthetic paper mill water, 340e346 anaerobic bacteria in paper mill wastewater, 346e347 reactions, 21 redox barriers, 273 soil remediation. See Electrokinetics (EK) techniques, 319e320 Electrochemical cleaning, 131 Electrochemical degradation, 133e135 FA degradation, 134e135 MB degradation, 133 of methyl orange dye, 349e350 ElectroChemical GeoOxidation (ECGO), 273 IP, 273e274 Electrochemical oxidation (EO), 82 index, 90 of organic compounds, 86e95 electrode materials, 91e95, 96te97t history of use, 90e91 theory and mechanism, 87e90 of organics in pulp and paper mill bleaching effluent, 349 of sulfide, 347e348 in water and wastewater treatment, 319e332 disinfection of wastewater and drinking water, 330e332 electrodes, 323e325 indirect electrooxidation of pollutants, 324t theory of electrooxidation, 321e323 treatment of different wastewaters, 326e330 Electrochemical remediation technologies (ECRTs), 273

Electrocoagulation (EC), 5, 27e29, 32te35t, 319e320 change of pH and conductivity, 48e50 concepts and theory of coagulation and flocculation, 5e16 destabilization mechanisms of colloids, 9e10 interface of colloidal particles, 5e8 metal salt coagulants, 10e16 stability of colloids in aqueous solutions, 8e9 dissolving of electrodes, 45e48 materials and methods, 38e44 analytical methods, 41e43 electrolysis parameters, 41t initial parameters of wastewaters and debarking effluents, 39t initial quality of surface waters, 39t statistical methods, 43e44 water samples and chemicals, 38e40 water treatment procedure, 40e41 practical considerations, 29e37 applications, 31 constructions, 29e31 economical and ecological considerations, 36e37 quality of statistical models, 44e45 removal percentages of resin acids and copper, 62t surface water treatment, 38, 50e57 theory, 16e29 main reactions, 17e19 side reactions, 19e21 sludge properties, 21e22 treatment parameters, 22e27 treatment, 253 wastewater treatment, 38, 57e65 and water treatment technologies, 31e36 Electrodes, 323e325, 333e335 comparison in EO treatment, 326t dissolving, 45e48 aluminum electrodes, 45e48 iron electrodes, 48 electrolysis reactions at, 185e186 materials, 22e24 in EO, 91e95, 96te97t physicochemical and electrochemical characterization, 118 potential, 23, 25e26 Electroflotation (EF), 20, 316, 319e320 Electrokinetics (EK), 260, 266, 279

circulation-enhanced, 274 electro-bioremediation, 264e266 Fenton process, 269e270, 320 with flushing agent enhancement, 266e269 impacts on soil health, 263e264 IP ECGO, 273e274 nonuniform EK and rotational mode, 275e276 oxidation-enhanced, 269e270 periodic electric potential application, 275 and permeable reactive barriers, 271e273 phenomena, 182e186 electrolysis reactions at electrodes, 185e186 electromigration, 184e185 electroosmosis, 182e184 electrophoresis, 184 principles, 260e264 remediation, 234 factors affecting EK, 262e263 of organic contamination, 260e276 UESR, 274e275 and ultrasonic treatment, 285e288 current progress and electroosmotic flow, 285e286 pH distribution, 286e287 POPs removal, 287e288 ultrasound-assisted EK remediation, 276 Electrolysis reactions at electrodes, 185e186 Electrolytic discharge of water, 88 Electromigration, 184e185, 260e261 Electroosmosis, 182e184, 260e261, 267, 272 Electroosmotic flow, 182e183, 285e286 rate, 261 Electroosmotic phenomenon, 167 Electrooxidation (EO), 319e321 theory of, 321e323 treatment of different pollutants, 328te329t in disinfection of different waters, 332t Electrophoresis, 184, 260 Electroreclamation. See Electrokinetics (EK)

366 Index

Electroremediation. See Electrokinetics (EK) Encapsulation, 232e233 Endocrine-disrupting compounds, 241e242 Energy consumption estimation requiring for oxidation processes, 143e144 efficiency control and kinetics, 121 Enmeshment, 10, 51e52 Environmental analysis, ultrasound in, 259e260 Environmental applications of ultrasound, 234e260 EO. See Electrochemical oxidation (EO); Electrooxidation (EO) EO/US degradation. See Sonoelectrochemical degradation (EO/US degradation) Epichlorohydrin/dimethylamine (Epi/ DMA), 176 EPSs. See Extracellular polymeric substances (EPSs) Escherichia coli, 254, 330e331 Ethylenediaminetetraacetic acid (EDTA), 244 Ex situ methods, 233e234 Excess sludge. See Secondary sludge Extracellular polymeric substances (EPSs), 170, 173e174

Flotation, 316 Fluoranthene (FLU), 277 Flushing agent enhancement, EK with, 266e269 Food to microorganism ratio (F/M ratio), 170 Formic acid (FA), 83e84 decarboxylation, 106 degradation in electrochemical degradation, 134e135 in sonochemical degradation, 136e138 in sonoelectrochemical degradation, 139e143 Free water, 172 Free-living bacteria, 316 Freeze-thaw conditioning, 178 effect on electro-dewatering, 203e206 Full factorial design, 43

F

Harmful substances, 165e166 HCB. See Hexachlorobenzene (HCB) Heat value of sludge, 22 Heating effect in ultrasonic treatment, 284e285 Heavy metals migration, 212e214 ultrasound-assisted heavy metals removal, 257 Hemin/graphite electrode, 330e331 Herbicides, 104, 258 Heterogeneous catalysts, 248 Hewlett Packard 6890 gas chromatograph, 338 Hexachlorobenzene (HCB), 270, 320 HFCVD. See Hot filament chemical vapor deposition (HFCVD) Higher frequency ultrasonication, 242 Hot filament chemical vapor deposition (HFCVD), 325 Hot spot concept, 236e237

F/M ratio. See Food to microorganism ratio (F/M ratio) FA. See Formic acid (FA) Faraday’s law, 18e19, 25e26, 45e46 Fenitrothion, 245e246 Fenton oxidation, 177e178, 233 Fenton process, 243e244 Fenton-like catalysts, 258 Fenton-like’s reagents, 245e246 Fenton’s reagent, 177, 244e246 Ferric chloride, 37 Ferric salts, 175 Ferrous and ferric salts, 318e319 Ferrous iron (FeII), 177 Ferrous sulfate (FeSO4), 191 Ferrous sulfide (FeS), 16 Finnish paper mill, 38 Fixed layer, 174e175 Flocculation, 4e5, 9 sludge, 174e177

G Galvanostatic test, 45e46 Gas chromatograph (GC), 338 GouyeChapman model, 6e8 Graphite, 186 Green rust, 19 Green technology, 320

H

hydrophobic gaseous nucleus, 101 model of cavitation, 101, 102f theory, 101 HPCD. See Hydroxypropyl cyclodextrin (HPCD) Human hearing limit, 234 Humic acid, 12e15, 15f Humic substances, 12e15 Humus coal, 12e15 Hydrochloric acid (HCl), 196 Hydrogen (H2), 19, 89e90 Hydrogen peroxide (H2O2), 103e104, 177, 196, 318e319, 322e323, 340 H2O2/Fenton/Fenton-like catalysts, 243e245 H2O2/ferrous ions, 177 Hydrogen sulfide (H2S), 16, 61 Hydrophilic compounds, degradation of, 105 Hydrophilic fraction of NOM, 12 Hydrophobic compounds, 266e267 degradation of, 105 Hydrophobic NOM, 12e15 Hydroxide removal, 31 Hydroxyl radicals, 88e89 Hydroxylamine hydrochloride (NH2OHeHCl), 196 Hydroxylation, 101 Hydroxypropyl cyclodextrin (HPCD), 267 Hypochlorite, 88, 322

I Ibuprofen, 116 IC. See Inorganic carbon (IC) ICAP. See Inductively coupled plasma optical emission spectrometry (ICAP) ICE. See Instantaneous current efficiency (ICE) ICP spectrometer. See Inductively coupled plasma spectrometer (ICP spectrometer) Ifosfamide (IF), 95 Igepal CA-720, 270 In situ generation of chlorine and hypochlorite, 114 In situ methods, 233e234 Incineration, 165e166 Indirect electrooxidation of pollutants, 324t

Index 367

Induced polarization (IP), 273 ECGO, 273e274 Inductively coupled plasma optical emission spectrometry (ICAP), 195e196 Inductively coupled plasma spectrometer (ICP spectrometer), 196 Industrial chemicals, 104 Industrial raw water treatment, metal salt coagulants in, 12e16 Inert electrodes, 24 Inorganic carbon (IC), 195 Inorganic compounds, 177 migration, 206e214 Instantaneous current efficiency (ICE), 90 Integrated sonoelectrochemical-Fenton processes, 247 Interparticle bridging, 9e10 Interrupted DC electro-dewatering, 187 voltage, 194e195 Interstitial water, 172 Intracellular water, 172 Investment costs, 36 IP. See Induced polarization (IP) IPA. See Isopimaric acid (IPA) Iridium oxide electrodes, 91e92 Iron, 16, 37, 243 electrodes, 23, 48 metal salts, 10e11 salts, 16 Isomorphous substitution, 5e6 Isopimaric acid (IPA), 39e40

K Kaolin, 277e278, 278t preparation, 279

L Lake water, 39e40 Land farming, 231e232 Landfilling of sludge, 165e166 Lasagna process, 272e273 Lead, 323 Lead dioxide, 323 Legionella, 331, 346 Linuron, 245e246 Liquid sample analysis, 120 Living organisms, 231e232 Londonevan der Waals forces of attraction, 8

Low-frequency ultrasonication, 242 Low-frequency ultrasound, 247, 253, 255

M Macrometal migration, 209e212 MaxwelleBoltzmann distribution, 6e8 MB. See Methylene Blue (MB) Mechanical cleaning, 316 Mechanical dewatering processes, 167, 179e182, 180t. See also Electro-dewatering process BFPs, 180e181 centrifuge process, 181e182 pressure filter press, 181 Mechanical vacuum filter, 179e180 Medium frequency ultrasonication, 242 Meiothermus, 315 Meiothermus silvanus, 333, 336, 344e346 Membrane filtration, ultrasound-assisted, 253 Membrane-based techniques, 317e318 Metal alloys, 93 cations and hydroxides, 4e5 ion removal, 31 oxideecoated titanium, 24 salt coagulants, 10e12, 50 in industrial raw water and wastewater treatment, 12e16 in research studies, 13te14t Methyl orange dye (MO dye) azo dye, 327 electrochemical degradation, 349e350 Methyl tert-butyl ether (MTBE), 337e338 Methylene Blue (MB), 83e84, 242e243, 251 degradation in electrochemical degradation, 133 in sonochemical degradation, 135e136 in sonoelectrochemical degradation, 138e139 Microfiltration (MF), 318 Microtox-Flash assay, 42e43 Microwave digester, 195e196 heating, 260 Mineralization of organic pollutants, 82 Mixed-metal oxide (MMO), 86e87

catalysts, 94 electrodes, 333e335 MLR. See Multiple linear regression (MLR) MMO. See Mixed-metal oxide (MMO) MODDE software, 43e44 Monopolar configuration, 30 MTBE. See Methyl tert-butyl ether (MTBE) Multiple linear regression (MLR), 43e44 Multistage electrode dewatering method, 188 Multivalent counterions, 8e9

N Nanofiltration (NF), 318 Nanoparticles of iron species, 244 Natural organic matter (NOM), 4e5, 12, 319 Natural waters, 5 Nernst potential, 6e8 Neutral amorphous metal hydroxides, 10e11 Neutral substances, 242 NF. See Nanofiltration (NF) Nitroaromatics, 115 4-Nitrophenol, 242 NMR. See Nuclear magnetic resonance (NMR) Noble metals, 186, 233 NOM. See Natural organic matter (NOM) “Nonactive” electrodes, 323 Nonionic hydroxides, 25 Nonpurgeable organic carbon (NPOC), 120 Nonuniform EK and rotational mode, 275e276 Novel Ti/SnO2eSb2O3/PTFE-La-Ce-bPbO2 anodes, 114 NPOC. See Nonpurgeable organic carbon (NPOC) Nuclear magnetic resonance (NMR), 171e172 Nutrients, 230

O OER. See Oxygen evolution reaction (OER) $OH radicals, 100

368 Index

Operating costs, 36e37 Organic compounds, 269 degradation, 104 migration, 206e214 contamination electrokinetic remediation, 260e276 soil remediation, 231e234, 231f material removal, 31 matter removal, 50e56, 58e60 polymers, 176 substances, 173 Organophosphorous pesticides, 241e242 ORP. See Oxidationereduction potential (ORP) Oxidants, 321, 337 chemicals, 170 Oxidation fraction, 196 Oxidation-enhanced electrokinetics, 269e270 Oxidationereduction potential (ORP), 17e18 Oxidative degradation, 236e237 Oxygen (O2), 47, 89e90, 177 Oxygen evolution reaction (OER), 88, 95 Ozonation, 233, 250e251 Ozone (O3), 177, 318e319, 322e323

P PACl. See Polyaluminium chloride (PACl) PAHs. See Polycyclic aromatic hydrocarbons (PAHs) PAP. See p-Aminophenol (PAP) Paper mill. See also Pulp and paper mill microbe communities, 315 wastewater anaerobic bacteria, 346e347 characteristics, 335, 335t Partial least squares regression (PLS), 43e44 Passivation, 83 Patented sonoxide ultrasonic water treatment technology, 254 PbO2 anodes, 90e91, 93 PCA. See Principal component analysis (PCA) PCBs. See Polychlorinated biphenyls (PCBs) PCE. See Perchloroethylene (PCE)

PCP. See Pentachlorophenol (PCP) PEIs. See Polyethylene imines (PEIs) Pentachlorophenol (PCP), 267 Perchloroethylene (PCE), 262 Periodate ions, 103 Periodic electric potential application, 275 Permeable reactive barriers (PRB), 271 EK and, 271e273 electrochemical redox barriers, 273 Lasagna process, 272e273 ZVMs PRB, 271e272 Persistent organic pollutants (POPs), 230e231 effectiveness of ultrasonic treatment, 276 electrokinetic and ultrasonic treatment, 285e288 electrokinetic remediation of organic contamination, 260e276 environmental applications of ultrasound, 234e260 materials equipment, 278e279 model POPs and clay, 277e278, 277t methodology experiments, 279e282, 280te281t, 282fe283f extraction and analysis, 282, 283f kaolin preparation, 279 ultrasonic treatment, 282e285 Persulfate (PS), 249e250 ions, 103 Pesticides, 104, 258 pH, 21 change of, 48e50 of solution, 23e25 water, 40 Pharmaceuticals, 104 removal of, 115e116 Phenanthrene (PHE), 277 Phenol decomposition, 105 Phenolic compound removal, 114e115 Phenolic pollutants, 243 Phenoxy radical, 105 Phosphorus, 15 removal from wastewater, 60 Physical conditioning method, 179 Physical sludge conditioning, 174, 178e179 Physicochemical treatment, 232

Phytoextraction, 231e232 Phytoremediation, 231e232 Plasma theory, 101 Platinum electrodes (Pt electrodes), 91e92, 321 Plexiglass cylinders, 278e279 PLS. See Partial least squares regression (PLS) Polarization, 83 Polish wood rosin, 39e40, 40t Pollutant concentration, 23 Polyaluminium chloride (PACl), 12, 36, 40, 191 Polychlorinated biphenyls (PCBs), 241e242 Polycyclic aromatic hydrocarbons (PAHs), 230, 277 Polyelectrolyte conditioning, 189 Polyelectrolytes, 176e177 Polyethylene imines (PEIs), 176 Polymers, 176, 191 conditioning effect on electro-dewatering, 203e206 of sludge, 182 polymeric flocculants, 318 Polynuclear species, 12 POPs. See Persistent organic pollutants (POPs) Potassium hydroxide (KOH), 196 Potentiostatic control, 110e111 Power amplifier, 119e120 power-modulated pulsed ultrasound, 243 ultrasound, 234e235 Praestol 855BS, 193 PRB. See Permeable reactive barriers (PRB) Precipitation, 10, 51e52 Pressure filter press, 181 Pressure-driven electro-dewatering, 202e203, 206e207 Primary sludge, 169e170 Principal component analysis (PCA), 197 Proteins, 173 PS. See Persulfate (PS) Pseudoetotal metal digestion, 195e196 Pseudoxanthomonas, 315 Pseudoxanthomonas taiwanensis, 333, 336, 344e346

Index 369

Pulp and paper mill. See also Paper mill bleaching effluent, 349 circulating waters and wastewaters, 314e315 experimental system for, 334f microorganisms in paper mill environment, 314e315 Purification of surface waters, 31 Pyrolysis, 101, 233

Q Quinones, 109e110

R Radical scavengers, 249 Radioactive wastewater treatment, ultrasound-assisted, 254e255 Reactive oxygen species (ROS), 322e323 Reactive red 195, 114 Reactors, 333e335 Rectified diffusion, 98 Repulsion force, 8 Residual metals, 56e57 Resin acids, 62e65, 63f Reverse osmosis (RO), 318 Revised BCR sequential extraction (V), 196 Rhizoremediation, 231e232 Rhodamine B dye, 246 Rhodotorula mucilaginosa, 252 RO. See Reverse osmosis (RO) ROS. See Reactive oxygen species (ROS) Ruthenium oxide electrodes, 91e92

S Scanning electron microscopy (SEM), 118 analyses, 123 SchulzeeHardy rule, 8e9 SDBS. See Sodium dodecylbenzene sulfonate (SDBS) Secondary sludge, 170 Sediment, ultrasound in, 256e258 SEM. See Scanning electron microscopy (SEM) Settleability of sludge, 22 Sewage sludge electro-dewatering, 173. See also Electro-dewatering process

annual sewage sludge production, 166t electro-dewatering process, 182e190 material and methods analytical methods, 195e196 experimental setup and procedure, 193e195 sludge conditioning, 191e193 sludge samples, 191 mechanical dewatering processes, 179e182 results current density during electro-dewatering, 197e199 electrical field effect on electro-dewatering, 199e201 migration of organic and inorganic compounds, 206e214 polymer and freeze/thaw conditioning effect on electro-dewatering, 203e206 sludge type effect on electro-dewatering, 201e203 sludge characterization affecting dewaterability, 169e174 sludge conditioning, 174e179 sludge dewaterability evaluation, 168e169 Sirius luminometer, 42e43 Sludge characterization affecting dewaterability, 169e174 EPSs, 173e174 water distribution in sludge, 171e173 conditioning, 165e166, 174e179, 191e193, 192t chemical, 174e178 physical, 178e179 dewaterability evaluation, 168e169 properties, 21e22 type, 169e171 digested sludge, 171 primary sludge, 169e170 secondary sludge, 170 ultrasound in sludge stabilization, 255e256 Sodium chloride (NaCl), 40 Sodium dodecylbenzene sulfonate (SDBS), 248 Sodium sulfate (Na2SO4), 40 Soil, 230 contamination, 230

EK impacts on soil health, 263e264 remediation of organic contamination, 231e234, 231f ultrasound in, 256e258 types, 264e265 vapor extraction, 232e233 washing, 232e233 Solar photo-assisted EO treatment for organic pollutant degradation, 330 Solid bowl centrifuge, 182 Solid retention time (SRT), 170 Solidification, 233 Soluble ions migration, 207e209 Solvents, 104 Sono-activated persulfate process, 249e250 Sono-Fentonelike process, 258 Sonocatalysis, 250 Sonocatalytic degradation of ARB dye, 244e245 Sonocatalytic/sono-assisted oxidation, 248e251 Sonochemical degradation, 135e138 FA degradation, 136e138 MB degradation, 135e136 Sonochemical reactions, 236 Sonochemical remediation, 104 Sonochemistry, 237 Sonoelectrochemical degradation (EO/ US degradation), 82e83, 138e143 efficiency in degradation of different organic compounds, 112te113t experimental setup, 119e120, 119f FA degradation, 139e143 MB degradation, 138e139 reactor types, 110e112 Sonoelectrochemical oxidation of organic compounds, 84e116 advantages and disadvantages of methods for small molecule removal, 85t EO of organic compounds, 86e95 sonochemical destruction methods of organic pollutants, 98e116 sonochemical degradation of organic compounds, 103e107 theory and mechanism, 98e103 Sonoelectrochemical remediation, 247e248

370 Index

Sonoluminescence, 100, 236 Sonolysis, 245 of organic compounds, 282 Sonophoto-Fenton system, 246e247 Sonophotocatalysis, 245e247 Sonophotolysis, 245e247 Sorptive barriers, 271 Specific resistance to filtration (SRF), 168 Specific UV absorbance (SUVA), 42 Spherical shock waves, 98 SPW. See Synthetic paper mill waters (SPW) SRB. See Sulfate-reducing bacteria (SRB) SRF. See Specific resistance to filtration (SRF) SRT. See Solid retention time (SRT) Standard thermodynamic potential of cell, 90 Standing wave formation, 106 Static equilibrium of bubbles in water, 98 Statistical methods, 43e44 Statistical models, quality of, 44e45 Statistics, 197 Sulfate, 29 Sulfate-reducing bacteria (SRB), 38e39 Sulfide, electrochemical oxidation of, 347e348 Sulfide precipitation, 61 Sulfuric acid, 37 Surface water, 172 EC in treatment, 50e57 organic matter removal, 50e56 residual metals and turbidity, 56e57 treatment by EC, 38 Surfactant-aided electrokinetics, 268 Surfactants, 104 SUVA. See Specific UV absorbance (SUVA) Synthetic paper mill waters (SPW), 335 aerobic bacteria in, 340e346 Synthetic polyelectrolytes, 318 Synthetic resin solutions, 39e40

T Tannins, 64e65 TCE. See Total current efficiency (TCE); Trichloroethylene (TCE) Temperature, 23, 27 Tetrachloride, 105

Textile industry dye, 349 Therapeutic ultrasound, 234e235 Thermal decomposition mechanism, 105 Thermal treatment, 233 Ti/PbO2 anodes, 90e91 Ti/Pt anodes, 90e91 Ti/Pt grid, 109e110 Ti/SnO2 anodes, 90e91 Ti/SnO2eSb2O5, 92, 95 Ti/Ta2O5eSnO2, 92, 117 electrode preparation, 117e118 Time to filter (TTF), 168 Titanium (Ti) substrate, 94 TOC. See Total organic carbon (TOC) TOCCATA catalyst, 319 Total current efficiency (TCE), 90 Total organic carbon (TOC), 90e91, 120, 195, 318 Toxic substances, 42e43 Toxic-chlorinated intermediates, 88 Toxicity analysis, 42e43 Toxicity removal, 62e65 Traditional homogenous Fenton’s reagent, 246e247 Transducer, 119e120 Transition metal salts, 177 Treatment time, 23, 25e26 Trichloroethylene (TCE), 241e242 Trichloroethylene, 105 2,4,6-Trichlorophenol (2,4,6-TCP), 270 TTF. See Time to filter (TTF) Turbidity, 56e57

U UESR. See Upward electrokinetic soil remediation (UESR) Ultrafiltration (UF), 318 Ultrasonic cavitation, 236 irradiation, 234 sinus wave generator, 119e120 surface cleaning, 254e255 treatment, 282e285 duration time and heating effect, 284e285 electrokinetic and, 285e288 POPs removal, 285 effect of ultrasonic power, 284 effect of water content, 283e284 ultrasonically enhanced electrokinetics, 320

Ultrasonication (US), 83, 243e245, 252, 279 efficiency in degradation of organic compounds, 107, 107te108t Ultrasound (US), 98, 234, 250 in air pollution control, 259 effect, 103e104 in environmental analysis, 259e260 environmental applications, 234e260, 238te241t sound frequency classification, 235f theoretical background, 235e237 in sediment and soil remediation, 256e258 ultrasound-assisted advanced oxidation of organic pollutants, 258 ultrasound-assisted heavy metals removal, 257 ultrasound-assisted organic desorption, 257 ultrasound-assisted organic destruction, 258 in sludge stabilization, 255e256 technology, 83 in water treatment, 237e255 and recycling, 259 Ultrasound-assisted adsorption, 251e252 advanced oxidation of organic pollutants, 258 biological treatment, 252 EK remediation, 276 membrane filtration, 253 organic desorption, 257 organic destruction, 258 radioactive wastewater treatment, 254e255 Ultraviolet germicidal irradiation (UV-C irradiation), 254 Ultraviolet radiation (UV radiation), 318e319 Untreated sewage sludge, 165e166 Upward electrokinetic soil remediation (UESR), 274e275 US. See Ultrasonication (US); Ultrasound (US) UV radiation. See Ultraviolet radiation (UV radiation) UV-C irradiation. See Ultraviolet germicidal irradiation (UV-C irradiation)

Index 371

V Vermiremediation, 231e232 Vibrio fischeri, 42e43 Vicinal water, 172 Visible light-responsive catalysts, 246 Volatile organic contamination, 232e233 Volatile solids (VSs), 195

W Washing, 316 Waste treatment and recycling, ultrasound in, 259 Wastewater treatment. See also Water treatment EC in, 38, 57e65 organic matter removal, 58e60 resin acids, copper, and toxicity removal, 62e65 sulfide precipitation, 61 metal salt coagulants in, 12e16 in pulp and paper mills, 315e319 characteristics of wastewater, 316t plant for treatment of paper mill effluent, 317f primary and secondary treatment, 316e317 tertiary treatment, 317e319 technologies, 4

Wastewater treatment plants (WWTPs), 165 Wastewaters, 5, 38e39 characteristics of finnish pulp and paper mills, 335, 336t disinfection of, 330e332 for experiments, 335 Water (H2O), 82 distribution in sludge, 171e173 pH, 40 samples and chemicals, 38e40 sonolysis, 235e236, 242 temperature, 45e46 water-holding, 172 Water treatment. See also Wastewater treatment procedure, 40e41 technologies, 31e36 types, 172 ultrasound in, 237e255 other sonocatalytic/sono-assisted oxidation, 248e251 sonoelectrochemical remediation, 247e248 sonophotolysis and sonophotocatalysis, 245e247 ultrasonication and H2O2/Fenton/ Fenton-like catalysts, 243e245 ultrasound-assisted adsorption, 251e252

ultrasound-assisted biological treatment, 252 ultrasound-assisted coagulation, 252e253 ultrasound-assisted disinfection, 253e254 ultrasound-assisted membrane filtration, 253 ultrasound-assisted radioactive wastewater treatment, 254e255 Wood extractives, 337e338 WWTPs. See Wastewater treatment plants (WWTPs)

X X-ray diffraction spectrometry (XRD), 118 analysis, 121e123

Z Zerovalent aluminum, 248 Zerovalent metals PRB (ZVMs PRB), 271e272 Zeta potential (z potential), 6e8, 174e175 of sludge, 184 of untreated water samples, 55 Zetasizer Nano ZS analyzer, 42