Electrochemical Technology Applied in Treatment of Wastewater and Ground Water [1 ed.] 9781620816820, 9781612097725

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ENVIRONMENTAL REMEDIATION TECHNOLOGIES, REGULATIONS AND SAFETY

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ELECTROCHEMICAL TECHNOLOGY APPLIED IN TREATMENT OF WASTEWATER AND GROUND WATER

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ENVIRONMENTAL REMEDIATION TECHNOLOGIES, REGULATIONS AND SAFETY

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ELECTROCHEMICAL TECHNOLOGY APPLIED IN TREATMENT OF WASTEWATER AND GROUND WATER CHUANPING FENG MIAO LI XU GUO CHAO ZHAO ZHENYA ZHANG AND

NORIO SUGIURA

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Nova Science Publishers, Inc. New York

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Independent verification should be sought for any data, advice or recommendations contained in this book. In addition, no responsibility is assumed by the publisher for any injury and/or damage to persons or property arising from any methods, products, instructions, ideas or otherwise contained in this publication. This publication is designed to provide accurate and authoritative information with regard to the subject matter covered herein. It is sold with the clear understanding that the Publisher is not engaged in rendering legal or any other professional services. If legal or any other expert assistance is required, the services of a competent person should be sought. FROM A DECLARATION OF PARTICIPANTS JOINTLY ADOPTED BY A COMMITTEE OF THE AMERICAN BAR ASSOCIATION AND A COMMITTEE OF PUBLISHERS. Additional color graphics may be available in the e-book version of this book.

LIBRARY OF CONGRESS CATALOGING-IN-PUBLICATION DATA Electrochemical technology applied in treatment of wastewater and ground water / Chuanping Feng ... [et al.]. p. cm. Includes bibliographical references and index.

ISBN:  (eBook)

1. Water--Purification--Technological innovations. 2. Groundwater--Purification--Technological innovations. 3. Water treatment plants--Pilot plants. I. Feng, Chuanping. TD430.E44 2011 628.1'66--dc23 2011035675

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CONTENTS vii 

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Preface Chapter 1

Introduction

1 

Chapter 2

Theories of Electrochemical Treatment

7 

Chapter 3

Electrochemical Treatment of Synthetic Industrial Water Containing Phenol and Ammonia

11 

Electrochemical Reduction of Nitrate Contained in Groundwater

33 

A Pilot Plant of Electrochemical Treatment System (0.3 m3/hr)

39 

Two Types of Practical Electrochemical Treatment Systems (Treatment Capacities of 4m3/hr and 0.5m3/hr)

51 

Chapter 4 Chapter 5 Chapter 6

References

61 

Index

67 

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PREFACE The electrochemical method is an attractive alternative to other methods of wastewater treatment because of its environmental respectability and ease of operation. This study introduces the application of the electrochemical method for treatment of domestic and industrial wastewater. Electrochemical degradation of phenol was evaluated at two typical anodes, Ti/RuO2-Pt and Ti/IrO2-Pt, as a treatment method in toxic aromatic compounds discharged by chemical plants, petroleum refineries and pharmaceutical factories. According to the results, the electrochemical process for effective removal of phenol and ammonium was achieved in the presence of NaCl, in which phenol could be readily mineralized. For electrochemical reduction of nitrate, the Cu-Zn alloy cathode and the Ti/IrO2-Pt anode were used in an undivided cell. The Cu-Zn alloy cathode has a significant capacity of electrochemical reduction of nitrate. Under the conditions of 10mA/cm2, addition of 500 mg/L NaCl and the parallel connection, the nitrate-N decreased from 50.0 mg/L to 9.8 mg/L for the 240-min electrolysis. Neither ammonia-N nor nitrite-N was detected at the end of electrolysis. Moreover, a series of experiments were performed with two types of practical electrochemical treatment systems (treatment capacities of 4m3/hr and 0.5m3/hr) using effluents from an anaerobic digester (EAD) of cattle wastewater, supernatants from primary sedimentation in a sewage plant (SPS) and domestic wastewater to evaluate the systems’ treatment abilities. As a result, for both EAD and SPS, the 4 m3/hr system was found to remove 87 to 91% of T-P, 74 to 96% of T-N, 70 to 94% of NH4-N, 88 to 91% of TOC and 75 to 87% of COD. Similarly, the 0.5 m3/hr system was able to remove 62 to 90% of T-P, 83 to 92% of T-N, 90 to 100% of NH4-N, 75to 83% of TOC and

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80 to 100% of COD. On the other hand, a pilot plant of an electrochemical treatment system (0.3 m3/hr) was successfully developed by treating domestic wastewater, pond water containing algae and wastewater from hog raising, in which electrocogulation and electrooxidation processes were used. In this system, the removal of T-N, T-P, NH4-N and COD from domestic wastewater and pond water containing algae was approximately 90%, while the removal of chlorophyll-a (chl-a) of algae was approximately 100%.

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Chapter 1

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INTRODUCTION With a rapidly growing economy and burgeoning populations, the world’s scarce water resources, such as rivers and lakes, are seriously affected by pollution from the vast discharges of industrial and domestic wastewater, indiscriminate solid waste disposal and runoff from an agricultural sector characterized by excessive use of fertilizer and pesticides and large-scale livestock breeding. Therefore, these wastewaters must be treated before being discharged. The increase in nitrogen and phosphorus from various sources of wastewater has resulted in serious eutrophication of natural bodies of water. High concentration of nitrate in potable water can cause several health problems such as “blue baby syndrome” in infants, liver damage and cancer. The maximum contaminant level (MCL) in wastewaters and potable water is 45 mg/L in the United States, while in the European Union it is 50 mg/L and 15 mg/L for infants (EEC, 1980; WHO, 1993). Aromatic compounds are common pollutants in the waste effluent discharged by many factories, such as chemical plants, petroleum refineries and pesticide and pharmaceutical factories. Of all the organic pollutants causing problems around the world, aromatic hydrocarbons and chlorinated hydrocarbons are among the most harmful (Comninellis et al., 1995; Drever et al., 1997). Phenolic products are toxic to humans and aquatic organisms and they are listed among the most common and serious environmental contaminants. A concentration greater than 2 mg/L phenol is toxic to fish, and concentrations between 10 and 100 mg/L result in death of aquatic life within 96 h (Gattrell et al., 1990). According to the 80/778/EEC Directive, the maximum admissible concentration (MAC) of phenols in drinking water should not exceed 0.5 pg/L. Many industrial processes produce toxic wastewaters, which are not easily

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Chuanping Feng, Miao Li, Xu Guo et al.

biodegradable and require costly physical or physico-chemical pretreatment (Pulgarin, 1994). In general, methods for treating wastewaters include biological methods or chemical treatment using ozone, chlorides or coagulants, ion exchange, reverse osmosis and electrochemical treatment. Although the biological treatment of wastewater is important, it cannot completely remove soluble components such as endocrine disrupters and pesticides. A large part of the nitrogen in domestic wastewater and stock-raising wastewater is found as ammonia, which is generally treated by biological nitrification-denitrification. However, the nitrification process requires a larger treatment system and longer treatment time than electrochemical treatment, resulting in higher treatment cost. Alternatively, chemical treatment involves the addition of high amounts of chemicals, resulting in unreacted chemicals being discharged in the treated wastewater (Mohan et al., 2001). In addition, biological and chemical methods may generate a considerable amount of sludge which itself requires treatment. Ion exchange and reverse osmosis processes (Samatya et al., 2006) cannot transfer nitrate into harmless compounds but only concentrate nitrate from water to brine, requiring further treatment. As environmental regulations become stringent, new and novel processes for efficient treatment of various kinds of wastewater at relatively low operating cost are needed. In recent years, electrochemical treatment has gained increasing interest because of its high treatment efficiency, ease of operation, small plant space requirement and relatively low investment costs. The processes can completely convert organic pollutants into gases such as N2, CO2, etc. Although the electrochemical method for treating a variety of industrial wastewaters has been extensively researched and reported (Sucre et al., 1981; Chin et al., 1985; Lin et al., 1996; Israilides et al., 1997; Vlyssides et al., 1998; Grimm et al., 1998), almost all the previous work concerning the electrochemical method was performed on a laboratory scale. There are only a few reports on practical industrial applications. This situation may be due to problems regarding electrode lifetime and energy consumption. On the other hand, a high voltage pulse has a low average current and large instantaneous current. Energy consumption is relatively low because the interval is long. Furthermore, the authors (Feng et al., 2000) have revealed that high voltage pulse can produce various radical species, which are more effective for the removal of NH4-N, and the applied voltage influences the RNO bleaching. The authors found that hexavalent chrominum was removed by an electrochemical process in the laboratory. As a result, the combination of electrochemical treatment and adsorption onto chitosan provides many

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Introduction

3

advantages, i.e., simplicity, low cost, high ion removal efficiency and ease of operation (Chuang et al., 2002). Moreover, Ti/SnO2 (Comninellis et al., 1993) and Ti/PbO2 (Wabner et al., 1985; Polcaro et al., 1999) were effective in removing various organic pollutants from wastewater, and the authors have demonstrated that a Ti/SnO2 anode can efficiently remove ammonia from wastewater (Feng et al., 2000). However, regulations are becoming stricter with respect to the permissible concentrations of metal ions in effluents. In general, noble metal oxides such as RuO2 have better performance than those of the corresponding noble metals, but it is difficult to utilize the dimensionally stable anode (DSA) electrodes coated with oxides in industrial applications because they are expensive and their service lifetimes are short. However, the high operating lifetimes of oxide coatings can be extended by the addition of either conducting oxides such as IrO2 or non-conducting oxides such as TiO2 to RuO2. TiO2 is especially attractive because of its lower cost (Comninellis et al., 1991). Houk et al. (1998) have shown that quaternary metal oxide films applied to Ti or Pt substrates exhibited high and persistent activity for the electrochemical incineration of benzoquinone. Motheo et al. (2000) reported that the Ti/RuO2-TiO2 electrode performs better than Ti/IrO2TiO2 or Ti/IrO2-RuO2-TiO2 electrodes when used in the electrochemical degradation of humic acid. For industrial water, electrochemical reaction can effectively oxidize toxic organics (Kötz et al., 1991; Brillas et al., 2000; Feng et al., 2003; Li et al., 2005). With unique features such as simplicity and robustness in structure and operation, it is possible that the electrochemical process can be developed as a cost-effective technology for the treatment of aromatic pollutants, particularly for low volume applications (Li et al., 2005). Electrochemical wastewater treatment effect mainly depends on the nature of the anodes that are used in the process (Stucki et al., 1991; Feng et al., 2003). The difference in the effectiveness and performance of different anode materials for wastewater treatment demonstrates the complexity of the EC reaction mechanisms involved. The current efficiency of traditional electrodes is very low in organic degradation, such as using graphite and nickel (Rodgers et al., 1999). Dimensionally stable anodes (DSAs), made by the deposition of a thin layer of metal oxides on a base metal, usually titanium, have been proved to be effective in organic degradation (Comninellis et al., 1994; Simod et al., 1997; Feng et al., 2003). Research (Comninellis et al., 1995; Feng et al., 2003; Li et al., 2005) shows that phenol is easy to be oxidized into maleic acid with RuO2 and Pt as anode, respectively. It has been reported that the oxidation of phenol can stop with such products as maleic acid and oxalic acid for Pt anodes

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(Gattrell et al., 1990; Comninellis et al., 1991). It is very difficult for maletic acid degradation on most researched anodes (Koile et al., 1975; Gattrell et al., 1990, 1993; Pulgarin et al., 1994; Schumann et al., 1998; Wang et al., 1998; Iniesta et al., 2001; Bahadir et al., 2003; Feng et al., 2003). However, with sufficient hydroxyl radicals, maleic acid could be oxidized directly to oxalic acid, which can be oxidized readily to CO2 (Feng et al., 2003; Li et al., 2005). Feng et al. (2003) have investigated the performance of Ti/TiO2-RuO2 for removal of ammonia, Tanaka et al. (2003) have demonstrated the effectiveness of Ti/RuO2 and Pt for degradation of orgnics contained in landfill leachate respectively. However, the anodes did not have long service time. With PbO2 electrodes, phenol can be completely removed. Whereas, the actual application of PbO2 electrodes will be limited due to the fact that the possible toxicity of Pb would leach from the working anode. And hence, it remains a subject to develop anodes with long service life and excellent performance for degradation of aromatic organics. In order to increase conductivity and current density, various electrolytes such as Na2SO4, H2SO4 and their mixtures were added to the medium during the electrochemical conversion of phenol (Gattrell et al., 1990; Ko¨ tz et al., 1991; Comninellis et al., 1993). However, very few investigators (Drever et al., 1997) have used NaCl as supporting electrolyte. In the presence of chloride ion during electrolysis, Cl2 gas is discharged on the anode above 2.1872 V potential (Hearst et al., 1976). Following the discharge of Cl2 gas, HOCl formation occurs with hydrolysis reaction (Do, 1996). HOCl is a strong oxidant, which oxidizes the phenol in solution, while the oxidant ability is decreased by formation of OCl- (Drever et al., 1997; Vlyssides et al., 1997). Meanwhile, Hypochlorous acid (HOCl) is replenished in the medium with the electrochemical reaction of NaCl. This represents the indirect oxidation of phenol, however, direct oxidation also occurs on the anode (Lin et al., 1998), most likely with a different mechanism. Phenol is easy to be oxidized into maleic acid with RuO2 and Pt as anode, respectively. The intermediate maleic acid has been proved and it is suggested that maleic acid is reduced to succinic acid at the cathode followed by oxidation to malonic and acetic acid, and then finally to carbon dioxide by anode oxidation (Comninellis et al., 1995; Feng et al., 2003). For electrochemical reduction of nitrate, the reduction of nitrate to the nontoxic nitrogen gas is proved to be difficult since it is one of the eight possible products (Li et al., 1988; Ureta-Zanartu et al., 1997). When the supporting electrolyte is acid nitrite, ammonia, hydroxylamine and hydrazine will be the by-products; furthermore, it was found (Dibyendu et al., 2000) that

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Introduction

5

the main by-products during electrochemical reduction of nitrate in basic or neutral electrolyte were nitrite and ammonia. As ammonia and nitrite, in general, are the main unfavorable reduction products, applications of the electrochemical process for denitrification are limited due to generation of them(Devkota et al., 2000; Lee et al., 2002; Cheng et al., 2005; Mácová et al., 2005; Brylev et al., 2007). However, if possible, the produced nitrite and ammonia, before their diffusion to the bulk, are oxidized to the initial nitrate and nitrogen at the anode, respectively. Recently, some researchers reported that high selectivity of nitrate reduction to gaseous nitrogen compounds was achieved by electrochemical method (Katsounaros et al., 2006; Li et al, 2009). This work deals with the electrochemical oxidation and reduction of contaminated water, which was divided into four parts. Firstly, electrochemical treatment of synthetic industrial water containing phenol and ammonia was studied. Secondly, electrochemical reduction of nitrate contained in groundwater in a laboratory scale was studied. Thirdly, a pilot plant of electrochemical treatment system (0.3 m3/hr) was studied for treatment of domestic wastewater, pond water containing algae and wastewater from hog raising. Finally, two types of practical electrochemical treatment systems (4m3/hr and 0.5m3/hr) were evaluated for treatment of effluents from an anaerobic digester (EAD) of cattle wastewater, supernatants from primary sedimentation in a sewage plant (SPS), and domestic wastewater.

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Chapter 2

THEORIES OF ELECTROCHEMICAL TREATMENT

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2.1. ELECTROCHEMICAL OXIDATION Generally, oxidation of organic matter by electrochemical treatment can be classified as direct oxidation at the surface of the anode and indirect oxidation distant from the anode surface; processes are influenced significantly by the anode material. Recently, oxides anode have been of interest because of higher conductivity and oxidizability. The mechanism of oxidation of organic matter at oxide anode (MOx) has been suggested by Comminellis (Comninellis et al., 1994). Water is electrolyzed by anodic catalysis to produce adsorbed hydroxyl radicals, given as equation (1); H2O + MOx → MOx[·OH] + H+ + e-

(1)

The adsorbed hydroxyl radicals may form chemisorbed active oxygen, as shown in equation (2); MOx[·OH] → MOx+1 + H+ + e-

(2)

Furthermore, another strong oxidant of hypochlorite may be produced in many wastewaters containing chlorides, as given in equation (3) (Gattrell et al., 1990; Szpyrkowicz, 2001); H2O + Cl- → HOCl + H+ + 2e-

(3)

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In addition, high voltage pulse can lead to the formation of a strong electric field and free radicals such as ·OH, ·O, H+ and H2O2, etc. (Sun, 2000); H2O →·OH, ·O, +H, H2O2

(4)

Organic matter (R) included in wastewater are oxidized by hydroxyl radicals (Comninellis et al., 1994), and the reactions are given in equations (5), (6) and (7) (Szpyrkowicz et al., 2001); R + MOx[·OH] → MOx + CO2 + zH+ + ze-

(5)

MOx+1 + R → MOx + RO

(6)

R + HOCl → product +Cl-

(7)

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Oxidation of organic matter depends upon the anode material, concentration of NaCl, and current and voltage applied. Although the influence of NaCl and anode material has been investigated in the electrochemical treatment applied DC power supply, the effect of anode material on pulse treatment has not been studied yet.

2.2. ELECTROCOAGULATION Electrocoagulation also occurs during the electrochemical treatment of wastewater. The electrocoagulation mechanisms have been proposed for the production of Fe(OH)3 or Fe(OH)2 (Yousuf et al., 2001). Mechanism 1 Anode: 4Fe → 4Fe2+ + 8e

(8)

4Fe2+ + 10H2O + O2 → 4Fe(OH)3 + 8H+

(9)

Cathode: 8H+ + 8e → 4H2

(10)

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Theories of Electrochemical Treatment

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Mechanism 2 Anode, Fe → Fe2+ + 2e

(11)

Fe2+ + 2OH- → Fe(OH)2

(12)

Cathode, 2H2O + 2e → H2 + 2OH-

(13)

Removal of various pollutants such as heavy metals (Carl et al., 1995), nitrate (Koparal et al., 2002) by electrocoagulation has been studied. In the present study, phosphorus included in wastewater was removed by the electrocoagulation mentioned above.

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2.3. ELECTROCHEMICAL REDUCTION The major electrochemical reactions involved in the electrochemical nitrate reduction are (Cheng et al., 2005): NO3- + H2O + 2e- = NO2- + 2OH-

(14)

NO3- + 3H2O + 5e- = 1/2N2 + 6OH-

(15)

NO2- + 5H2O + 6e- = NH3 + 7OH-

(16)

NO2- + 4H2O + 4e- = NH2OH + 5OH-

(17)

2NO2- + 4H2O + 6e- = N2 + 8OH-

(18)

2NO2- + 3H2O + 4e- = N2O + 6OH-

(19)

NO2- + H2O + 2e- = NO + 2OH-

(20)

N2O + 5H2O + 4e- = 2NH2OH + 4OH-

(21)

2H2O + 2e- = H2 + 2OH- (side reaction)

(22)

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Chapter 3

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ELECTROCHEMICAL TREATMENT OF SYNTHETIC INDUSTRIAL WATER CONTAINING PHENOL AND AMMONIA The aim of this part of experiments is to find anodes with high performance of electrochemical phenol degradation and long service time. In the present study, phenol, which is the basic unit of aromatic compounds, was used as the model organic for electrochemical degradation. Anodes of Ti coated with RuO2 and doped with Pt (Ti/ RuO2-Pt) and Ti coated with IrO2 and doped with Pt (Ti/IrO2-Pt) were examined for their performance in electrochemical phenol degradation.

3.1. EXPERIMENTAL PROCEDURES A continuous electrochemical cell was designed in our laboratory with a net working volume of 300 mL (Figure 1). A Cole Parmer model peristaltic pump was used to circulate the phenol solution from a 1000-mL beaker to the electrochemical cell at a speed of 100 mL/min. The 300-mL electrolysis cell was made of acryl plates with four outer spots for the electrodes assembled. For each cell, two types of electrodes, Ti/RuO2-Pt, Ti/IrO2-Pt of 75 cm2 (15 cm×5 cm) prepared by TohoTech company(Japan) were used as the anode and Ti plate with the same area as the cathode with a distance of 10 mm between the two electrodes. The emerging areas of the anode and cathode in the treated solution were same, which was 50 cm2. A DC potentiostat (HuanAn YaGuang Electronic Company) with a voltage range of 0–50 V and a current range of 0–

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5 A was employed as power supply for electrochemical organic degradation. A magnetic stirrer was used to homogenize the phenol solution.

Figure 1. Schematic diagram of the electrochemical apparatus.

During experiments, phenol solutions with different initial concentrations of 8, 20, 40 and 80 mg/L were prepared for electrolysis experiments. To investigate the effect of sodium chloride (NaCl) dosage on the phenol degradation, the NaCl of 0.1, 0.3, 0.5 mg/L (w/v) were added into the phenol solutions as the supporting electrolyte, respectively. To investigate the effect of (NH4)2SO4, phenol and (NH4)2SO4 was used to prepare synthetic wastewater with different concentration of both. Electrolysis experiments were performed under galvanostatic control at different current densities of 10, 20 mA/cm2 respectively. 1000 mL of synthetic phenol solution prepared as above was poured into the beaker, the reaction started with the application of specified current density and the solution was mixed at a constant 300 rpm by the magnetic stirrer. At 10 min intervals, 10 mL of samples were drawn from the beaker for phenol and TOC analysis. The electrolysis was ceased when either 99% of initial phenol was converted or 2 h elapsed.

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Electrochemical Treatment of Synthetic Industrial Water …

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3.2. ANALYTICAL METHODS 3.2.1. Detection of Free Radicals Species and Oxidizing Substance

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To measure the production of free radicals and oxidizing substance formed during the electrochemical treatment, NaCl solution containing 50 μmol L-1 RNO (p-Nitrosodimethylaniline) was used because RNO reacts rapidly with hydroxyl radicals selectively. The bleaching of RNO solution by hydroxyl radicals was measured by absorbance changes at 440 nm (Comninellis et al., 1994). RNO solution of 1000 mL was poured in the beaker, and circulated at the flow rate of 100 mL/min. NaCl at different dosage of 0.1, 0.3, 0.5 and 1.0 g was mixed in distilled water as electrolyte, respectively. While electrolysis was performed under galvanostatic control at 0.5 and 1.0 A, respectively, resulting in current densities of 10 and 20 mA/cm2. Samples were taken at intervals of 1 or 5 min and absorbance of RNO solution was measured by 722S spectrophotometer (Shangai Delicacely Scientific Facility Co., Ltd).

3.2.2. Cyclic Voltammetry (CV) In order to investigate the behavior of Ti/RuO2-Pt and Ti/IrO2-Pt anodes during the electrolysis of the phenol solution, the cyclic voltammetry experiments were operated by a computer controlled CS300 electrochemical workstation using a three-electrode cell (Company of Huazhong electronic), Pt is chosen as the counter electrodes, and Hg/HgCl electrode as the reference electrode, and the working anodes are Ti/RuO2-Pt and Ti/IrO2-Pt having a size of 1.0 cm×1.0 cm. The electrolyte of 0.05 M NaCl solution containing 8 mg/L phenol was used. The potential was scanned at the scan rate of 100 mV/s, starting from 0 V, and the scan range was run from -2.5 V to 2.5 V.

3.2.3. Others Concentration of phenol was measured by 4-amido-antipyrine titration, and COD was determined by K2CrO7 oxidation according to Chinese Method of Water and Wastewater measurement (Wei et al., 2002). pH was measured

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by pH S-3Cprecision pH/mV meter. The TOC was measured by a TOC analyzer (1030, Aurora) based on the combustion-infrared method. Surface morphology of cathode was characterized ex situ by atomic force microscopy (Digital Instruments, DimensionTM3000, US).

3.3. RESULTS AND DISCUSSION

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3.3.1. Phenol Electrolysis Performance of Ti/RuO2-Pt and Ti/IrO2-Pt for phenol degradation with different NaCl dosages is shown in Figure 2. It can be seen from Figure 2 that Ti/RuO2-Pt and Ti/IrO2-Pt have similar effect in the electrochemical phenol degradation with different dosages of NaCl at the same current density, i.e., the concentration of phenol decreased with respect to treatment time, and the phenol removal ratios increased with the increasing of NaCl dosages. However, there were a little slower about the phenol degradation rate with Ti/IrO2-Pt anode than with Ti/RuO2-Pt anode. Furthermore, with Na2SO4 as supporting electrolyte, electrochemical degradation of phenol was very slow at both anodes, which was similar with study of Li et al. (2005), who found that with the Ti/RuO2 and Pt anodes only 40% or less of the TOC was removed after a long treatment period. It was suggested that NaCl, which could be oxidized to form a strong oxidant of HOCl, could promote the degradation of phenol. As shown in Figure 2, with 0.1 g/L NaCl as supporting electrolyte, the phenol concentration almost did not decrease in about the initial 5 min, while the phenomenon was not observed with 0.3 g/L and 0.5 g/L NaCl as supporting electrolyte. This is due to the fact that HOCl could not be produced enough with lower NaCl (0.1 g/L) at the beginning. Although complete phenol removal was achieved by all of the anodes, the performance of Ti/RuO2-Pt was better than that of Ti/IrO2-Pt at same experiment condition. For the Ti/RuO2-Pt anode, phenol concentration decreased from around 8 mg/L to zero after 30 min of electrolysis with 0.5 g/L and 0.3 g/L NaCl as supporting electrolyte at the current density of 10mA/cm2 (Figure 2A), while it reached zero after approximately 20 min at 20 mA/cm2 (Figure 2C). On the other hand, the Ti/IrO2-Pt showed slower removal ratio compared to the Ti/RuO2-Pt. At different current density of 10 mA/cm2 and 20 mA/cm2, phenol concentration decreased from around 8 mg/L to zero after 50 min (Figure 2B) and 30 min during electrolysis (Figure 2D), respectively. It was found that the potential increased during the electrolysis, although the

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Electrochemical Treatment of Synthetic Industrial Water …

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potential did not increase significantly. That must due to the fact that the NaCl was consumed to produce HOCl with respect of time, which resulted into the decrease of conductivity. Although both about the same voltage input and the same dosage of NaCl solution for the two types of anodes in the electrolysis cells, different times for complete phenol removal were taken, i.e., less electric energy with the Ti/RuO2-Pt than that with the Ti/IrO2-Pt was consumed to achieve complete phenol removal. Considering the economy factor, it is found that at current density of 10 mA/cm2 with 0.3 g/L NaCl as supporting electrolyte was best for degradation of 8 mg/L phenol contaminated water by Ti/RuO2-Pt anode. Figure 3 shows the variation of TOC during the phenol electrolysis with NaCl as supporting electrolyte. TOCs decreased to zero in the given treatment period for both Ti/RuO2-Pt and Ti/IrO2-Pt, the similar tendency of change in TOC and phenol concentration was observed, meaning that phenol was completely oxidized into CO2 in the end of the electrolysis.

Figure 2. Electrochemical degradation of 8 mg/L phenol in 1000 mL electrolyte as a function of time for different dosage of NaCl as supporting electrolyte. (A) I = 10

mA/cm2, Ti/RuO2-Pt; (B) I = 10 mA/cm2, Ti/IrO2-Pt. (C) I = 20 mA/cm2, Ti/RuO2-Pt; (D) I = 20 mA/cm2, Ti/IrO2-Pt. (---♦---) 0.5 g/L NaCl; (-•-) 0.3 g/L NaCl; (▲ ) 0.1 g/L NaCl; (ÆÅ) 0.5 g/LNa2SO4.

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Figure 3. TOC as a function of time with 0.3 g/L NaCl as supporting electrolyte. (A) Ti/RuO2-Pt; (B) Ti/IrO2-Pt; I = 20 mA/cm2. (▲ ) 8 mg/L; (•) 20 mg/L.

Passivation of Electrodes Since the formation of films on the surface of Ti/IrO2-Pt anode was observed during the electrolysis, the change of performance of anodes with the service time was investigated. Figure 4. shows that the electrochemical oxidation ability of Ti/IrO2-Pt anode dropped quickly after serviced for 10 hours at the current density of 10mA/cm2 with 0.3 g/L NaCl as supporting electrolyte. The phenol concentration decreased from around 8 mg/L to zero in 40 min and 90 min by the first service and the 10 hours service time, respectively. However, no formation of polymeric film and no change of performance on the Ti/RuO2-Pt anode were observed after servicing for 10 hours(data not shown), which demonstrated that the Ti/RuO2-Pt was more suitable for phenol removal than Ti/IrO2-Pt. It was found that some yellow substances were deposit on the Ti/IrO2-Pt anode. Some substances were appeared to be adhered to the surface of the anode after electrolysis (Figure 5). The yellow substances could be partly washed off by dilute H2SO4 solution. After washing the electrode, the electrodes remain yellow-brown. It is can be see from Figure 5 that after using the surface of Ti/IrO2-Pt anode was more rough than that of unused. The washed-off-solution was analyzed by HPLC. Benzoquinone and hydroquinone were found in a higher quantity in the solution (data not shown). It is well known that the ability of phenol to foul electrodes and the tarry deposit forming on electrodes during phenol oxidation is attributed to phenolic polymerization products. The oxidation of phenolic compounds at solid electrodes produces phenoxy radicals, which are responsible for coupling to form a passivating polymeric film on the electrodes (Gattrell et al., 1993).

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Figure 4. Electrochemical degradation of 8 mg/L phenol in 1000 mL electrolyte as a function of time with 0.3 g/L NaCl as supporting electrolyte with different service time. (A) Ti/IrO2-Pt; (B) Ti/RuO2-Pt; I = 10 mA/cm2. (▲) First service, (•) service 10 h later.

Figure 5. AFM photograph of Ti/IrO2-Pt anode (A) unused and (B) used for electrolysis.

According to the cyclic voltammograms (Figure 6), the Ti/IrO2-Pt anode had a lower oxygen evolution potential than the Ti/RuO2-Pt anode, with values of 1.20 and 1.58 V (Vs the standard Hg/HgCl) respectively. The low overpotential of the Ti/IrO2-Pt anode would shorten the lifetime of the hydroxyl radicals, and would hinder the oxygen to transfer from the radicals to organic oxidation which leads to the accumulation of polymeric products in electrochemical oxidation. The organic film formed on the Ti/IrO2-Pt would deactivate the anode oxidizability. Since high oxygen evolution potential can decrease the unwanted power loss to oxygen generation, it will increase the current efficiency for electrochemical degradation. The radical reaction

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forming molecular oxygen was probably restrained because of the high overpotential of the Ti/RuO2-Pt anode, which was favorable to organic oxidation by hydroxyl radicals formed at the electrode surface. It was ascertained from Figure 6 that the Ti/RuO2-Pt anode was more effective than the Ti/IrO2-Pt anode during the phenol electrolysis. Various strategies have been developed to address these surface fouling problems including the use of electrochemical pre-treatment (Koile et al., 1979), laser activation (Poon et al., 1986), chemical or electrochemical treatment and reducing the concentration and lifetime of phenoxy radicals (Gattrell et al., 1990; Houk et al., 1998). The formation rate of tar depends on the concentration of phenoxy radicals, which can be limited by decreasing the concentration of phenol and minimizing the current density. The phenoxy radical lifetime can be decreased by decreasing pH in solution because the phenol oxidation potential decreases with pH while the phenoxy radical oxidation potential remains unchanged (Polcaro et al., 1999). As some yellow substances were deposit on the Ti/IrO2-Pt anode and no substance was deposit on Ti/RuO2-Pt anode, which may be that reason that shapes of voltammetric lines were different between two anodes.

Figure 6. Cyclic voltammograms of the anode materials Ti/RuO2-Pt, and Ti/IrO2-Pt obtained at a scan rate of 100 mV/s in 8 mg/L phenol, 0.05M NaCl, 25oC, pH=7.6 (E vs Hg/HgCl).

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Effect of Initial Concentrations of Phenol on the Degradation Efficiency In order to investigate the treatment efficiency on high concentration of phenol, the experiments of electrochemical degradation of 20, 40 and 80 mg/L phenol solutions were carried out with a selecting current density and NaCl concentration. It is clear from Figure 7 that complete phenol removal was achieved on high concentration of phenol as the electrolysis time was extended. At current density of 10mA/cm2 with 1.0 g/L NaCl as supporting electrolyte, the phenol concentration decreased from around 20, 40 and 80 mg/L to zero after 30 min, 60 min, and 130 min respectively. The electrolysis time for complete removal of phenol was proportional to the concentration of phenol. In conclusion, the Ti/RuO2-Pt anode also performs well for electrochemical degradation of high concentration of phenol solution with appropriate current density and NaCl as supporting electrolyte.

Figure 7. Electrochemical degradation of 20,40 and 80 mg/L phenol in 1000 mL electrolyte as a function of time with 1.0 g/L NaCl as supporting electrolyte. Ti/RuO2Pt, I = 10 mA/cm2. ( ♦ ) 20 mg/L, ( • ) 40 mg/L, ( ▲ ) 80 mg/L.

Change of pH during Phenol Electrolysis Figure 8 shows the change of pH during the electrolysis process at different conditions, where the pHs were not controlled. As shown in Figure 6, the similar trend of the pH change was observed at 0.3 NaCl-8 mg/L phenols, 0.3 NaCl-20 mg/L phenol, i.e., the pH increased at the beginning, and then decreased, finally remained nearly constant. However, the pH reached the maximum value of 9.3 at 0.3 NaCl-8 mg/L phenol in the first 5 min., and then

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the pH began to drop sharply with the electrolysis time. On the other hand, the pH in the higher concentration phenol solution (20 mg/L) increased to 9.0 in the first 5 min., and then increased gradually to 9.2 in 20 min, and after that dropped in accordance with the electrolysis time. Initial pH rise could be attributed to the production of hydroxyl ions on the cathode, and the drop of pH was caused by gradual consumption of hydroxyl anions as well as production of hydrogen cations in dissociation reactions of HOCl and OClalong with the reaction of HOCl in the treated phenol solution (Bahadir et al., 2003). Since HOCl was immediately consumed by oxidation of phenol in the higher concentration phenol solution, the pH was kept in the range of 9.0-9.3 for a longer period. Moreover, formation of different organic acids also resulted in the pH drop. The pH reached approximately 7.6 when the phenol was completely removed from the two concentrations of phenol solution (Comninellis et al., 1991, 1994; Houk et al., 1998; Feng et al., 2003; Li et al., 2005;).

Figure 8. The change of pH with time during electrochemical degradation of 8 and 20 mg/L phenol in 1000 mL electrolyte as a function of time with 0.3 g/L NaCl as supporting electrolyte. I = 10 mA/cm2,Ti/RuO2-Pt. (♦) 8 mg/Lphenol, (▲) 20 mg/L phenol.

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3.3.2. ICE with Electrolysis Time The COD method was used for the determination of the current efficiency. In this method, the COD was measured during electrolysis and the instantaneous current efficiency (ICE) was calculated using the following equations (Comninellis et al., 1991).

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ICE=

[CODt − CODt +Δt ]FV 8IΔt

Where CODt and CODt+⊿t are the COD (in gO2/L) at times t and t+⊿t(in s), respectively, I is the current intensity (A), F is the Faraday constant(96487 C/mol), t is the time(in s),V is the volume of the electrolyte (L) and 8 is a dimensional factor for unit consistence. It can be observed from Figure 9 that ICE was relatively higher at the initial step of 0–10 min than the 10–20 min and the final step of 20–40 min, and the ICE for higher concentration of phenol solution was larger compared to lower concentration of phenol solution. Comninellis et al. (1991) and Polcaro et al. (1999) indicated that the initial organic concentration influenced the current efficiency, and that the current efficiency at greater initial organic concentrations was relatively high. This behavior was explained by the side reaction of oxygen evolution, i.e., the radicals formed could react rapidly with the organic matters before the oxygen evolution in a high initial organic concentration. The initial reaction at 0–10 min is the oxidation of phenol, which might be the oxidation of phenol to benzoquinone and further oxidation to a series of intermediate products with the aromatic ring opening. The ICE decreases rapidly in 10–20 min. and the main reaction in this period is the oxidation of the intermediates formed in 0– 10 min. obviously, they were more difficult for electro-oxidation than phenol or benzoquinone. It is worthwhile to note that ICE increased again in the last step at both phenol concentrations, which is probably because of the formation of HOCl that promote the oxidation of intermediate products.

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Figure 9. Change in ICE with respect to treatment time for electrochemical degradation of 8 and 20 mg/L phenol solution. I = 10 mA/cm2, Ti/RuO2-Pt, 0.3 g/LNaCl.(A)8 mg/L (B) 20 mg/L.

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3.3.3. Free Radicals and Oxidizing Substance Production Mechanism of Phenol Oxidation The mechanism of electrochemical degradation of organic matter at anodes has been suggested by Comninellis (1994). Water is electrolyzed by anodic catalysis to produce adsorbed hydroxyl radicals. H2O + MOx→ MOx [·OH] + H+ + e-.

(1)

The adsorbed hydroxyl radicals may form chemisorbed active oxygen. MOx [·OH] → MOx+1 + H+ + e-

(2)

Meanwhile, the hydroxyl radicals will react with each other to form molecular oxygen to complete the electrolysis of the water molecules. MOx [·OH] →M +O2+ H+ +e-

(3)

Furthermore, another strong oxidant of hypochlorite may be produced in many wastewaters containing chlorides (Israilides et al., 1997; Szpyrkowicz et al., 2001).

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(4)

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Organic matter (R) included in wastewater are oxidized by hydroxyl radicals (Szpyrkowicz et al., 2001). R + MOx [·OH] → MOx + CO2 + z H+ + z e-

(6)

MOx+1 + R → MOx + RO

(7)

R + HOCl → product + Cl-

(8)

The mechanism of electrochemical oxidation of phenol have been carried out in the previous works (Comninellis et al., 1991; Feng et al., 1991), It is well known that phenol oxidation starts with a one-electron transfer, leading to a phenoxy radical reaction (Lund et al., 1991), and some possible reactions of phenoxy radicals relates to radical–radical coupling, radical disproportionation, radical elimination or radical oxidation to cation and then followed by benzoquinone formation. Benzoquinone is absorbed onto the electrode surface and gives up an electron, and an adsorbed ·OH radical would attack the benzoquinone (Feng et al., 1991; Lund et al., 1991). Further oxidation of these intermediates yields harmless end-products CO2 and H2O. Li et al. (2005) showed that the difference in the oxygen evolution potential for the different anodes could lead to different pathways of electrochemical phenol degradation. With sufficient hydroxyl radicals formed on the anode, maleic acid could be oxidized directly to oxalic acid, which can be oxidized readily to CO2. HOCl formed in the presence of NaCl would dominate the oxidization of phenol to CO2 during the electrochemical degradation of phenol in the present experiments.

Free Radicals and Oxidizing Substance Production during Electrolysis Figure 10 shows the absorbance of RNO sharply decreased in the initial 10 min. at 10 mA/cm2 and 20 mA/cm2 with 0.3 g/LNaCl addition, and the bleaching ratio was up to 87%. It was clear that bleaching rate was quicker at the current density of 20 mA/cm2 than 10 mA/cm2, and it was quicker on Ti/RuO2-Pt anodes than on Ti/IrO2-Pt anodes under the same condition. However, it shows that the absorbance had almost no change without NaCl, indicating that formation of hypochlorous acid was an important bleaching factor during the electrolysis. As the ring of benzoquinone produced by the oxidation of phenol could be opened by hydroxyl radicals but not

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hypochlorous acid to form maleic acid and another small organic compound, which is believed to be important for phenol degradation, and the poor ability of Pt for producing hydroxyl radicals (Feng et al., 2003), the existing of RuO2 or IrO2 coating were indispensable. However, the existing of Pt would keep the electrode for a long service time. Comninellis et al. (1994) suggested that hydroxyl radicals react selectively with RNO, but hypochlorous acid play a very important role in the RNO bleaching from the present study, which would promote the phenol degradation. Li et al.(2005) showed that phenol in Na2SO4 solution decreased from 490 mg/L to zero in 35 hours with Ti/RuO2 at current density of 20 mA/cm2, but phenol solution with 1.0 g/L NaCl as supporting electrolyte was degraded from 80 mg/L to zero in 130 minutes at current density of 10 mA/cm2 in present study (Figure 6). Therefore, the presene of NaCl is necessary for efficient degradation of phenol.

Figure 10. Electrochemical bleaching of 50 μ mol/L RNO in 0.3 g/L NaCl solution as a function of treatment time with Ti/RuO2-Pt and Ti/IrO2-Pt anodes. (■) I = 10 mA/cm2,Ti/RuO2-Pt; (•) I = 20 mA/cm2,Ti/RuO2-Pt; (▬) I = 10 mA/cm2,Ti/IrO2-Pt; (▲) I = 20 mA/cm2,Ti/IrO2-Pt; (*) I = 10 mA/cm2,Ti/RuO2-Pt,0 mg/L NaCl; (♦) I = 10 mA/cm2,Ti/IrO2-Pt, 0 mg/L NaCl.

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3.3.4. Effect of the Concentration of Ammonia Nitrogen to the Degradation of Phenol

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Figure 11. The removal of phenol and ammonia nitrogen at different initial concentration of ammonia nitrogen, NaCl 0.8 g/L, I= 23.7mA/cm2, C represents the concentration of ammonia nitrogen. (a) The removal of phenol, (b) the removal of ammonia nitrogen.

As shown in Figure 11, ammonia nitrogen greatly affects the electrochemical oxidation of phenol, without adding ammonia nitrogen after 2 hours of degradation, the removal efficiency of phenol finally reached 99.6%. While the concentration of ammonia nitrogen is 100mg/L, by 2 hours of degradation, the removal of phenol is merely 26.6%. However, it is not the case that the higher ammonia nitrogen concentration is, the lower removal of phenol remains. When the concentration of ammonia nitrogen is 75 mg/L, the removal of phenol is eventually to 76.6%, and when the concentration of ammonia nitrogen is 50mg/L, the removal of phenol is eventually only 31.7%. In fact, in the chloride system, the effectiveness of direct oxidation on the anode at the electrode surface and the effectiveness of indirect oxidation of free hydroxyl can be basically neglected. The main oxidant is HClO, and the oxidation of ammonia nitrogen mainly follows the following reaction equation (White et al., 1999): 2Cl− → Cl2 + 2e−

(9)

Cl2 + H2O → HOCl + H+ + Cl−

(10)

HOCl + (2/3)NH3 → (1/3)N2 + H2O + H+ +Cl−

(11)

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(12)

HOCl + (1/4)NH4+ → (1/4)NO3− + (1/4)H2O + (3/2)H+ + Cl−

(13)

HOCl + (1/2)OCl− → (1/2)ClO3− + H+ + Cl−

(14)

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In the reactions, Cl- served as the catalyst to continue the degradation of ammonia nitrogen. Under the ideal condition, most of the amount of ammonia nitrogen should be degraded into nitrogen gas as shown in Eq. 11 and Eq. 12 and gets free from the solution to obtain the goal of denitrification. But it has been reported that The reaction equations show as (Cheng et al., 2005): NO3- + 2H+ + 2e- → NO2- + H2O

(15)

NO3- + 6H+ + 5e- → (1/2)N2 + 3H2O

(16)

NO3- + 9H+ + 8e- → NH3 + 3H2O

(17)

NO2- + 4H+ + 3e- → (1/2)N2 + 2H2O

(18)

NO2- + 7H+ + 6e- → NH3 + 2H2O

(19)

NO2- + 5H+ + 4e- → NH2OH + H2O

(20)

As shown in Eq. 17 and Eq. 19, nitrate will be reduced to ammonia nitrogen. For this, we assume that as the reaction continued, the huge amount of electrons will loss with the oxidation of phenol on the anode, while at the same time the huge amount of electrons will exist on the cathode to provide with Eq. 17 and Eq. 19 in which, much more electrons are consumed to allow ammonia nitrogen to be oxidized into nitrate via HClO and then, reduced to ammonia nitrogen on the cathode. The ammonia nitrogen produced by reduction will be continued to oxidized to nitrate via HClO and a vicious cycle was formed. The great amount of electric power and HClO will be consumed to cause significant decrease of HClO which is acted as the main oxidant during the degradation of phenol, and the removal rate of phenol is lowered. This speculation is also certified in Figure 11B. Figure 11B shows that at the different initial concentration of ammonia nitrogen, the removal is fluctuated along with the time period. When the initial concentration of ammonia nitrogen is 50 mg/L, for 1 hour of electrolysis, the

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fluctuation of ammonia nitrogen removal is very high. This shows that after being oxidized on the anode the ammonia nitrogen is again reduced on the cathode, which made the less final removal of phenol shown in Figure 11B. However, when the initial concentration of ammonia nitrogen is 75mg/L, the fluctuation of ammonia nitrogen removal is not high, thus the final removal of phenol is relatively higher as shown in Figure 11A. When the concentration of ammonia nitrogen is 100 mg/L the final removal of phenol is the lowest. The concentration of ammonia nitrogen will be nearly returned to its initial state after several fluctuations. This is because of the concentration of ammonia nitrogen is much higher compared with that of phenol. The electric power is all consumed to the oxidation of ammonia nitrogen and the reduction of nitrate which decreases the removal of phenol. When the initial concentration of ammonia nitrogen is less, the situation would change. When the initial concentration of ammonia nitrogen is 25 mg/L, the removal of phenol is little lower compared with that of no ammonia nitrogen being added, but the removal rate is faster within the time period of 90 minutes. It is speculated that NH2Cl produced from the reaction of ammonia nitrogen with HOCl speeds up the oxidation of phenol. Based on the theory of “breakpoint chlorination “.The reaction equation is as follows: HOCl + NH4+ → NH2Cl + H2O + H+

(21)

HOCl + NH2Cl → NHCl2 + H2O

(22)

NHCl2 + H2O → NOH + 2H+ + Cl-

(23)

NHCl2 + NOH → N2 + HOCl + H+ + Cl-

(24)

NH2Cl produced by Eq.21 is the disinfector frequently used in water disposal technology with its strong oxidizability. Some people once used potassium permanganate and chloramine to pre-oxidize the refractory organics in the water, and obtained a better results (Yan et al., 2007). Therefore we can speculate that chloramine can served as an oxidant to oxidize phenol which speeds up the degradation of phenol.

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3.3.5. Effect of Cathode on the Degradation

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We have found in the experiment that when phenol and ammonia nitrogen are in the simultaneous degradation, a film of black substance is formed on the titanium plate which served as the cathode. However, this phenomenon can not be found when only phenol is degraded after the experiment. Although the stability of anode has been testified by cyclic voltammetry experiments before, yet the function of the cathode may be altered by the substance adhered on the cathode to make the higher reducibility of the cathode, and nitrate produced by the oxidation of ammonia nitrogen was again reduced to ammonia nitrogen. In the following experiments, the cathode was soaked in 1: 9 nitric acid for one hour to remove the black substance and rinsed by DI water before the start of each experiment. Three sets of degradation experiments were carried out with the initial concentration of phenol fixed as 50mg/L and the different initial concentration of ammonia nitrogen. The results are shown in Figure 12.

Figure 12. The removal of phenol and ammonia nitrogen at different initial concentration of ammonia nitrogen, NaCl 0.8g/L, I= 23.7mA/cm2, C represents the concentration of ammonia nitrogen,The removal of phenol, (b) The removal of ammonia nitrogen.

The results proved the black substance adhered to the cathode weighs heavier influence on the degradation of both phenol and ammonia nitrogen. After the black substance being removed, the removals of phenol are all up to over 80% by three sets of experiments, and the removal of ammonia nitrogen appears no fluctuation. When the initial concentration of ammonia nitrogen is 50mg/L, the removals of phenol and ammonia nitrogen are all up to more than 98%. This demonstrates that there is little nitrate reduced on the cathode, and the removal results of both ammonia nitrogen and phenol are fairly good. But

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the disposal results are not ideal when the initial concentration of ammonia nitrogen is 25 mg/L and 75 mg/L respectively. It is supposed that this is because of the effect of chloramine produced in Eq.13 and the existence of chloramine could accelerate the degradation of phenol and made the fast degradation of ammonia nitrogen itself.

3.3.6. Effect of Initial pH on the Degradation

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As various organic acids are generated from the degradation of phenol, the pH in the solution will change. Accordingly, the initial pH may have influence on the degradation of phenol and ammonia nitrogen. To adjust pH into 3 and 11 respectively with the saturated solution of NaOH and H2SO4 and with other conditions remaining unchanged the experiment were carried. The results are shown in Figure 13.

Figure 13. The removal of phenol and ammonia nitrogen at different initial pH, NaCl 0.8g/L, I= 23.7mA/cm2. (a) The removal of phenol, (b) the removal of ammonia nitrogen.

As shown in Figure 6, comparing initial pH as 3 and 11 with the pH unadjusted, the removals obviously decreased. The removal of phenol merely reached 60%, while the removal of ammonia nitrogen is only 50%. When initial pH is 3, the removal of ammonia nitrogen is only 30%, and the removal exhibited fluctuations along with the time period. This shows that ammonia nitrogen was oxidized to nitrate on the anode and the nitrate again reduced to ammonia nitrogen on the cathode. This has been testified by analyzing the amount of nitrate content in the water.

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Figure 14. The content of ammonia nitrogen and nitrate at different initial pH, NaCl 0.8 g/L, I= 23.7 mA/cm2. (a) pH=3, (b) pH=11.

As shown in Figure 14, when initial pH was 3, the curve that the content of ammonia nitrogen as function of time, and the curve that the content of nitrate as function of time are just totally coincide. The content of ammonia lessened, the content of nitrate expanded. While the content of ammonia nitrogen expanded, the content of nitrate lessened. The curve of the removal of ammonia nitrogen exhibited a fluctuating uprising tendency, while the curve of nitrate always kept fluctuation without obvious sign of rising tendency. This result fully shows that under the condition of initial pH=3, ammonia nitrogen was oxidized into nitrate on the anode, and among which, a portion of nitrate again reduced into ammonia nitrogen on the cathode. This vicious recycle leaded to a great deal of oxidants generated on the anode wasted in which phenol could not obtain sufficient oxidant and the degradation of phenol slowed down. In fact, from Eq. 15 to Eq. 20, it can be seen that H+ would be consumed to reduce nitrate into ammonia nitrogen. The decrease of pH will allow the increase of H+ in the solution which accelerates the reduction of nitrate. In addition, from Eq. 9 and Eq. 10 we can find that the increase of H+ will go against C12 to be dissolved in the solution, and this lessened the concentration of HOCl which is unfavorable to the electrochemical oxidation. When initial pH was 11, the removals of phenol and ammonia nitrogen were slightly better than that of initial pH was 3, but there is a much big gap compared with the pH unadjusted. From Figure 14b, we can see that there is no large fluctuation of nitrate concentration along with the time period which explains that nitrate is seldom reduced on the cathode. This is also because of pH going up to lessen the content of H+ in the solution to restrain the reduction of nitrate. However, Eq. 14 shows that when pH goes up, at the same time ClO3- accelerates to generate and lower the concentration of HOCl. This result

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is similar to that of Lin and Wu (1996), but different from that of Vlyssides et al. (2002).

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3.4. CONCLUSIONS The electrochemical approach is a novel process for effective removal of phenol. In the present study, effect of current density, dosage of NaCl, initial phenol concentration on the performance of phenol electrolysis were investigated using Ti/RuO2-Pt and Ti/IrO2-Pt anodes Phenol could be readily mineralized at the Ti/RuO2-Pt and Ti/IrO2-Pt anodes; however, its degradation was considerably slower at the Ti/IrO2-Pt anode. Complete TOC removal was achieved at both anodes with NaCl as supporting electrolyte. It was found that Ti/RuO2-Pt anode had a higher oxygen evolution potential than Ti/IrO2-Pt anode, which will increase the current efficiency for electrochemical degradation. Moreover, HOCl formed quickly with sodium chloride as supporting electrolyte and that would play an important role on the oxidation of phenol. With NaCl as supporting electrolyte, specific anode surface treatment of the RuO2-Pt coating provided the anode with an apparent catalytic function for rapid organic oxidation that was mainly brought about by HOCl generated during electrolysis process. The ICE was relatively higher at the initial step and the final stepand. The ICE for higher concentration of phenol solution was larger compared to lower concentration of phenol solution. Considering the economy factor, using Ti/RuO2-Pt anode, at current density of 10 mA/cm2 with 0.3 g/L NaCl as supporting electrolyte was best for electrochemical degradation of 8 mg/L phenol contaminated water. The initial concentration of ammonia nitrogen remains great influence on the degradation of phenol by electrochemical oxidation, and the influence mechanism is much complicated. The initial pH had heavy influence on the effectiveness of degradation. Neither acid pH nor alkaline pH is beneficial to the degradation of phenol and ammonia nitrogen. Under acid condition, the increase of H+ will strengthen the reducibility of the cathode to allow nitrate to be reduced to ammonia nitrogen and lower the removal of ammonia nitrogen while under alkaline condition, the formation of ClO3- will block the formation of HOCl which will result in the inferior effectiveness of electrochemical oxidation.

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Chapter 4

ELECTROCHEMICAL REDUCTION OF NITRATE CONTAINED IN GROUNDWATER

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The aim of this part of the experiments is to completely remove nitrate in polluted water using the electrochemical reduction in an undivided cell, and to find a proper condition to perform both cathodic reduction of nitrate and anodic oxidation of the produced ammonia and nitrite.

4.1. EXPERIMENTAL PROCEDURES A vitreous beaker having a net volume of 1 L was used in all experiments as a tank. The electrolysis cell (2x13x5.2=135.2cm3) was manufactured by acryl plates with outer spots for the electrodes assembled. Cu/Zn(Cu: 62.2 wt. %; Zn: 37.8 wt. %) plate of 75 cm2 (13 cm×5 cm) was used as the cathode and Ti/IrO2-Pt electrode (TohoTech company, Japan) with the same area as the anode, a distance of 1.0cm between the cathode and the anode was set. In this experiment, there was two divers electrode connection. One was conventional connection (Figure 15a). Another one was parallel connection that a Cu-Zn alloy cathode was put in the middle of the cell and two Ti/IrO2-Pt anodes were put beside the cathode (Figure 15b). A DC power supply with a voltage range of 0–50 V and a current range of 0–5 A was employed as power supply. The cycling rate of the pump was 100mL/min.

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Figure 15. Set-up of the electrochemical system.

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4.2. ANALYSIS In this study, KNO3 was put into the 1L distilled water simulating the contaminated water. To duplicate the ion concentration of groundwater, 0.5g/L Na2SO4 was added to the water as electrolyte, and NaCl if needed. At specific time intervals, samples were withdrawn from the electrochemical tank. All analyses were done according to standard methods. NO3- -N and NO2- -N was determined by standard colorimetric method using spectrophotometer. NH4+ N was determined by titration method. Dissolved cooper and zinc content of the filtered samples were determined by Inductively Coupled Plasma Mass Spectrometry (ICPMS). The possible formation of hydrazine and hydroxylamine was not investigated because the treated solutions were changed into basic after electrolysis, in which hydrazine and hydroxylamine will not be produced (De, 2000).

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4.3. RESULTS AND DISCUSSION 4.3.1. Parallel Connection and Conventional Connection

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Figure 16. Nitrate converted to nitrite, ammonia and nitrogen at current density of 10 mA/cm2, 500 mg/L Na2SO4. (a) conventional connection, (b) parallel connection.

In this study, two different cathode connections were chosen, the parallel connection (Figure 15a) and the conventional connection (Figure 15b). Figure 16 shows the variation of total nitrogen, nitrate-N, nitrite-N, and ammonia-N. A current density of 10 mA/cm2 was used in most electrolysis nitrate reduction as it exhibited a relatively high reduction rate in our experiments. In both the experiments, the 500mg/L Na2SO4 was added into the solution to simulate the ion concentration of groundwater. It can be seen from Figure 16 that the electrochemical nitrate reduction had different behavior with the two conditions at the current density of 10 mA/cm2; moreover, the effect on byproduct production was also different. Under the condition of conventional connection (Figure 16a), the concentration of nitrate decreased with respect to treatment time, it decreased from 50.0 to 22.7 mg/L in 240 min. On the other hand, the ammonia-N increased from 0 to 8.3 mg/L; and the nitrite-N increased to 1.3 mg/L at the first 90 min, then decreased to 0mg/L at the 240 min. The total nitrogen decreased from 50 to 31.7 mg/L. While under the condition of parallel connection (Figure 3b), the nitrate-N decreased from 50.0 to 6.6mg/L. The ammonia-N increased rapidly from 0 to 20.0mg/L at the 120min and then maintained stability. At the end of the electrolysis, the ammonia-N was 16.3mg/L. The nitrite-N increased sharply from 0 to 2.4mg/L and at the end of the electrolysis, it decreased to 0mg/L. TN decreased from 50.0mg/L to 23.4mg/L. As shown above, the parallel connection was more

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efficient than the conventional connection in the electrochemical reduction of nitrate. The parallel connection can make full use of the both sides of the CuZn alloy cathode which could reduce the nitrate, as a result of which, the cathode area increases. And the experiments demonstrated that the parallel connection could make the pressure in the cell decline, which weaken the hydrogen production side reaction. Therefore it improved the efficiency of the electrochemical reduction of nitrate. It can be seen from Figure 3 that under the both conditions, nitrite-N increased sharply in the beginning and decreased to 0 in the end. It is because that the nitrite is the intermediate product of the electrolysis. Cu-Zn alloy cathode is important. Cu converts the nitrate to nitrite, while Zn makes the nitrite to NH3 and other products (Mácová et al., 2005). It is obvious that the ammonia-N increased rapidly in the beginning of the electrolysis, and maintained the stable after 120min. This is because that the cathodic reductive rate of nitrate is faster than the anodic oxidition rate of ammonia. The cathodic reductive nitrate rate declined with the respect of electrolysis time, while the anodic oxidition rate of ammonia increased. That proved that under no NaCl add, anodic direct oxidation can not remove the ammonia completely which was the product of cathodic reduction. After the 240min electrolysis, the concentration of Cu2+ and Zn2+ were 0.033mg/L and 0.016mg/L respectively which were far lower than the standard 1mg/L and 1mg/L.

4.3.2. Influences of Addition Time of NaCl The effect of different addition time of NaCl on nitrate reduction is shown in Figure 17. At the current density of 10 mA/cm2, the nitrate-N decreased from 50.0 to 9.9, 7.5, and 10.5 mg/L in 240 min at the addition time of 0, 120, 180 min, respectively. Without adding NaCl, the nitrate-N decreased from 50.0 to 6.6mg/L in 240min. It was obvious that the nitrate reduction rate without NaCl addition was a little higher than that in the presence of NaCl. It was due to that nitrate reduction was retarded in the presence of chloride ion. Dash et al. (2005) used titanium as cathode to electrochemically reduce the nitrate, found that the nitrate reduction in the presence of 250 mg/L sodium chloride was retarded, whereas sulfate had little effect on the nitrate reduction. However, under on NaCl condition, the ammonia-N maintained a high concentration, while adding NaCl, after 240min electrolysis, no ammonia-N and nitrite-N could be detected.

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According to Rajeswar et al. (1997), oxidizing hypochlorite ion will be formed during electrolysis in the presence of chloride ion. 2Cl-=Cl2 + 2e-

(10)

Cl2+H2O=HClO+H+ + Cl-

(11)

HClO =ClO-+H+

(12)

Figure 17. Nitrate converted to nitrite, ammonia and nitrogen at different NaCl addition times (10mA/cm2, 500mg /L NaCl). (a) 0 min, (b) 60 min, (c) 120 min, (d) no NaCl.

The hypochlorite acid formed during the electrolysis would oxidize the by-products of ammonia and nitrite, which were assumed to be oxidized into nitrogen gas and nitrate, respectively (Pressley et al., 1972). NH4++ HClO=N2+H2O+H++Cl-

(13)

NO2-+ HClO=NO3-+H2O+Cl-

(14)

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Although no obvious difference of total nitrogen was observed after 240 min treatment at the different NaCl addition times, changes in the nitrate-N and ammonia-N during electrolysis were rather different with different NaCl addition times; nitrate-N was rapidly reduced and a large amount of ammonia was produced before addition of NaCl, and nitrite was also produced during electrolysis. After the addition of NaCl, both ammonia and nitrite were removed.

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4.4. CONCLUSIONS In order to develop a method of electrochemical reduction of nitrate, at the same time removing the byproducts ammonia and nitrite, the Cu-Zn alloy cathode and the Ti/IrO2-Pt anode were used for the synthetic nitrate solution in the undivided cell. The Cu-Zn alloy cathode has a significant capacity of electrochemical reduction of nitrate. Under the conditions of 10mA/cm2, addition of 500mg/L NaCl and the parallel connection, the nitrate-N decreased from 50.0mg/L to 9.8mg/L for the 240-min electrolysis. No ammonia-N and nitrite-N could be detected in the end. In the present study, the nitrate was removed completely by electrochemical approach using Cu/Zn cathode and Ti/IrO2-Pt anode with the addition of NaCl, it is a worthy method for treatment of nitrate-polluted water.

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Chapter 5

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A PILOT PLANT OF ELECTROCHEMICAL TREATMENT SYSTEM (0.3 M3/HR) In this part of the experiments, a pilot plant scale electrochemical treatment system (0.3 m3/hr) using pulse voltage has been successfully developed, and its performance evaluated using domestic wastewater, pond water containing algae and wastewater from hog raising. To investigate the effect of anode materials on their oxidizability for organic matter, measurement of anode behavior and detection of hydroxyl radicals formed in the electrochemical process were investigated in the laboratory.

5.1. EXPERIMENTAL PROCEDURES 5.1.1. Laboratory Experiments As shown in Figure 18, the electrochemical apparatus consists of a power supply, flat-plate anode and cathode (4x12 cm) mounted in an acrylic cell (200 mL), peristaltic pump and reservoir (500 mL). A high voltage pulse generator (YHPG-5K-50MTR, Yamabishi Electric Co., Ltd.) was used as the power supply. Cooling was provided by ice to keep the temperature in the solution below 30ºC. In this study, the cathode was made from a titanium sheet, and the anodes were a sheet of titanium, platinum and titanium coated with Ti/ RuO2-TiO2. The ratio of RuO2 to TiO2 was 30:70 (v/v). The preparation of Ti/ RuO2-TiO2 used procedures suggested by Motheo et al. (2000). That is, the titanium

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supported oxide layer was coated by thermal decomposition of precursors in isopropyl alcohol. The solvent was evaporated at a low temperature and the dried layer heated at 450ºC for 5 min under a 5 L/min O2 flux. Finally, the electrode was annealed at 450ºC for 1 hr. The distance between cathode and anode was set at 4 cm.

Figure 18. Schematic diagram of the electrochemical apparatus.

The electrochemical pilot plant consisted of a screen, wastewater tank, and reactor A, reactor B, two sedimentation tanks, DC power supply and pulse generator; its schematic diagram is shown in Figure 19. Wastewater was pumped through the screen to separate some large solid particles. It then went to the reactor A for treatment for 15 min, and subsequently to the first sedimentation tank for 1 hr. The effluent from the first sedimentation tank was transferred to the reactor B for a 15 min treatment. Finally, the treated wastewater was settled in the secondary sedimentation tank for 1 hr, and discharged.

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5.1.2. Pilot Plant of Electrochemical Treatment System

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Figure 19. Schematic diagram of the pilot plant experiments.

The cathodes in reactor A and B were made of stainless steel, the anode in the reactor A was made of iron, and the anode in reactor B was Ti/RuO2-TiO2 fabricated by the method mentioned above. The cathodes and anodes were cone-shaped. The dimensions of the anodes were φ21xφ36xH73 cm, and the distance between the cathode and anode was 2 cm. The cone electrode shape was used to disturb the wastewater flow and promote the pollutant mass transfer between the electrodes and wastewater. DC (FX060-100, TAKASAGO, Co., LTD) was applied the reactor A; the current density was 3 mAcm-2. High voltage pulse (YHPG-0.8K-100A, Yamabishi Electric Co., LTD) was applied to the reactor B, the voltage and frequency were 500 V and 25 kHz, respectively.

5.2. ANALYSIS 5.2.1. Detection of Hydroxyl Radicals and Oxidizing Substance To measure hydroxyl radicals and oxidizing substance formed during the electrochemical treatment, an NaCl solution (0.02%, w/w) containing 50 μmol/L RNO was used because RNO reacts rapidly with hydroxyl radicals

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selectively. The bleaching of RNO solution by hydroxyl radicals was measured by absorbance changes at 440 nm (Wabner et al., 1985). RNO solution of 500 mL was poured in the reservoir, and circulated at the flow rate of 87 mL/min. The applied voltage, peak current and frequency were 600 V, 1.0 A and 1 KHz, respectively. Samples were taken at intervals of 5 min and absorbance of RNO solution was measured by HACH DR/4000 spectrophotometer.

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5.2.2. Electrocatalytic Properties of Anodes In order to investigate the electrocatalytic activities of anodes, cyclic voltammetry measurements were performed in three electrolytes of 100 mM phosphate buffer solution, 100 mM phosphate buffer solution which contained 50 mM NH4Cl and 50 mM NH3, respectively. A platinum wire was used as the counter electrode and a saturated calomel electrode (SCE) was employed as a reference. Working electrodes were a flat-plate made of titanium, Ti/RuO2TiO2 and platinum; their dimensions were 4x12 cm. The geometrical area immersed in the electrolytes was 20 cm2. A potential sweep unit (Model1114, BAS co.) was applied to the potential waveform, and cyclic voltammetry was recorded with a X-Y recorder (MEMORY HiCORDER 8807, HIOKIE. E. Co.). The potential was scanned at 20 mV/s, starting from 0 V, and the scan range was –2 V~2 V.

5.2.3. Others Samples were taken from untreated and treated wastewaters. The evolved gas was collected from the reactor B in a TEDLAR pack (1 L) usually used for gas sampling. The TEDLAR packs were purged using argon gas before starting the experiments, and then connected to the reactor B by polyethylene tube. Concentrations of T-P, T-N, TOC, COD, BOD, chl-a in the wastewater and T-P, T-N, TOC in the sludge were determined according to Japanese sewage test method. O2, N2, CO2, in the collected gases were identified by gas chromatography (GC-8A, Shimadzu, Co., Japan). The detector was TCD, and the columns were Shimalite Q and Porapak Q.

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5.3. MATERIALS Domestic wastewater and pond water containing algae (700 L in each case) were used in the experiments. To evaluate the treatment efficiency on the wastewater containing high concentration of pollutants, raw wastewater from hog raising and its effluent of biologically treated wastewater were used. The effluent was obtained from raw wastewater treated by screening, coagulation with polyaluminum chloride (PAC), primary aeration and sedimentation, secondary aeration and sedimentation.

5.4. RESULTS AND DISCUSSION

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5.4.1. Electrocatalytic Properties of Anodes

Figure 20. Cyclic voltammogram obtained in 100 mM phosphate buffer, scan rate 20 mV/s; reference saturated calomel electrode: (a) titanium, (b) Ti/RuO2–TiO2, (c) platinum.

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Figure 21. Cyclic voltammogram obtained in 100mM phosphate buffer containing 50 mM NH4Cl, scan rate: 20 mV/s; reference: saturated calomel electrode: (a) titanium, (b) Ti/RuO2–TiO2, (c) platinum.

Figure 22. Cyclic voltammogram obtained in 100 mM phosphate buffer containing 50 mMNH3, scan rate: 20 mV/s; reference: saturated calomel electrode: (a) titanium, (b) Ti/RuO2–TiO2, (c) platinum.

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De Vooys et al. (2001) conducted the comparative experiments to discuss the electrocatalytic oxidation of ammonia and the intermediates formed during the reaction using a series of transition metals, and shown that the activity of platinum electrode for ammonium oxidation was larger than others transition metals such as palladium, rhodium, iridium electrodes. In our present study, cyclic voltammograms for titanium, Ti/ RuO2-TiO2 and platinum electrodes were used, as shown in Figure 20–Figure 22. Titanium was almost no electrochemical activity during all experiments, but the reduction current was largest in phosphate buffer solutions. On the other hand, in the case of Ti/ RuO2-TiO2, the oxidation current increased gradually from 0 V, and increased sharply from 0.9V, and reached approximately 1.5 mA at 2 V in phosphate buffer solution. The cyclic voltammogram obtained was similar to the one reported by Pelegrini et al. (Pelegrini et al., 1999), i.e., the charge was distributed over a wide range of potential, and the band located at 0.9 V was observed corresponding to various solid-state redox transitions of rhodium. However, no oxidation current was observed before 0.9 V, and 1 mA at 2 V in the case of platinum. Furthermore, the oxidation current at Ti/ RuO2-TiO2 was slightly larger than that found using platinum; especially, the oxidation current became smaller and rather low at potentials lower than 0.9 v when added NH4Cl or NH3. Therefore, the oxygen-containing products like N2O and NO may be formed because the selectivity to N2 is 100% at potentials lower than 0.8 V (De Vooys et al., 2001).

5.4.2. Effect of Anode Materials on Radical and Oxidizing Substance Production The difference between titanium, platinum and Ti/ RuO2-TiO2 can be clearly seen in Figure 23. The absorbance of RNO using Ti/ RuO2-TiO2 sharply decreased up to the treatment for 30 min, and then decreased slowly. The bleaching ratio was approximately 88% at 30 min and 97% at the end of the experiment. Compared to Ti/ RuO2-TiO2 anode, absorbance in the case of platinum decreased slowly during all the experiments, and the bleaching ratio reached only 36% at the termination of the experiment. On the other hand, a decrease in absorbance value was not observed for Ti anode, but it rather increased. This result was due to Ti ions electrochemically dissolving and entering into the RNO solution, resulting in increased turbidity of RNO solution. Corrosion of the titanium surface was observed after 1hr. Obviously, radicals such as hydroxyl radicals, which were formed at the Ti/ RuO2-TiO2

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electrode were larger than at the Pt and Ti electrode. Comninellis (1993) reported that hydroxyl radicals formed at platinum anode were almost zero. It was proposed that platinum anode favors selective oxidation to Ti/SnO2, but not complete combustion. Anodes made of noble metal such as platinum have high oxidative activity for pollutants in wastewater, but the reaction rate is slow, and it is easily inactive (Takasu et al., 2001). The results obtained in the pulse voltage treatment confirmed the same phenomenon.

Figure 23. Electrochemical bleaching of 50 µmol/l in NaCl solution (0.02%, w/w) as a function of treatment time for different anode materials. Pulse voltage: 600V; peak current: 1.0 A; frequency: 1 kHz.

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5.4.3. Water Quality during Pilot Plant of Electrochemical Treatment Table 1. Electrochemical treatment using pilot plant (0.3 m3/hr) (a) Domestic wastewater T-N NH4-N (mg/L) (mg/L) Untreated* 33.03 23.09 Treated** 8.86 4.35 (b) Pond water containing algae T-N T-P Treatment (mg/L) (mg/L) Untreated* 4.4 0.296 Treated** 0.73 0.03 Treatment

T-P (mg/L) 4.5 0.045

COD (mg/L) 36.5 5

COD (mg/L) 46 3.5

BOD (mg/L) 10 2.2

SS (mg/L) 68 2

chl-a (μg/L) 270 0.6

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Note: * indicates before electrochemical treatment. ** indicates after electrochemical treatment. Current density: 3 mAcm-2; pulse voltage: 500 V; frequency: 25 kHz .

The quality of domestic wastewater and pond water containing algae is shown in Table 1. The concentrations of T-N, NH4-N, T-P, and COD in domestic wastewater after treatment by the electrochemical treatment system were reduced by 73%, 81%, 99% and 86%, respectively. The concentrations of T-N, T-P, BOD, COD, SS in pond water containing algae reduced by 83%, 90%, 78%, 92% and 97%, respectively, while chl-a removed was almost totally removed. It can be seen that the treatment system has a high efficiency when treating domestic wastewater and wastewater containing algae. The T-P and algae were removed mainly by coagulation due to the formation of metal ions by electrolysis at the electrodes in reactor A, as shown in equations (8)– (13). Other soluble substances such as NH4-N were decomposed by direct oxidation at the anode and/or indirectly oxidized in reactor B as shown in equations (5)–(7). Although ammonia oxidation at electrode-liquid interfaces occurred (Gootzen et al., 1998; De Vooys et al., 2001), the removal of ammonia may also through an indirect oxidation route (Szpyrkowicz et al., 1995) because many radical species such as hydroxyl radicals, hypochlorite might be formed during electrochemical treatment of wastewater containing chlorides. It may be necessary to examine the role of direct and indirect oxidation for ammonia in the future Both treated domestic wastewater and pond water became transparent, and the sludge was usable as a fertilizer because no chemicals were used.

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5.4.4. Mass Balance The results for N and P were calculated by N and P concentrations in the wastewater multiplied the volume of wastewater used in the experiments. For domestic wastewater, approximately 91% of P was contained in sludge, while 100% of P was coagulated in sludge when pond water containing algae was treated. The P included in the sludge from the treatment of domestic wastewater treatment was 0.46% (d.b.), and 0.29% (d.b.) for pond water containing algae. On the other hand, approximately 45% of the N from domestic wastewater and 28% from algae wastewater were transferred to sludge, respectively. The N contained in the sludge resulted from domestic wastewater treatment were 1.68% (d.b.) and 1.31% in the sludge from algae wastewater treatment. It can be seen that 38.6% of N from domestic wastewater and 55.2% from algae wastewater were removed as nitrogen gas. Experimentally, 74.2% of N2 and 0.1% of CO2 in evolved gas was detected. The production of N2 and CO2 follows the reactions (5)–(7).

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5.4.5. Quality of Wastewater from Hog Raising Table 2. Electrochemical treatment of wastewater from hog raising using pilot plant (0.3m3/hr) Materials

Treatment

Not biologically treated Biologically treated

Untreated*

T-N (mg/L) 852.0

Treated** Untreated Treated**

*

T-P (mg/L) 50.3

NH4-N (mg/L) 807.0

BOD (mg/L) 1200.0

COD (mg/L) 730.0

SS (mg/L) 850.0

724.2

6.3

650.0

690.0

390.0

260.0

831.8 272.8

123.0 22.1

94.1 41.8

16.0 13.0

140.0 84.0

19.0 8.0

Note: * indicates before electrochemical treatment. ** indicates after electrochemical treatment. Current density: 3 mAcm-2; pulse voltage: 500 V; frequency: 25 kHz..

The removal of T-N, T-P, NH4-N, BOD, COD and SS from raw wastewater from hog raising were 15%, 87.5%, 19.5%, 42.5%, 46.6% and 69.4%, respectively; while the removal from biologically treated wastewater was 67.2%, 82.0%, 55.6%, 18.8%, 40% and 57.9%, respectively (Table 2). Obviously, relatively larger removals of T-N and NH4-N were obtained in biologically treated wastewater. This result can be explained by suspended solids in wastewater limiting the electrochemical oxidation of pollutants. For

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this reason, electrochemical treatment has been proved uneconomical when utilized with raw wastewater containing high concentrations of suspended solids. It is necessary to biologically pretreat if the electrochemical treatment system is to be utilized for wastewater from hog raising.

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5.5. CONCLUSIONS On the basis of laboratory experiments examining the electrochemical catalytic activity of anodes and detection of the production of hydroxyl radicals and oxidizing substance during pulse voltage treatment, a pilot plant electrochemical treatment system using high voltage pulse has been developed successfully. The treatment performance was evaluated using domestic wastewater, pond water containing algae and wastewater from hog raising. The findings of this study are as follows: (1) no difference in electrocatalytic activity for ammonia oxidation between platinum and Ti/ RuO2-TiO2 was observed from the cyclic votammogram; (2) hydroxyl radicals formed in the electrochemical process were detected using a RNO solution and the production of hydroxyl radicals with Ti/ RuO2-TiO2 anode was larger than found using platinum and titanium anode; (3) excellent removal of T-N, NH4N, T-P, and COD from domestic wastewater and pond water containing algae was achieved by the pilot plant of electrochemical treatment system. Furthermore, chl-a was almost completely removed. On the other hand, the electrochemical treatment system was not feasible for treating raw wastewater from hog raising. However, the excellent treatment results were obtained in treating biologically treated wastewater from hog raising. It was suggested that a biological pretreatment might be necessary for treating wastewater containing high concentration of suspended solids.

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Chapter 6

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TWO TYPES OF PRACTICAL ELECTROCHEMICAL TREATMENT SYSTEMS (TREATMENT CAPACITIES OF 4M3/HR AND 0.5M3/HR) In this part of the experiments, electrochemical treatment systems for practical use based on the results of the fundamental experiments reported in previous work were developed, and the systems’ performance was evaluated. A series of experiments were carried out with two types of practical electrochemical treatment systems (treatment capacities of 4m3/hr and 0.5m3/hr) using effluents from an anaerobic digester (EAD) of cattle wastewater, supernatants from primary sedimentation in a sewage plant (SPS), and domestic wastewater to evaluate the systems’ treatment abilities.

6.1. EXPERIMENTAL PROCEDURES Electrochemical treatment systems with treatment capacities of 4m3/hr (batch type) and 0.5 m3/hr (continuous type) were used as the experimental apparatus. A schematic diagram is displayed in Figure 24. The two treatment systems are composed of a pretreatment tank for removing suspended solids such as fallen leaves, a buffer tank for assuring a constant supply of water to the next process, reactor A (8 m3 in 4m3/hr system and 0.5 m3 in 0.5m3/hr system), reactor B (8 m3 in 4m3/hr system and 0.5 m3 in 0.5m3/hr system), and a sedimentation tank. The A and B reactors were key processes, in which the

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cathode and anode electrodes were placed. The cathode in both reactors, A and B, was a stainless steel cylinder 50 cm in diameter and 100 cm in height. The anode in reactor A was an iron cylinder (diameter 30 cm, height 100 cm), and the anode in reactor B was titanium coated with RuO2-TiO2 to obtain better catalysis and durability (Feng et al., 2003). DC (FX060-100, Takasago, LTD) was applied to reactor A, and a current density of 3mA/cm2 was achieved. The electrode polarities in reactor A were changed periodically to remove the solids adhering to the anode. High-voltage pulses (YHPG-0.8K-100A, Yamabishi Electric Co., LTD) were applied to reactor B (500V, 50KHz). Additionally, an ultrasonic signal (600 KHz, 300w) was applied to reactor B to enhance the treatment efficiency for organic matter, and to clean the anode. Three ultrasonic elements were installed equidistantly around reactor B.

Figure 24. Schematic diagram of the experimental apparatus.

Batch tests were conducted in the 4m3/hr system. An EAD of 8m3 was treated for 2 hours in reactor A, and then transferred to reactor B for 2 hours of treatment. An SPS of 8m3 was simultaneously treated for 2 hours in reactors A and B to investigate their behavior. Finally, the supernatants of EAD and SPS from reactor B were allowed to settle for 2 hours in the sedimentation tank.Sampling from the reactors was conducted immediately when each process was completed. The evolved gases from reactors A and B were collected by a TEDLAR pack (1L), usually used for gas sampling. A polyethylene cylinder (13 cm diameter and 65cm length) was put in each reactor, with 30cm exposed over the water surface. The TEDLAR packs were

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connected to the cylinder, and then purged by argon gas before starting the experiments. The 0.5m3/hr system was established as a discharge outlet of domestic wastewater, and the experiments were performed continuously for one month. The treatment time was 1 hr in the reactors A and B, and the sedimentation time was 1 hr at the flow rate of 0.5m3/hr. To investigate the relationships among the flow rate, treated water quality, and electric power consumption, the flow rate was set to 0.5m3/hr, 1.0 m3/hr, and 1.5 m3/hr. The water samples were taken from the discharge outlet and sedimentation tank. The procedure for collecting evolved gas was the same as for the 4m3/hr system.

6.2. ANALYSIS

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T-P, T-N, TOC, COD and total solids (TS) in the wastewater, and T-P, TN, and TOC in the sludge were analyzed based on the Japanese methods of sewage testing. O2, N2, CO2, CH4, and H2 (in the collected gases) were identified by gas chromatography (TCD detector, Shimalite Q and Porapak Q column), and N2O was also analyzed (ECD detector, Porapak Q column).

6.3. MATERIALS The effluent from an anaerobic digester (EAD) of cattle wastewater (fermentation temperature: 36°C; hydraulic retention time: 30 days) and supernatant of primary sedimentation in a sewage plant (SPS) were used in the 4m3/hr system, and domestic wastewater was used in the 0.5m3/hr system.

6.4. RESULTS AND DISCUSSION 6.4.1. Water Quality The results obtained from the 4m3/hr system are illustrated in Table 3. For EAD, reactor A removed 34% of T-P, 52% of T-N, and 28% of TOC; for reactor B these values were 40%, 59%, and 53%. After 2 hours sedimentation, 87% of T-P, 96% of T-N, and 91% of TOC were removed. Reactor A removed 26% of the COD, and reactor B,40%. The sedimentation tank, however,

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removed 75% of the COD. Furthermore, reactor A removed 20% of the NH4N, and reactor B, 33%. The sedimentation tank, however, removed 94%. Table 3. Water treatment quality in wastewater with 4 m3/hr system T-P

T-N

NH4-N

TOC

COD

Wastewater

Reactor

(mg/L)

(mg/L)

(mg/L)

(mg/L)

(mg/L)

EAD

Untreated

3780 2510 (34%) 2270 (40%) 490 (87%)

4750 2290 (52%) 1950 (59%) 200 (96%)

2471 1976 (20%) 1647 (33%) 150 (94%)

16862 12080 (28%) 7886 (53) 1494 (91%)

2573 1916 (26%) 1540 (40%) 643 (75%)

6.03 2.47 (59%) 3.08 (49%) 0.56 (91%)

21.8 9.5 (56%) 12.7 (42%) 5.7 (74%)

11.0 7.1 (36%) 7.4 (33%) 3.3 (70%)

108.0 74.8 (31%) 53.1 (51%) 12.7 (88%)

31.0 27.5 (11%) 26.5 (15%) 4.1 (87%)

Reactor A Reactor B Effluent SPS

Untreated Reactor A Reactor B Effluent

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( ) indicates removal ratio.

Table 4. Water quality in wastewater with 0.5 m3/hr system Flow rate

T-P

T-N

NH4-N

TOC

COD

(m3/h)

Wastewater

(mg/L)

(mg/L)

(mg/L)

(mg/L)

(mg/L)

0.5

Untreated

0.60 0.23 (62%)

2.22 0.22 (90%)

0.32 0.00 (100%)

4.69 0.82 (83%)

1.50 0.00 (100%)

1.30 0.13 (90%)

4.32 0.34 (92%)

0.37 0 (100%)

4.68 0.81 (83%)

2.31 0.46 (80%)

0.53 0.07 (87%)

3.12 0.53 (83%)

0.21 0.02 (90%)

2.81 0.69 (75%)

4.32 0.00 (100%)

Effluent 1.0

Untreated Effluent

1.5

Untreated Effluent

( ) indicates removal ratio.

For SPS, reactor A removed 59% of T-P, 56% of T-N, and 31% of TOC; for reactor B, these values were 49%, 42% and 51%. In contrast, the sedimentation tank removed 91% of T-P, 74% of T-N and 88% of TOC. For

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NH4-N, reactor A removed 36%, reactor B 33%, and the sedimentation tank 70%. Although no change in COD was observed in reactors A and B, the sedimentation process removed 87%. As indicated in Table 4, although the design capacity of the electrochemical treatment system was 0.5m3/hr, the system removed 62 to 90% of T-P, 83 to 92% of T-N,90 to 100% of NH4-N, 75 to 83% of TOC, 80 to 100% COD in the range of 0.5 to 1.5m3/hr. Although the quality of untreated wastewater fluctuated, the removal of pollutants in wastewater remained constant, and was not influenced by increasing the flow rate from 0.5 m3/hr to 1.5 m3/hr. Moreover, the electric energy consumption for the 4 m3/hr system was 16.0 kwh, and 6.9 kwh was consumed at 0.5 m3/hr, 7.0 kwh at 1.0 m3/hr, and 8.6 kwh at 1.5 m3/hr, i.e., the electric power consumption increased slightly with an increased flow rate from 0.5 m3/hr to 1.5 m3/hr.

6.4.2. Composition of Evolved Gases

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Table 5. Composition of evolved gas with 4 m3/hr system CO2

O2

N2

H2

CH4

N2O

Wastewater

Reactor

(%)

(%)

(%)

(%)

(%)

(%)

EAD

ReactorA

6.1

51.7

37.7

0.0

4.0

0.5

ReactorB

7.3

39.7

40.3

0.0

12.2

0.4

ReactorA

0.4

78.1

15.6

0.6

0.3

5.0

ReactorB

0.1

90.1

8.4

0.1

0.0

1.2

SPS

Table 6. Composition of evolved gas with o.5 m3/hr System Flow rate (m3/h)

Reactor

N2 (%)

O2 (%)

CO2 (%)

N2O (%)

0.5

Reactor A

71.0

26.3

2.1

0.6

Reactor B

68.3

27.9

2.8

1.1

Reactor A

76.1

22.8

0.7

0.4

Reactor B

74.7

22.0

1.1

2.2

Reactor A

74.5

23.7

1.4

0.4

Reactor B

71.8

24.8

1.8

1.6

1.0

1.5

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The composition of evolved gases in the 4m3/hr system is listed in Table 5. For EAD, reactor A had 6.1% CO2, 51.7% O2, 37.7% N2, 0.0% H2, 4.0% CH4, and 0.5% N2O; for reactor B, these values were 7.3% CO2, 39.7% O2, 40.3% N2, 0.0% H2, 12.2% CH4, and 0.4% N2O. It is obvious that the ratios of CO2, N2, and CH4 from reactor B exceeded those from reactor A, but that O2 from reactor B was less than that of reactor A. However, for SPS, reactor A had 0.4% CO2, 78.1% O2, 15.6% N2, 0.6% H2, 0.3% CH4, and 5.0% N2O, while reactor B had 0.1% CO2, 90.1% O2, 8.4% N2, 0.1% H2, 0.0% CH4, and 1.2% N2O. The ratios of CO2, N2, and CH4 when treated with the SPS were less than in EAD. This is probably due to the fact that more organic matter was contained in EAD than in SPS. Furthermore, the SPS was opposite to the EAD, i.e., the ratios of CO2, N2, H2, and CH4 from reactor B were less than in reactor A, and the O2 from reactor B was more than that from reactor A. This may be due to the difference in experiment conditions and compositions between EAD and SPS, and should be examined in the future. The composition of evolved gases in the 0.5m3/hr system is presented in Table 6. Within the flow rate range of 0.5 to 1.5m3/hr, the gas composition was 68.3 to 76.1% N2, 22.0 to 27.9% O2, 0.7 to 2.8% CO2, and 0.4 to 2.2% N2O.

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6.4.3. Composition of Sludge Table 7. Composition in sludge with 4 m3/hr system Wastewater EAD

SPS

Reactor Reactor A Reactor B Sedimentation tank Reactor A Reactor B Sedimentation tank

T-P (%) 11.3 9.6 6.3 0.3 0.7 1.2

T-N (%) 5.3 8.1 5.3 4.0 12.8 6.1

TOC (%) 25.8 23.1 27.2 14.1 24.2 23.8

TS (g/L) 53.0 35.3 53.0 60.0 20.0 43.0

Table 7 indicates the composition of sludge in the 4m3/hr system. The order of TOC, T-P, and T-N in the sludge was TOC>T-P>T-N for EAD, and TOC>T-N>T-P for SPS. TOC in sludge was greatest in both EAD and SPS. Furthermore, TS in reactor A was larger than in reactor B.

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6.4.4. Discussion The above results demonstrate that both electrochemical treatment systems could efficiently treat wastewater such as EAD, SPS, and domestic wastewater. In the 4 m3/hr system, the removal of T-P in reactor A was greater than in reactor B for SPS, and slightly increased in reactor B for EAD. However, the removals of TOC, NH4-N, and BOD in reactor A were less than in reactor B. Additionally, TS in the sludge obtained from both EAD and SPS in reactor A exceeded that in reactor B. We can conclude that T-P was removed mainly by electrocoagulation in the sludge due to the formation of metal ions from electrodes. Sedimentation was subsequently enhanced in reactor B, resulting in greater removal of T-P in the effluent. This was clarified by estimating the mass balance (no data shown). TOC and NH4-N were decomposed mainly by electrochemical and ultrasonic treatment. Organic matter in wastewater was decomposed during electrochemical treatment because of oxidant formation such as ozone (Peter et al., 1982), and H2O2 (Brillas et al., 1995). Generally, hydroxyl radicals (・OH) (Comninellis et al., 1994; Polcaro et al., 1999) are produced by anodic catalyst or highvoltage pulses (Sun et al., 2000). This reaction is given as equation (1).

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H2O → ·OH + H+ + e-

(1)

The reaction given in equation (2)( Vega-Mercado et al., 1997; Vlyssides et al., 1999) may occur because of the presence of chlorides in EAD and SPS. H2O + Cl- →·ClOH + H+ + 2e-

(2)

Furthermore, ultrasound applied in an aqueous solution causes cavitations, which grow and collapse at high pressure. These collapses produce physical and chemical effects. The physical effect cleans the surfaces of electrodes, enhancing treatment efficiency. However, radical species such as ·OH, ·O, HOO· are produced by the chemical effect (Peterier, 1997), as illustrated in equations (3), (4), (5) and (6). H2O → H· + ·OH

(3)

O2 → 2O·

(4)

H· + O2 → HOO·

(5)

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O· + H2O → 2· OH

(6)

ClOH or ·OH were also confirmed using p-nitrosodimethylaniline (RNO) in the previous works [17]. Therefore, organic matter (R) is oxidized by ·OH or ·ClOH according to equations (7) and (8). Reactions (7) and (8) will progress continuously if the electrochemical process continues. Finally, the organic matter was completely mineralized to gases such as CO2, N2 or N2O. R +・OH → RO + H+ + e-

(7)

R +・ClOH → RO + H+ + Cl- + e-

(8)

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In addition, ammonia in wastewater was oxidized by ·OH or ·ClOH (Szpyrkowicz, 1995). 2NH4 + 6 ·OH → N2 + 6H2O + 5H+ + 6e-

(9)

2NH4+ + 3·ClOH → N2 + 3H2O + 5H+ + 3Cl-

(10)

Observing the composition of evolved gases during the treating of EAD and domestic wastewater reveals why reactor B removed more NH4-N and TOC than did reactor A. However, SPS, CO2, N2, CH4 and N2O contained in the evolved gases from reactor B were less than those from reactor A. This is probably due to the fact that the suspended solid in wastewater hindered the action of the pulse and ultrasonic action when the experimental conditions changed. In other words, the electrochemical treatment system needs to put reactor A first, then reactor B, and then the sedimentation process. Furthermore, NH4-N and TOC were efficiently removed in the sedimentation tank, which may be due to the residue effect of the radical species. This should be investigated in detail in the future. In addition, electrocoagulation occurs during electrochemical treatment for wastewater. The reactions may take place when using a metal anode such as iron (Koparal et al., 2002).

At the cathode,

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59

(11)

At the anode, Fe → Fe2+ + 2e.

(12)

In the solution, Fe2+ + 2OH- → Fe(OH)2.

(13)

Fe(OH)2 reacts with dissolved oxygen in wastewater, Fe(OH)2 → Fe(OH)3.

(14)

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Removal of various pollutants such as heavy metals (Carl, 1995) and nitrate (Koparal, 2002) by electrocoagulation has been studied. In this study, phosphorus included in wastewater was removed by the electrocoagulation mentioned above.

6.5. CONCLUSIONS Two types of electrochemical wastewater treatment systems incorporating an ultrasonic process have been developed, and their treatment abilities were demonstrated using EAD, SPS and domestic wastewater. The mechanism of the treatment system was also analyzed by the quality of treated wastewater and the composition of evolved gas. As a result, higher removals of T-P, T-N, NH4-N, TOC and COD were obtained by electrochemical treatment. Furthermore, the high-voltage pulse and ultrasonic treatment effectively decomposed NH4-N and TOC due to the formation of radical species such as hydroxyl radical and hypochlorite, and T-P was removed by the electrocoagulation. Furthermore, the two treatment systems must be compared with various biological processes through economic analysis for popularizing the electrochemical process in the future.

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INDEX

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A abatement, 65 acetic acid, 4 acid, 3, 4, 23, 28, 31, 37, 64 active oxygen, 7, 22 adsorption, 2 advantages, 3 AFM, 17 agricultural sector, 1 algae, viii, 5, 39, 43, 47, 48, 49 ammonia, vii, 2, 3, 4, 5, 25, 26, 27, 28, 29, 30, 31, 33, 35, 36, 37, 38, 45, 47, 49, 58, 62, 63, 64 ammonium, vii, 45, 66 argon, 42, 53 aromatic compounds, vii, 11 Aromatic compounds, 1 aromatic hydrocarbons, 1 atomic force, 14

B beet molasses, 66 biological processes, 59 bismuth, 63 bleaching, 2, 13, 23, 24, 42, 45, 46 blue baby, 1 breeding, 1 by-products, 4, 37

C cancer, 1 carbon, 4, 62, 63, 64 carbon dioxide, 4 catalysis, 7, 22, 52, 65 catalyst, 26, 57 catalytic activity, 49 cation, 23 cattle, vii, 5, 51, 53 China, 66 chlorides, 2, 7, 22, 47, 57 chlorinated hydrocarbons, 1 chlorination, 27, 65 chlorine, 63 chlorine distribution, 63 chlorophyll, viii chromatography, 42, 53 CO2, 2, 4, 8, 15, 23, 42, 48, 53, 55, 56, 58, 65 coagulants, 2 coatings, 3 combustion, 14, 46, 62 complexity, 3 composition, 56, 58, 59 compounds, 1, 2, 16 conductivity, 4, 7, 15 consumption, 2, 20, 53, 55, 62 contaminant, 1

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Index

68 copper, 64 cost, 2, 3 cyclic voltammetry, 13, 28, 42, 63 cycling, 33

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D degradation, vii, 3, 11, 12, 14, 15, 17, 19, 20, 22, 23, 24, 25, 26, 27, 28, 29, 30, 31, 62, 63, 64, 65 degradation rate, 14 denitrification, 2, 5, 26, 62 deposition, 3 detection, 39, 49 detoxification, 65 diffusion, 5 dimensionally stable anode (DSA), 3 discharges, 1, 65 disinfection, 63 dissociation, 20 dissolved oxygen, 59 distilled water, 13, 34 dosage, 12, 13, 15, 31 drinking water, 1 durability, 52 dyes, 64

E economy, 1, 15, 31 Efficiency, 19 effluent, 1, 40, 43, 53, 57, 61, 64 effluents, vii, 3, 5, 51, 62 effluents from an anaerobic digester (EAD), vii, 5, 51 electric field, 8 electrochemical cell, 11 Electrochemical degradation, vii, 15, 17, 19, 64, 65 electrochemical method, vii, 2, 5, 64 electrochemical treatment systems, vii, 5, 51, 57 electrochemistry, 65 electrocogulation, viii

electrodes, 3, 11, 13, 16, 33, 41, 42, 45, 47, 52, 57, 61, 62, 64, 66 electrolysis, vii, 4, 11, 12, 13, 14, 15, 16, 17, 18, 19, 21, 22, 23, 26, 31, 33, 34, 35, 36, 37, 38, 47, 61, 63 electrolyte, 4, 12, 13, 14, 15, 16, 17, 19, 20, 21, 24, 31, 34, 61, 63, 65 electron, 23 electrons, 26 electrooxidation, viii electroreduction, 63 endocrine, 2 energy consumption, 2, 55 environmental regulations, 2 equipment, 62 European Union, 1 experimental condition, 58

F factories, vii, 1 fermentation, 53 fiber, 62 films, 3, 16, 63 fish, 1 fluctuations, 27, 29 fouling, 18, 66 free radicals, 8, 13

G galvanostatic control, 12, 13 graphite, 3 groundwater, 5, 34, 35

H health problems, 1 heavy metals, 9, 59, 61 height, 52 hexavalent chrominum, 2 hog raising, viii, 5, 39, 43, 48, 49 hydrazine, 4, 34 hydrogen, 20, 36, 62

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Index hydrogen peroxide, 62 hydrolysis, 4 hydroquinone, 16 hydroxyl, 4, 7, 8, 13, 17, 20, 22, 23, 25, 39, 41, 45, 47, 49, 57, 59 Hydroxylamine, 67 Hypochlorous acid (HOCl), 4

I ice, 39 ideal, 26, 29 Inductively Coupled Plasma Mass Spectrometry (ICPMS), 34 infants, 1 initial state, 27 instantaneous current efficiency (ICE), 21 ion exchange, 2, 65 ions, 3, 20, 45, 47, 57, 61, 64 iridium, 45 iron, 41, 52, 58

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J

69

metals, 45, 61 meter, 14 microscopy, 14 molecular oxygen, 18, 22 molecules, 22 monitoring, 65 morphology, 14

N Na2SO4, 4, 14, 24, 34, 35 NaCl, vii, 4, 8, 12, 13, 14, 15, 16, 17, 18, 19, 20, 23, 24, 25, 28, 29, 30, 31, 34, 36, 37, 38, 41, 46, 62 nickel, 3, 64 nitrate, vii, 1, 2, 4, 5, 9, 26, 27, 28, 29, 30, 31, 33, 35, 36, 37, 38, 59, 61, 62, 63, 64, 65, 66 nitrification, 2 nitrogen, 1, 2, 4, 25, 26, 27, 28, 29, 30, 31, 35, 37, 38, 48, 63, 64, 65 nitrogen compounds, 5 nitrogen gas, 4, 26, 37, 48 noble metals, 3

Japan, 11, 33, 42, 62, 63

O L lakes, 1 lifetime, 2, 17 liquid interfaces, 47 liver, 1 liver damage, 1 livestock, 1 LTD, 41, 52

oil, 63 organic matter, 7, 8, 21, 22, 39, 52, 56, 58 osmosis, 2 oxidation, 3, 4, 5, 7, 13, 16, 17, 20, 21, 23, 25, 26, 27, 28, 30, 31, 33, 36, 45, 46, 47, 48, 49, 61, 62, 63, 65, 66 oxide electrodes, 62 oxidizability, 7, 17, 27, 39 oxygen, 17, 21, 23, 31, 45, 61 ozone, 2, 57

M maximum admissible concentration (MAC), 1 maximum contaminant level (MCL), 1 media, 63 metal oxides, 3

P palladium, 45 parallel, vii, 33, 35, 38 passivation, 63 pathways, 23, 64

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Index

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70

performance, 3, 11, 14, 16, 31, 39, 49, 51, 62, 64, 65 peristaltic pump, 11, 39 pesticide, 1 pesticides, 1, 2 petroleum refineries, vii, 1 pH, 13, 18, 19, 20, 29, 30, 31 pharmaceutical factories, vii, 1 phenol, vii, 1, 3, 4, 5, 11, 12, 13, 14, 15, 16, 17, 18, 19, 20, 21, 22, 23, 25, 26, 27, 28, 29, 30, 31, 61, 62, 63, 64, 65, 66 phenol oxidation, 16, 18, 23, 66 phosphorus, 1, 9, 59 plants, vii, 1 platinum, 39, 42, 43, 44, 45, 49, 62, 63, 66 pollution, 1, 65 polyaluminum chloride (PAC), 43 polymer, 61 polymeric products, 17 polymerization, 16 potable water, 1 potassium, 27 potential waveform,, 42 PTFE, 61 purification, 63

S screening, 43 sedimentation, vii, 5, 40, 43, 51, 52, 53, 54, 58 sedimentation in a sewage plant (SPS), vii, 5, 51, 53 selectivity, 5, 45 service life, 3, 4 sewage, vii, 5, 42, 51, 53 shape, 41 sludge, 2, 42, 47, 48, 53, 56, 57 sodium, 12, 31, 36, 64 sodium hydroxide, 64 solid waste, 1 species, 2, 47, 57, 58, 59 spectrophotometer, 13 speculation, 26 steel, 41, 52 substrates, 3, 63, 65 Sun, 8, 57, 64, 65 surface treatment, 31 syndrome, 1

T Q quartz, 66

R radicals, 4, 7, 8, 13, 16, 17, 21, 22, 23, 39, 41, 45, 47, 49, 57 reaction mechanism, 3 reaction rate, 46 reactions, 8, 9, 20, 23, 26, 48, 58, 63 removals, 28, 29, 30, 48, 57, 59 resins, 65 reverse osmosis, 2 rhodium, 45 RNO bleaching, 2, 24 runoff, 1

tanks, 40 tar, 18 TEDLAR pack, 42, 52 temperature, 39, 40, 53 testing, 53 thermal decomposition, 40 tin, 63 titanium, 3, 28, 36, 39, 42, 43, 44, 45, 49, 52, 63 total solids (TS), 53 toxicity, 4 transition metal, 45, 62

U ultrasound, 57 unique features, 3

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Index

V vinasse, 66

71

wastewater treatment, vii, 3, 48, 59, 62, 63, 64, 65 water quality, 53 water resources, 1 workstation, 13

W Z zinc, 34, 64

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waste, 1, 61, 65 wastewater, vii, 1, 2, 3, 5, 8, 9, 12, 23, 39, 40, 41, 42, 43, 46, 47, 48, 49, 51, 53, 54, 55, 57, 58, 59, 61, 62, 63, 64, 65, 66

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