**Meet the editor**

Dr. Robina Farooq has been involved in teaching, research, management and academic work in numerous distinguished universities of *Britain*, *China* and *Pakistan for the last 27 years*. Currently, she is serving the COMSATS Institute of Information Technology, Lahore, Pakistan. She discovered innovative and low-cost processes for the treatment of wastewater. She is the author of scientific manu-

scripts, books and book chapters and was granted patents by USPTO, USA. She is the recipient of Best Innovator, Best University Teacher and Productive Scientist Awards. She worked on projects including ultrasonic decomposition of pollutants, phytoremediation of wastewater, bioelectrochemical synthesis of renewable fuel, bioelectrochemical decomposition of wastewater and energy recovery, recovery of heavy metals from effluents, microbial fuel cell technology for wastewater remediation and retrieval of precious metals from printed circuit boards.

Dr. Zaki Ahmad is a Professor Emeritus of King Fahd University of Petroleum and Minerals, Saudi Arabia, and an adjunct professor at COMSATS Institute of Information Technology, Lahore. He is the fellow of IOM3, UK, and a chartered engineer of the UK Engineering Council. He is a member of the European Federation of Corrosion. He is the recipient of best researcher award by Energy Exchange

in 2011. He is the author and editor of six books including the popular text book entitled Principles of Corrosion Engineering and Corrosion Control published by Elsevier at international level and over 150 research papers. His projects in nanotechnology, green engineering, and harvesting water from air incorporate human values.

## Contents



Magdalena Urbaniak and Anna Wyrwicka

#### **X** Contents


Guzmán and Karla Alejandra Suárez-Meraz

Chapter 12 **Social Perspectives on the Effective Management of Wastewater 253**

Dalia Saad, Deirdre Byrne and Pay Drechsel

## Preface

**Section 3 Resource Recovery and Management 133**

**Resource Recovery 153**

Taner Yonar

**VI** Contents

**Technologies 213**

**By-Products 239**

**Wastewater 253**

Chapter 7 **Phosphorus Recovery by Struvite Crystallization from Livestock Wastewater and Reuse as Fertilizer: A Review 135**

Tao Zhang, Rongfeng Jiang and Yaxin Deng

Ciro Bustillo-Lecompte and Mehrab Mehrvar

Chapter 10 **Wastewater Treatment through Low Cost Adsorption**

Guzmán and Karla Alejandra Suárez-Meraz

Chapter 12 **Social Perspectives on the Effective Management of**

Dalia Saad, Deirdre Byrne and Pay Drechsel

Gonzalo Montes-Atenas and Fernando Valenzuela

Chapter 11 **Integral use of Nejayote: Characterization, New Strategies for Physicochemical Treatment and Recovery of Valuable**

**Oxidation Processes (AOPs) 175**

Chapter 9 **Treatment of Antibiotics in Wastewater Using Advanced**

Ayse Kurt, Berna Kiril Mert, Nihan Özengin, Özge Sivrioğlu and

Eduardo Alberto López-Maldonado, Mercedes Teresita Oropeza-

Chapter 8 **Slaughterhouse Wastewater: Treatment, Management and**

This volume covers the physico-chemical treatment technologies for wastewater treatment and recovery of materials, energy and water as a resource. It mainly consists of processes including oxidation, adsorption and management of persistent pollutants in wastewater. It consists of three sections. Oxidation of contaminants in wastewater is well placed in Section 1. The chapter 'Removal of BTX Contaminants with O3 and O3/UV Process' describes a promising procedure for the removal of BTX components from aquatic environment. A kinetic model for pollutant degradation by UV/H2O2 is developed and validated in the chapter 'Kinetic Modelling of the UV/H2O2 Process: Determining the Effective Hydroxyl Radical Concentration'. Fenton reagent and electrolysis process are used to generate hydroxyl radicals in the chapter 'Electrooxidation-Ozonation: A Synergistic Sustainable Treatment Process'. Main parameters to control the proc‐ ess are addressed with examples. Iron-containing active phase that was deposited on natural layered silicate (vermiculite) using several techniques is discussed in the chapter 'Removal of Phenol from Wastewater Using Fenton-Like Reaction over Iron Oxide–Modified Silicates'. This chapter also explains experimental methods for the treatment of model industrial effluent of phenol. The chapter 'Treatment of Antibiotics in Wastewater Using Advanced Oxidation Proc‐ ess' presents an overview of the literature on antibiotics and their removal from water by ad‐ vanced oxidation processes. It was found that most of the investigated AOP methods for the oxidation of antibiotics in water are direct and indirect photolyses with the combinations of H2O2, TiO2, ozone and Fenton's reagent.

Section 2, 'Micropollutants and Their Removal Processes', describes studies about the micro‐ pollutants entering into the environment after conventional wastewater treatment facilities. Its chapter 'Micropollutants in Wastewater: Fate and Removal Processes' explains the esti‐ mation of micropollutants by physical properties instead of using tedious and cost-intensive analysis. The section includes information concerning wastewater treatment plants function‐ ing in respect to PCDDs/PCDFs and PCBs. Fate and impact of these pollutants are extensive‐ ly elaborated in 'PCDDs/PCDFs and PCBs in Wastewater and Sewage Sludge'.

Section 3, 'Resource Recovery and Management', is designated for recovery of chemicals and management of wastewater. The study about 'phosphorus recovery by struvite crystal‐ lization from livestock wastewater and reuse as fertilizer' represents that struvite crystalliza‐ tion is a promising tool for recovering phosphorus from livestock wastewater.

The chapter 'Slaughterhouse Wastewater: Treatment, Management and Resource Recovery' draws the attention of readers and researchers towards the wastewater of meat processing industry. In this chapter, the regulatory framework relevant to the SWW management, envi‐ ronmental impacts, health effects and treatment methods is discussed. The chapter 'Waste‐ water Treatment via Low-Cost Adsorption Technologies' addresses the wastewater treatment of mining residues for removal of metals via adsorption methodologies, along with classic models of adsorption thermodynamics and kinetics that are well presented.

The chapter 'The Integral Use of Nejayote: Characterization, New Strategies for Physicochemical Treatment and Recovery of Valuable Products' describes physico-chemical treat‐ ment strategy. It addresses the problem of *nejayote* and its reuse using waste from shrimp

shells. The chapter 'Effective Management of Wastewater for Future Water Security "Social Perspectives"' discusses how the adopting of a holistic methodology that acknowledges so‐ ciological factors including community participation and public involvement, social percep‐ tion, attitudes, impression, gender roles and public acceptance would lead to improvements in wastewater management practice.

I would like to express my gratitude to Prof. Dr. Zaki Ahmad who started to work with me as coeditor. Prof. Zaki passed away during his work on this book. His efforts and contribu‐ tions are highly appreciated, and his services as book editor are highly acknowledged.

I would like to thank Ms. Martina Usljebrka, Publishing Process Manager, for enabling me to publish this book. I want to thank my husband, Prof. Dr. Saleem Farooq Shaukat, my daughter Kinza Farooq, sons Abdul Basit and Faisal Farooq and grandchildren Zoha Fatima and Aarib Basit who kept me motivated to accomplish this work. I am grateful for my father Mr. Muhammad Mukhtar and my mother Mrs. Rafia Mukhtar, my sisters and my brothers who always supported and encouraged me throughout my life.

> **Prof. Dr. Robina Farooq** Department of Chemical Engineering COMSATS Institute of Information Technology Lahore, Pakistan

> > **Prof. Dr. Zaki Ahmad**

COMSATS Institute of Information Technology Lahore, Pakistan

#### **Note from the publisher**

It is with great sadness and regret that we inform the contributing authors and future read‐ ers of this book that the editor, Prof. Zaki Ahmad, passed away during his work on publica‐ tions and before having a chance to see them.

Prof. Ahmad was InTech's long-term collaborator and edited his first book with us in 2011 *Recent Trends in Processing and Degradation of Aluminium Alloys*, followed by publications 'Aluminium Alloys - New Trends in Fabrication and Applications', 'New Trends in Alloy Development, Characterization and Application' and 'High Temperature Corrosion'. This fruitful collaboration continued until his final days when he was acting as a coeditor on the books *Biological Wastewater Treatment and Resource Recovery* and *Physico-chemical Treatment of Wastewater and Resource Recovery*.

We would like to acknowledge Dr. Zaki Ahmad's contribution to open access scientific pub‐ lishing, which he made during 6 years of dedicated work on edited volumes, and express our gratitude for his always pleasant cooperation with us.

InTech Book Department Team

**Oxidation of Contaminants in Wastewater**

shells. The chapter 'Effective Management of Wastewater for Future Water Security "Social Perspectives"' discusses how the adopting of a holistic methodology that acknowledges so‐ ciological factors including community participation and public involvement, social percep‐ tion, attitudes, impression, gender roles and public acceptance would lead to improvements

I would like to express my gratitude to Prof. Dr. Zaki Ahmad who started to work with me as coeditor. Prof. Zaki passed away during his work on this book. His efforts and contribu‐ tions are highly appreciated, and his services as book editor are highly acknowledged.

I would like to thank Ms. Martina Usljebrka, Publishing Process Manager, for enabling me to publish this book. I want to thank my husband, Prof. Dr. Saleem Farooq Shaukat, my daughter Kinza Farooq, sons Abdul Basit and Faisal Farooq and grandchildren Zoha Fatima and Aarib Basit who kept me motivated to accomplish this work. I am grateful for my father Mr. Muhammad Mukhtar and my mother Mrs. Rafia Mukhtar, my sisters and my brothers

It is with great sadness and regret that we inform the contributing authors and future read‐ ers of this book that the editor, Prof. Zaki Ahmad, passed away during his work on publica‐

Prof. Ahmad was InTech's long-term collaborator and edited his first book with us in 2011 *Recent Trends in Processing and Degradation of Aluminium Alloys*, followed by publications 'Aluminium Alloys - New Trends in Fabrication and Applications', 'New Trends in Alloy Development, Characterization and Application' and 'High Temperature Corrosion'. This fruitful collaboration continued until his final days when he was acting as a coeditor on the books *Biological Wastewater Treatment and Resource Recovery* and *Physico-chemical Treatment of*

We would like to acknowledge Dr. Zaki Ahmad's contribution to open access scientific pub‐ lishing, which he made during 6 years of dedicated work on edited volumes, and express

**Prof. Dr. Robina Farooq**

Lahore, Pakistan

Lahore, Pakistan

**Prof. Dr. Zaki Ahmad**

Department of Chemical Engineering

COMSATS Institute of Information Technology

COMSATS Institute of Information Technology

who always supported and encouraged me throughout my life.

in wastewater management practice.

VIII Preface

**Note from the publisher**

*Wastewater and Resource Recovery*.

InTech Book Department Team

tions and before having a chance to see them.

our gratitude for his always pleasant cooperation with us.

#### **Removal of BTX Contaminants with O3 and O3/UV Processes Removal of BTX Contaminants with O<sup>3</sup> and O<sup>3</sup> /UV Processes**

Ján Derco, Katarína Šimovičová, Jozef Dudáš and Mária Valičková Ján Derco, Katarína Šimovičová, Jozef Dudáš and Mária Valičková

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65889

#### **Abstract**

The legal basis for the monitoring of priority and priority hazardous substances in water, sediment, and biota follows from Directive 2008/105/EC which defines the good chemical status to be achieved by all Member States together with the Water Framework Directive 2000/60/EC. The BTX compounds are considered to be the most toxic components of gasoline. Thus, organic petroleum components can induce a serious problem to public health and the aquatic environment. The effect of ozone and ozone/UV on degradation of the BTX in a model water was studied. The results indicate that the highest BTX removal rates were observed during the first 5 min of the process for all investigated pollutants. The treatment efficiencies above 90% were observed in all investigated pollutants after 40 min of ozonation. The results show a significant proportion of stripping in the removal of BTX components. Higher overall efficiency was observed by O<sup>3</sup> /UV process after abstracting share of stripping process. Application of investigated processes appears to be a promising procedure for removal of petrol aromatic hydrocarbons from aquatic environment. However, for practical application, an improvement of process removal efficiency and investigation of impact of ozonation intermediates and products on aquatic microorganisms are required.

**Keywords:** BTX, jet‐loop reactor, ozone, O<sup>3</sup> /UV, wastewater treatment

## **1. Introduction**

The adoption of the Framework Directive on water [1] provides a policy tool that enables sustainable protection of water resources. The Decision No. 2455/2001/EC of the European Parliament and the Council of November 2001 [2] established the list of 33 priority substances

and reproduction in any medium, provided the original work is properly cited.

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

or group of substances, including the priority hazardous substances, presenting a significant risk to water pollution or *via* the aquatic environment including such risks to water used for the abstraction of drinking water.

Hazardous substances are defined as substances or groups of substances that are toxic, persistent, and liable to bioaccumulation, and other substances or groups of substances which give rise to an equivalent level of concern. The EC member countries have extended this list with relevant pollutants for individual countries. Thus, in the supplement of the Water Act [3] there have been identified altogether 59 relevant substances for SR.

The BTEX contaminants consist of benzene, ethyl benzene, toluene, and three isomers of xylene. These compounds are the volatile organic compounds (VOCs) found in petroleum derivatives such as petrol (gasoline). They represent one of the main groups of soluble organic compounds that is present in wastewater from refinery. They are the most toxic components of gasoline. These substances can lead to serious health problems ranging from irritation of eyes, skin, and mucous membranes and ending with weakened nervous system, decreased bone marrow function, and cancer. Benzene in particular is highly toxic. The World Health Organization classifies the substance as carcinogenic. It is also on the list of the priority substances [4].

Many oil substances have acute toxic effect on aquatic microorganisms with possible chronic consequences [5].

Commonly used wastewater treatment processes usually apply physical and physiochemical processes. Thus, the discharge of organic pollutants may create some environmental problems, particularly at microlevel. The aromatic oil fraction consists mainly of polyaromatic hydrocarbons (PAHs) and is more toxic and persistent than the aliphatic hydrocarbons [6]. Leakages including release of petroleum products, e.g., gasoline, diesel fuel, lubricating, and heating oil, from leaking oil tanks are the most frequent sources of soil and groundwater contaminations with BTEX substances. They are polar and readily soluble, and thus they are able to penetrate into soil and groundwater and cause serious environmental problems. These compounds dispose of acute and long‐term toxic effects [4].

BTEX are among the most frequently detected contaminants in US public drinking‐water systems that rely on groundwater sources [7]. These organic compounds make up a significant percentage of petroleum. The most contaminated locality of hazardous BTEX substances in the Slovak Republic is the airport at Sliač, Sliač‐Vlkanová territory, contaminated by the Soviet Army, and the gas station in Rajecké Teplice where BTEX contaminants were identified as dominant in the groundwater. BTEX were also found in groundwater in Bratislava due to poor technical conditions of technological equipment (old stocks of aviation fuel) and the subsequent uncontrolled release of oil into the rock mass at the Airport of M. R. Štefanik [8].

Ozone is a very powerful oxidizing agent (*E*° = 2.07 V). Ozone may react with organic compounds in two ways: by direct reaction as molecular ozone or by indirect reaction through formation of secondary oxidants like free radical species [9–11]. In practice, both mechanisms may occur depending on the type of chemical wastewater pollution.

At low pH, the predominant reaction mechanisms are the direct electrophilic attack by molecular ozone [12], i.e., ozonolysis.

Under such conditions, ozone is a selective oxidant and reacts with multiple bonds (C=C, C=N, N=N, etc.) , but only at low rates with single bonds (C-C, C-O, O-H). At high pH, indirect reaction occurs, i.e., organics are degraded by secondary oxidants/chain reaction involving powerful radicals including OH, which are produced by ozone decomposition. These radicals are very strong and nonselective oxidants. Hydroxyl radicals can be formatted by increasing pH or by decomposition of O<sup>3</sup> with homogeneous and heterogeneous catalysts.

The main goal of our research was to study the feasibility of ozone and combination of O3 /UV processes for removal of selected benzene, toluene, and xylenes (BTX) from water/ wastewater. Investigation of process kinetics and stripping of volatile substances were also performed.

## **2. Experimental methods**

or group of substances, including the priority hazardous substances, presenting a significant risk to water pollution or *via* the aquatic environment including such risks to water used for

Hazardous substances are defined as substances or groups of substances that are toxic, persistent, and liable to bioaccumulation, and other substances or groups of substances which give rise to an equivalent level of concern. The EC member countries have extended this list with relevant pollutants for individual countries. Thus, in the supplement of the Water Act [3]

The BTEX contaminants consist of benzene, ethyl benzene, toluene, and three isomers of xylene. These compounds are the volatile organic compounds (VOCs) found in petroleum derivatives such as petrol (gasoline). They represent one of the main groups of soluble organic compounds that is present in wastewater from refinery. They are the most toxic components of gasoline. These substances can lead to serious health problems ranging from irritation of eyes, skin, and mucous membranes and ending with weakened nervous system, decreased bone marrow function, and cancer. Benzene in particular is highly toxic. The World Health Organization classifies the substance as carcinogenic. It is also on the list of the priority sub-

Many oil substances have acute toxic effect on aquatic microorganisms with possible chronic

Commonly used wastewater treatment processes usually apply physical and physiochemical processes. Thus, the discharge of organic pollutants may create some environmental problems, particularly at microlevel. The aromatic oil fraction consists mainly of polyaromatic hydrocarbons (PAHs) and is more toxic and persistent than the aliphatic hydrocarbons [6]. Leakages including release of petroleum products, e.g., gasoline, diesel fuel, lubricating, and heating oil, from leaking oil tanks are the most frequent sources of soil and groundwater contaminations with BTEX substances. They are polar and readily soluble, and thus they are able to penetrate into soil and groundwater and cause serious environmental problems. These

BTEX are among the most frequently detected contaminants in US public drinking‐water systems that rely on groundwater sources [7]. These organic compounds make up a significant percentage of petroleum. The most contaminated locality of hazardous BTEX substances in the Slovak Republic is the airport at Sliač, Sliač‐Vlkanová territory, contaminated by the Soviet Army, and the gas station in Rajecké Teplice where BTEX contaminants were identified as dominant in the groundwater. BTEX were also found in groundwater in Bratislava due to poor technical conditions of technological equipment (old stocks of aviation fuel) and the subsequent uncontrolled release of oil into the rock mass at the Airport of M. R. Štefanik [8]. Ozone is a very powerful oxidizing agent (*E*° = 2.07 V). Ozone may react with organic compounds in two ways: by direct reaction as molecular ozone or by indirect reaction through formation of secondary oxidants like free radical species [9–11]. In practice, both mechanisms

there have been identified altogether 59 relevant substances for SR.

compounds dispose of acute and long‐term toxic effects [4].

may occur depending on the type of chemical wastewater pollution.

the abstraction of drinking water.

4 Physico-Chemical Wastewater Treatment and Resource Recovery

stances [4].

consequences [5].

#### **2.1. Experimental equipment**

Ozonation trials were performed in a laboratory ozonation reactor. A schematic of the ozonation apparatus is illustrated in **Figure 1** [13]. Ozonation jet‐loop reactor was operated in batch mode with regard to wastewater and in continuous mode with regard to gas. Active volume of the reactor was 3.5 L. The treated wastewater was transferred into ozonation reactor before starting operation of the reactor. A membrane pump was used to maintain external circulation of liquid reaction mixture. Pulsation of recirculated external flow was minimized with diaphragm pulsation damper (SERA 721.1, Seybert & Rahier, Immenhausen, Germany). Ozone was generated using a Lifetech generator with maximum production of 5 g h-1. Ozone was generated at 50% of the maximum ozone generator's power and maintaining continuous oxygen flow of 60 L h-1. A mixture of O<sup>3</sup> and O<sup>2</sup> was injected into a wastewater sample through a Venturi ejector. At the same time, the ejector sucked the mixture of O<sup>3</sup> and O<sup>2</sup> from the reactor headspace.

This, together with external circulation, should improve the efficiency of ozone utilization in the ozonation reactor. Pen‐Ray UV lamp with wavelength 254 nm was used to generate hydroxyl radicals in the reactor. The outfall of reaction-gas mixture was transported into a destruction glass column by a fine‐bubble porous distributive device. The destructive reactive column contained a potassium iodide solution. The active volume of the destructive reactive column was 1.0 dm3 . An excess O<sup>3</sup> was destructed in this device [13].

#### **2.2. Analytical methods**

For determination BTEX compounds in model wastewater, gas chromatography was used with MS detector in connection with headspace autosampler.

**Figure 1.** Schematic scheme of the ozonation equipment. 1—O<sup>2</sup> , 2—O<sup>3</sup> generator, 3—O<sup>2</sup> /O<sup>3</sup> gas flow, 4—UV detector of O3 , 5—O<sup>3</sup> reactor, 6—Venturi nozzle, 7—UV source, 8—tempering jacket, 9—recirculation pump, 10—external reaction mixture recirculation loop, 11—sucking of gas from head space, 12—sampling device, 13—outfall of gas, 14—destruction device for residual O<sup>3</sup> .

Gas chromatographic method after static headspace extraction was used for quantification of the BTX in water. Headspace part was analyzed by gas chromatography with MS detector (Agilent Technologies 7890A GC Systems). All substances used for preparation of model wastewater and standard stock solutions were purchased from Dr. Ehrenstorfer.

#### **2.3. Processing of experimental data**

Desorption of volatile pollutants from water solution is proportional to its concentration in water solution

$$\frac{dS}{dt} = -k \ . \ S \tag{1}$$

where k/h-1 is desorption rate.

Experimental data of BTX degradation were fitted by the zero (Eq. (2))‐, first (Eq. (3)‐, and second (Eq. (4))‐order reaction kinetic models. For a batch reaction system, under the assumption of a constant reaction volume, the following relationships were obtained.

$$S\_r = S\_0 - k\_\text{o}t \tag{2}$$

Removal of BTX Contaminants with O3 and O3/UV Processes http://dx.doi.org/10.5772/65889 7

$$S\_r = S\_0 \exp\left(-k\_\mathrm{i} t\right) \tag{3}$$

$$\text{L } S\_i = \frac{S\_0}{\left(1 + S\_0 k\_2 t\right)}\tag{4}$$

where S*<sup>t</sup>* (g m–3) stands for the content of BTX substances in model wastewater in time *t*, S0 (g m–3) is the beginning content of BTX substances in model wastewater, *k*<sup>0</sup> (g m–3 h–1), *k*<sup>1</sup> (*h*–1), and *k*<sup>2</sup> (g–1 m3 h–1) are the rate constants for the kinetics of the zero, the first, and the second reaction order, respectively [13].

The grid search optimization method was applied to calculate values of parameters of the used mathematical models. The objective function was defined as the sum of squares between the measured and calculated values of BTX components divided with the number of measurements reduced by the number of estimated parameters [13, 14].

#### **3. Results of the work**

Gas chromatographic method after static headspace extraction was used for quantification of the BTX in water. Headspace part was analyzed by gas chromatography with MS detector (Agilent Technologies 7890A GC Systems). All substances used for preparation of model

, 2—O<sup>3</sup>

 reactor, 6—Venturi nozzle, 7—UV source, 8—tempering jacket, 9—recirculation pump, 10—external reaction mixture recirculation loop, 11—sucking of gas from head space, 12—sampling device, 13—outfall of gas, 14—destruction

generator, 3—O<sup>2</sup>

/O<sup>3</sup>

gas flow, 4—UV detector of

Desorption of volatile pollutants from water solution is proportional to its concentration in

. *dS = kS dt* <sup>−</sup>

Experimental data of BTX degradation were fitted by the zero (Eq. (2))‐, first (Eq. (3)‐, and second (Eq. (4))‐order reaction kinetic models. For a batch reaction system, under the assumption

*<sup>t</sup>* 0 0 *S =S kt* <sup>−</sup> (2)

(1)

wastewater and standard stock solutions were purchased from Dr. Ehrenstorfer.

of a constant reaction volume, the following relationships were obtained.

**2.3. Processing of experimental data**

.

**Figure 1.** Schematic scheme of the ozonation equipment. 1—O<sup>2</sup>

6 Physico-Chemical Wastewater Treatment and Resource Recovery

where k/h-1 is desorption rate.

water solution

device for residual O<sup>3</sup>

O3 , 5—O<sup>3</sup> The removals of studied compounds with the ozonation time are presented in **Figure 2**. The initial concentrations of benzene, toluene, o-xylene, and p-xylene were 800, 1600, 800, and 600 µg l-1, respectively. From **Figure 2**, it is obvious that for all studied pollutants measured, the highest removal rates were observed within the first 5 min of the process. The highest affinity of ozone was measured toward p‐xylene (59.6% removal efficiency). On the other hand, the lowest treatment efficiency was measured for benzene (20%) within the same ozonation time.

**Figure 2.** BTX (dimensionless values) removal during ozonation of model wastewater ( benzene, toluene, p-xylene, o‐xylene).


**Table 1.** Removal efficiency values of BTX compounds with ozone.

The treatment efficiencies of BTX components increased with the increase of ozonation time [15]. The highest treatment efficiency was observed for p‐xylene (81.3% during the first 20 min of ozonation). The second highest treatment efficiency (72.2%) was measured for o‐xylene. Final removal efficiencies of BTX constituents were observed in the range from 86.4 to 90%. The values of removal efficiency during O<sup>3</sup> process are summarized in **Table 1**.

The best fit of experimental degradation data of all studied pollutants was obtained by the first‐order kinetic model.

The removals of studied BTX by O<sup>3</sup> /UV treatment time are presented in **Figure 3**. It is obvious that the removal rates of BTX by O<sup>3</sup> /UV process are higher in comparison with the removal rates observed with ozone alone. This is confirmed also by the values of removal efficiency given in **Table 2**.

Comparisons of p‐xylene and o‐xylene removals using O<sup>3</sup> only for oxidation and O<sup>3</sup> /UV treatment of model wastewater are presented in **Figures 4** and **5**, respectively. Slightly higher removal rates of these two pollutants were measured when treated with ozone alone.

**Figure 3.** BTX (dimensionless values) removal by O<sup>3</sup> /UV treatment of model wastewater ( benzene, toluene, p-xylene, o‐xylene).


**Table 2.** Removal efficiency values of BTX compounds by O<sup>3</sup> /UV process.

The treatment efficiencies of BTX components increased with the increase of ozonation time [15]. The highest treatment efficiency was observed for p‐xylene (81.3% during the first 20 min of ozonation). The second highest treatment efficiency (72.2%) was measured for o‐xylene. Final removal efficiencies of BTX constituents were observed in the range from 86.4 to 90%.

*t* **[min] Benzene [%] Toluene [%] p‐Xylene [%] o‐Xylene [%]**

5 20.1 26.4 59.6 44.4 20 53.9 68.3 85.7 86.4 60 84.6 89.6 90.0 89.9

The best fit of experimental degradation data of all studied pollutants was obtained by the

rates observed with ozone alone. This is confirmed also by the values of removal efficiency

ment of model wastewater are presented in **Figures 4** and **5**, respectively. Slightly higher

removal rates of these two pollutants were measured when treated with ozone alone.

process are summarized in **Table 1**.

/UV treatment time are presented in **Figure 3**. It is obvious

/UV process are higher in comparison with the removal

/UV treatment of model wastewater ( benzene, toluene, p-xylene,

only for oxidation and O<sup>3</sup>

/UV treat-

The values of removal efficiency during O<sup>3</sup>

**Table 1.** Removal efficiency values of BTX compounds with ozone.

8 Physico-Chemical Wastewater Treatment and Resource Recovery

Comparisons of p‐xylene and o‐xylene removals using O<sup>3</sup>

first‐order kinetic model.

given in **Table 2**.

The removals of studied BTX by O<sup>3</sup>

that the removal rates of BTX by O<sup>3</sup>

**Figure 3.** BTX (dimensionless values) removal by O<sup>3</sup>

o‐xylene).

**Figure 4.** p‐Xylene concentration profiles during O<sup>3</sup> and O<sup>3</sup> /UV treatment of model wastewater ( p‐xylene‐O<sup>3</sup> , p‐xylene‐O<sup>3</sup> /UV).

**Figure 5.** o‐YLENE concentration profiles during O<sup>3</sup> and O<sup>3</sup> /UV treatment of model wastewater ( o‐xylene‐O<sup>3</sup> , o‐xylene‐O<sup>3</sup> /UV).

However, insignificant differences in the removal rates and treatment efficiencies follow from the treatment of the investigated BTX compounds using O<sup>3</sup> and O<sup>3</sup> /UV. The data presented above represent overall removal of pollutants during O<sup>3</sup> , or O<sup>3</sup> /UV treatment, i.e., an effect of stripping of pollutants content is also included in the data.

The effect of gas striping of the investigated BTX compounds at the conditions of ozonation and O<sup>3</sup> /UV trials was also studied. Volatility of substances depends on the size of molecules as well as on the vapor pressure [16]. With the increase of the molecular weight, the solubility of substance in water decreases. Important factor influencing solubility in water is hydrophobicity of substances. Solubility in water decreases with increase hydrophobicity of substance [17]. The information on hydrophobicity gives octanol‐water partitioning coefficient value. Evaporation of substances correlates with vapor pressure [17] and is strongly influenced by temperature and the pressure of the system [18].

Volatility of substances can be quantified by the values Henry low. With the increase of the Henry's constant value, solubility of substance in water decreases. The values of basic physical-chemical properties of BTX components are given in **Table 3** [19].

Comparison of o‐xylene, p‐xylene, benzene, and toluene concentration profiles measured during the stripping only and ozonation treatment of model water are presented in **Figures 6**–**9**, respectively. As it was already mentioned, 10 min of ozonation corresponds to input of 45 mg O3 per liter of active volume of the jet-loop ozonation reactor.

It is obvious from the presented results that stripping can significantly contribute to the removal of investigated compounds during the ozonation and O<sup>3</sup> /UV treatments. The higher contribution of stripping to the overall removal of the component during the ozonation is for benzene. This observation correlates very well with the physical properties of the different components (**Table 3**). Similar results were obtained for toluene (36.8 μg L-1 min-1) in comparison to benzene (32.0 µg L-1 min-1).


*Notes:* AA‐EQS, Annual Average Value of Environmental Quality Standard (EQS).

a European Commission (2008) List of EQS [20].

b Regulation of the Slovak Government [21] List a EQS of total xylenes of 10 μg L-1.

**Table 3.** Basic characteristics for studied organic components of BTX.

Concentration profiles of the studied BTX substances during ozonation, i.e., after excluding removal of the compounds by stripping process during ozonation of the model wastewater, are presented in **Figure 10**. In other words, the values plotted in **Figure 10** were obtained by subtraction of the concentrations of individual components due to stripping by oxygen flow from the total concentrations obtained during ozonation of the model wastewater.

However, insignificant differences in the removal rates and treatment efficiencies follow from

The effect of gas striping of the investigated BTX compounds at the conditions of ozonation

Volatility of substances can be quantified by the values Henry low. With the increase of the Henry's constant value, solubility of substance in water decreases. The values of basic physi-

Comparison of o‐xylene, p‐xylene, benzene, and toluene concentration profiles measured during the stripping only and ozonation treatment of model water are presented in **Figures 6**–**9**, respectively. As it was already mentioned, 10 min of ozonation corresponds to input of 45 mg

It is obvious from the presented results that stripping can significantly contribute to the

contribution of stripping to the overall removal of the component during the ozonation is for benzene. This observation correlates very well with the physical properties of the different components (**Table 3**). Similar results were obtained for toluene (36.8 μg L-1 min-1) in compari-

Molecular weight[g mol-1] 78.11 92.13 106.16 106.16 Water solubility[mg L-1] 1700 515 175 198 Vapor pressure(at 20°C) [kPa] 12.6923 3.7863 0.8799 0,8799 Boling point[°C] 80.1 110.6 144.4 138.4 Octanol‐water partition coefficient (at 20°C)[log Kow] 2.13 2.69 2.77 3.15

AA-EQS [µg L-1] 10a 100a 10b 10 b

/UV trials was also studied. Volatility of substances depends on the size of molecules as well as on the vapor pressure [16]. With the increase of the molecular weight, the solubility of substance in water decreases. Important factor influencing solubility in water is hydrophobicity of substances. Solubility in water decreases with increase hydrophobicity of substance [17]. The information on hydrophobicity gives octanol‐water partitioning coefficient value. Evaporation of substances correlates with vapor pressure [17] and is strongly influenced by

and O<sup>3</sup>

, or O<sup>3</sup>

/UV. The data presented

/UV treatment, i.e., an effect of

/UV treatments. The higher

**Benzene Toluene o‐Xylene p‐Xylene**

mol-1] 0.55 0.67 0.50 0.71

the treatment of the investigated BTX compounds using O<sup>3</sup>

above represent overall removal of pollutants during O<sup>3</sup>

10 Physico-Chemical Wastewater Treatment and Resource Recovery

temperature and the pressure of the system [18].

son to benzene (32.0 µg L-1 min-1).

Henry's law constant(at 25°C) [kPa m<sup>3</sup>

European Commission (2008) List of EQS [20].

cal-chemical properties of BTX components are given in **Table 3** [19].

per liter of active volume of the jet-loop ozonation reactor.

removal of investigated compounds during the ozonation and O<sup>3</sup>

*Notes:* AA‐EQS, Annual Average Value of Environmental Quality Standard (EQS).

Regulation of the Slovak Government [21] List a EQS of total xylenes of 10 μg L-1.

**Table 3.** Basic characteristics for studied organic components of BTX.

and O<sup>3</sup>

O3

a

b

stripping of pollutants content is also included in the data.

**Figure 6.** o‐Xylene concentration profiles during stripping and ozonation of model wastewater ( o-xylene-striping, o‐xylene‐ozonation).

**Figure 7.** p‐Xylene concentration profiles during stripping and ozonation of model wastewater ( p-xylene-striping, p‐xylene‐ozonation).

**Figure 8.** Benzene concentration profiles during stripping and ozonation of model wastewater ( benzene-striping, benzene‐ozonation).

**Figure 9.** Toluene concentration profiles during stripping and ozonation of model wastewater ( toluene-striping, toluene‐ozonation).

Ozonation and stripping experimental trials were performed at the same operational conditions except presence of ozone in the system for stripping tests. The removal efficiencies for BTX components due to ozone oxidation of the model wastewater sample are given in **Table 4**.

Experimental data were processed using kinetic models to evaluate an order of the reaction. The calculated concentration profiles obtained by kinetic models corresponded to the best fit of experimental data (**Table 4**) for ozone oxidation, i.e., after excluding contribution of stripping to overall BTX concentrations during ozonation.

**Figure 10.** Experimental ( benzene, toluene, p-xylene, o‐xylene) and calculated (lines) BTX concentration profiles during ozonation of model wastewater.


**Table 4.** Removal efficiency values of BTX compounds by O<sup>3</sup> (stripping excluded).


**Table 5.** Removal efficiency values of BTX compounds by O<sup>3</sup> /UV.

Ozonation and stripping experimental trials were performed at the same operational conditions except presence of ozone in the system for stripping tests. The removal efficiencies for BTX components due to ozone oxidation of the model wastewater sample are given in **Table 4**. Experimental data were processed using kinetic models to evaluate an order of the reaction. The calculated concentration profiles obtained by kinetic models corresponded to the best fit of experimental data (**Table 4**) for ozone oxidation, i.e., after excluding contribution of strip-

**Figure 9.** Toluene concentration profiles during stripping and ozonation of model wastewater ( toluene-striping,

**Figure 8.** Benzene concentration profiles during stripping and ozonation of model wastewater ( benzene-striping,

ping to overall BTX concentrations during ozonation.

toluene‐ozonation).

benzene‐ozonation).

12 Physico-Chemical Wastewater Treatment and Resource Recovery

The highest removal rates were observed during the first 5 minutes of ozonation for all investigated pollutants. However, there is approximately 10% difference in removal efficiency (**Table 1**) and it is caused by ozonation reaction only (**Table 4**).

The best fit of experimental degradation data of benzene and toluene was obtained by the first‐order kinetic model. On the other hand, the second‐order kinetic model was more appropriate for description degradation of xylenes. The rate constant values and the values of the correlation coefficient *XY r* corresponding to treated kinetic data after subtracting volatilized

portions due to oxygen stripping are summarized both for O<sup>3</sup> and O<sup>3</sup> /UV treatments in **Table 6**.


**Table 6.** Kinetic parameters and statistical characteristics–striping excluded.

**Figure 11.** Experimental ( benzene, toluene, p-xylene, o‐xylene) and calculated (lines) BTX concentration profiles during O<sup>3</sup> /UV treatment of model wastewater.

Similar concentration profiles for the studied BTX compounds during the O<sup>3</sup> /UV treatment of the model wastewater are presented in **Figure 11**. The values presented in **Figure 11** were also obtained by subtraction of concentrations of individual components caused by gas stripping. The calculated concentration profiles were also obtained by kinetic models corresponding to the best fit of experimental values (**Table 6**) for O<sup>3</sup> /UV treatment, i.e., after excluding contributions of stripping of individual compounds. The removal efficiencies values are given in **Table 5**.

Removal efficiencies are very close to those given in **Table 2**. In the case of o‐xylene and p‐ xylene, the best fit was obtained by the second‐order kinetic model (**Table 6**). The first‐order kinetic models for benzene and toluene may indicate significant influence of gas stripping on total removal of these compounds from solution and the process is probably determined by physical phenomena rather than chemical.

**Figure 12.** Ozone concentration profiles in the jet‐loop reactor during treatment of model wastewater with O<sup>3</sup> ( ) and O3 /UV ( ).

**Figure 12** illustrates ozone concentration profiles in the jet‐loop reactor during treatment of model wastewater with O<sup>3</sup> and O<sup>3</sup> /UV.

Lower ozone concentration in water or higher ozone consumption in the system is obvious for O<sup>3</sup> /UV treatment. Lower experimental ozone concentration for O<sup>3</sup> /UV reaction system can be explained by decomposition and hydroxyl radicals formation. However, insignificant increase of removal rates was observed as a result of radical reaction mechanism. On the other hand, the values of kinetic constant for all compounds (**Table 6**) are slightly higher when ozone alone was applied in comparison to O<sup>3</sup> /UV treatment. Thus, the higher removal rates for ozone treatment of BTX were observed in comparison to the O<sup>3</sup> /UV treatment process.

#### **4. Conclusion**

/UV treatment of

/UV treatment, i.e., after excluding contributions

Similar concentration profiles for the studied BTX compounds during the O<sup>3</sup>

best fit of experimental values (**Table 6**) for O<sup>3</sup>

/UV treatment of model wastewater.

during O<sup>3</sup>

**Pollutants** *k***<sup>1</sup>**

14 Physico-Chemical Wastewater Treatment and Resource Recovery

O3

O3

O3

O3

**Table 6.** Kinetic parameters and statistical characteristics–striping excluded.

physical phenomena rather than chemical.

the model wastewater are presented in **Figure 11**. The values presented in **Figure 11** were also obtained by subtraction of concentrations of individual components caused by gas stripping. The calculated concentration profiles were also obtained by kinetic models corresponding to the

**Figure 11.** Experimental ( benzene, toluene, p-xylene, o‐xylene) and calculated (lines) BTX concentration profiles

 **[h–1]** *k***<sup>2</sup>**

/UV 5.42 × 10–2 – 0.9770

/UV 4.56 × 10–2 – 0.9977

/UV – 1.78 × 10–4 0.9636

/UV – 3.82 × 10–4 0.9980

Benzene O<sup>3</sup> 3.92 × 10–2 – 0.9964

Toluene O<sup>3</sup> 5.55 × 10–2 – 0.9885

o-Xylene O<sup>3</sup> – 2.38 × 10–4 0.9793

p-Xylene O<sup>3</sup> – 5.78 × 10–4 0.9839

 **[g m–3 h–1]**

of stripping of individual compounds. The removal efficiencies values are given in **Table 5**.

Removal efficiencies are very close to those given in **Table 2**. In the case of o‐xylene and p‐ xylene, the best fit was obtained by the second‐order kinetic model (**Table 6**). The first‐order kinetic models for benzene and toluene may indicate significant influence of gas stripping on total removal of these compounds from solution and the process is probably determined by Effect of ozone and O<sup>3</sup> /UV treatments on BTX components were investigated in this study. Investigation of stripping of volatile substances was also performed.

The highest removal rates for all investigated BTX components were observed during the first 5 min of processing for both ozonation and O<sup>3</sup> /UV treatment processes.

Ozone showed the highest affinity to p‐xylene. The lowest removal efficiency was measured for benzene. Treatment efficiencies above 90% were observed for all investigated pollutants after 40 min of ozonation. Longer ozonation time resulted in very low enhancements of removal efficiencies of both ozonation and O<sup>3</sup> /UV treatment processes.

Application of O<sup>3</sup> /UV treatment had no significant effect in comparison with ozonation only, particularly for benzene. In case of o‐xylene and p‐xylene removal efficiencies, over 90% were observed after 20 min of the process. Forty minutes of the process were needed for more than 90% removal efficiency of ethylbenzene.

Due to high volatility of BTX components their removal from liquid phase can be significantly influenced by stripping. According to physical characteristics, the highest stripping can be expected for benzene and toluene [15].

From the processing of experimental data by simple kinetic models [13, 14], removal of oxylene and p‐xylene was best achieved by the second‐order kinetic model. On the other hand, best fit of experimental data for benzene and toluene was obtained using the first‐order kinetic model.

From the result of the study one can conclude that ozonation is a prospective process and a promising procedure for the removal of BTX components from aquatic environment. However, further research should be performed to enhance process efficiency and to study the impact of reaction intermediates and products on aquatic ecosystem.

## **Acknowledgements**

This work was supported by the Slovak Research and Development Agency under the contract No. APVV‐0656‐12. The authors would like to thank also for the support from the VEGA Grant 1/0859/14.

## **Author details**

Ján Derco<sup>1</sup> \*, Katarína Šimovičová<sup>2</sup> , Jozef Dudáš<sup>1</sup> and Mária Valičková<sup>1</sup>

\*Address all correspondence to: jan.derco@stuba.sk

1 Institute of Chemical and Environmental Engineering, Faculty of Chemical and Food Technology, Slovak University of Technology, Bratislava, Slovak Republic

2 Water Research Institute, Bratislava, Slovak Republic


[3] Water Act (2004). Act No. 364/2004 on water (Water Act) (in Slovak).

observed after 20 min of the process. Forty minutes of the process were needed for more than

Due to high volatility of BTX components their removal from liquid phase can be significantly influenced by stripping. According to physical characteristics, the highest stripping can be

From the processing of experimental data by simple kinetic models [13, 14], removal of oxylene and p‐xylene was best achieved by the second‐order kinetic model. On the other hand, best fit of experimental data for benzene and toluene was obtained using the first‐order

From the result of the study one can conclude that ozonation is a prospective process and a promising procedure for the removal of BTX components from aquatic environment. However, further research should be performed to enhance process efficiency and to study

This work was supported by the Slovak Research and Development Agency under the contract No. APVV‐0656‐12. The authors would like to thank also for the support from the VEGA

1 Institute of Chemical and Environmental Engineering, Faculty of Chemical and Food

[1] European Commission (2000). Directive 2000/60/EC of the European Parliament and of the Council of 23 October 2000 establishing a framework for community action in the

[2] European Commission (2001). Decision No. 2455/2001/EC of the European Parliament and of the Council establishing the list of priority substances in the field of water policy

and amending Directive 2000/60/EC, 2001. Off. J. Eur. Union, OJ L 44, pp. 1–5.

and Mária Valičková<sup>1</sup>

, Jozef Dudáš<sup>1</sup>

Technology, Slovak University of Technology, Bratislava, Slovak Republic

field of water policy. Off. J. Eur. Communities, L 327/1.

the impact of reaction intermediates and products on aquatic ecosystem.

90% removal efficiency of ethylbenzene.

16 Physico-Chemical Wastewater Treatment and Resource Recovery

expected for benzene and toluene [15].

kinetic model.

**Acknowledgements**

Grant 1/0859/14.

**Author details**

\*, Katarína Šimovičová<sup>2</sup>

\*Address all correspondence to: jan.derco@stuba.sk

2 Water Research Institute, Bratislava, Slovak Republic

Ján Derco<sup>1</sup>


#### **Kinetic Modeling of the UV/H2O2 Process: Determining the Effective Hydroxyl Radical Concentration Kinetic Modeling of the UV/H2O2 Process: Determining the Effective Hydroxyl Radical Concentration**

Ainhoa Rubio-Clemente, E. Chica and Gustavo A. Peñuela Ainhoa Rubio-Clemente, E. Chica and Gustavo A. Peñuela

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65096

#### **Abstract**

[16] Bedient PB. Groundwater contamination, transport and remediation. Englewood Cliffs,

[17] Srijata M, Pranab R. BTEX : a serious ground‐water contaminant. *Res. J. Environ. Sci.*

[18] Ullmann F. *Ullmann's Encyclopedia of Industrial Chemistry* (6th, completely rev. ed.) Electronic Release. Weinheim, Germany: Wiley – VCH. 2011. ISBN: 9783527303878. [19] Matěju V. ed. The Compendium of Clean‐up Technologies. 2006; Vodní zdroje Ekomonitor

[20] European Commission (2008). Directive 2008/105/EC of the European Parliament and of 10 the Council of 16 December 2008 on environmental quality standards in the field of water policy, amending and subsequently repealing Council Directives 82/176/EEC, 83/513/EEC, 84/156/EEC, 84/491/EEC, 86/280/EEC and amending Directive 2000/60/EC.

[21] Regulation of the government of the Slovak Republic N° 269/2010: Quality objectives of surface waters and on limit values for pollution indicators of waste waters and special

New Jersey: Prentice Hall PTR. 1994.

18 Physico-Chemical Wastewater Treatment and Resource Recovery

waters (in Slovak).

2011; **5:** 394–398*.* DOI: 10.3923/rjes.2011.394.398.

spol. s.r.o., Chrudim. ISBN: 80‐86832‐15‐5 (in Czech).

A kinetic model for pollutant degradation by the UV/H2O2 system was developed. The model includes the background matrix effect, the reaction intermediate action, and the pH change during time. It was validated for water containing phenol and three different ways of calculating HO° level time-evolution were assumed (non-pseudo-steady, pseudo-steady and simplified pseudo-steady state; denoted as kinetic models A, B and C, respectively). It was found that the kind of assumption considered was not significant for phenol degradation. On the other hand, taking into account the high levels of HO2° formed in the reaction solution compared to HO° concentration (~10–7 M >>>> ~10–14 M), HO2° action in transforming phenol was considered. For this purpose, phenol-HO2° reaction rate constant was calculated and estimated to be 1.6x103 M-1 s-1, resulting in the range of data reported from literature. It was observed that, although including HO2° action allowed slightly improving the kinetic model degree of fit, HO° developed the major role in phenol conversion, due to their high oxidation potential. In this sense, an effective level of HO° can be determined in order to be maintained throughout the UV/ H2O2 system reaction time for achieving an efficient pollutant degradation.

**Keywords:** UV/H2O2 process, matrix background, kinetic model, reaction rate constant

#### **1. Introduction**

Nowadays, one of the major problems associated with the presence of toxic and persistent pollutants in the aquatic environment is the unfeasibility of conventional treatments for the effective removal of those substances [1-3]. Hence the application of alternative technologies,

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

such as advanced oxidation processes, is needed [4]. Among these techniques, the UV/H2O2 system is included. It consists of the photolysis of hydrogen peroxide (H2O2) by applying ultraviolet (UV) radiation resulting in the generation of hydroxyl radicals (HO°) [5, 6]. This process may be performed at room temperature and pressure, it has no mass transfer problems, it is easy to maintain and operate, no sludge requiring a subsequent treatment and disposal is produced, and it may achieve a complete pollutant mineralization [5]. Therefore, the UV/H2O2 system seems to be a promising alternative for the treatment of water containing toxic and recalcitrant substances. However, this kind of technology can be expensive due to the associated electrical and oxidant costs [7].

In order to reduce costs and make the process more feasible for industrial applications, the UV/H2O2 system optimization is required [8] and kinetic models can be considered as functional tools for this purpose. Up to date, several kinetic models have been proposed for describing the UV/H2O2 process and predicting different pollutant removal rates [6–15]. In some of these models [13, 15], the proposed set of ordinary differential equations (ODE) defining the studied pollutant degradation rate can be simplified into a pseudo-first–order kinetic expression, whose solution is an exponential one. In such as models, experimental results are fit to that solution. Subsequently, model predictions agree well with laboratory data. In that kind of models the calculated reaction rate constants for the tested pollutant degradation are apparent reaction rate constants (*K*app), which include pollutant removal reaction rate constants and the values of parameters such as the quantum yield for the oxidant, the conjugate base (HO2 − ), and the contaminant photolysis, the initial level of the chemical species involved in pollutant oxidation, the optical path length of the system, the UV-light intensity, and the molar extinction coefficients of H2O2/HO2 − and pollutant, among others. Therefore, knowing those parameters is not required.

On the other hand, there are dynamic kinetic models that try to solve the considered ODE set, for which the values of the mentioned variables are required, increasing the complexity of the kinetic model. In order to solve the proposed ODE set, the pseudo-steady state approximation assumption for reactive intermediates, such as HO°, is invoked by arguing that these chemical species are as transient ones as their concentration can be presumed to be at a pseudo-steady state [10, 12, 14]. In other models [6–15], on the contrary, the non-pseudo-steady state premise in free radical rate expressions for predicting the degradation of the probe compound in a more accurate way is applied. However, although hydroperoxyl radicals (HO2°) are involved in those models, none of them, excluding Huang and Shu [11] and Liao and Gurol [12] models, includes HO2° in the target pollutant oxidation final expression. Additionally, some of these models cannot be reproduced unless a conversion factor is included, as demonstrated by Audenaert et al. [8].

In this sense, the aim of this work was to develop a kinetic model based on the main chemical and photochemical reactions for pollutant degradation in water systems by the UV/H2O2 process taking into account the decomposition of the pollutant through direct photolysis, HO° oxidation, and HO2° and superoxide radical (O2°− ) transformation. Furthermore, HO° scavenging effects of carbonate (CO3 2−), bicarbonate (HCO3 − ), sulfate (SO4 2−), and chloride (Cl− ) ions were considered. pH changes in the bulk and the detrimental action of the organic matter (OM) and the reaction intermediates in shielding UV and quenching HO° were studied. The influence of the pseudo-steady and non-pseudo-steady state hypothesis for determining HO° concentration-evolution with time was also examined and the second-order HO2° reaction rate constant for the studied pollutant was determined. MATLAB software was used to solve the ODE set that characterizes the current model and the results were validated by using experimental data obtained from the literature for phenol (PHE) degradation by the UV/H2O2 process in a completely mixed batch photoreactor.

## **2. Experimental model approach**

such as advanced oxidation processes, is needed [4]. Among these techniques, the UV/H2O2 system is included. It consists of the photolysis of hydrogen peroxide (H2O2) by applying ultraviolet (UV) radiation resulting in the generation of hydroxyl radicals (HO°) [5, 6]. This process may be performed at room temperature and pressure, it has no mass transfer problems, it is easy to maintain and operate, no sludge requiring a subsequent treatment and disposal is produced, and it may achieve a complete pollutant mineralization [5]. Therefore, the UV/H2O2 system seems to be a promising alternative for the treatment of water containing toxic and recalcitrant substances. However, this kind of technology can be expensive due to the associated

In order to reduce costs and make the process more feasible for industrial applications, the UV/H2O2 system optimization is required [8] and kinetic models can be considered as functional tools for this purpose. Up to date, several kinetic models have been proposed for describing the UV/H2O2 process and predicting different pollutant removal rates [6–15]. In some of these models [13, 15], the proposed set of ordinary differential equations (ODE) defining the studied pollutant degradation rate can be simplified into a pseudo-first–order kinetic expression, whose solution is an exponential one. In such as models, experimental results are fit to that solution. Subsequently, model predictions agree well with laboratory data. In that kind of models the calculated reaction rate constants for the tested pollutant degradation are apparent reaction rate constants (*K*app), which include pollutant removal reaction rate constants and the values of parameters such as the quantum yield for the oxidant, the conjugate

), and the contaminant photolysis, the initial level of the chemical species involved

and pollutant, among others. Therefore, knowing

) transformation. Fur-

−

), sulfate

2−), bicarbonate (HCO3

) ions were considered. pH changes in the bulk and the detrimen-

in pollutant oxidation, the optical path length of the system, the UV-light intensity, and the

On the other hand, there are dynamic kinetic models that try to solve the considered ODE set, for which the values of the mentioned variables are required, increasing the complexity of the kinetic model. In order to solve the proposed ODE set, the pseudo-steady state approximation assumption for reactive intermediates, such as HO°, is invoked by arguing that these chemical species are as transient ones as their concentration can be presumed to be at a pseudo-steady state [10, 12, 14]. In other models [6–15], on the contrary, the non-pseudo-steady state premise in free radical rate expressions for predicting the degradation of the probe compound in a more accurate way is applied. However, although hydroperoxyl radicals (HO2°) are involved in those models, none of them, excluding Huang and Shu [11] and Liao and Gurol [12] models, includes HO2° in the target pollutant oxidation final expression. Additionally, some of these models cannot be reproduced unless a conversion factor is included, as demonstrated by

In this sense, the aim of this work was to develop a kinetic model based on the main chemical and photochemical reactions for pollutant degradation in water systems by the UV/H2O2 process taking into account the decomposition of the pollutant through direct

photolysis, HO° oxidation, and HO2° and superoxide radical (O2°−

thermore, HO° scavenging effects of carbonate (CO3

−

electrical and oxidant costs [7].

20 Physico-Chemical Wastewater Treatment and Resource Recovery

base (HO2

−

Audenaert et al. [8].

2−), and chloride (Cl−

(SO4

molar extinction coefficients of H2O2/HO2

those parameters is not required.

A mathematical model for predicting pollutant degradation and the concentrations of the main species involved in a UV/H2O2 system was developed. The developed model describes radical chain reactions occurring during the UV/H2O2 process in the presence of HO° scavengers and UV-radiation absorbers, such as dissolved organic matter (DOM), anions and reaction intermediate products. Pollutant degradation mechanisms, direct UV photolysis and radical attack by HO°, HO2°, O2°− and other anion radicals (CO3°− , SO4°− , H2ClO°, HClO°, Cl°, and Cl2°− ) were included. Additionally, the model incorporates the competitive UV-radiation absorption by H2O2, the parent compound and the DOM in terms of dissolved organic carbon (DOC), as well as the formation and disappearance of intermediate products, also considered as DOC. Moreover, it accounts for the solution pH change due to the mineralization of organic compounds and the formation of acids.

## **2.1. UV/H2O2 system fundamentals**

The UV/H2O2 system initiates with the primary photolysis of H2O2 or HO2 − , producing HO° according to Eqs. (1) and (2) [6, 13]. Based on the Beer-Lambert law and quantum yield definition, the reaction rates for H2O2/HO2 − direct photodegradation and HO° generation are obtained through Eqs. (3)—(5), respectively.

$$\mathrm{H\_2O\_2\to2HO^\circ}\tag{1}$$

$$\mathrm{H}\mathrm{O}\_{2}^{-} + \mathrm{H}\_{2}\mathrm{O} \xrightarrow{h\nu} 2\mathrm{HO}^{\circ} + \mathrm{OH}^{-} \tag{2}$$

$$\frac{d\left[\text{H}\_2\text{O}\_2\right]}{dt} = -\phi\_{\text{H}\_2\text{O}\_2}I\_{a,\text{H}\_2\text{O}\_2} \tag{3}$$

$$\frac{d\left[\text{HO}\_2^-\right]}{dt} = -\phi\_{\text{HO}\_2^-} I\_{a,\text{HO}\_2^-} \tag{4}$$

$$\frac{d\left[\overline{\rm HO}^{\circ}\right]}{dt} = 2\phi\_{\rm H\_2O\_2}I\_{a,\,\rm H\_2O\_2} + 2\phi\_{\rm HO\_2^{-}}I\_{a,\,\rm HO\_2^{-}}\tag{5}$$

where *φ*H2O2 and *φ*HO− 2 (mol Ein−1) are the quantum yields of the photochemical reactions of H2O2 and HO2 − . *Ia*, H2O2 and *Ia*, HO− 2 (Ein L−1 s−1) refer to the UV-radiation intensities absorbed by H2O2 and HO2 − , respectively, calculated according to Eqs. (6) and (7), where *f*H2O2 and *f*HO− 2 are the fractions of the UV-radiation absorbed by H2O2 and HO2 − , respectively (Eqs. (8) and (9)) [7, 8].

$$I\_{a,H\_2O\_2} = l\_0 f\_{H\_2O\_2} \left\{ 1 - \exp \left[ -2.3 l \left( \varepsilon\_{H\_2O\_2} [H\_2O\_2] + \varepsilon\_{HO\_2^-} [HO\_2^-] + \varepsilon\_C [C] + \varepsilon\_{DOC} [DOC] \right) \right] \right\} \tag{6}$$

$$I\_{a, \mathrm{HO}\_2^-} = I\_0 f\_{\mathrm{HO}\_2^-} \left\{ 1 - \exp \left[ -2.3 l \left( \varepsilon\_{\mathrm{H}\_2\mathrm{O}\_2} \left[ \mathrm{H}\_2 \mathrm{O}\_2 \right] + \varepsilon\_{\mathrm{HO}\_2^-} \left[ \mathrm{HO}\_2^- \right] + \varepsilon\_{\mathrm{C}} \left[ \mathrm{C} \right] + \varepsilon\_{\mathrm{DOC}} \left[ \mathrm{DOC} \right] \right) \right] \right\} \tag{7}$$

$$f\_{\rm H\_2O\_2} = \frac{\varepsilon\_{\rm H\_2O\_2} \left[\rm H\_2O\_2\right]}{\varepsilon\_{\rm H\_2O\_2} \left[\rm H\_2O\_2\right] + \varepsilon\_{\rm HO\_2^-} \left[\rm HO\_2^-\right] + \varepsilon\_{\rm C} \left[\rm C\right] + \varepsilon\_{\rm DOC} \left[\rm DOC\right]}\tag{8}$$

$$f\_{\rm HO\_2^-} = \frac{\varepsilon\_{\rm HO\_2^-} \left[ \rm HO\_2^- \right]}{\varepsilon\_{\rm H\_2O\_2} \left[ \rm H\_2O\_2 \right] + \varepsilon\_{\rm HO\_2^-} \left[ \rm HO\_2^- \right] + \varepsilon\_{\rm C} \left[ \rm C \right] + \varepsilon\_{\rm DOC} \left[ \rm DOC \right]} \tag{9}$$

in which [H2O2], [HO2 − ], [C], and [DOC] are H2O2, HO2 − , contaminant and DOM, in terms of DOC, concentrations, respectively. εH2O2 , εHO2 –, εC and εDOC (M−1 m−1) are the molar extinction coefficients of H2O2, HO2 − , the pollutant and the DOC, respectively. In turn, (mm) is the photoreactor path length and <sup>0</sup> (Ein s−1), the incident UV-light intensity.

In addition to the oxidant photolysis, the target pollutant (C) may interact with the UVradiation, undergoing degradation and producing reaction intermediates. A fraction of those by-products can be dissolved in the solution [16]. This fraction is denoted as DOC (Eq. (10)) [6]. The reaction rate for the contaminant direct photolysis is obtained through Eq. (11).

$$\text{C} \xrightarrow{h\nu} \text{DOC} + \text{?} \tag{10}$$

Kinetic Modeling of the UV/H2O2 Process: Determining the Effective Hydroxyl Radical Concentration http://dx.doi.org/10.5772/65096 23

$$\frac{d\left[\text{C}\right]}{dt} = -\phi\_{\text{c}} I\_{a,\text{C}}\tag{11}$$

where c is the quantum yield of pollutant photolysis and , C is the amount of UV-light absorbed by the contaminant, calculated through Eqs. (12) and (13).

2 2

ë û = - (4)

(Ein L−1 s−1) refer to the UV-radiation intensities absorbed

 ee


−

, contaminant and DOM, in terms of

–, εC and εDOC (M−1 m−1) are the molar extinction

¾¾® + (10)

and

(6)

(8)

(9)

, respectively (Eqs.

HO , HO

22 22 2 2 HO , H O HO , HO

 f - é ù° ë û = + (5)

(mol Ein−1) are the quantum yields of the photochemical reactions of

, respectively, calculated according to Eqs. (6) and (7), where *f*H2O2

*I I*

2 2 *<sup>a</sup> <sup>a</sup>*

*I*

f- -

*a*

2

HO


*dt*

f

are the fractions of the UV-radiation absorbed by H2O2 and HO2

( 2 2 ) 2 2 <sup>2</sup> ,HO HO <sup>0</sup> HO 2 2 HO <sup>2</sup> DOC *<sup>a</sup>* 1 2.3 H O HO C DOC *<sup>C</sup> I I f exp l* - -

e

2 2

HO 2 2 HO 2 C DOC

2


HO 2 2 HO 2 C DOC


, εHO2

The reaction rate for the contaminant direct photolysis is obtained through Eq. (11).

C DOC ? *<sup>h</sup>*u

e

HO 2

 é ù

In addition to the oxidant photolysis, the target pollutant (C) may interact with the UVradiation, undergoing degradation and producing reaction intermediates. A fraction of those by-products can be dissolved in the solution [16]. This fraction is denoted as DOC (Eq. (10)) [6].

HO

H O HO C DOC

+ ++ é ù é ù éù é ù ë û ëû ë û ë û

 ee

<sup>0</sup> (Ein s−1), the incident UV-light intensity.

−

, the pollutant and the DOC, respectively. In turn, (mm) is the

 e

HO 2 2

H O H O HO C DOC

 ee



+ ++ é ù é ù éù é ù ë û ëû ë û ë û

ì ü é ù - = -- + ++ é ù í ý é ù éù é ù ë û ëû ë û ê ú ë û î þ ë û (7)

2

e

2 2 2

2 2 2

 e

ë û <sup>=</sup>

 e

], [C], and [DOC] are H2O2, HO2

é ù ë û <sup>=</sup>

*d*

HO

*dt*

and *Ia*, HO−

*d*

22 Physico-Chemical Wastewater Treatment and Resource Recovery

and *φ*HO−

− . *Ia*, H2O2

2

−

2 2

2

DOC, concentrations, respectively. εH2O2


HO

*f*

in which [H2O2], [HO2

coefficients of H2O2, HO2

photoreactor path length and

e

e

−

−

H O

*f*

where *φ*H2O2

*f*HO− 2

H2O2 and HO2

by H2O2 and HO2

(8) and (9)) [7, 8].

$$I\_{a,\\_\text{C}} = I\_0 f\_\text{C} \left\{ 1 - \exp \left[ -2.3l \left( \varepsilon\_{\text{H}\_2\text{O}\_2} \left[ \text{H}\_2\text{O}\_2 \right] + \varepsilon\_{\text{HO}\_2^-} \left[ \text{HO}\_2^- \right] + \varepsilon\_{\text{C}} \left[ \text{C} \right] + \varepsilon\_{\text{DOC}} \left[ \text{DOC} \right] \right) \right] \right\} \tag{12}$$

$$f\_{\rm C} = \frac{\varepsilon\_{\rm C} \boxed{\rm C}}{\varepsilon\_{\rm H\_2O\_2} \boxed{\rm H\_2O\_2} + \varepsilon\_{\rm HO\_2^-} \boxed{\rm HO\_2^-} + \varepsilon\_{\rm C} \boxed{\rm C} + \varepsilon\_{\rm DOC} \boxed{\rm DOC}} \tag{13}$$

Once HO° are produced, they rapidly react with the pollutant of interest, degrading it to form reaction intermediates (Eq. (14)), which subsequently can be attacked by HO° and undergo further degradation to produce final products, such as CO2, H2O, and mineral acids (Eq. (15)) [7]. In this model, intermediate substances were considered as HO° scavengers as well as UVlight absorbers. Additionally, it was assumed that the pH of the solution decreased due to the conversion of the target pollutant, and consequently the reaction intermediates, into carbon dioxide (i.e., H2CO3 \* in the aqueous phase); although it must be highlighted that not all the DOC is mineralized, since carboxylic acids are also formed during the oxidation process, making the pH of the bulk decreases as well [8]. Under this presumption, Eq. (15) is simplified as Eq. (16). The mass balances for the evolution of the pollutant, the dissolved organic fraction of the formed by-products, HO° and the H2CO3 \* concentrations with time are shown by Eqs. (17)–(20), respectively.

$$\text{C} + \text{HO}^{\text{o}} \xrightarrow{k\_{\text{C,HO}^{\text{o}}}} \text{DOC} + \text{?} \tag{14}$$

$$\text{DOC} + \text{HO}^{\circ} \xrightarrow{k\_{\text{DOC}, \text{HO}^{\circ}}} \text{CO}\_{2} + \text{H}\_{2}\text{O} + \text{mineral acids} \tag{15}$$

$$\text{DOC} + \text{HO}^{\circ} \xrightarrow{k\_{\text{DO}^{\circ}, \text{HO}^{\circ}}} \text{H}\_{2}\text{CO}\_{3}^{\bullet} + \text{H}\_{2}\text{O} \tag{16}$$

$$\frac{d\left[\text{C}\right]}{dt} = -k\_{\text{C,HO}^{\cdot}}\left[\text{HO}^{\circ}\right]\text{[C]}\tag{17}$$

$$\frac{d\left[\text{DOC}\right]}{dt} = k\_{\text{C,HO}^{-}}\left[\text{HO}^{\circ}\prod \text{C}\right] - k\_{\text{DOC,HO}^{-}}\left[\text{HO}^{\circ}\prod \text{DOC}\right] \tag{18}$$

$$\frac{d\left[\overline{\text{HO}^{\circ}}\right]}{dt} = -k\_{\text{C,HO}^{\circ}}\overline{\left[\text{HO}^{\circ}\right]\text{C}}\left[-k\_{\text{DOC,HO}^{\circ}}\overline{\left[\text{HO}^{\circ}\right]\text{DOC}}\right] \tag{19}$$

$$\frac{d\left[\text{H}\_2\text{CO}\_3^\*\right]}{dt} = k\_{\text{DOC}, \text{HO}^\*} \left[\text{HO}^\*\right] \text{DOC} \left[\text{DOC}\right] \tag{20}$$

where [C], [DOC], [HO°], and [H2CO3 \* ] correspond to pollutant, dissolved by-products, HO° and H2CO3 \* levels, respectively. *k*C,HO∘ and *k*DOC,HO∘ are the rate constants of Eq. (14) and (15) or (16).

In the UV/H2O2 system, recombination of HO° can occur to produce H2O2. However, these free radicals can also react with H2O2 and HO2 − , particularly when the oxidant is in excess, to produce HO2°. Although HO2° are less reactive than HO° (E° = 0.98 and 2.8 V, respectively) [8], they can also be involved in pollutant degradation, especially if these radicals are produced in high amounts in the system. Furthermore, HO2° can produce O2°− , which subsequently can participate in pollutant degradation and mineralization [17]. Therefore, the role of these reactive oxygen species was included in the proposed kinetic model.

It is important to note that as the oxidation process develops, the pH of the solution generally goes down, and consequently, some chemical species appear while other species vanish. In order to consider the change of chemical species inside the bulk according to the pH of the solution in the kinetic model, a correction factor (*δRi* ) was introduced for the photolysis of H2O2, HO2 − , and for the reaction between H2O2 and HO°. This correction factor can adopt two values (0 and 1). When *δRi* =1, reaction *Ri* is promoted (i.e., the time-varying concentrations of the chemical species involved in reaction *Ri* are taken into account in the model). When *δRi* =0, reaction *Ri* is not considered in the model and, subsequently, the chemical species taking part in reaction *Ri* are neglected.

On the other hand, species commonly present in water, such as DOM and inorganic anions (e.g., CO3 2−, HCO3 − , SO4 2−, and Cl− , among others) may also have a significant effect because of their ability to absorb UV-light and/or to scavenge HO°. The HO° scavenging effect of matrix constituents drastically limits the oxidation action of HO°, leading to a decrease in the performance of the system [18].

C,HO

C,HO DOC,HO

*k k*

*k k*

\* 2 3

high amounts in the system. Furthermore, HO2° can produce O2°−

reactive oxygen species was included in the proposed kinetic model.

=1, reaction *Ri*

are neglected.

the solution in the kinetic model, a correction factor (*δRi*

tions of the chemical species involved in reaction *Ri*

2−, and Cl−

H CO

*dt*

*d*

é ù

° ° HO C HO DOC é ù ë û = °- ° é ùé ù é ùé ù ë ûë û ë ûë û

C, HO DOC, HO ° ° HO C HO DOC é ù° ë û =- ° - ° é ùé ù é ùé ù ë ûë û ë ûë û

DOC,HO

*k*

\*

ë û = ° é ùé ù ë ûë û

In the UV/H2O2 system, recombination of HO° can occur to produce H2O2. However, these free

produce HO2°. Although HO2° are less reactive than HO° (E° = 0.98 and 2.8 V, respectively) [8], they can also be involved in pollutant degradation, especially if these radicals are produced in

participate in pollutant degradation and mineralization [17]. Therefore, the role of these

It is important to note that as the oxidation process develops, the pH of the solution generally goes down, and consequently, some chemical species appear while other species vanish. In order to consider the change of chemical species inside the bulk according to the pH of

On the other hand, species commonly present in water, such as DOM and inorganic anions

their ability to absorb UV-light and/or to scavenge HO°. The HO° scavenging effect of matrix

, and for the reaction between H2O2 and HO°. This correction factor can adopt

is not considered in the model and, subsequently, the chemical spe-

, among others) may also have a significant effect because of

−

*k*

° HO C é ù ë û =- ° é ùé ù ë ûë û

*dt* (17)

] correspond to pollutant, dissolved by-products, HO°

, particularly when the oxidant is in excess, to

, which subsequently can

) was introduced for the photolysis of

are taken into account in the model).

is promoted (i.e., the time-varying concentra-

(20)

*dt* (18)

*dt* (19)

° HO DOC

levels, respectively. *k*C,HO∘ and *k*DOC,HO∘ are the rate constants of Eq. (14) and (15) or

C

*d*

DOC

24 Physico-Chemical Wastewater Treatment and Resource Recovery

*d*

*d HO*

where [C], [DOC], [HO°], and [H2CO3

radicals can also react with H2O2 and HO2

and H2CO3

H2O2, HO2

When *δRi*

(e.g., CO3

−

two values (0 and 1). When *δRi*

=0, reaction *Ri*

cies taking part in reaction *Ri*

2−, HCO3 − , SO4

(16).

\*

Taking into account all the mentioned processes and in order to give a more realistic view of what happens in the UV/H2O2 process, the kinetic equation describing pollutant degradation can be expressed as Eq. (21).

$$-\frac{d\mathbf{C}}{dt} = \phi\_c I\_{a,\text{C}} + k\_{\text{I}4} \left[ \mathbf{C} \right] \left[ \mathbf{H} \mathbf{O}^o \right] + k\_{\text{I}5} \left[ \mathbf{C} \right] \left[ \mathbf{H} \mathbf{O}\_2^\circ \right] + k\_{\text{I}6} \left[ \mathbf{C} \right] \left[ \mathbf{O}\_2^\circ \right] + \sum\_{l=35}^{40} k\_l \left[ \mathbf{C} \right] \left[ \mathbf{A} \mathbf{R}\_l \right] \tag{21}$$

where the terms , C, 14[C][HO°], 15 <sup>C</sup> HO2 ° , 16 <sup>C</sup> O2 ° − and ∑ = 35 <sup>40</sup> <sup>C</sup> AR represent the specific contributions of UV-radiation, the oxidation of HO°, HO2°, O2°− and the formed anion radicals (AR) (including CO3°− , SO4°− , H2ClO°, HClO°, Cl°, and Cl2°− ) to the overall pollutant degradation, respectively.

In order to quantitatively evaluate the contribution of the cited terms in the contaminant removal, parameters *d, f, g*, and *h* were introduced in Eq. (21), as described by Eq. (22).

$$-\frac{d\mathbf{C}}{dt} = \phi\_c I\_{a,\,\,\mathbf{C}} + k\_{\rm l4} \left[\mathbf{C}\right] \overline{\left[\mathbf{H}\mathbf{O}^o\right]} \overline{\mathbf{d}} + k\_{\rm l5} \left[\mathbf{C}\right] \overline{\left[\mathbf{H}\mathbf{O}\_2^\circ\right]} \overline{\mathbf{f}} + k\_{\rm l6} \left[\mathbf{C}\right] \overline{\left[\mathbf{O}\_2^\circ\right]} \overline{\mathbf{g}} + \sum\_{l=\rm 5\mathcal{S}}^{\rm 40} k\_l \left[\mathbf{C}\right] \overline{\left[\mathbf{A}\mathbf{R}\_l\right]} \overline{\mathbf{h}} \tag{22}$$

When *d* = *f* = *g* = *h* = 0 (i.e., when the initial concentrations of oxidant radical species are equal to zero), the kinetic model only describes the degradation of the pollutant by photolysis. If *d* = *h* = 1 and *f* = *g* = 0, in addition to the photolysis conversion, in a deionized water, the model can predict pollutant transformation through HO° and CO3°− (the latter from the reaction between HO° and HCO3 − or CO3 2−). When *d* = *f* = *h* = *1* and g = 0, contaminant removal by photolysis and the action of HO° and HO2°, as well as CO3°− , is described. When all the parameters are equal to 1 (i.e., *d* = *f* = *g* = *h* = 1) and there are no inorganic anions different from HCO3 − and CO3 2− in the studied water, the kinetic model accounts for the degradation of the pollutant by direct photolysis and oxidation through HO°, HO2°, O2°− , and CO3°− . As the model includes the reactions where Cl− , SO4 2−, CO3 2−, and HCO3 − are involved, it can be used for the treatment of different types of water.

Based on the reactions illustrated in **Figures 1** and **2**, and the different involved parameters, the mass balances and the corresponding ODE of the species of interest (C, DOC, H2O2, HO2 - , HO°, HO2°, O2°− , CO3°− , SO4°− , HClO°− , H2ClO°, HClO°, Cl°, Cl2°− , OH− , H+ , H2CO3 \* , CO3 2−, HCO3 − , HSO4 − , SO4 2−, and Cl− ) are summarized in **Table 1**.

**Figure 1.** Reaction network of the UV/H2O2 process for the degradation of the parent compound without considering the effect of matrix constituents. The species above and below the arrows are the co-reactants and the co-products, respectively.

**Figure 2.** Reaction network describing the effect of the matrix background in the UV/H2O2 system with co-reactants and co-products above and below the arrows, respectively.

#### 26 Physico-Chemical Wastewater Treatment and Resource Recovery Kinetic Modeling of the UV/H2O2 Process: Determining the Effective Hydroxyl Radical Concentration http://dx.doi.org/10.5772/65096 27


**Figure 1.** Reaction network of the UV/H2O2 process for the degradation of the parent compound without considering the effect of matrix constituents. The species above and below the arrows are the co-reactants and the co-products, re-

**Figure 2.** Reaction network describing the effect of the matrix background in the UV/H2O2 system with co-reactants

and co-products above and below the arrows, respectively.

spectively.


**Table 1.** Set of the ODE used in the kinetic model [9, 18, 27, 30–39].

#### **2.2. Proposed kinetic model**

**Table 1.** Set of the ODE used in the kinetic model [9, 18, 27, 30–39].

28 Physico-Chemical Wastewater Treatment and Resource Recovery

In the developed kinetic model, three different ways of calculating the evolution of HO° concentration during time were presumed: (a) a non-pseudo-steady or transient state (i.e., the net formation rate of HO° is different from zero); (b) a pseudo-steady state; and (c) a simplified pseudo-steady state; correspondingly denoted as kinetic model A, B, and C.

In the prediction model A the concentration of HO° can be calculated with Eq. (23). From Eq. (23), the HO° concentration can be written as Eq. (24) (prediction model B).

22 22 3 4 7 2 2 8 HO , H O HO ,HO 5 22 5 22 6 2 7 8 2 9 2 11 2 2 2 12 2 2 2 HO 2 2 H O HO H O HO HO HO HO HO HO HO HO O H O HO H O O *a R a R R R d I Ik dt k k kk kk k* f df d d d - - - ° °- ° ° é ù° ë û <sup>=</sup> <sup>+</sup> - °- é ùé ù ë ûë û é ù é ù é ùé ù é ù é ùé ù é ù ° - ° - ° °- ° - ë ûë û ë û ë ûë û ë û ë û ë û éù é ù éù éù é ù é ù °+ + - ëû ë û ë û ëû ë û ëû <sup>14</sup> 2 17 18 3 19 3 23 4 26 30 2 C HO DOC HO CO HO HCO HO HSO HO Cl HO Cl HO *k d kk k k k k* --- - ° é ùé ù° - ë ûë û éù é ù é ù é ùé ù é ù é ù é ù ° - ° - ° - ° - ë ûë û ë û ë û ë û ëû ë û ë û éù é ù éù éù ° - ° ëû ëû ëû ë û (23)

$$\begin{aligned} \left[\text{HO}^{\circ}\right] &= \left(2\phi\_{\text{H}\_{2}\text{O}\_{2}}I\_{a,\text{H}\_{2}\text{O}\_{2}}\delta\_{R\_{3}} + 2\phi\_{\text{HO}\_{2}^{\circ}}I\_{a,\text{HO}\_{2}^{\circ}}\delta\_{R\_{4}} + k\_{1}\right[\text{H}\_{2}\text{O}\_{2}]\left[\text{HO}\_{2}^{\circ}\right] + k\_{12}\left[\text{H}\_{2}\text{O}\_{2}\right]\left[\text{O}\_{2}^{\circ -}\right] \\ &\left(k\_{5}\left[\text{H}\_{2}\text{O}\_{2}\right]\delta\_{R\_{7}} + k\_{5}\left[\text{H}\_{2}\text{O}\_{2}\right]\delta\_{R\_{8}} + k\_{6}\left[\text{HO}\_{2}^{-}\right] + k\_{8}\left[\text{HO}\_{2}^{\circ}\right] + k\_{9}\left[\text{O}\_{2}^{\circ -}\right] + k\_{14}\left[\text{C}\right]d + \\ &\left(k\_{17}\left[\text{DOC}\right] + k\_{18}\left[\text{CO}\_{3}^{2-}\right] + k\_{19}\left[\text{HCO}\_{3}^{-}\right] + k\_{23}\left[\text{HSO}\_{4}^{-}\right] + k\_{26}\left[\text{C}\Gamma\right] + k\_{30}\left[\text{C}\right]^{\circ} \right) \end{aligned} \tag{24}$$

Considering that the oxidant is in a high level (i.e., *k5*[H2O2]δR7 + *k5*[H2O2]δR8 ≫**∑**ki [Xi ], where **∑**ki [Xi ] = *k*6[HO2 – ] + *k*8[HO2 ° ]*k*9[O2 °–] + *k*14[*C*]*d* + *k*17[*DOC*]+*k*18[*CO*<sup>3</sup> 2–] + *k*19[*HCO*<sup>3</sup> – ] + *k*23[*HSO*<sup>4</sup> – ] + *k*26[*Cl*– ] + *k*30[*Cl*<sup>2</sup> °–]) and 2*φH2O2 Ia,H2O2 δR3* + 2*φHO2 –Ia,HO2 –δR4* ≫**∑**(*k11*[H2O2][HO2 ° ] +*k12*[H2O2][O2 °–], Eq. (24) can be simplified to Eq. (25) (prediction model C). ki and [Xi ] are the reaction rate constants between HO° and species *i*, and the concentration of species *i*, respectively.

$$\mathbb{E}\left[\text{HO}^{\text{o}}\right] = \frac{2\phi\_{\text{H}\_{2}\text{O}\_{2}}I\_{a,\text{H}\_{2}\text{O}\_{2}}\,\delta\_{\text{R}\_{3}} + 2\phi\_{\text{HO}\_{2}^{-\text{o}}}I\_{a,\text{HO}\_{2}^{-\text{o}}}\,\delta\_{\text{R}\_{4}}}{k\_{\text{5}}\text{[H}\_{2}\text{O}\_{2}\text{]}\,\delta\_{\text{R}\_{7}} + k\_{\text{5}}\text{[H}\_{2}\text{O}\_{2}\text{]}\,\delta\_{\text{R}\_{8}}}\tag{25}$$

The initial values of DOC and inorganic anionic species (CO3 2−, HCO3 − , SO4 2−, Cl− , etc.) are set according to the conditions of the water to be treated.

#### **2.3. Numerical solution of the proposed kinetic models**

The ODE system compiled in **Table 1** was solved applying MATLAB software and ODE15S function. For simultaneously solving the ODE set of the proposed kinetic models, it was necessary to define several photochemical parameters such as *φ*H2O2 , *φ*pollutant, *φ*DOC, *ε*H2O2 , *ε*HO2 –, *ε*pollutant, *ε*DOC, *I*0 and *l*. Additionally, the initial concentrations of all the species involved in the system and the rate constants of the chemical reactions between those species were required (**Table 1**). During setting up the model, differential rate equations describing the time dependence of the concentration of the variety of the considered species were defined and plotted.

## **3. Results and discussion**

In order to validate the proposed kinetic models, experimental data for PHE degradation by the UV/H2O2 process were used from Alnaizy and Akgerman [19] study. These authors conducted a set of experiments in a completely mixed batch cylindrical photoreactor made on Pyrex glass. The used photochemical parameters and the kinetic reaction rate constants of PHE with HO°, O2°− , and CO3°− are presented in **Table 2**.


**Table 2.** Values of the parameters used in the kinetic model validation for phenol degradation.

#### **3.1. Assumptions taken into consideration**

As stated previously, the developed kinetic models A, B, and C employed the non-pseudosteady, the pseudo-steady, and the simplified pseudo-steady state assumption, respectively, to estimate HO° concentration. In the proposed models, the impact of UV radiation individually and/or the combined action of H2O2 and UV light (including the effect of HO°, HO2°, O2°− , and CO3°− ) on PHE degradation was studied.

The initial values of DOC and inorganic anionic species (CO3

necessary to define several photochemical parameters such as *φ*H2O2

The ODE system compiled in **Table 1** was solved applying MATLAB software and ODE15S function. For simultaneously solving the ODE set of the proposed kinetic models, it was

*ε*pollutant, *ε*DOC, *I*0 and *l*. Additionally, the initial concentrations of all the species involved in the system and the rate constants of the chemical reactions between those species were required (**Table 1**). During setting up the model, differential rate equations describing the time dependence of the concentration of the variety of the considered species were defined and plotted.

In order to validate the proposed kinetic models, experimental data for PHE degradation by the UV/H2O2 process were used from Alnaizy and Akgerman [19] study. These authors conducted a set of experiments in a completely mixed batch cylindrical photoreactor made on Pyrex glass. The used photochemical parameters and the kinetic reaction rate constants of PHE

**Parameters Notation Numerical values References**

phenol = εC 51 600 M−1 m−1 [19]

<sup>0</sup> 1.516 × 10−6 Ein L−1 s−1

° − = 16 5.8 × 103

° − = 35 2.2 × 107

Quantum yield phenol = C 0.07 mol Ein−1 [23]

phenol, 2

phenol, CO3

As stated previously, the developed kinetic models A, B, and C employed the non-pseudosteady, the pseudo-steady, and the simplified pseudo-steady state assumption, respectively, to estimate HO° concentration. In the proposed models, the impact of UV radiation individually

are presented in **Table 2**.

Path length 63.5 mm

Kinetic rate constant phenol-HO° phenol, HO° = 14 6.6 × 109

**Table 2.** Values of the parameters used in the kinetic model validation for phenol degradation.

according to the conditions of the water to be treated.

30 Physico-Chemical Wastewater Treatment and Resource Recovery

**3. Results and discussion**

, and CO3°−

Molar extinction coefficient ε

Kinetic rate constant phenol-O2°−

Kinetic rate constant phenol-CO3°−

**3.1. Assumptions taken into consideration**

Incident UV-light intensity (radiation of 254 nm > 90% and power = 15 W)

with HO°, O2°−

**2.3. Numerical solution of the proposed kinetic models**

2−, HCO3 − , SO4

2−, Cl−

, *φ*pollutant, *φ*DOC, *ε*H2O2

M−1 s−1 [40]

M−1 s−1 [41]

M−1s−1 [42]

, etc.) are set

, *ε*HO2 –, A PHE concentration of 2.23 × 10−3 M and a H2O2/PHE ratio of 495 were selected for validating the model. The used PHE solution was prepared by adding the appropriate amount of pollutant solution to deionized water [19]. Therefore, the effect of inorganic anions, excluding HCO3 − and CO3 2− was not taken into account. In this sense, in the PHE degradation rate expression (ODE1) the contribution of inorganic anion radicals, such as SO4°− , H2ClO°, HClO°, Cl°, and Cl2°−, was not studied. It was assumed that the other terms included in ODE1 contributed to pollutant degradation.

As the treated water was deionized, the presence of OM different from the parent compound in the initial solution was neglected ([DOC]0 = 0 M). Hence, the DOC in the solution came from PHE photolysis and free radical (HO°, HO2°, O2°− , and CO3°− ) oxidation.

On the other hand, several authors agree that OM reduction by direct photolysis in a UV/ H2O2 oxidation process can be neglected [7–9]. That is the reason why this was not included in the proposed kinetic model. However, it is highlighted that the OM is able to absorb UVlight, preventing UV-penetration into the bulk and avoiding H2O2/HO2 − and pollutant direct photolysis. Therefore, the detrimental effect of UV-shielding in PHE degradation was taken into consideration. For including this effect in the kinetic model, OM molar extinction coefficient, referred as DOC molar extinction coefficient (ԑDOC), must be previously known. Although this parameter has already been measured [7, 8], its value is not a universal one, since DOM is a complex group of aromatic and aliphatic hydrocarbon structures with attached functional groups [20, 21]. As Alnaizy and Akgerman [19] did not measure this variable, a mean value from Peuravuori and Pihlaja [22] study was presumed. This value corresponded to *ԑ*DOC(280 nm) = 35 967 M−1 m−1, which was in the same order of magnitude than *ɛ*PHE. This number could be acceptable due to the formed aromatic intermediates during PHE conversion [19] conserve structural similarities with the parent compound, like the aromatic ring, responsible for the molecule excitation.

Additionally, it is widely known that the quantum yield of a compound is dependent on the excitation wavelength and the pH of the solution. For 254 nm, PHE quantum yield in an aqueous solution was found to be in the range of 0.02– 0.12 mol Ein−1 at pH 1.6– 3.2 [23]. Therefore, an average value (0.07 mol Ein−1) was taken as PHE quantum yield.

Moreover, it is widely recognized that the solution pH decreases as the process proceeds. This variation in the pH can cause difficulties in modeling studies, since the presence of radical species such as HO2°, and O2°− , among other chemical species involved in the oxidation system, is significantly dependent on the pH of the medium. Therefore, to give a more realistic view of what happens inside the reaction medium, H2O2 and HO2 − photolysis reactions (Eqs. (1) and (2), correspondingly), as well as reactions expressed in Eqs. (26) and (27) were discriminated in the model according to the solution pH time evolution, as it is simplified in **Figure 3**. For selecting the suitable reactions with regard to the pH changes over time, previous information about the evolution of the pH during the performance of the process is required. However, in some occasions this is not provided. In this case, the initial pH of the solution was 6.8 [19]. At this pH, one of the predominant species into the bulk was H2O2, since the pKa of the H2O2/ HO2 − equilibrium is 11.6. Therefore, the photolysis of HO2 − was neglected in the model performance (i.e., *δR3* and *δR4* ). Additionally, the initial concentrations of HO° and other considered species were assumed to be zero, with the exception of the pollutant and H2O2. On the other hand, it was found that the pH of the medium rapidly dropped from 6.8 to 4.7–4.2 (4.5 as a mean value) within the first 30 min of radiation [19]. Therefore, and according to **Figure 3**, Eq. (26) was considered during the first 30 min of the reaction while Eq. (27), after that time (i.e., *δR7* = 1, *δR8* = 0 and *δR7* = 0, *δR8* = 1, before and after the first 30 min of the process, respectively).

**Figure 3.** (a) Reaction discrimination diagram as a function of the solution pH and (b) pH evolution of the medium during the UV/H2O2 process in a completely mixed batch photoreactor. Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

$$\text{HO}\_2^{\circ-} + \text{H}^+ \to \text{HO}\_2^{\circ} \tag{26}$$

$$\text{H}\_2\text{O}\_2 + \text{HO}^\circ \rightarrow \text{O}\_2^{\circ -} + \text{H}\_2\text{O} + \text{H}^+\tag{27}$$

#### **3.2. Kinetic model validation**

Initially, the associated ODE sets with model A, B, and C were solved. Parameters *f* = *g* = 0 and *d* = *h* = 1 were considered in order to solely investigate the influence of the direct photolysis, HO°, and CO3°− on PHE degradation.

**Figure 4** compares the simulation results of the proposed kinetic models A, B, and C with the experimental data for 2.23 × 10−3 M PHE with a H2O2/PHE ratio of 495 and a reaction time of 220 min. The measured and simulated results for PHE direct photolysis alone are also presented. It is observed that more than 90% of the initial PHE concentration was removed after 220 min due to both direct photolysis and indirect degradation (primarily due to HO° attack). The effect of CO3°− could be seen as marginal because of the reduced number of those radicals, in the range of 10−15 M, and the low reaction rate constant with PHE compared to HO°. In addition, it is shown that PHE was not completely removed by direct UV photolysis under the tested photochemical conditions (**Table 2**), since approximately 50% of the total PHE degradation was attributed to UV photolysis, as it was experimentally determined by Alnaizy and Akgerman [19]. On the other hand, the figure demonstrates that the prediction kinetic model C was in good agreement with the available experimental data with a relative high correlation factor (*R*<sup>2</sup> =99.34%). For prediction models A and B, *R*<sup>2</sup> were 97.41 and 97.37%, respectively. As a result of the subtle differences between kinetic models A and B, the depicted line describing model A overlaps model B line. An acceptable agreement between model predictions considering only the UV radiation and experimental data was also verified (*R2* = 98.25%).

this pH, one of the predominant species into the bulk was H2O2, since the pKa of the H2O2/

considered species were assumed to be zero, with the exception of the pollutant and H2O2. On the other hand, it was found that the pH of the medium rapidly dropped from 6.8 to 4.7–4.2 (4.5 as a mean value) within the first 30 min of radiation [19]. Therefore, and according to **Figure 3**, Eq. (26) was considered during the first 30 min of the reaction while Eq. (27), after

**Figure 3.** (a) Reaction discrimination diagram as a function of the solution pH and (b) pH evolution of the medium during the UV/H2O2 process in a completely mixed batch photoreactor. Operating conditions: H2O2/PHE = 495; [C]0 =

Initially, the associated ODE sets with model A, B, and C were solved. Parameters *f* = *g* = 0 and *d* = *h* = 1 were considered in order to solely investigate the influence of the direct photolysis,

O H HO 2 2

−

). Additionally, the initial concentrations of HO° and other

= 1, before and after the first 30 min of the process,

°- + ° + ® (26)

H O HO O H O H 2 2 2 2 ° °- <sup>+</sup> + ®+ + (27)

was neglected in the model

equilibrium is 11.6. Therefore, the photolysis of HO2

= 0, *δR8*

and *δR4*

32 Physico-Chemical Wastewater Treatment and Resource Recovery

= 0 and *δR7*

= 1, *δR8*

HO2 −

performance (i.e., *δR3*

that time (i.e., *δR7*

2.23 × 10−3 M; *t* = 220 min.

**3.2. Kinetic model validation**

on PHE degradation.

HO°, and CO3°−

respectively).

**Figure 4.** Comparison of the kinetic model predictions (lines) versus experimental data (o) and (\* ). Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

Furthermore, **Figure 4** clearly presents that the removal rate of the target pollutant was not significantly dependent on the hypothesis assumed to estimate the HO° level (non-pseudosteady, pseudo-steady and simplified pseudo-steady state assumptions). This could be explained from the relatively low concentration of those reactive species in the solution (with a magnitude order of 10−14 M) when compared to the level of other species involved in the system, such as HO2° and O2°− , whose concentrations were in the range of 10−7 M. The number of HO° remaining in the solution is in concordance with the low final HO° levels found in the literature [24, 25] and even higher than those reported by Ray and Tarr [26].

In **Figure 5**, the evolution of the HO°, HO2°, O2°− , DOC and H2CO3 \* normalized estimated concentrations using the prediction model A is depicted. It is observed that HO2° number increased as the oxidation system proceeded, while O2°− level decreased. Typically, the effect of HO2° and O2°− radicals are neglected in the UV/H2O2 system [6–8, 15, 17, 27, 28] since they are found to be less reactive than HO°, as stated previously. However, when they are produced in a high level, they could also participate in the contaminant oxidation. That is the current case of HO2°, as [HO2°] >>>> [HO°]. Therefore, the contribution of HO2° to PHE degradation should be studied. On the other hand, in this work the action of O2°- in PHE conversion can be omitted since O2°- level decreased with the reaction time, as expected, because of the drop in the pH solution.

**Figure 5.** Concentration-time profiles of HO° (---), HO2° (-.-.-), O2°− (-Δ-), DOC (⋅⋅⋅) and H2CO3 \* (—) using the prediction kinetic model A. (\* ) represents PHE experimental evolution. Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

Furthermore, generally, there is a drop in the pH of the medium as the system progresses. This is probably due to acidic compound formation, such as carboxylic acids and H2CO3 \* resulting from pollutant degradation and mineralization. In this study the decrease of the pH in the solution was predicted via DOC conversion and the sole generation of H2CO3 \* by model A. As presented in **Figure 5**, DOC generation was progressively increasing as PHE was being degraded up to a certain point (117 min, corresponding to [DOC]max = 2.061 × 10−3 M, and equivalent to ca. 93% of PHE degradation). From this point, DOC started to decrease until [DOC]f = 1.879 × 10−3 M. That breakpoint represented the moment at which PHE was almost completely transformed into by-products. In addition, at this point, PHE mineralization began to be more evident, since H2CO3 \* level rose approximately in a linear way, up to a final level of 6.493 × 10−14 M, with the subsequent pH decrease. Similarities between the pattern of this DOC profile and that of the formed intermediate curves reported in Alnaizy and Akgerman [19] research are highlighted.

On the other hand, it is worth noting that HO° level evolution with the reaction time was different when comparing the kinetic prediction model A or B with C. Obviating HO2° contribution to PHE degradation, **Figure 6** shows that the highest final HO° level was achieved in the prediction model A, with a maximum HO° concentration equal to 9.435 × 10−14 M. This value was similar to HO° final level in the prediction model B (9.400 × 10−14 M) and different to that of the prediction model C, where HO° final concentration was 4.486 × 10−14 M (ca. 48% lower than the obtained in the kinetic model A or B).

In **Figure 5**, the evolution of the HO°, HO2°, O2°−

34 Physico-Chemical Wastewater Treatment and Resource Recovery

of HO2° and O2°−

in the pH solution.

tion kinetic model A. (\*

10−3 M; *t* = 220 min.

[DOC]f

to be more evident, since H2CO3

[19] research are highlighted.

increased as the oxidation system proceeded, while O2°−

, DOC and H2CO3

radicals are neglected in the UV/H2O2 system [6–8, 15, 17, 27, 28] since they

) represents PHE experimental evolution. Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 ×

Furthermore, generally, there is a drop in the pH of the medium as the system progresses. This

from pollutant degradation and mineralization. In this study the decrease of the pH in the

presented in **Figure 5**, DOC generation was progressively increasing as PHE was being degraded up to a certain point (117 min, corresponding to [DOC]max = 2.061 × 10−3 M, and equivalent to ca. 93% of PHE degradation). From this point, DOC started to decrease until

of 6.493 × 10−14 M, with the subsequent pH decrease. Similarities between the pattern of this DOC profile and that of the formed intermediate curves reported in Alnaizy and Akgerman

On the other hand, it is worth noting that HO° level evolution with the reaction time was different when comparing the kinetic prediction model A or B with C. Obviating HO2° contribution to PHE degradation, **Figure 6** shows that the highest final HO° level was achieved

 = 1.879 × 10−3 M. That breakpoint represented the moment at which PHE was almost completely transformed into by-products. In addition, at this point, PHE mineralization began

level rose approximately in a linear way, up to a final level

is probably due to acidic compound formation, such as carboxylic acids and H2CO3

solution was predicted via DOC conversion and the sole generation of H2CO3

\*

concentrations using the prediction model A is depicted. It is observed that HO2° number

are found to be less reactive than HO°, as stated previously. However, when they are produced in a high level, they could also participate in the contaminant oxidation. That is the current case of HO2°, as [HO2°] >>>> [HO°]. Therefore, the contribution of HO2° to PHE degradation should be studied. On the other hand, in this work the action of O2°- in PHE conversion can be omitted since O2°- level decreased with the reaction time, as expected, because of the drop

**Figure 5.** Concentration-time profiles of HO° (---), HO2° (-.-.-), O2°− (-Δ-), DOC (⋅⋅⋅) and H2CO3

\*

level decreased. Typically, the effect

\*

\*

(—) using the predic-

\*

by model A. As

resulting

normalized estimated

**Figure 6.** Estimated HO° concentration using the prediction models A (–––), B (----), and C (-.-.-). Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

Approximately, in the first 40 min of the process a larger number of HO° in the aqueous medium with the developed kinetic model C was evidenced. Apparently, this amount of HO° was sufficient to degrade about 60% of PHE initial level under the studied experimental conditions. In contrast, models A and B, whose lines are overlapped, produced a lower number of HO° and the theoretical conversion of PHE remained above the experimental data. One possible reason for this discrepancy can be ascribed to DOM and dissolved oxygen positive effects in producing reactive oxygen species (ROS), as HO° [26], which were not considered. After 40 min of reaction, the HO° level was higher in the kinetic models A and B than in model C. This larger amount of HO° might lead to a faster conversion of the pollutant in comparison with the predicted model C, since the hypothetical depletion curve of PHE was below the measured data. Nevertheless, this rapid pollutant degradation did not occur actually. Therefore, there was an amount of HO° produced in excess that was not reacting with the contaminant. This surplus of HO° could be involved in free radical scavenging reactions. As model A and B consider the detrimental effect of HO° consuming reactions, it is suggested that their kinetic rate constants are higher than those ones used in this paper for these reactions to have a larger weight in the system. Additionally, the contradictory outcome between the actual situation and the theoretical one in the first and second stage of the process can also be attributed to the fact that just a fraction of the concentration of the species involved in the whole kinetic equations of the predicted models was actually reacting. Consequently, the real level of the species implicated in each kinetic reaction should be considered. However, it is rather difficult to determine which amount of the chemical species is exactly involving in each reaction for each time step, especially due to the high reactivity of radicals as the oxidation system progresses. In this context, further studies are required to overcome this limitation.

From these findings, it is suggested that there was an effective level of the formed HO°. Below that level, there was a lack of HO° for an efficient pollutant conversion; and above it, an excessive number of HO° was generated. That HO° effective level could be of relevance for industrial applications in order to be maintained throughout the reaction time, allowing an efficient pollutant degradation.

#### **3.3. Estimation of PHE-HO2° reaction rate constant**

In order to study the action of HO2° for pollutant degradation in the UV/H2O2 system, PHE-HO2° rate constant was calculated. For this purpose, the kinetic model A was used and it was estimated through a non-linear least-square objective function. The objective function for minimizing the error between the predicted and the measured data was defined as Eq. (28) [17].

$$\text{Minimize}: f = \sum \left( \mathbb{C} \right)\_{\text{predicted}} - \left[ \mathbb{C} \right]\_{\text{measured}} \right)^2 \tag{28}$$

where [C]measured and [C]predicted correspond to the evolution of experimental and calculated pollutant concentration, respectively. This expression is a function of PHE-HO2° rate constant. The optimum value for PHE-HO2° second-order rate constant was found to be 1.6 × 103 M−1 s−1, which is consistent with the range of the reported values by Kozmér et al. ((2.7±1.2) × 103 M−1 s−1) [29]. The results of running the new prediction kinetic model A (with and without the contribution of HO2° to pollutant conversion) and the experimental data are presented in **Figure 7**. The figure shows that the prediction model A with the estimation

**Figure 7.** Comparison of experimental (symbols) vs predicted data (lines) using the kinetic model A with (—) and without (⋅⋅⋅) the action of HO2°. Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

of PHE-HO2° rate constant was in a stronger agreement with the experimental data (*R2* = 99.0%) than the previous predicted kinetic model A studied in section 3.2 (*R2* = 97.41%). The same conclusion can be drawn for kinetic model B and C, with a *R2* of 98.78 and 99.57% using the calculated PHE-HO2° rate constant, in comparison with 97.37 and 99.34%, respectively. Therefore, in order to achieve a better fit between experimental and predicted data, HO2° action in pollutant degradation should be considered. Nonetheless, HO2° contribution to PHE degradation is non-significant in comparison with the role developed by HO° attack.

## **4. Conclusions**

should be considered. However, it is rather difficult to determine which amount of the chemical species is exactly involving in each reaction for each time step, especially due to the high reactivity of radicals as the oxidation system progresses. In this context, further

From these findings, it is suggested that there was an effective level of the formed HO°. Below that level, there was a lack of HO° for an efficient pollutant conversion; and above it, an excessive number of HO° was generated. That HO° effective level could be of relevance for industrial applications in order to be maintained throughout the reaction time, allowing an

In order to study the action of HO2° for pollutant degradation in the UV/H2O2 system, PHE-HO2° rate constant was calculated. For this purpose, the kinetic model A was used and it was estimated through a non-linear least-square objective function. The objective function for minimizing the error between the predicted and the measured data was defined as Eq. (28) [17].

where [C]measured and [C]predicted correspond to the evolution of experimental and calculated pollutant concentration, respectively. This expression is a function of PHE-HO2° rate constant. The optimum value for PHE-HO2° second-order rate constant was found to be 1.6 ×

M−1 s−1, which is consistent with the range of the reported values by Kozmér et al.

and without the contribution of HO2° to pollutant conversion) and the experimental data are presented in **Figure 7**. The figure shows that the prediction model A with the estimation

**Figure 7.** Comparison of experimental (symbols) vs predicted data (lines) using the kinetic model A with (—) and

without (⋅⋅⋅) the action of HO2°. Operating conditions: H2O2/PHE = 495; [C]0 = 2.23 × 10−3 M; *t* = 220 min.

M−1 s−1) [29]. The results of running the new prediction kinetic model A (with

( )

predicted measured Minimize : *<sup>f</sup>* <sup>=</sup> éù éù C C - å ëû ëû (28)

2

studies are required to overcome this limitation.

36 Physico-Chemical Wastewater Treatment and Resource Recovery

**3.3. Estimation of PHE-HO2° reaction rate constant**

efficient pollutant degradation.

103

((2.7±1.2) × 103

A kinetic model for studying pollutant degradation by the UV/H2O2 system was developed, including the background matrix effect in scavenging free radicals and shielding UV-light and the reaction intermediate action, as well as the change of the pH as the UV/H2O2 process proceeds. Three different ways for calculating HO° level time evolution were assumed (nonpseudo-steady, pseudo-steady and simplified pseudo-steady state; denoted as kinetic models A, B, and C, respectively). It was found that the assumption of pseudo-steady (simplified or not) or transient state for determining the HO° level evolution with time was not significant in PHE degradation rate due to the relatively low HO° level present into the bulk (~10−14 M). On the other hand, taking into account the high levels of HO2° formed in the reaction solution compared to HO° concentration (~10−7 M >>>> ~10-14 M), HO2° action in transforming PHE was considered. For this purpose, PHE-HO2° reaction rate constant was calculated and estimated to be 1.6 × 103 M−1 s−1, resulting in the range of data reported from literature. It was observed that, although including HO2° action allowed slightly improving the kinetic model degree of fit, HO° developed the major role in PHE conversion, due to their high oxidation potential.

Additionally, it was found that there was an effective level of the HO° formed in solution. Below that level, there was a lack of HO° for an efficient pollutant conversion; and above it, an excessive number of HO° was generated. That HO° effective level calculated from kinetic model C could be of relevance for industrial applications in order to be maintained throughout the reaction time, allowing an efficient pollutant degradation.

In this study, there was an attempt to contemplate a wide range of the chemical reactions involved in the UV/H2O2 process and although high correlation factors were obtained, it is suggested to include the positive effect of the OM and the dissolved oxygen in generating ROS, as well as the effect of other anions naturally present in water bodies, as phosphate and nitrate, for the model to be a more accurate approximation of reality.

## **Acknowledgements**

This work was financially supported by the Colombian Administrative Department of Science, Technology and Innovation (COLCIENCIAS).

## **Author details**

Ainhoa Rubio-Clemente1,2\*, E. Chica3 and Gustavo A. Peñuela1

\*Address all correspondence to: ainhoarubioclem@gmail.com

1 Grupo GDCON, Facultad de Ingeniería, Sede de Investigaciones Universitarias (SIU), Universidad de Antioquia UdeA, Colombia

2 Departamento de Ciencia y Tecnología de los Alimentos, Facultad de Ciencias de la Salud, Universidad Católica San Antonio de Murcia UCAM, Spain

3 Departamento de Ingeniería Mecánica, Facultad de Ingeniería, Universidad de Antioquia UdeA, Colombia


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**Provisional chapter**

## **Electrooxidation-Ozonation: A Synergistic Sustainable Wastewater Treatment Process Electrooxidation-Ozonation: A Synergistic Sustainable Wastewater Treatment Process**

Carlos E. Barrera-Diaz and Nelly González-Rivas Additional information is available at the end of the chapter

Carlos E. Barrera-Diaz and Nelly González-Rivas

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65887

#### **Abstract**

Advanced oxidation processes (AOPs) have shown to be very useful technologies for application in different wastewater treatment areas. These processes use the very strong oxidizing power of hydroxyl radicals to oxidize organic compounds to carbon dioxide and water. These procedures usually involve the use of O<sup>3</sup> , H2 O2 , Fenton's reagent and electrolysis to generate the hydroxyl radicals. However, some recent investigations have found that the use of a coupled processes using O3 /electrooxidation increases the effectiveness of the process and also could reduce the operating costs associated to the application of AOPs. In this chapter, there is a description of our work in the treatment of wastewater using an ozonation-electrooxidation combined process. The main parameters to control for having a successful application of such method are discussed. Several examples for different kinds of polluted water are addressed.

**Keywords:** organic pollutants, degradation, mass transfer, mineralization, removal

## **1. Introduction**

Traditional wastewater treatments involve the addition of chemicals or the use of microorganisms to treat polluted water. However, in both processes, there is always a residue known as sludge. The sludge management and final disposal could represent up to 50% of the total wastewater treatment plant cost. Therefore, novel ways to deal with this issue should be developed. In this way, the use of advanced oxidation processes (AOPs), in which the HO• radical production is favored, could represent an interesting option for treat wastewater with less or without sludge production.

The final goal of AOPS is the complete degradation of the pollutants present in wastewater, aiming its final mineralization, yielding as final products: carbon dioxide, water and inorganic

compounds. These methodologies solve the problem of the final disposal of sludge; because when they are well developed, there is no production of sludge. Obviously, not always is possible the complete mineralization of contaminates. Nevertheless, most of the times, the final products of the destruction of contaminants are harmless compared to the original ones.

Electrochemical techniques use one of the cleanest reagents: "the electron." Thus, since the main reactive used in the oxidation process is green, this becomes sustainable. Oxidation of the organic compounds could occur at the interface of the anode/aqueous solution or in solution via intermediates. Electrochemical oxidation consists in the application of an external source of energy into an electrochemical cell that contains one or more pairs of electrodes. At the cathode, a reduction reaction occurs and the oxidation reactions takes places at the anode. The use of boron diamond doped anodes (BDD) allows the generation of HO• radicals, which reacts with organic compounds.

Electrochemical oxidation is considered a robust technology and is easy to use, for those reasons, it has been used for a diversity of wastewater treatments. The main advantages of this technology over other conventional treatments are as follows: the main reagent is the electron; many processes occur in the electrochemical cell; the addition of chemicals is not required; and the process is carried out at room temperature and atmospheric pressure.

Ozone is a powerful oxidant produced in gas phase, and by means of a diffuser, a mass transfer occurs to aqueous solution. A main advantage of ozone is that it oxidizes organic compounds without producing residual sludge. However, it was found that some compounds are ozone-resistant such as iopamidol, sucralose and atrazine-desethyl.

Combination of the ozonation processes with others, such as ozone/hydrogen peroxide, ozone with sand filtration and activated carbon filtration, have been used to remove ozone resistant contaminants; this allows enhancing the removal efficiency and reducing ozone dosage. Recent reports indicates that there are two reaction mechanisms for ozone oxidation: direct ozonation that takes place at acidic solutions and indirect HO• radical ozonation at basic solutions.

One of the major limitations for the use of ozone is the mass transfer from the gas phase to the liquid phase, the same behavior is observed at direct electrooxidation in which the HO• generation takes place at the anode surface. Thus, when both processes take place at the same time a synergy occurs, the process reaction time is decreased, this implies that the ozone and electricity consumption is also reduced.

In this chapter, there is a description of our work in the treatment of wastewater using an ozonation-electrooxidation combined process. The main parameters to control for having a successful application of such method are discussed. Several examples for different kinds of polluted water are addressed.

## **2. Characteristics of the hydroxyl radical**

Advanced oxidation processes rely on the hydroxyl radical formation. The hydroxyl radical, OH• is a highly reactive radical, able to react unselectively and rapidly with organic  pollutants, including recalcitrant organic compounds, such as aromatic, chlorinated and phenolic compounds [1].

There are different technologies to produce hydroxyl radicals. Nevertheless, the most "greener" one is its electrochemical production direct from the treated water. Among all the electrodes used in the production of hydroxyl radicals, the BDD anodes have shown to be perhaps the most efficient ones. They also have some other useful characteristics that allowed his use in the wastewater treatment, such as the radicals are loosely retained in the surface of the electrode allowing its oxidative action close to the surface area.

No matter from which source the hydroxyl radicals are generated, they have some special characteristics that make the extremely useful in the treatment of wastewater polluted with recalcitrant organic compounds. Some of the characteristics of the hydroxyl radical are as follows:


compounds. These methodologies solve the problem of the final disposal of sludge; because when they are well developed, there is no production of sludge. Obviously, not always is possible the complete mineralization of contaminates. Nevertheless, most of the times, the final products of the destruction of contaminants are harmless compared to the original ones. Electrochemical techniques use one of the cleanest reagents: "the electron." Thus, since the main reactive used in the oxidation process is green, this becomes sustainable. Oxidation of the organic compounds could occur at the interface of the anode/aqueous solution or in solution via intermediates. Electrochemical oxidation consists in the application of an external source of energy into an electrochemical cell that contains one or more pairs of electrodes. At the cathode, a reduction reaction occurs and the oxidation reactions takes places at the anode. The use of boron diamond doped anodes (BDD) allows the generation of HO• radi-

Electrochemical oxidation is considered a robust technology and is easy to use, for those reasons, it has been used for a diversity of wastewater treatments. The main advantages of this technology over other conventional treatments are as follows: the main reagent is the electron; many processes occur in the electrochemical cell; the addition of chemicals is not required;

Ozone is a powerful oxidant produced in gas phase, and by means of a diffuser, a mass transfer occurs to aqueous solution. A main advantage of ozone is that it oxidizes organic compounds without producing residual sludge. However, it was found that some compounds are

Combination of the ozonation processes with others, such as ozone/hydrogen peroxide, ozone with sand filtration and activated carbon filtration, have been used to remove ozone resistant contaminants; this allows enhancing the removal efficiency and reducing ozone dosage. Recent reports indicates that there are two reaction mechanisms for ozone oxidation: direct ozonation that takes place at acidic solutions and indirect HO• radical ozonation at basic solutions.

One of the major limitations for the use of ozone is the mass transfer from the gas phase to the liquid phase, the same behavior is observed at direct electrooxidation in which the HO• generation takes place at the anode surface. Thus, when both processes take place at the same time a synergy occurs, the process reaction time is decreased, this implies that the ozone and

In this chapter, there is a description of our work in the treatment of wastewater using an ozonation-electrooxidation combined process. The main parameters to control for having a successful application of such method are discussed. Several examples for different kinds of

Advanced oxidation processes rely on the hydroxyl radical formation. The hydroxyl radical, OH• is a highly reactive radical, able to react unselectively and rapidly with organic

and the process is carried out at room temperature and atmospheric pressure.

ozone-resistant such as iopamidol, sucralose and atrazine-desethyl.

cals, which reacts with organic compounds.

44 Physico-Chemical Wastewater Treatment and Resource Recovery

electricity consumption is also reduced.

**2. Characteristics of the hydroxyl radical**

polluted water are addressed.


The hydroxyl radical has a high oxidation potential as shown in **Table 1**, it can be generated, chemically, electrochemically or by UV radiation combined with the presence of suitable catalyst. The major failure of the electrochemical oxidation is the high-energy consumption during the process of mineralization of pollutants.


**Table 1.** Oxidation potential for some common oxidants [2].

## **3. Ozone**

Ozone is a pale blue gas with a pungent odor. It is generated from oxygen. The electric discharge method is the most common process for the preparation of ozone on laboratory and industrial scale. The electrical discharge generates ionized oxygen atoms that react with oxygen molecules to producing ozone.

Ozone is a powerful oxidant and very highly unstable. For this reason, the gas should be produce *in situ* prior to its use on the wastewater treatment. Once the ozone is produced, a diffuser is used to transport the gas to aqueous solution trough a mass transfer process. A major advantage of ozone is that it fully degrades organic materials, leaving no residual sludge.

Ozone can oxidize and destroy the organics through two different pathways, the direct and the indirect ones. In the first, one of the molecules react directly with the ozone molecules. In the second case, the ozone reacts to generate oxidant species such as hydroxyl radicals that carried out the oxidation process. The oxidation pathway that operates in a particular oxidation depends on the reaction rate of ozone and the organic. Sometimes the product generated in the reaction could promote or inhibit the ozone decomposition modifying the initial oxidation mechanism.

It has been found that some compounds are resistant to the oxidation by ozone, such as iopamidol, sucralose and atrazine-desethyl. In order to overcome this limitations, the use of combined ozone processes, such as the UV light, metal oxides catalyst and hydrogen peroxides, have been proposed. In many cases, a remarkable procedure improvement was found [3–5]

**Figure 1** shows an ozone bubble column reactor in which ozone is feed in the bottom part of the reactor, O3 passes through a diffuser that allows the generation of small bubbles which reaches the wastewater contained inside the reactor.

**Figure 1.** Schematic representation of a column bubble reactor in which ozone is feed in the bottom part.

The decomposition of ozone in water to form hydroxyl radicals, which occurs as are shown in Eqs. (**1**)–(**6**) [6]:

$$\text{O}\_3 + \text{OH}^- \rightarrow \text{O}\_2 + \text{HO}\_2^- \tag{1}$$

$$\rm{O}\_3 + \rm{HO}\_2^- \rightarrow \rm{HO}\_2^- + \rm{O}\_3^- \tag{2}$$

$$\text{HO}\_2^\cdot \rightarrow \text{H}^+ + \text{O}\_2^- \tag{3}$$

$$\rm O\_2^{\cdot -} + O\_3 \rightarrow O\_2 + O\_3^{\cdot -} \tag{4}$$

$$\rm{O}\_{3}^{-} + \rm{H}^{+} \rightarrow \rm{HO}\_{3}^{-} \tag{5}$$

$$\rm{HO}\_3^\cdot \rightarrow \rm{OH}^\cdot + \rm{O}\_2^\cdot \tag{6}$$

As observed, it takes six reactions to form one hydroxyl radical; now, a mass transfer from the gas phase to the aqueous phase should take place in order to have available hydroxyl radicals in aqueous solution. The process is often limited since only a part of ozone is effectively converted to hydroxyl radicals [7].

#### **4. Electrooxidation**

industrial scale. The electrical discharge generates ionized oxygen atoms that react with oxy-

Ozone is a powerful oxidant and very highly unstable. For this reason, the gas should be produce *in situ* prior to its use on the wastewater treatment. Once the ozone is produced, a diffuser is used to transport the gas to aqueous solution trough a mass transfer process. A major advantage of ozone is that it fully degrades organic materials, leaving no residual sludge.

Ozone can oxidize and destroy the organics through two different pathways, the direct and the indirect ones. In the first, one of the molecules react directly with the ozone molecules. In the second case, the ozone reacts to generate oxidant species such as hydroxyl radicals that carried out the oxidation process. The oxidation pathway that operates in a particular oxidation depends on the reaction rate of ozone and the organic. Sometimes the product generated in the reaction could promote or inhibit the ozone decomposition modifying the initial oxida-

It has been found that some compounds are resistant to the oxidation by ozone, such as iopamidol, sucralose and atrazine-desethyl. In order to overcome this limitations, the use of combined ozone processes, such as the UV light, metal oxides catalyst and hydrogen peroxides, have been proposed. In many cases, a remarkable procedure improvement was found [3–5]

**Figure 1** shows an ozone bubble column reactor in which ozone is feed in the bottom part of

**Figure 1.** Schematic representation of a column bubble reactor in which ozone is feed in the bottom part.

passes through a diffuser that allows the generation of small bubbles which

gen molecules to producing ozone.

46 Physico-Chemical Wastewater Treatment and Resource Recovery

tion mechanism.

the reactor, O3

reaches the wastewater contained inside the reactor.

Electrochemical oxidation is considered a robust technology and easy to use, for that reasons, it has been used for a diversity of wastewater treatment areas. The main advantages of this technology over other conventional treatments are as follows:


In 2003, Marselli et al. demonstrated that the production of hydroxyl radicals during conductivediamond electrolysis of aqueous wastes is possible. Consequently, a very new class of oxidation processes, the electrochemical advanced oxidation processes (EAOP) were discovered [8].

In the direct electrooxidation, pollutants in the bulk of the wastewater must reach the electrode surface and the oxidation reaction takes places once they are adsorbed onto this surface. Thus, the electrode materials influence the selectivity and efficiency of the oxidation process and mass transfer becomes a very important process. **Table 2** shows some anodic materials that have been investigated for the oxidation of organic compounds.


**Table 2.** Oxygen evolution potential of some electrodic materials [9].

**Figure 2** shows the general scheme of an electrochemical reactor. It contains an anode made of boron diamond doped in which hydroxyl radicals are produced. The cathode is made of stainless steel and allows water reduction.

**Figure 2.** An electrooxidation reactor in which hydroxyl radicals are produced in the anode and water reduction takes place in the cathode.

The main reactions involved in the hydroxyl radicals production are shown in Eqs. (**7**)–(**9**)

$$\bullet \, DBB + \mathrm{H}\_{2}\mathrm{O} \rightarrow \mathrm{D}BB\Big(\bullet \, \mathrm{OH}\Big) + \mathrm{H}^{\*} + \mathrm{e}^{-} \tag{7}$$

$$\text{ADDB} \left( \bullet \text{OH} \right) + \text{R} \rightarrow \text{DBB} + \text{migration products} + \text{H}^+ + \text{e}^- \tag{8}$$

$$\text{ODBB} \Big( \bullet \text{OH} \Big) \to \text{DBB} + \frac{1}{2} \text{O}\_2 + \text{H}^+ + \text{e}^- \tag{9}$$

Diamond anodes exhibits three outstanding properties as compared with other advanced oxidation technologies and with electrolysis with other anodes [10]:


However, as can be observed in **Figure 2**, the hydroxyl formation is limited to the anodic surface and also by the mass transfer from the liquid to the electrode.

#### **5. Integrated ozonization-electrooxidation reactor**

In order to have a synergistic effect of the two previously described processes, a couple treatment consisting in introducing electrodes inside the ozone reactor has been proposed.

**Figure 3** shows an ozone bubble column reactor in which two electrodes are introduced. As can be observed, the bubbles generated by the addition of ozone in the bottom part of the reactor allowing a complete mixing of the solution.

In this reactor, the ozone and the electrooxidation reaction takes place at the same time, thus the hydroxyl radical concentration is enhanced, the mass transfer is limited and a large amount of bubbles provides an excellent mixing inside the reactor. There are several variables to control in the integrated process aiming to obtain a complete degradation of pollutants:


**Figure 2** shows the general scheme of an electrochemical reactor. It contains an anode made of boron diamond doped in which hydroxyl radicals are produced. The cathode is made of

**Material Oxygen evolution potential**

Pt 1.60 Graphite 1.70 SnO<sup>2</sup> 1.90 PbO2 1.90 Boron doped diamond 2.30

48 Physico-Chemical Wastewater Treatment and Resource Recovery

**Table 2.** Oxygen evolution potential of some electrodic materials [9].

The main reactions involved in the hydroxyl radicals production are shown in Eqs. (**7**)–(**9**)

**Figure 2.** An electrooxidation reactor in which hydroxyl radicals are produced in the anode and water reduction takes

*DBB DBB* H O •OH H e <sup>2</sup> ( ) + − + → ++ (7)

stainless steel and allows water reduction.

place in the cathode.


**Figure 3.** An electrooxidation reactor in which hydroxyl radicals are produced in the anode and water reduction takes place in the cathode.

With a set of well-optimized parameters, there is always in improvement in results compared with the two separated techniques. In **Table 3**, some samples are gathered in which the integrated ozone-electrooxidation process has been applied to different kinds of wastewater. As it is possible to observe, there is a significant improvement in the quality of the treated wastewater. In all the cases, the chemical oxygen demand (cod) is almost eliminated, and some other parameters are also decreased in an important amount. The most used parameters to control the quality of treated wastewater are as follows: conductivity, total organic carbon (TOC), color, turbidity and biochemical oxygen demand (BOD5 ), the values of these parameters obtained before and after of the coupled treatment demonstrate the suitability of the proposed procedure to treat wastewater from different sources.



**Table 3.** Examples of the electrooxidation-ozonation process applied to wastewater treatment.

## **6. Conclusions**

Electrooxidation-ozonation is an efficient process for the treatment of different kinds of wastewater, since there is always a large reduction in COD, color, and turbidity, conductivity and BOD5 . The coupled process always has a superior performance compared with the application of separated processes. It is also noteworthy to mention that the coupled process is green, as it does not produce residual sludge. This coupled process has the potential to be used in wastewater in which other processes do not work well, including those with recalcitrant pollutants.

## **Acknowledgements**

The authors wish to acknowledge financial support from the UAEM, and the financial support from the CONACYT through Sistema Nacional de Investigadores, which is greatly appreciated.

#### **Nomenclature**

With a set of well-optimized parameters, there is always in improvement in results compared with the two separated techniques. In **Table 3**, some samples are gathered in which the integrated ozone-electrooxidation process has been applied to different kinds of wastewater. As it is possible to observe, there is a significant improvement in the quality of the treated wastewater. In all the cases, the chemical oxygen demand (cod) is almost eliminated, and some other parameters are also decreased in an important amount. The most used parameters to control the quality of treated wastewater are as follows: conductivity, total organic carbon

**Figure 3.** An electrooxidation reactor in which hydroxyl radicals are produced in the anode and water reduction takes

eters obtained before and after of the coupled treatment demonstrate the suitability of the

COD reduction was observed to be 80% after 44 min of treatment. Initial

**Results References**

), the values of these param-

[11]

[12]

(79%), color (95%),

(TOC), color, turbidity and biochemical oxygen demand (BOD5

50 Physico-Chemical Wastewater Treatment and Resource Recovery

place in the cathode.

proposed procedure to treat wastewater from different sources.

pH was 7.5 at all experiments

Industrial wastewater Integration of the two processes at pH 7 and 20 mA cm−2 of current density greatly improved the reduction in COD (84%), BOD5

Industrial wastewater In the integrated electrochemical-ozone process with energy pulses, the

turbidity (96%) and total coliforms (99%)

BOD5 Biochemical oxygen demand


## **Author details**

Carlos E. Barrera-Diaz\* and Nelly González-Rivas

\*Address all correspondence to: cbd0044@gmail.com

Joint Center for Research in Sustainable Chemistry, UAEM-UNAM, Toluca Estado de México, México


[13] García-Morales MA, Roa-Morales G, Barrera-Díaz C, Martínez Miranda V, Balderas Hernández P, Pavón Silva TB. Integrated advanced oxidation process (ozonation) and electrocoagulation treatments for dye removal in denim effluents. International Journal of Electrochemical Science. 2013;**8**:8752–8763.

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[4] Khuntia S, Majumder SK, Ghosh P. Quantitative prediction of generation of hydroxyl radicals from ozone microbubbles. Chemical Engineering Research and Design,

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[7] Barrera Díaz CE, González-Rivas N. The use of Al, Cu, and Fe in an integrated electrocoagulation-ozonation process. Journal of Chemistry. 2015; 6 (Article ID 158675).

[8] Marselli B, Garcia-Gomez J, Michaud PA, Rodrigo MA, Comninellis C. Electrogeneration of hydroxyl radicals on boron-doped diamond electrodes. Journal of the Electrochemical

[9] Panizza M, Cerisola G. Direct and mediated anodic oxidation of organic pollutants.

[10] Rodrigo MA, Cañizares P, Sánchez-Carretero A, Sáez C. Use of conductive-diamond electrochemical oxidation for wastewater treatment. Catalysis Today. 2010;**15**:173–177.

[11] Bernal-Martínez LA, Barrera-Díaz C, Solís-Morelos C, Natividad R. Synergy of electrochemical and ozonation processes in industrial wastewater treatment. Chemical

[12] Bernal-Martínez LA, Barrera-Díaz C, Natividad R, Rodrigo MA. Effect of the continuous and pulse in situ iron addition onto the performance of an integrated electrochemical–ozone reactor for wastewater treatment. Fuel. 2013;**110**:133–140. doi:10.1016/j.

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#### **Removal of Phenol from Wastewater Using Fenton-Like Reaction over Iron Oxide–Modified Silicates Removal of Phenol from Wastewater Using Fenton-Like Reaction over Iron Oxide–Modified Silicates**

Agnieszka Węgrzyn Agnieszka Węgrzyn

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65097

#### **Abstract**

Iron-containing active phase was deposited on natural layered silicate (vermiculite) using several techniques such as ion exchange, precipitation, and forced hydrolysis during hydrothermal digestion. Tuning of the synthesis conditions resulted in preparation of the catalysts with different loading of active phase and physicochemical properties. The composite materials were characterized with respect to their structure (X-ray diffraction), agglomeration state of Fe (diffuse reflectance UV-vis spectroscopy), and chemical composition. Catalytic tests were performed in semi-batch reactor under atmospheric pressure. Aqueous solution of phenol was used as a model industrial effluent, and hydrogen peroxide was added as an oxidant. Spectral techniques were used for identification of intermediate oxidation products. Spent catalysts were also characterized, and structural and chemical changes were determined, e.g., leaching degree of active phase.

**Keywords:** Fenton-like process, advanced oxidation processes, catalysis, silicate, vermiculite, nanocrystalline iron oxide, phenol

## **1. Introduction**

Refractory organic compounds, such as dyes, phenols, or endocrine disrupting compounds (EDC), are characterized with high toxicity, carcinogenic properties, and this poses a serious hazard to aquatic living organisms. Difficulty of contaminations' removal is caused by their resistance to aerobic digestion, stability to light, heat, and oxidizing agents. Technologies used currently for wastewater treatment, however, used widely, suffer from design shortcomings or are very expensive. Emerging technologies, so-called advanced oxidation processes (AOP), is

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

a large group of methods based on oxidation using strong oxidants, such as ozone or hydrogen peroxide. In AOP methods, higher conversion levels may be obtained at atmospheric pressure and temperatures lower compared to other oxidation processes [1–6]. Moreover, chemical oxidants may be accompanied by catalysts or physical agents such as sunlight, UV or γ radiation, ultrasounds, microwave, or cavitation, increasing efficiency of the reaction [7, 8].

The catalysts used in the Fenton-like system are, among others, natural iron-bearing earth materials, such as goethite, hematite, magnetite, or ferrihydrite [9–11]. Modification of iron oxides to improve their performance in organic pollutant degradation can be achieved by substitution with other transition metals [12]; however, introduction of heavy metals may be questionable from the point of view of the secondary contamination with catalytic leachates. It is also known that nanoscale materials are characterized with different properties compared to their bulk phase [8, 13, 14]. Nanocatalysts offer higher specific surface areas and few or no mass-transfer limitations. It is expected that reaction rate will be higher for nanomaterials. Also, diffusion of large organic molecules (organic dyes, pharmaceuticals) will be no longer problematic as it is observed in microporous materials. On the other hand, the separation and recycling of nanocatalysts at a technical scale still present a challenge.

To circumvent the costly catalyst separation process, magnetic properties of some iron oxides may be exploited [15, 16]. The other possibility is the immobilization on solid support. The most popular materials in this group are activated carbon, silica, and aluminum oxide [17– 21]; however, more advanced technologies are also studied employing graphene oxide [22]. The encapsulation of iron oxide nanoparticles in polymer matrix or carbonized sewer sludge was reported as another possibility to stabilize oxide nanoparticles [23, 24].

Facing much more stringent environmental regulations, new waste-free technologies must be developed, based on cheaper, non-toxic materials. Clays proposed as starting materials fulfill all requirements for low-cost, ecological precursors for industrial technologies or large-scale applications. Such materials could be used as catalysts in a large group of emerging technologies consisting on oxidation processes, such as wet oxidation, catalytic wet air oxidation, and advanced oxidation processes. Natural clay minerals provide with excellent support for Fecontaining nanocrystalline active phase of the Fenton-like reaction. Vermiculite, which was used in presented work as a catalytic support, is natural clay mineral belonging to phyllosilicates. It is characterized with high thermal and mechanical stability. Moreover, its properties may be easily modified to obtain efficient adsorbents or catalysts [25, 26].

## **2. Materials and methods**

Commercial expanded vermiculite (South Africa), fraction size 0.5–2 mm, was provided by Romico Polska Sp. z o.o. The silicate was pulverized in electrical blender, and fraction below 180 μm was separated (sample S0). Such prepared vermiculite was used as a support for deposition of nanocrystalline iron oxides.

Two standard procedures [27, 28] were applied to obtain well-defined oxide structures. Pure 2-line ferrihydrite was prepared by precipitation from 0.1 M Fe(NO3)3·9H2O (p.a., POCh) solution using 1 M KOH (p.a., POCh). Potassium hydroxide solution was added dropwise at RT and constant stirring until pH was equal to 7. Product was centrifuged, washed with water, and freeze dried. Similar procedure was used to obtain vermiculite-supported ferrihydrite. Suspension of 5 g of vermiculite (S0) was prepared in 150 mL of distilled water, then 100 mL of Fe(NO3)3·9H2O solution was added dropwise. Suspension was stirred for the next 2 h to allow ion exchange. In the next step, 1 M KOH solution was added to raise pH up to 7. Crystallization was continued for the next 30 min, product was centrifuged, washed, and dried. Sample codes, depending on Fe/vermiculite ratio, were S2, S3, and S4 (**Table 1**).


**Table 1.** Intended Fe/vermiculite ratio in vermiculite-supported Fe oxide catalysts.

a large group of methods based on oxidation using strong oxidants, such as ozone or hydrogen peroxide. In AOP methods, higher conversion levels may be obtained at atmospheric pressure and temperatures lower compared to other oxidation processes [1–6]. Moreover, chemical oxidants may be accompanied by catalysts or physical agents such as sunlight, UV or γ radiation,

The catalysts used in the Fenton-like system are, among others, natural iron-bearing earth materials, such as goethite, hematite, magnetite, or ferrihydrite [9–11]. Modification of iron oxides to improve their performance in organic pollutant degradation can be achieved by substitution with other transition metals [12]; however, introduction of heavy metals may be questionable from the point of view of the secondary contamination with catalytic leachates. It is also known that nanoscale materials are characterized with different properties compared to their bulk phase [8, 13, 14]. Nanocatalysts offer higher specific surface areas and few or no mass-transfer limitations. It is expected that reaction rate will be higher for nanomaterials. Also, diffusion of large organic molecules (organic dyes, pharmaceuticals) will be no longer problematic as it is observed in microporous materials. On the other hand, the separation and

To circumvent the costly catalyst separation process, magnetic properties of some iron oxides may be exploited [15, 16]. The other possibility is the immobilization on solid support. The most popular materials in this group are activated carbon, silica, and aluminum oxide [17– 21]; however, more advanced technologies are also studied employing graphene oxide [22]. The encapsulation of iron oxide nanoparticles in polymer matrix or carbonized sewer sludge

Facing much more stringent environmental regulations, new waste-free technologies must be developed, based on cheaper, non-toxic materials. Clays proposed as starting materials fulfill all requirements for low-cost, ecological precursors for industrial technologies or large-scale applications. Such materials could be used as catalysts in a large group of emerging technologies consisting on oxidation processes, such as wet oxidation, catalytic wet air oxidation, and advanced oxidation processes. Natural clay minerals provide with excellent support for Fecontaining nanocrystalline active phase of the Fenton-like reaction. Vermiculite, which was used in presented work as a catalytic support, is natural clay mineral belonging to phyllosilicates. It is characterized with high thermal and mechanical stability. Moreover, its properties

Commercial expanded vermiculite (South Africa), fraction size 0.5–2 mm, was provided by Romico Polska Sp. z o.o. The silicate was pulverized in electrical blender, and fraction below 180 μm was separated (sample S0). Such prepared vermiculite was used as a support for

Two standard procedures [27, 28] were applied to obtain well-defined oxide structures. Pure 2-line ferrihydrite was prepared by precipitation from 0.1 M Fe(NO3)3·9H2O (p.a., POCh)

ultrasounds, microwave, or cavitation, increasing efficiency of the reaction [7, 8].

56 Physico-Chemical Wastewater Treatment and Resource Recovery

recycling of nanocatalysts at a technical scale still present a challenge.

was reported as another possibility to stabilize oxide nanoparticles [23, 24].

may be easily modified to obtain efficient adsorbents or catalysts [25, 26].

**2. Materials and methods**

deposition of nanocrystalline iron oxides.

Pure hematite with crystal size of 4 nm was prepared by forced hydrolysis. 3.32 g of Fe(NO3)3·9H2O was dissolved in preheated HCl (p.a., POCh) solution (0.002 M, 400 mL) to obtain Fe concentration of 0.02 M.. Solution was transferred into polypropylene bottle fitted in autoclave and heated at 98°C for 7 days. Product was centrifuged, washed with water, and freeze dried. Synthesis of vermiculite-supported nano-hematite was performed using acidified iron nitrate solutions with the addition of 20 g of vermiculite. Sample codes, depending on Fe/ vermiculite ratio, were HS0 (no Fe salt was added), HS1, HS2, and HS3 (**Table 1**).

Phenol removal was studied as a test reaction, and semi-batch reactor was used to minimize formation of side products [26]. Round-bottom flask equipped with reflux condenser was heated to 70°C on magnetic stirrer. Each time reactor was charged with 340 mL of phenol solution (pH = 5.4) and 600 mg of catalyst. Hydrogen peroxide (30%, p.a., POCh) was added into the reaction mixture in 13-min intervals (10 min of non-disturbed reaction and 3 min for sample withdrawal and next injection). Phenol concentration was studied spectrophotometrically (Thermo SCIENTIFIC EVOLUTION 220) as a complex with 4-aminoantipyrine. H2O2 concentration (using VO3 − in 8 M H2SO4), Fe dissolved in reaction mixture (SCN− complex), and colored intermediate products (sample quenched with methanol) were also determined spectrophotometrically. In each interval, pH was measured. Reaction conditions were summarized in **Table 2**.


**Table 2.** Reaction conditions of phenol removal over vermiculite-supported iron oxide catalysts.

The conversion, *X* (%), of model pollutant (phenol) was calculated according to Eq. (1):

$$X = \frac{C\_0 - C}{C\_0} \cdot 100\% \tag{1}$$

where *C*0 is the starting concentration and *C* is the concentration at a given reaction time.

Fresh and spent catalysts were characterized by X-ray diffraction method (XRD) using a powder diffractometer (Bruker, D2 PHASER) equipped with CuKα radiation source. The Sherrer equation (2) was used for determination of nano-hematite crystal size:

$$D = \frac{0.89\lambda}{\beta \cos \theta} \tag{2}$$

where *λ* is the X-ray wavelength, *β* is the line broadening at half the maximum intensity, and *θ* is the Bragg angle.

The coordination and aggregation of iron present in the catalysts were studied by diffuse reflectance-UV-vis spectroscopy (DRS-UV-vis). The measurements were performed in the range of 190–900 nm with a resolution of 2 nm using an Evolution 600 (Thermo) spectrophotometer. Content of iron was measured using spectrophotometric technique at wavelength *λ* = 510 nm (Thermo SCIENTIFIC EVOLUTION 220) as a complex with 1,10-phenanthroline after leaching of metal cations in 6 M HCl.

### **3. Results and discussion**

#### **3.1. Catalytic tests**

Three types of catalytic tests were carried out in semi-batch reactor: concentrated (R1—1 g/L) and diluted (R01—0.1 g/L) phenol solutions with the addition of significant excess of oxidant (six times 2 mL) and diluted phenol solution with the minimum amount of oxidant added (six times 0.2 mL). In each series of catalytic tests, it was observed that initiation phase is the first step, as in the case of free-radical reactions, especially for experiments carried out in concentrated phenol solution (**Figure 1A**). Initial 10–20 min are characterized with slow increase in pollutant conversion. After 30–50 min of the reaction over iron oxide-containing catalysts, conversion rapidly increased reaching values above 95%. Non-modified silicates, on the other hand, presented very low activity. Sample submitted to hydrothermal treatment (HS0) slightly increased phenol oxidation compared to non-catalytic process; however, in the latter case, conversion was not higher than 8% after 75 min. On the contrary, the addition of starting vermiculite (S0) to reaction mixture resulted in slow increase in phenol conversion up to 42%. Reduction in particle size was the only preparation step in this case; therefore, contaminations present in the starting materials, such as interlayer and adsorbed transition metal cations as well as naturally occurring iron oxides and carbonates, may be responsible for the observed catalytic effect.

The conversion, *X* (%), of model pollutant (phenol) was calculated according to Eq. (1):

where *C*0 is the starting concentration and *C* is the concentration at a given reaction time.

Fresh and spent catalysts were characterized by X-ray diffraction method (XRD) using a powder diffractometer (Bruker, D2 PHASER) equipped with CuKα radiation source. The

l

where *λ* is the X-ray wavelength, *β* is the line broadening at half the maximum intensity, and

The coordination and aggregation of iron present in the catalysts were studied by diffuse reflectance-UV-vis spectroscopy (DRS-UV-vis). The measurements were performed in the range of 190–900 nm with a resolution of 2 nm using an Evolution 600 (Thermo) spectrophotometer. Content of iron was measured using spectrophotometric technique at wavelength *λ* = 510 nm (Thermo SCIENTIFIC EVOLUTION 220) as a complex with 1,10-phenanthroline after

Three types of catalytic tests were carried out in semi-batch reactor: concentrated (R1—1 g/L) and diluted (R01—0.1 g/L) phenol solutions with the addition of significant excess of oxidant (six times 2 mL) and diluted phenol solution with the minimum amount of oxidant added (six times 0.2 mL). In each series of catalytic tests, it was observed that initiation phase is the first step, as in the case of free-radical reactions, especially for experiments carried out in concentrated phenol solution (**Figure 1A**). Initial 10–20 min are characterized with slow increase in pollutant conversion. After 30–50 min of the reaction over iron oxide-containing catalysts, conversion rapidly increased reaching values above 95%. Non-modified silicates, on the other hand, presented very low activity. Sample submitted to hydrothermal treatment (HS0) slightly increased phenol oxidation compared to non-catalytic process; however, in the latter case, conversion was not higher than 8% after 75 min. On the contrary, the addition of starting vermiculite (S0) to reaction mixture resulted in slow increase in phenol conversion up to 42%. Reduction in particle size was the only preparation step in this case; therefore, contaminations present in the starting materials, such as interlayer and adsorbed transition metal cations as


<sup>=</sup> (2)

0 0 100% *C C <sup>X</sup> C*

58 Physico-Chemical Wastewater Treatment and Resource Recovery

Sherrer equation (2) was used for determination of nano-hematite crystal size:

*θ* is the Bragg angle.

**3.1. Catalytic tests**

leaching of metal cations in 6 M HCl.

**3. Results and discussion**

0.89 *<sup>D</sup> cos*

b q

**Figure 1.** Conversion of phenol in oxidation reaction over vermiculite-supported iron oxide catalysts; A—initial concentration of phenol 1 g/L, volume of H2O2 injection = 2 mL; B—initial concentration of phenol 0.1 g/L, volume of H2O2 injection = 2 or 0.2 mL.

When diluted solution of phenol was used (**Figure 1B**) and accompanied by small excess of oxidant (R01m—0.2 mL), almost no effect was observed within assigned experimental time. Only one sample, doped with nano-hematite, HS2, showed catalytical properties after 50 min of reaction. The non-catalytical reaction performed with large excess of oxidant (R01—2 mL) resulted in quite significant conversion equal to 48% within 75 min. Slightly higher activity was observed when ferrihydrite-doped samples, S2 and S3, and non-modified silicates were added as catalysts. After constant gradual increase of conversion, it reached 60–75% within 75 minutes. Only one sample with the highest loading of ferrihydrite and samples doped with nano-hematite allowed to reach the conversion level above 95%. Nevertheless, it should be stressed that after initial increase in conversion, it was inhibited and much slower at longer reaction times in the case of removal of concentrated pollutant. Similar effect of the reaction stagnation, due to accumulation of the reaction products, was also observed in homogeneous Fenton reaction [29].

It may be observed that conversion of phenol was more effective with higher doping with ferrihydrite. On the contrary, regardless reaction conditions, in the series of nano-hematitecontaining catalysts, an optimum amount of iron oxide results in higher efficiency of the reaction. The best sample in this case was HS2 doped with 3.36 wt.% of iron in the form of nano-hematite.

According to results of the catalytic tests described above, 2-line ferrihydrite supported on vermiculite is less active than analogous materials containing hematite. Additional experiment, performed in concentrated phenol solution (1 g/L) and using active sample HS3 as catalyst, provided information about reaction path and intermediate products. UV-vis spectra for samples withdrawn during experiment, quenched with methanol or mixed additionally with VO3 − /H2SO4 solution, allowed to distinguish between transition products formed in the course of the reaction. After 29 min of the reaction, which corresponds to 43% of phenol conversion, colored products were formed. In the UV-vis spectrum recorded in methanol (**Figure 2A**) bands assigned to phenol (278 and 284 nm), hydroquinone (300 nm) and benzoquinone/quinhydrone (245, 255, and 300 nm) were identified. However, bands assigned to catechin were strongly overlapped by other strong peaks, and it cannot be excluded that this product was also formed. After 62 min of the experiment phenol conversion reached 97%, no colored products were recorded, and reaction was completed.

**Figure 2.** Identification of transition products of phenol oxidation over HS3 catalyst (reaction conditions: PhOH = 1 g/L, volume of H2O2 injection = 2 mL): A—derivative UV-vis spectra recorded in MeOH, B—derivative UV-vis spectra recorded in VO3 − /H2SO4; P—phenol; Q—qiunhydrone; B—benzoquinone; H—hydroquinone; C—catechin.

However, in the spectra measured after the reaction of the sample of effluent with VO3 − / H2SO4 mixture (**Figure 2B**), peaks below 230 nm, assigned to unidentified organic compounds, were recorded at the end of the test. Evolution of pH followed opposite trend as phenol conversion, and at 29 and 62 min, it was equal to 2.87 and 2.48, respectively. Those observations confirm that final products are not only H2O and CO2 but also organic acids.

#### **3.2. Characterization of as received and spent catalysts**

Iron oxide-bearing catalysts were obtained by direct deposition of formed oxide on the silicate support. Vermiculite was selected due to its mechanical and thermal stability. On the contrary to montmorillonite, it is not exfoliating rapidly in contact with water, and swelling is limited to changes of number of water molecules in the interlayer space. Moreover, mineral itself contains significant amount of iron.

As it was shown in **Figure 3**, both expected oxide structures were formed [27, 28]. XRD pattern of 2-line ferrihydrite consists of two broad reflections, while nano-hematite is characterized by the presence of several sharp but not intense peaks. Ferrihydrite structure was not observed after deposition on the support due to inherent poor ordering of the structure and low content in the composite material. On the other hand, using the Sherrer equation, it was confirmed that crystal size of pure nano-hematite phase was 4 nm. Only traces of nano-hematite could be identified in two vermiculite-supported samples with the highest loading of deposited phase —HS2 (3.36 Fe wt.%) and HS3 (6.72 Fe wt.%). Therefore, it was not possible to determine precise crystal parameters for oxide phase.

Removal of Phenol from Wastewater Using Fenton-Like Reaction over Iron Oxide–Modified Silicates http://dx.doi.org/10.5772/65097 61

conversion, colored products were formed. In the UV-vis spectrum recorded in methanol (**Figure 2A**) bands assigned to phenol (278 and 284 nm), hydroquinone (300 nm) and benzoquinone/quinhydrone (245, 255, and 300 nm) were identified. However, bands assigned to catechin were strongly overlapped by other strong peaks, and it cannot be excluded that this product was also formed. After 62 min of the experiment phenol conversion reached 97%, no

**Figure 2.** Identification of transition products of phenol oxidation over HS3 catalyst (reaction conditions: PhOH = 1 g/L, volume of H2O2 injection = 2 mL): A—derivative UV-vis spectra recorded in MeOH, B—derivative UV-vis spectra

However, in the spectra measured after the reaction of the sample of effluent with VO3

confirm that final products are not only H2O and CO2 but also organic acids.

**3.2. Characterization of as received and spent catalysts**

H2SO4 mixture (**Figure 2B**), peaks below 230 nm, assigned to unidentified organic compounds, were recorded at the end of the test. Evolution of pH followed opposite trend as phenol conversion, and at 29 and 62 min, it was equal to 2.87 and 2.48, respectively. Those observations

Iron oxide-bearing catalysts were obtained by direct deposition of formed oxide on the silicate support. Vermiculite was selected due to its mechanical and thermal stability. On the contrary to montmorillonite, it is not exfoliating rapidly in contact with water, and swelling is limited to changes of number of water molecules in the interlayer space. Moreover, mineral itself

As it was shown in **Figure 3**, both expected oxide structures were formed [27, 28]. XRD pattern of 2-line ferrihydrite consists of two broad reflections, while nano-hematite is characterized by the presence of several sharp but not intense peaks. Ferrihydrite structure was not observed after deposition on the support due to inherent poor ordering of the structure and low content in the composite material. On the other hand, using the Sherrer equation, it was confirmed that crystal size of pure nano-hematite phase was 4 nm. Only traces of nano-hematite could be identified in two vermiculite-supported samples with the highest loading of deposited phase —HS2 (3.36 Fe wt.%) and HS3 (6.72 Fe wt.%). Therefore, it was not possible to determine precise

/H2SO4; P—phenol; Q—qiunhydrone; B—benzoquinone; H—hydroquinone; C—catechin.

− /

colored products were recorded, and reaction was completed.

60 Physico-Chemical Wastewater Treatment and Resource Recovery

recorded in VO3

−

contains significant amount of iron.

crystal parameters for oxide phase.

**Figure 3.** Structure of vermiculite-supported ferrihydrite- (A) and nano-hematite-containing (B) catalysts; F—ferrihydrite; H—nano-hematite; 0, 1, 2—basal reflections of vermiculite corresponding to 0, 1, and 2 layers of interlayer water; i—interstratified vermiculite phases.

Changes in vermiculite structure reflected chemical modifications performed in each synthesis. Starting material (S0) was characterized with complex pattern typical for vermiculites both containing in the interlayer divalent cations and collapsed structure (0 layers of water). Moreover, the interlayer cations are accompanied with 1 or 2 layers of water. Additional peaks below 8° 2*θ* were assigned to interstratified contracting and non-contracting phases [30].

Upon hydrothermal treatment in the sample HS0 intensity of peak corresponding to one water layer increased, while disappeared peak assigned to the collapsed structure. In the synthesis of 2-line ferrihydrite composite (**Figure 3A**), a first step consisted on an ion exchange of interlayer anions for iron. As a result, peak at 1.40–1.43 nm may be observed; however, it was shifted to lower values for higher loadings of iron oxide: 1.39 and 1.37 nm for S3 and S4, respectively. Described phenomenon is a result of partial dehydration of the interlayer gallery and formation of so-called HIV—hydroxy-interlayered vermiculites [31, 32]. Similar shift was observed also for the sample with the highest nano-hematite content: HS3—1.39 nm.

Application of potassium hydroxide, however, resulted in a deeper rearrangement of interlayer space. Both peaks assigned to interstratification and one water layer almost disappeared. On the other hand, intercalation of K+ resulted in a large increase in peak intensity at 1 nm [33]. On the contrary, in the samples doped with nano-hematite in hydrothermal conditions (**Figure 3B**), peak assigned to 0 layers of interlayer water decreased with increasing amount of oxide.

In the structure of spent catalysts, traces of hematite were still possible to identify; however, other changes concerning catalyst properties were noticed. Vermiculite support upon reaction in concentrated solution was transformed into Mg2+/Fe3+ intercalated structure containing 2 layers in water molecules (**Figure 4A**). In the samples doped with ferrihydrite, only traces of interlayer potassium were preserved, and interlayer spaces were occupied with di- and trivalent cations released from silicate matrix. Hydration state and the number of water molecules strongly depended on the initial amount of iron oxide—the lower doping level the easier rehydration proceeded. Similar dependence was observed also for nano-hematite deposited samples. Such phenomenon should be explained as a result of blocking of interlayer spaces with iron hydroxides. Moreover, iron oxide particles, which were grown near the edges of vermiculite layers, may act as cementing agent, preventing structure swelling. It was also observed that rehydration of the structure depends on the reaction conditions (**Figure 4B**): the higher concentration of phenol and hydrogen peroxide, the easier intercalation of water molecules. As it was shown in **Figure 4C/D**, swelling intensity, which may be expressed as peaks 1.42 (2 layers of water) and 1.20 nm (1 layer of water) intensity ratio, increased at higher concentration of substrates. It cannot be excluded that acidic reaction products also enhanced structural changes of vermiculite support.

**Figure 4.** Structure of spent catalysts: A—vermiculite-supported iron oxide catalysts after reaction with phenol concentration 1 g/L; B and C—evolution of basal spacings of HS1 sample at different reaction conditions; D—evolution of basal spacings of nano-hematite-containing catalysts after reaction with phenol concentration 1 g/L; H—nano-hematite; 0, 1, 2—basal reflections of vermiculite corresponding to 0, 1, and 2 layers of interlayer water; i—interstratified vermiculite phases.


\*2.4-2.6 nm (low-intensity peak).

In the structure of spent catalysts, traces of hematite were still possible to identify; however, other changes concerning catalyst properties were noticed. Vermiculite support upon reaction in concentrated solution was transformed into Mg2+/Fe3+ intercalated structure containing 2 layers in water molecules (**Figure 4A**). In the samples doped with ferrihydrite, only traces of interlayer potassium were preserved, and interlayer spaces were occupied with di- and trivalent cations released from silicate matrix. Hydration state and the number of water molecules strongly depended on the initial amount of iron oxide—the lower doping level the easier rehydration proceeded. Similar dependence was observed also for nano-hematite deposited samples. Such phenomenon should be explained as a result of blocking of interlayer spaces with iron hydroxides. Moreover, iron oxide particles, which were grown near the edges of vermiculite layers, may act as cementing agent, preventing structure swelling. It was also observed that rehydration of the structure depends on the reaction conditions (**Figure 4B**): the higher concentration of phenol and hydrogen peroxide, the easier intercalation of water molecules. As it was shown in **Figure 4C/D**, swelling intensity, which may be expressed as peaks 1.42 (2 layers of water) and 1.20 nm (1 layer of water) intensity ratio, increased at higher concentration of substrates. It cannot be excluded that acidic reaction products also enhanced

**Figure 4.** Structure of spent catalysts: A—vermiculite-supported iron oxide catalysts after reaction with phenol concentration 1 g/L; B and C—evolution of basal spacings of HS1 sample at different reaction conditions; D—evolution of basal spacings of nano-hematite-containing catalysts after reaction with phenol concentration 1 g/L; H—nano-hematite; 0, 1, 2—basal reflections of vermiculite corresponding to 0, 1, and 2 layers of interlayer water; i—interstratified

structural changes of vermiculite support.

62 Physico-Chemical Wastewater Treatment and Resource Recovery

vermiculite phases.

**Table 3.** Interlayer distances of iron oxide-modified vermiculite-based catalysts before and after reaction.

The basal spacings calculated for modified vermiculites (**Table 3**) show that synthesis consisting on 2-line ferrihydrite precipitation resulted in the formation of hydroxy-interlayered phase and disappearance of peaks related to interstratified phases. Moreover, vermiculite was also partially intercalated with potassium. After the reaction, almost all peak positions returned to the initial values similar to the starting material. In nano-hematite modified samples, characteristic peaks for interstratification remained in their positions. However, in the course of phenol oxidation, first peak (~2.5 nm) became less noticeable.

It may be concluded that deposited iron oxide phases changed properties of the support; however, alteration was reversible in reaction conditions. Although XRD patterns do not allow to follow degradation of active phase directly, some indications of that process may be observed through properties of vermiculite.

More data considering properties of the deposited iron oxides were provided by DRS-UV-vis spectroscopy. As it was mentioned before, vermiculite itself contains iron [25] and UV-vis spectrum recorded for solid-state samples consisted of several characteristic bands. Isolated Fe3+ cations in the tetrahedral coordination give rise to peaks at 224 nm in both silicate materials (S0 and HS0), and cations in the octahedral coordination may be identified by the presence of band at 260 nm [34, 35]. The bands at 319 and 358 nm are characteristic for small oligonuclear Fe*x*O*y* clusters. Formation of bulk Fe2O3 particles gave characteristic bands above 400 nm [34].

Ferrihydrite and hematite were characterized by multiple bands, revealed by the second derivative spectra (results not shown), and assigned to the electronic transitions [27, 36]. The spectra of both oxides consisted of peaks at 260–264 nm, which should be assigned to charge transfer. The bands at 401 and 424 nm for ferrihydrite and hematite, respectively, resulted from 6 A1 → <sup>4</sup> E; 4 A1 transition, 519 and 550 nm—2(6 A1) → 2(4 T1) (electron pair transition, EPT), 690– 718 and 665 nm—6 A1 → <sup>4</sup> T2. Additionally, for nano-hematite, the following bands were assigned to 6 A1 → <sup>4</sup> T1 transitions at 310 and 840 nm and 6 A1 → <sup>4</sup> E transitions at 384 nm.

**Figure 5.** Agglomeration state of iron species in vermiculite-supported ferrihydrite- (A) and nano-hematite-containing (B) catalysts (DRS-UV-vis spectra).

Due to possible release of the cations from vermiculite and the contamination of deposited iron oxides during synthesis, the octahedra may be distorted, and consequently, ligand field and band positions may be changed. For 2-line ferrihydrite-containing catalysts, DRS-UV-vis peaks were shifted to 296, 456–478, 527–556, and 675 nm (**Figure 5A**). Similar result, with peak positions at 294, 442, 476, 544, 679, and 840 nm, was obtained for nano-hematite deposited on silicate (**Figure 5B**).

Further changes in the catalyst structure took place in the course of phenol oxidation. In the spectrum of starting silicate, S0, new band in the range of 360–480 nm was formed (**Fig‐ ure 6A**). It is possible that adsorbed on the surface and interlayer Fe3+ cations present in original material were released and redeposited in the form of larger clusters. Catalysts modified with ferrihydrite after reaction with diluted phenol solution (R01) were depleted with active phase, and DRS-UV-vis spectra have shown minimum at 300 and 480 nm. Much larger minimum was registered in the differential spectrum of sample S3 after reaction with concentrated phenol solution. The shape and positions of minima (390, 453, and 524–550 nm) reflected distribution of absorption peaks in fresh catalyst. It may be expected that degradation of the catalyst is significant, although it is mechanical rather than chemical in nature.

Removal of Phenol from Wastewater Using Fenton-Like Reaction over Iron Oxide–Modified Silicates http://dx.doi.org/10.5772/65097 65

Fe3+ cations in the tetrahedral coordination give rise to peaks at 224 nm in both silicate materials (S0 and HS0), and cations in the octahedral coordination may be identified by the presence of band at 260 nm [34, 35]. The bands at 319 and 358 nm are characteristic for small oligonuclear Fe*x*O*y* clusters. Formation of bulk Fe2O3 particles gave characteristic bands above 400 nm [34].

Ferrihydrite and hematite were characterized by multiple bands, revealed by the second derivative spectra (results not shown), and assigned to the electronic transitions [27, 36]. The spectra of both oxides consisted of peaks at 260–264 nm, which should be assigned to charge transfer. The bands at 401 and 424 nm for ferrihydrite and hematite, respectively, resulted from

**Figure 5.** Agglomeration state of iron species in vermiculite-supported ferrihydrite- (A) and nano-hematite-containing

Due to possible release of the cations from vermiculite and the contamination of deposited iron oxides during synthesis, the octahedra may be distorted, and consequently, ligand field and band positions may be changed. For 2-line ferrihydrite-containing catalysts, DRS-UV-vis peaks were shifted to 296, 456–478, 527–556, and 675 nm (**Figure 5A**). Similar result, with peak positions at 294, 442, 476, 544, 679, and 840 nm, was obtained for nano-hematite deposited on

Further changes in the catalyst structure took place in the course of phenol oxidation. In the spectrum of starting silicate, S0, new band in the range of 360–480 nm was formed (**Fig‐ ure 6A**). It is possible that adsorbed on the surface and interlayer Fe3+ cations present in original material were released and redeposited in the form of larger clusters. Catalysts modified with ferrihydrite after reaction with diluted phenol solution (R01) were depleted with active phase, and DRS-UV-vis spectra have shown minimum at 300 and 480 nm. Much larger minimum was registered in the differential spectrum of sample S3 after reaction with concentrated phenol solution. The shape and positions of minima (390, 453, and 524–550 nm) reflected distribution of absorption peaks in fresh catalyst. It may be expected that degradation of the catalyst is

significant, although it is mechanical rather than chemical in nature.

A1) → 2(4

T2. Additionally, for nano-hematite, the following bands were

A1 → <sup>4</sup>

T1) (electron pair transition, EPT), 690–

E transitions at 384 nm.

6 A1 → <sup>4</sup>

E; 4

assigned to 6

718 and 665 nm—6

A1 → <sup>4</sup>

(B) catalysts (DRS-UV-vis spectra).

silicate (**Figure 5B**).

A1 transition, 519 and 550 nm—2(6

T1 transitions at 310 and 840 nm and 6

A1 → <sup>4</sup>

64 Physico-Chemical Wastewater Treatment and Resource Recovery

**Figure 6.** Leaching of iron species from vermiculite-supported iron oxide-containing catalysts upon phenol oxidation reaction (DRS-UV-vis spectra).

In nano-hematite-containing catalysts, degradation proceeded differently for each sample. At the lowest loading of active phase (**Figure 6B**, HS1), leaching was the most noticeable compared to the other samples, which were used in the reaction with concentrated substrates (R1). Moreover, the largest minimum was observed at 360–400 nm, while at 454 and 530 nm, two smaller features were observed. When the amount of hematite was increasing, minima recorded in DRS-UV-vis spectra were smaller and shifted to higher wavelengths (**Figure 6C/D**). Therefore, it may be concluded that optimization of the active phase loading is more important for hematite-containing composites, both in terms of catalyst stability and its activity. Surprisingly, although degradation of the catalysts is less noticeable in the reaction using lower concentration of phenol, the addition of lower excess of oxidant may also increase leaching of active components (e.g., **Figure 6B**). This feature may result in olation-oxolation processes, proceeding differently in the presence of H2O2.

On the basis of catalyst characterization, the following model was proposed for more active silicate-based nano-hematite-modified materials (**Figure 7**). In optimum conditions of about 3.36 wt.% of iron, which corresponds to 4.8 wt.% of deposited iron oxide, interlayer spaces of vermiculite are not blocked by hydroxides and are free to accommodate Fe3+ cations. On the surface of the layered support, patches of nanocrystalline phase are formed. Below the optimum hematite loading, besides well-defined nanocrystals, also oligomeric clusters of iron oxide are deposited, which may be easily dissolved by the reaction substrates and products in the course of the reaction. The interlayer space of vermiculite is still available for an ionexchange process. Above the optimum loading of the active phase, interlayer spaces of vermiculite are blocked by hydroxy-compounds, which may be removed during the reaction. Deposited nano-hematite phase remains almost intact during the reaction.

**Figure 7.** Simplified structure of nano-hematite-containing vermiculite-supported catalysts.

#### **3.3. Catalytic activity vs. catalyst degradation**

Changes in the catalyst chemical composition were followed during the reaction and correlated with catalytical results. In **Table 4**, it was presented that ferrihydrite-containing catalysts were more susceptible to Fe leaching. Surprisingly, the lower was oxide doping, the higher percentage of active phase was dissolved. No such straight relationship was observed for nanohematite-containing catalysts. Apparently, small oligoclusters and interlayered hydroxyspecies described in model in Section 3.2, indeed, contributed significantly to dissolved species. It was also observed that catalytic activity should not be attributed completely to homogeneous reaction. Reaction mixtures over ferrihydrite-doped catalysts were characterized with higher concentration of Fe available for homogeneous reaction. Times, required to obtain phenol conversion equal 40 and 50%, were longer for ferrihydrite-containing catalysts in comparison to hematite-doped materials. Moreover, in the latter case Fe concentrations in the reaction mixtures were relatively low. As it was described in Section 3.1, when diluted phenol solution was used for the reaction activity stagnated due to product accumulation. Another explanation could be recombination of radicals formed over the catalysts. Therefore, time for 50% phenol conversion is more or less 10 min delayed compared to 40% conversion. On the other hand, time difference for the reactions performed in concentrated phenol solution is closer to 1–3 min. Another conclusion may be formed on the basis of the analysis of residual phenol concentrations. Within 75 min of the reaction, phenol concentration is reduced to 3–31 and 2– 6 mg/L for ferrihydrite- and hematite-containing catalysts, respectively, in reactions using starting solution equal to 100 mg/L. When 1 g/L phenol solution was used, final concentrations were equal to 8–11 and 15–30 mg/L for both iron containing series of catalysts. In this way, it was confirmed that dispersed pollutants are more difficult to remove efficiently then concentrated.


n.d., not determined.

oxide are deposited, which may be easily dissolved by the reaction substrates and products in the course of the reaction. The interlayer space of vermiculite is still available for an ionexchange process. Above the optimum loading of the active phase, interlayer spaces of vermiculite are blocked by hydroxy-compounds, which may be removed during the reaction.

Deposited nano-hematite phase remains almost intact during the reaction.

**Figure 7.** Simplified structure of nano-hematite-containing vermiculite-supported catalysts.

Changes in the catalyst chemical composition were followed during the reaction and correlated with catalytical results. In **Table 4**, it was presented that ferrihydrite-containing catalysts were more susceptible to Fe leaching. Surprisingly, the lower was oxide doping, the higher percentage of active phase was dissolved. No such straight relationship was observed for nanohematite-containing catalysts. Apparently, small oligoclusters and interlayered hydroxyspecies described in model in Section 3.2, indeed, contributed significantly to dissolved species. It was also observed that catalytic activity should not be attributed completely to homogeneous reaction. Reaction mixtures over ferrihydrite-doped catalysts were characterized with higher concentration of Fe available for homogeneous reaction. Times, required to obtain phenol conversion equal 40 and 50%, were longer for ferrihydrite-containing catalysts in comparison to hematite-doped materials. Moreover, in the latter case Fe concentrations in the reaction mixtures were relatively low. As it was described in Section 3.1, when diluted phenol solution was used for the reaction activity stagnated due to product accumulation. Another explanation could be recombination of radicals formed over the catalysts. Therefore, time for 50% phenol conversion is more or less 10 min delayed compared to 40% conversion. On the other hand, time difference for the reactions performed in concentrated phenol solution is closer to 1–3 min. Another conclusion may be formed on the basis of the analysis of residual phenol concentrations. Within 75 min of the reaction, phenol concentration is reduced to 3–31 and 2– 6 mg/L for ferrihydrite- and hematite-containing catalysts, respectively, in reactions using starting solution equal to 100 mg/L. When 1 g/L phenol solution was used, final concentrations were equal to 8–11 and 15–30 mg/L for both iron containing series of catalysts. In this way, it

**3.3. Catalytic activity vs. catalyst degradation**

66 Physico-Chemical Wastewater Treatment and Resource Recovery

\* In solution after 75 min of reaction.

\*\*Percentage of initial content.

**Table 4.** Comparison of Fe content in catalysts and reaction solutions, residual concentration of phenol and time of 40 and 50% phenol conversion.

## **4. Conclusions**

Depending on the experimental conditions, a nanocrystalline phase of hematite was formed in the hydrothermal synthesis. On the other hand, precipitation resulted in the formation of ferrihydrite phase. It was demonstrated that the latter phase is less active than nano-hematite; moreover, it was shown that optimum loading of the active phase is required to obtain the highest reaction efficiency: fast and high phenol conversion with minimum amount of side products as well as limited catalyst degradation. Among the transition products, formation of quinones was confirmed using derivative UV-vis spectroscopy. Physicochemical techniques also confirmed that nano-hematite-containing catalysts were more stable in studied reaction —only limited changes were observed in agglomeration state of Fe-containing materials, and leaching of iron was reduced. It was also shown that each group of catalysts is in different extents susceptible to degradation. However, the observed catalytic effect cannot be attributed only to homogeneous reaction. It was confirmed that dispersed pollutants are more resistant to degradation.

## **Author details**

Agnieszka Węgrzyn

Address all correspondence to: a.m.wegrzyn@uj.edu.pl; a.m.wegrzyn@gmail.com

Faculty of Chemistry, Jagiellonian University in Krakow, Kraków, Poland


[5] Kim K-H, Ihm S-K. Heterogeneous catalytic wet air oxidation of refractory organic pollutants in industrial wastewaters: a review. Journal of Hazardous Materials. 2011;186:16–34. DOI:10.1016/j.jhazmat.2010.11.011

**4. Conclusions**

68 Physico-Chemical Wastewater Treatment and Resource Recovery

to degradation.

**Author details**

Agnieszka Węgrzyn

**References**

69018

10.1016/S0920-5861(99)00112-1

Depending on the experimental conditions, a nanocrystalline phase of hematite was formed in the hydrothermal synthesis. On the other hand, precipitation resulted in the formation of ferrihydrite phase. It was demonstrated that the latter phase is less active than nano-hematite; moreover, it was shown that optimum loading of the active phase is required to obtain the highest reaction efficiency: fast and high phenol conversion with minimum amount of side products as well as limited catalyst degradation. Among the transition products, formation of quinones was confirmed using derivative UV-vis spectroscopy. Physicochemical techniques also confirmed that nano-hematite-containing catalysts were more stable in studied reaction —only limited changes were observed in agglomeration state of Fe-containing materials, and leaching of iron was reduced. It was also shown that each group of catalysts is in different extents susceptible to degradation. However, the observed catalytic effect cannot be attributed only to homogeneous reaction. It was confirmed that dispersed pollutants are more resistant

Address all correspondence to: a.m.wegrzyn@uj.edu.pl; a.m.wegrzyn@gmail.com

[1] Luck F. Wet air oxidation: past, present and future. Catalysis Today. 1999;53:81–91. DOI:

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70 Physico-Chemical Wastewater Treatment and Resource Recovery

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**Micropollutants and their Removal Process**

**Provisional chapter**

## **Micropollutants in Wastewater: Fate and Removal Processes Micropollutants in Wastewater: Fate and Removal Processes**

Sreejon Das, Nillohit Mitra Ray, Jing Wan, Adnan Khan, Tulip Chakraborty and Madhumita B. Ray Adnan Khan, Tulip Chakraborty and Madhumita B. Ray Additional information is available at the end of the chapter

Sreejon Das, Nillohit Mitra Ray, Jing Wan,

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65644

#### **Abstract**

The occurrence of micropollutants (MPs) in various streams of municipal wastewater treatment plants (WWTPs), and their fate and removal processes are discussed. The fate of MPs in WWTPs largely depends on adsorption on suspended particulates, primary and secondary sludge and dissolved organic carbon, and removal occurs due to coagulation-flocculation, and biodegradation. The log Kow (>2.5) and pKa are the dominant properties of the MPs, and the concentration, organic fraction, and surface charge of suspended particulates dictate the extent of adsorption of MPs. Most of the conventional WWTPs do not remove complex MPs by biodegradation or biotransformation effectively (kbio ≤0.0042 L/gss/h), and the removal varies widely for different compounds, as well as for the same substance, due to operational conditions such as aerobic, anaerobic, anoxic, sludge retention time (SRT), pH, redox potential, and temperature. Membrane bioreactor performs better for moderately biodegradable compounds due to the diverse nature of microorganisms as well as greater adaptability due to longer SRT. Ozone and UV-based advanced oxidation processes, membrane filtration can be used for tertiary treatment due to their high rate as well as easy implementation. Various partition coefficients and rate constants values for different MPs are also provided for design and application.

**Keywords:** micropollutants, wastewater, fate and removal, adsorption, coagulation, biodegradation, membrane filtration, advanced oxidation processes

Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons

## **1. Introduction**

The widespread presence of micropollutants (MPs) in aquatic systems is a major concern all across the globe. For example, about 143,000 compounds were registered in European market in 2012; many of which would end up in water systems at some point of their lifecycle. Most of them are not eliminated or biotransformed in traditional wastewater treatment plants, can be persistent in aquatic system or form new chemical species reacting with background humic substances in sunlight, can be bioactive, and can bioaccumulate [1–5]. Although they are present in almost undetectable (low to subparts per billion (ppb)) concentrations, their existence in aquatic systems has been connected to various detrimental effects in organisms such as estrogenicity, mutagenicity, and genotoxicity [6].

While no compound-specific regulation exists anywhere for the removal of MPs in wastewater plants, some regulations are there for the presence in water for compounds such as pesticides, lindane, nonylphenol, and synthetic hormones [7]. The MPs fall into several categories as pharmaceuticals, personal care products (PPCPs), household chemicals, and industrial agents. A comprehensive list of 242 chemicals is provided in EU FP7 Project [8] of which about 70% are pharmaceuticals and personal care products and 30% are industrial agents including perfluoro compounds, pesticides, herbicides, and food additives. Since a significant majority of the MPs in municipal wastewater belong to the class of pharmaceuticals and personal care products (PPCP), fate and removal processes of these compounds are discussed in detail in this chapter.

## **2. Commonly found PPCP in wastewater effluent and surface water**

About 70% of the pharmaceuticals in the wastewater originates from household, 20% comes from livestock farming, 5% is from hospital effluent, and rest 5% comes in runoff from nonparticular sources [9]; however, seasonal and geographical variations typically occur. The fate of MPs in wastewater plant depends on the physical properties such as solubility, octanolwater partition coefficient, and Henry's constant. A list of commonly found pharmaceuticals, personal care products, and biocides and their concentration in wastewater effluent and surface water and physical properties are presented in **Table 1**. The solubility of MPs varies in a wide range of 0.15 mg/L (maprotiline, C10 H23 N, an antidepressant drug) to 588,000 mg/L (acesulfame, C4H4KNO4S, and artificial sweetener), which is also in accordance with their concentration in the effluent.



**1. Introduction**

estrogenicity, mutagenicity, and genotoxicity [6].

76 Physico-Chemical Wastewater Treatment and Resource Recovery

concentration in the effluent.

Disinfectants, pharmaceuticals (prescriptions, over-

**Type MP Application Average**

The widespread presence of micropollutants (MPs) in aquatic systems is a major concern all across the globe. For example, about 143,000 compounds were registered in European market in 2012; many of which would end up in water systems at some point of their lifecycle. Most of them are not eliminated or biotransformed in traditional wastewater treatment plants, can be persistent in aquatic system or form new chemical species reacting with background humic substances in sunlight, can be bioactive, and can bioaccumulate [1–5]. Although they are present in almost undetectable (low to subparts per billion (ppb)) concentrations, their existence in aquatic systems has been connected to various detrimental effects in organisms such as

While no compound-specific regulation exists anywhere for the removal of MPs in wastewater plants, some regulations are there for the presence in water for compounds such as pesticides, lindane, nonylphenol, and synthetic hormones [7]. The MPs fall into several categories as pharmaceuticals, personal care products (PPCPs), household chemicals, and industrial agents. A comprehensive list of 242 chemicals is provided in EU FP7 Project [8] of which about 70% are pharmaceuticals and personal care products and 30% are industrial agents including perfluoro compounds, pesticides, herbicides, and food additives. Since a significant majority of the MPs in municipal wastewater belong to the class of pharmaceuticals and personal care products (PPCP), fate and removal processes of these compounds are discussed in detail in this chapter.

**2. Commonly found PPCP in wastewater effluent and surface water**

About 70% of the pharmaceuticals in the wastewater originates from household, 20% comes from livestock farming, 5% is from hospital effluent, and rest 5% comes in runoff from nonparticular sources [9]; however, seasonal and geographical variations typically occur. The fate of MPs in wastewater plant depends on the physical properties such as solubility, octanolwater partition coefficient, and Henry's constant. A list of commonly found pharmaceuticals, personal care products, and biocides and their concentration in wastewater effluent and surface water and physical properties are presented in **Table 1**. The solubility of MPs varies in a wide range of 0.15 mg/L (maprotiline, C10 H23 N, an antidepressant drug) to 588,000 mg/L (acesulfame, C4H4KNO4S, and artificial sweetener), which is also in accordance with their

> **concentration (ng/L) [10, 11]**

> > **WWTP effluent**

**water**

Atenolol β-blocker 205 843 0.3 0.16 9.6 1.37 ×

Azithromycin Antibiotic 12 175 <1 at 25°C 4.02 8.74 5.30 ×

**Solubility\* (mg/mL)**

**Surface mole)\***

**log** *K***ow\*** **pKa\* Henry's constant (atm-m3 /**

E-18

E-29



**Type MP Application Average**

78 Physico-Chemical Wastewater Treatment and Resource Recovery

Fluconazole Antifungal

Repaglinide Antidiabetic

Sulfamethoxazole (VITO)

Sulfamethoxazole

(JRC)

Haloperidol Psychiatric

Diltiazem Antiarrhythmic

Gemfibrozil Lipid‐lowering

drugs

medication

agents

Flecainide Antiarrhythmic

medication

medications

agents

**concentration (ng/L) [10, 11]**

> **WWTP effluent**

**water**

Diphenhydramine Antihistamine – 11.7 0.0752 3.44 8.98 3.70 ×

Bisoprolol β‐blockers – 41.6 0.0707 2.3 14.09 2.89 ×

Oxazepam Anxiolytics – 161.7 0.0881 2.24 10.61 5.53 ×

Carbamazepine Antiepileptic drugs – 832.3 0.152 2.45 15.96 1.08 ×

Diclofenac Analgesics 65 647 0.00447 4.98 4 4.73 ×

Orphenadrine Antihistamine – 3.9 0.03 3.77 8.91 4.08 ×

Citalopram Antidepressant – 33.8 – – – –

Fexofenadine Antihistamine – 165.0 0.00266 5.6 4.04 –

Fluoxetine Antidepressant – 2.1 0.0017 4.05 9.8 8.90 ×

Terbutaline Antiasthmatics – 1.1 5.84 0.90 8.86 1.65 ×

Clindamycin Antibiotics – 70.4 3.1 2.16 12.16 2.89 ×

Telmisartan Antihypertensives – 367.5 0.0035 7.7 3.65 – Eprosartan Antihypertensives – 226.8 0.00866 3.9 3.63 –

Zolpidem Hypnotics – 1.5 0.0313 3.15 6.2 – Hydroxyzine Antihistamine – 1.1 0.0914 3.43 15.12 – Ketoprofen Analgesics – 86.0 0.0213 3.12 4.45 2.12 ×

Ranitidine Antihistamine – 68 0.0795 0.27 8.08 3.42 ×

Antibiotics – 280.2 0.459 0.89 6.16 –

Antibiotics – 142.3 0.459 0.89 6.16 –

Alfuzosin Alpha‐blockers – 2.8 0.282 2.02 14.64 – Bupropion Antidepressant – 1.0 312 3.6 18.29 – Ciprofloxacin Antibiotics – 96.3 1.35 0.28 6.09 5.09 ×

**Solubility\* (mg/mL)**

– 108.2 1.39 0.58 12.71 –

– 3.1 0.00294 5.05 3.68 –

– 45.5 0.0324 2.98 13.68 5.75 ×

– 32.2 0.00446 4.30 8.66 2.26 ×

– 10.7 0.0168 3.09 12.86 8.61 ×

– 137.7 0.0278 3.4 4.42 –

**Surface mole)\***

**log** *K***ow\*** **pKa\* Henry's constant (atm-m3 /**

E‐09

E‐13

E‐15

E‐19

E‐10

E‐10

E‐12

E‐09

E‐14

E‐17

E‐08

E‐18

E‐22

E‐11

E‐15



"–": Data are not available in the literature. \*Solubility, log *K*ow, pKa, and Henry's law constant for selected micropollutants are found in http://www.drugbank.ca/, http://chem.sis.nlm.nih.gov/chemidplus/ and https:// pubchem.ncbi.nlm.nih.gov/.

**Table 1.** Commonly found MPs in municipal wastewater effluent and surface water.

**Type MP Application Average**

80 Physico-Chemical Wastewater Treatment and Resource Recovery

Diethyltoluamide (DEET)

Terbutylazine‐ desethyl

**concentration (ng/L) [10, 11]**

> **WWTP effluent**

Insecticide 135 593 0.912 2.80 2.08 ×

**water**

Dimethoate Insecticide 22 – 25 0.78 1.05 ×

MCPA Insecticides – 149.9 0.63 3.25 3.13 1.33 ×

Carbendazim Fungicide 16 81 0.029 1.52 4.2 2.12 ×

Diuron Herbicide 54 201 0.042 2.68 5.04 ×

Glyphosate Herbicide 373 – 12 −3.40 0.8 4.08 ×

MCPA Herbicide 40 25 0.63 3.25 3.13 1.33 ×

Mecoprop‐p Herbicide 45 424 0.62 3.13 3.1 1.82 ×

Triclosan Microbiocide 20 116 0.010 4.76 7.9 4.99 ×

Terbutylazine Herbicide – 90.6 0.0085 3.21 2 3.72 ×

Atrazine Herbicide – 4.2 0.0347 2.61 1.7 2.36 ×

Isoproturon Herbicide – 10.1 0.065 2.87 – 1.12 ×

Bentazone Herbicide – 9.6 0.5 2.34 2.92 2.18 ×

Metolachlor Herbicide – 12.4 0.53 3.13 – 9 × E‐09 Dichlorprop Herbicide – 9.6 0.35 3.43 3.1 8.68 ×

Simazine Herbicide – 26.3 0.0062 2.18 1.62 9.42 ×

Atrazine‐desethyl Herbicide – 13.8 3.2 1.51 – 1.53 ×

Chlortoluron Herbicide – 3.2 0.07 2.41 – – Hexazinone Herbicide – 0.8 33 1.85 – 2.26 ×

Linuron Herbicide – 40.1 0.075 3.20 – –

Herbicide – 68.8 – – – –

Irgarol (cybutryne) Herbicide 3 30 – – – – Isoproturon Herbicide 315 12 0.065 2.87 1.12 ×

Carbaryl Insecticide – 1.6 0.11 2.36 10.4 –

Acesulfame Food additive 4010 22500 588 −1.33 5.67 – Sucralose Food additive 540 4600 22.7 −1.00 4.2 –

Pesticides [10] Diazinon Insecticide 15 173 0.04 3.81 2.6 1.13 ×

Biocides [10] 2, 4‐D Herbicide 67 13 0.012 2.81 2.73 1.59 ×

**Solubility\* (mg/mL)**

**Surface mole)\***

**log** *K***ow\*** **pKa\* Henry's constant (atm-m3 /**

E‐07

E‐08

E‐10

E‐09

E‐07

E‐11

E‐10

E‐19

E‐10

E‐09

E‐08

E‐09

E‐08

E‐09

E‐10

E‐09

E‐11

E‐10

E‐09

E‐12

## **3. Fate and removal processes of MPs in wastewater**

The municipal wastewater treatment plants (WWTP) are designed to remove most of the suspended solids, dissolved organics, and nutrients from the wastewater. WWTPs employ primary, secondary, and occasional tertiary treatment processes to optimally treat the incoming wastewater. In primary treatment, coagulants such as alum, ferric chloride, and polymers and polymeric coagulant aids are used to remove colloidal and suspended particulates. In the process, organics attached with dissolved humic substances and particles can also be removed. In secondary treatment, dissolved organics are removed aerobically by a consortium of microorganisms in suspension. The thickened sludge from both primary and secondary clarifiers is digested anaerobically (biosolids) prior to disposal. In some places, tertiary treatment processes such as activated carbon adsorption, ozonation, or filtration are adopted for final treatment of effluent to remove trace concentration of the organics.

The fate processes for MPs in a typical WWTP include adsorption on suspended particulates, dissolved humic substances, primary and secondary sludge, while the removal processes include coagulation and sedimentation, biodegradation, adsorption, advanced oxidation, and membrane filtration as shown in **Figure 1**. Volatilization of the MPs during any of the treatment steps is negligible due to their very low Henry's constant (<10−5 atm‐m3 /mol) as shown in **Table 1**.

#### **3.1. Fate: adsorption of micropollutants**

Adsorption on suspended solids in both primary and secondary treatment units is an impor‐ tant fate process for MPs in wastewater. Adsorption may occur due to the hydrophobic interactions between the aliphatic and aromatic groups of the compounds with the fat and lipid fractions in primary sludge and the lipophilic cell membrane of the microorganisms in secondary sludge, respectively. Electrostatic interactions also occur between the positively charged groups in the MPs and the negatively charged microorganisms in secondary sludge. Many acidic pharmaceuticals are negatively charged at neutral pH, and their sorption on sludge is negligible.

**Figure 1.** Conceptual model of fate and removal processes of a micropollutant in a typical WWTP.

With a nonpolar core and polar moiety, the properties of pharmaceuticals and antibiotics vary widely, making it difficult to estimate their sorption on sludge. Kinney et al. [12] analyzed nine different biosolids produced by municipal wastewater treatment plants in seven different states in U.S. for 87 different MPs, and the measured concentrations of the contam‐ inants in various sludge were in the range of 64–1811 mg/kg dry weight. Nineteen different pharmaceuticals were detected in these biosolids, representing a wide range of physico‐ chemical properties, including compounds with low log *K*ow and high water solubility values. Adsorption of MPs on biosolids did not exhibit any particular trend, and no correlation was found between organic carbon‐normalized MPs concentrations in biosolids with log *K*ow, suggesting that organic carbon content of the biosolids may not be the only factor controlling MPs adsorption. It is generally expected that compounds with low water solubilities and large log *K*ow values will more likely to be present in organic‐rich biosolids compared to highly soluble organics; however, this study indicated significant presence of water soluble phar‐ maceuticals in all nine biosolids. The 25 MPs detected in all nine biosolids had water solubility ranging from 1.3 × 10−5 to 8.28 × 104 mg/L, and log *K*ow from 1.50 to 9.65 indicating complex nature of the process. Other factors, such as the quantity of organics entering the influent stream (which typically varies temporally and spatially), volume of influent, biosolids/water ratio, and sludge retention time (SRT), all affect the distribution of the MPs in different phases. Increasing sludge age had detrimental effect on the adsorption of lindane [13] on activated sludge, adsorption of pentachlorophenol reduced from 40 to 60% at sludge ages below 4 days to less than 10% at sludge ages above 25 days [14].

lipid fractions in primary sludge and the lipophilic cell membrane of the microorganisms in secondary sludge, respectively. Electrostatic interactions also occur between the positively charged groups in the MPs and the negatively charged microorganisms in secondary sludge. Many acidic pharmaceuticals are negatively charged at neutral pH, and their sorption on

**Figure 1.** Conceptual model of fate and removal processes of a micropollutant in a typical WWTP.

With a nonpolar core and polar moiety, the properties of pharmaceuticals and antibiotics vary widely, making it difficult to estimate their sorption on sludge. Kinney et al. [12] analyzed nine different biosolids produced by municipal wastewater treatment plants in seven different states in U.S. for 87 different MPs, and the measured concentrations of the contam‐ inants in various sludge were in the range of 64–1811 mg/kg dry weight. Nineteen different pharmaceuticals were detected in these biosolids, representing a wide range of physico‐ chemical properties, including compounds with low log *K*ow and high water solubility values. Adsorption of MPs on biosolids did not exhibit any particular trend, and no correlation was found between organic carbon‐normalized MPs concentrations in biosolids with log *K*ow, suggesting that organic carbon content of the biosolids may not be the only factor controlling MPs adsorption. It is generally expected that compounds with low water solubilities and large log *K*ow values will more likely to be present in organic‐rich biosolids compared to highly soluble organics; however, this study indicated significant presence of water soluble phar‐

sludge is negligible.

82 Physico-Chemical Wastewater Treatment and Resource Recovery


The concentrations of some of the commonly found MPs in sludge are summarized in **Table 2**.



**Table 2.** Concentrations of commonly found MPs in sludge.

**MP Type/application Concentration**

84 Physico-Chemical Wastewater Treatment and Resource Recovery

Clarithromycin 0.7–67

Chlortetracycline 0.069 –0.346 Oxytetracycline 0.052–0.743 Demeclocycline 0.036–0.131

Acetaminophen Analgesic 0.013–0.419 Carbamezipine 0.011–0.042

**(mg/kg)**

[20] Ciprofloxacin 0.22–3.1

Norfloxacin 1.06–7.23 Sludge (primary, excess

Azithromycin 2.5–64 Sludge (excess activated,

Erythromycin 0.030–0.041 Sludge, Class A & B

Roxythromycin 0.337–1.446 Anaerobically digested

Sulfamethoxazole 0.019–68 Sludge (excess activated,

Sulfapyridine 0.1–28 Sludge (excess activated,

Sulfamethazine 0.026–0.128 Anaerobically digested

Sulfamerazine 0.112–0.669 Biosolids from sewage

Trimethoprim 0.017–41 Sludge (excess activated,

Lincomycin 0.006–0.174 Municipal sludge,

Tylosin 1.074–1.958 Anaerobically digested

**Source Reference**

Jia et al. [19], Golet et al.

Kinney et al. [12], Ding

Gobel et al. [21], Nieto et al. [23], Ding et al. [22]

Nieto et al. [23], Ding et

Gobel et al. [21], Nieto et

Ding et al. [22], Subedi

Gobel et al. [21]

et al. [22]

Nieto et al. [23]

Gobel et al. [21]

Ding et al. [22]

al. [22]

al. [23]

et al. [25]

Nieto et al. [23]

activated, anaerobically digested, dewatered)

anaerobically digested)

biosolids

biosolids

biosolids

sludge

Tetracycline 0.282–1.914 McCellan & Halden [24],

Ding et al. [22] Doxycycline 0.225–0.966

Clindamycin nd–0.006 Municipal sludge Subedi et al. [25]

Tiamulin nd–0.7 Agricultural Field soil Schlusener et al. [26]

dewatered sludge

anaerobically digested),

anaerobically digested)

anaerobically digested)

biosolids

dewatered sludge

dewatered sludge,

Although a complex process as described above, the extent of MPs adsorption on sludge is traditionally modeled using linear equilibrium model as

$$\mathbf{C\_{ads}} = \mathbf{K\_{d}} \, \mathbf{C\_{ss}} \mathbf{C\_{dis}} \tag{1}$$

where ads is the adsorbed concentration of the MP (g/L), ss is the suspended particulate concentration (g/L), dis (g/L) is the dissolved concentration, and is the adsorption constant (L/gss), which is also known as the partition coefficient of the compound between the solids and water. has been proposed as a relatively accurate indicator of adsorption [29, 30]; for compounds with a value below 300 L/kg (log = 2.48), the sorption onto secondary sludge is insignificant. Polar compounds typically have higher values in secondary sludge compared to primary sludge. Typical values are presented in **Table 3**. of a compound can be correlated to more fundamental properties such as *K*ow.



"–": Data are not available in the literature. \*log *K*ow for selected MPs are found in http://www.drugbank.ca/. # log *Kd* and log *K*oc values are collected from references (Ref.) as follows: (a) [31], (b) [30], (c) [32].

**Table 3.** log *Kd* and log *K*oc values of some commonly found MPs.

Although a complex process as described above, the extent of MPs adsorption on sludge is

where ads is the adsorbed concentration of the MP (g/L), ss is the suspended particulate concentration (g/L), dis (g/L) is the dissolved concentration, and is the adsorption constant (L/gss), which is also known as the partition coefficient of the compound between the solids and water. has been proposed as a relatively accurate indicator of adsorption [29, 30]; for compounds with a value below 300 L/kg (log = 2.48), the sorption onto secondary sludge is insignificant. Polar compounds typically have higher values in secondary sludge compared to primary sludge. Typical values are presented in **Table 3**. of a compound

Diclofenac 4.98 1.2041 – b Estradiol 4.01 2.2304 – c

Ibuprofen 3.5 0.8513 – b Estriol 2.45 1.7324 – c

DEET 2.18 1.91 2.27 a Diphenhydramine 3.27 2.5 2.86 a

Clofibric acid 2.57 0.6812 – b Estrone 3.13 2.2304 – c

Ifosfamide 0.86 0.1461 – b Ethinylestradiol 3.67 2.4997 – c

Carbamazepine 2.45 1.95 2.31 a Fenoprofen 3.1 1.415 – c

Hydrocodone 2.16 2.03 2.38 a Fluoxetine 4.05 0.699 – c

Cyclophosphamide 0.63 0.3802 – b Amitriptyline 4.92 2.87 3.21 a

Gemfibrozil 4.77 2.11 2.47 a Gemfibrozil 3.4 1.2856 – c

Diazepam 2.82 1.3222 – b Hydrocodone 1.2 2.0294 – c

Diazepam 2.82 2.14 2.53 a Fluoxetine 4.05 3.08 3.43 a

Ethinylestradiol 3.9 2.5428 – b Indomethacine 4.27 1.4472 – c

Naproxen 3.2 2.16 2.56 a Ketoprofen 3.12 1.2041 – C

Perfluorooctanoic acid 6.3 2.3424 – c Mefenamic acid 5.12 2.6375 – C

Diclofenac 4.51 2.18 2.54 a Methadone 3.93 1.8808 – C

Perfluorononanoic acid 5.48 3.0934 – c Metoprolol 1.88 1.8129 – C

*C KCC* ads ss dis = *<sup>d</sup>* (1)

**. Micropollutants log** *K***\***

**ow log** *Kd* **log** *K***oc Ref#**

**.**

traditionally modeled using linear equilibrium model as

86 Physico-Chemical Wastewater Treatment and Resource Recovery

can be correlated to more fundamental properties such as *K*ow.

**Micropollutants log** *K***ow\* log** *Kd* **log** *K***oc Ref#**

As mentioned before, the sorption to sludge is not significant for compounds with log *K*ow < 2.5, moderate sorption for log *K*ow between 2.5 and 4, and high sorption for log *K*ow > 4.0 is expected. In absence of experimental data, to relate with *K*ow, Eqs. (2) and (3) are given by Matter-Muller et al. [33] and Dobbs et al. [34], respectively:

$$\log K\_d = 0.67 \times \log K\_{\text{ow}} + 0.39 \tag{2}$$

$$
\log K\_d = 0.58 \times \log K\_{\rm{ow}} + 1.14 \tag{3}
$$

 can also be estimated using Eq. (4) (Fetter [35]) and Eq. (5) (Jones et al. [36]) if the fraction of organic carbon of the solids is known as

$$K\_d = f\_{\rm oc} \times \frac{10^{0.72 \times \log K\_{\rm ov} + 0.49}}{1000} \tag{4}$$

$$K\_d = f\_{\rm oc} \times 0.41 \times K\_{\rm ov} \tag{5}$$

values of and oc versus ow for MPs from the literature are plotted in **Figure 2** showing slightly lower linear dependence of and oc on ow as compared to Eqs. (2) and (3). In addition, the goodness of fit as indicated by *R*<sup>2</sup> is in the range of 0.45–0.48, indicating possible influence of other parameters than only oc or ow.

**Figure 2.** Correlation between log *Kd* versus log *K*ow and log *K*oc versus log *K*ow for MPs listed in **Table 3**.

MP adsorption on sludge mostly follow linear isotherm such as Fruendlich:

$$q\_e = K\_f.C\_e^{1/n} \tag{6}$$

where e = mass adsorbed per unit mass of adsorbent at equilibrium (mg/g)

*Ce* = concentration of MP in water at equilibrium (mg/L)

1 = strength of adsorption (dimensionless)

log =´ + 0.58 log ow 1.14 *K K <sup>d</sup>* (3)

*K f <sup>d</sup>* (4)

=´ ´ oc ow 0.41 *Kf K <sup>d</sup>* (5)

1/ <sup>=</sup> . *<sup>n</sup> e fe q KC* (6)

is in the range of 0.45–0.48, indicating possible

can also be estimated using Eq. (4) (Fetter [35]) and Eq. (5) (Jones et al. [36]) if the fraction

oc 10

= ´

ow 0.72 log 0.49

*K*

1000 ´ +

values of and oc versus ow for MPs from the literature are plotted in **Figure 2** showing slightly lower linear dependence of and oc on ow as compared to Eqs. (2) and (3). In

**Figure 2.** Correlation between log *Kd* versus log *K*ow and log *K*oc versus log *K*ow for MPs listed in **Table 3**.

MP adsorption on sludge mostly follow linear isotherm such as Fruendlich:

where e = mass adsorbed per unit mass of adsorbent at equilibrium (mg/g)

*Ce* = concentration of MP in water at equilibrium (mg/L)

of organic carbon of the solids is known as

88 Physico-Chemical Wastewater Treatment and Resource Recovery

addition, the goodness of fit as indicated by *R*<sup>2</sup>

influence of other parameters than only oc or ow.

= adsorption capacity at unit concentration (mg/g)(L/mg)1/

The values of and 1/*n* for MPs on sludge varied from 0.0052 to 4.40 (mg/g) (L/mg)1/*<sup>n</sup>* and 0.51 to 1.0076, respectively [37–40]. Larger values indicate higher affinity of adsorption for a particular sludge and closer the value of 1/*n* around 1.0, greater is the indication of compa‐ ratively strong adsorption bond. Typically, adsorption equilibrium is achieved within 24 hours with almost 90% removal from dissolved phase occurs in an hour; for example, at 3.6 g/L mixed liquor suspended solids (MLSSs) concentration, 95% of oxytetracycline was removed from water within only 1 hour and the concentration at equilibrium remained unchanged over 24 hours [40].

Colloidal particles are a relatively small fraction of the total waterborne particle mass (<10%) in typical wastewater but possess large surface areas which can enable covalent, electrostatic, and hydrophobic binding of MPs depending on their polarity. The magnitude of sorption depends on the molecular weight distribution and aromatic content of the colloids fraction, which also depends on the sewage composition, strength, and sludge age [41]. Similar to adsorption on suspended particulates, adsorption on colloidal particles can be quantified using a distribution coefficient *K*coc. Holbrook et al. [41] determined *K*coc using pyrene as a model MP and colloidal fractions from two biological wastewater plants; sorption coefficients (*K*coc) for pyrene ranged from 1 × 103 L/kg colloids to 80 × 103 L/kg colloids and were comparable to values obtained in the literature for natural organic matter. Good correlation was obtained between *K*coc and the aromaticity of the colloidal particles.

#### **3.2. Removal processes**

#### *3.2.1. Coagulation and sedimentation of micropollutants*

Coagulation‐flocculation processes are typically used for improving efficiency of wastewater treatment plants promoting the removal of suspended solids, colloids, and some dissolved organics, which do not settle spontaneously. The coagulation process works by destabilizing the colloids/emulsions using coagulants such as metal salts and/or synthetic organic polymers following any of the mechanisms such as double‐layer compression, adsorption and charge neutralization, entrapment of particles in precipitate, adsorption and interparticle bridging. The parameters that affect the performance of coagulation are coagulant dosage, pH, and ionic strength of the solution. Based on the type of coagulant such as aluminum sulfate, ferrous sulfate, and ferric chloride, optimum pH range for coagulation varies between 4.0 and 8.5. In case of polymeric coagulants, the active group (carboxyl, amino group, etc.) present on the polymer influences the change of charge with pH [42].

In general, removal of MPs by coagulation‐flocculation processes is not very effective for most of the compounds studied with a few exceptions. Earlier studies on removal of MPs by coagulation were reported for simulated drinking water treatment processes [43–47], and percent removal of various MPs varied from 15 to 75% using alum and iron salts, and excess lime/soda ash softening. Vieno et al. [46] evaluated the role of dissolved organic matter, mainly the humic substances in the coagulation process. In the presence of dissolved humic matter, diclofenac, ibuprofen, and bezafibrate could be removed by ferric sulfate coagulation. The removal of diclofenac reached a maximum of 77%, while 50% of ibuprofen, and 36% of bezafibrate were removed. Hence, a high amount of high‐molecular‐weight dissolved organic matter enhanced the removal of ionizable pharmaceuticals. However, contradictory results were reported by Choi et al. [43] where removal of seven tetracycline classes of antibiotic (TAs) from synthetic and river water using coagulation was achieved. TAs were assumed to be removed through the charge neutralization of zwitterionic or negative TAs by cationic Al (III) and sweep coagulation using poly‐aluminum chloride (PACl). Aluminum hydroxide precip‐ itates were formed in the presence of sufficient alkalinity, and TAs were removed by being enmeshed into or adsorbed onto the precipitates. It was suggested that the presence of dissolved organic matter, especially the low‐molecular‐weight fractions, resulted in possible inhibition of MP removal. This was due to preferential removal of the organic matter by the coagulant.

Huerta‐Fontela et al. [48] performed coagulation with alum‐coagulants, flocculation with a diallyldimethyl ammonium chloride homopolymer (poly‐DADMAC), followed by clarifica‐ tion through sand filters. Of the 55 pharmaceutical compounds present, only five compounds (chlordiazepoxide, zolpidem, bromazepam, clopidogrel, and doxazosin) were completely removed, while warfarin, betaxolol, and hydrochlorothiazide accounted for removals higher than 50%. For some pharmaceuticals such as irbesartan, losartan, or carbamazepine epoxide, negligible removals were obtained.

Suarez et al. [49] evaluated the performance of coagulation‐flocculation process for the pretreatment of hospital effluent, both in a batch mode and continuous pilot scale. Highest removal efficiency (>90%) was reported for PPCPs such as galaxolide, tonalide, and synthetic musk (ADBI); these are lipophilic compounds, carrying high negative charge, which facilitates their coagulation in the presence of higher fat content in wastewater. Asakura and Matsuto [50] studied the effect of coagulation for treating landfill leachate. Out of the various EDCs, only nonylphenol showed a removal of >90%, whereas diethylhexylphthalate (DEHP) removal was about 70%. Other EDCs such as diethylphthalate (DEP), dibutylphthalate (DBP), butylben‐ zylphthalate (BBP), 4‐t‐octylphenol (4tOP), and 4‐n‐octylphenol (4nOP) showed poor removal (<50%) by coagulation, with the lowest removal of 20% for bisphenol A.

Few studies have reported the removal of MPs due to coagulation and flocculation in waste‐ water (**Table 4**). Matamoros and Salvadó [51] evaluated several MPs removal in a coagulation/ flocculation‐lamellar clarifier for treating secondary effluent. The hydrophobicity of the compounds (log *K*ow) was found to be a major factor in determining the removal efficiency with coagulation‐flocculation. The highest removal of 20–50% was observed for the compounds with log *K*ow ≥ 4 at pH 7–8. Since adsorption of MPs on the suspended solids and colloids is the precursor step for their removal during coagulation, the removal efficiency can be tied with the removal efficiency of suspended solids as

Micropollutants in Wastewater: Fate and Removal Processes http://dx.doi.org/10.5772/65644 91

$$\% \text{removal} = \frac{K\_d \ C\_{\text{ss}}}{1 + K\_d \ C\_{\text{ss}}} E\_{\text{Tss}} \tag{7}$$

where *E*TSS is the efficiency of TSS removal (%) during coagulation.

percent removal of various MPs varied from 15 to 75% using alum and iron salts, and excess lime/soda ash softening. Vieno et al. [46] evaluated the role of dissolved organic matter, mainly the humic substances in the coagulation process. In the presence of dissolved humic matter, diclofenac, ibuprofen, and bezafibrate could be removed by ferric sulfate coagulation. The removal of diclofenac reached a maximum of 77%, while 50% of ibuprofen, and 36% of bezafibrate were removed. Hence, a high amount of high‐molecular‐weight dissolved organic matter enhanced the removal of ionizable pharmaceuticals. However, contradictory results were reported by Choi et al. [43] where removal of seven tetracycline classes of antibiotic (TAs) from synthetic and river water using coagulation was achieved. TAs were assumed to be removed through the charge neutralization of zwitterionic or negative TAs by cationic Al (III) and sweep coagulation using poly‐aluminum chloride (PACl). Aluminum hydroxide precip‐ itates were formed in the presence of sufficient alkalinity, and TAs were removed by being enmeshed into or adsorbed onto the precipitates. It was suggested that the presence of dissolved organic matter, especially the low‐molecular‐weight fractions, resulted in possible inhibition of MP removal. This was due to preferential removal of the organic matter by the

Huerta‐Fontela et al. [48] performed coagulation with alum‐coagulants, flocculation with a diallyldimethyl ammonium chloride homopolymer (poly‐DADMAC), followed by clarifica‐ tion through sand filters. Of the 55 pharmaceutical compounds present, only five compounds (chlordiazepoxide, zolpidem, bromazepam, clopidogrel, and doxazosin) were completely removed, while warfarin, betaxolol, and hydrochlorothiazide accounted for removals higher than 50%. For some pharmaceuticals such as irbesartan, losartan, or carbamazepine epoxide,

Suarez et al. [49] evaluated the performance of coagulation‐flocculation process for the pretreatment of hospital effluent, both in a batch mode and continuous pilot scale. Highest removal efficiency (>90%) was reported for PPCPs such as galaxolide, tonalide, and synthetic musk (ADBI); these are lipophilic compounds, carrying high negative charge, which facilitates their coagulation in the presence of higher fat content in wastewater. Asakura and Matsuto [50] studied the effect of coagulation for treating landfill leachate. Out of the various EDCs, only nonylphenol showed a removal of >90%, whereas diethylhexylphthalate (DEHP) removal was about 70%. Other EDCs such as diethylphthalate (DEP), dibutylphthalate (DBP), butylben‐ zylphthalate (BBP), 4‐t‐octylphenol (4tOP), and 4‐n‐octylphenol (4nOP) showed poor removal

Few studies have reported the removal of MPs due to coagulation and flocculation in waste‐ water (**Table 4**). Matamoros and Salvadó [51] evaluated several MPs removal in a coagulation/ flocculation‐lamellar clarifier for treating secondary effluent. The hydrophobicity of the compounds (log *K*ow) was found to be a major factor in determining the removal efficiency with coagulation‐flocculation. The highest removal of 20–50% was observed for the compounds with log *K*ow ≥ 4 at pH 7–8. Since adsorption of MPs on the suspended solids and colloids is the precursor step for their removal during coagulation, the removal efficiency can be tied with

(<50%) by coagulation, with the lowest removal of 20% for bisphenol A.

coagulant.

negligible removals were obtained.

90 Physico-Chemical Wastewater Treatment and Resource Recovery

the removal efficiency of suspended solids as

Carballa et al. [52] observed that during coagulation-flocculation of primary wastewater, lipophilic compounds such as musks were adsorbed in the lipid fractions of the sludge with two different fat concentrations of 60 and 150 mg/L, while acidic compounds such as diclofenac were adsorbed due to electrostatic interaction. Compounds with high sorption properties (galaxolide and tonalide) and diclofenac were significantly removed during coagulation-flocculation with efficiencies around 70%. Compounds with lower *Kd* values, such as diazepam, carbamazepine, ibuprofen, and naproxen, were reduced to a lesser extent (up to 25%).



"–": Data are not available in the literature. \*log *K*ow for selected MPs are found in http://www.drugbank.ca/. # log *Kd* and log *K*oc values are collected from references (Ref.) as follows: (a) [31], (b) [30], (c) [32].

**Table 4.** Removal of MPs by coagulation/flocculation process from various effluents.

#### *3.2.2. Biodegradation of micropollutants in secondary treatment*

Most of the conventional municipal WWTPs do not remove complex MPs by biodegrada‐ tion and/or biotransformation effectively. Observed removal efficiencies vary in a wide range for different compounds, as well as for the same substance, due to operational condi‐ tions such as aerobic, anaerobic, anoxic, sludge retention time (SRT), pH, redox potential, and water temperature. Membrane bioreactors (MBRs) seem to be more effective than con‐ ventional‐activated sludge (CAS) process as MBR process combines biological treatment with membrane filtration (micro and ultrafiltration). In addition, due to higher SRT at MBRs compared to CAS, biodiversity of the microorganisms in MBR is greater than CAS, and opportunity for adaptation of specific microorganisms to the persistent compounds is greater in MBR than in CAS. Removal of 29 antibiotics in a CAS process was reviewed by Verlicchi et al. [54], where removal of compounds such as sulfamethoxazole, ciprofloxacin, roxithromycin, norfloxacin, erythromycin, etc., varied in a wide range of 0 (spiramycin) and 98% (cefaclor) in CAS and between 15 (azithromycin) and 94% (ofloxacin) in MBRs. Only 1 (azithromycin) out of 10 compounds investigated in both systems exhibited higher average removal efficiency in CAS than in MBR. Trinh et al. [55] traced 48 MPs including steroidal hormones, xenoestrogens, pesticides, caffeine, pharmaceuticals, and personal care products (PPCPs) in a MBR with >90% removal for many of the compounds. However, amitriptyline, carbamazepine, diazepam, diclofenac, fluoxetine, gemfibrozil, omeprazole, sulphamethoxazole, and trimethoprim were only partially removed in MBR with the re‐ moval efficiencies of 24–68% [55]. Similar results were obtained in a pilot‐scale MBR oper‐ ated for a Swiss hospital effluent for 1 year [56, 57]. Among the 56 pharmaceuticals, an overall load elimination of all pharmaceuticals and metabolites in the MBR was only 22% due to the presence of persistent iodinated contrast media (almost 80% of the total organic load). Weiss and Reemtsma [58] reported that major advantage of MBR lies for the com‐ pounds with moderate removal in CAS; MBR showed no advantages for both well‐degrad‐ able and recalcitrant compounds. For polar compounds, MBR does not provide significant benefits, because effluent quality is improved only gradually and the most critical compo‐ nents of high aerobic stability remain almost unaltered [58].

**Coagulant Dosage(ppm) with pH**

92 Physico-Chemical Wastewater Treatment and Resource Recovery

Ferric sulfate 78.5–pH 4.5 Dichlofenac Lake water with

– – Celestolide Secondary effluent

**compound Source Removal**

dissolved humic acid

Salvadó [51] Tricholsan <sup>24</sup>

Heptachlor 36 Aldrin 49 Benzofluoranthine 70 Benzopyrene 72

Ibuprofen 50 Bezafibrate 36 Carbamazepine <10 Sulfamethoxazole <10

from WWTP

Octylphenol 50 Tonalide 24 DMP 19 Galaxolide 16 Ibuprofen 4 Carbamazepine 2

"–": Data are not available in the literature. \*log *K*ow for selected MPs are found in http://www.drugbank.ca/. #

Most of the conventional municipal WWTPs do not remove complex MPs by biodegrada‐ tion and/or biotransformation effectively. Observed removal efficiencies vary in a wide range for different compounds, as well as for the same substance, due to operational condi‐ tions such as aerobic, anaerobic, anoxic, sludge retention time (SRT), pH, redox potential, and water temperature. Membrane bioreactors (MBRs) seem to be more effective than con‐ ventional‐activated sludge (CAS) process as MBR process combines biological treatment with membrane filtration (micro and ultrafiltration). In addition, due to higher SRT at MBRs compared to CAS, biodiversity of the microorganisms in MBR is greater than CAS, and opportunity for adaptation of specific microorganisms to the persistent compounds is greater in MBR than in CAS. Removal of 29 antibiotics in a CAS process was reviewed by Verlicchi et al. [54], where removal of compounds such as sulfamethoxazole, ciprofloxacin, roxithromycin, norfloxacin, erythromycin, etc., varied in a wide range of 0 (spiramycin) and 98% (cefaclor) in CAS and between 15 (azithromycin) and 94% (ofloxacin) in MBRs. Only 1 (azithromycin) out of 10 compounds investigated in both systems exhibited higher average removal efficiency in CAS than in MBR. Trinh et al. [55] traced 48 MPs including steroidal hormones, xenoestrogens, pesticides, caffeine, pharmaceuticals, and personal care products (PPCPs) in a MBR with >90% removal for many of the compounds. However, amitriptyline, carbamazepine, diazepam, diclofenac, fluoxetine, gemfibrozil, omeprazole,

and log *K*oc values are collected from references (Ref.) as follows: (a) [31], (b) [30], (c) [32].

**Table 4.** Removal of MPs by coagulation/flocculation process from various effluents.

*3.2.2. Biodegradation of micropollutants in secondary treatment*

**(%)** 

**Reference**

77 Vieno et al. [46]

50 Matamoros and

log *Kd*

Longer SRT as required for nitrogen removal also played an important role in reducing the concentrations of certain MPs [59, 60], and a SRT > 10 days was recommended. Longer SRTs resulted in diverse growth of the microbial community including the growth of nitrifying bacteria. Nitrifying bacteria had shown potential for cometabolic degradation of MPs [61, 62]. However, much longer SRT (49 days) was required for 61% removal of iopromide compared to zero removal in CAS [61]. Mixed bacterial cultures also have proved to be quite effective in removing MPs such as triclosan, BPA, and ibuprofen in river [63, 64] and WWTP [65, 66]. While MPs such as quaternary ammonium compounds are biodegraded as single compound, their biodegradation is inhibited in a mixture using *Pseudomonas* sp. isolated from returned activated sludge [67].

Although an important process variable, hydraulic retention time (HRT) shows varied results for the removal of MPs in WWTP indicating that further research is required on this. A study conducted by Wever et al. [57] reported that decreasing the HRT in a CAS resulted in increasing the concentrations of MPs such as 2, 6 and 1, 6 NDSA; however, it did not affect the percent removal of these compounds in a MBR. In case of pharmaceutical and fragrance compounds, Joss et al. [29] reported that HRT played a very minor role when considering a time period of 0.7 hours for fixed bed reactor, 13 hours for a MBR, and 17 hours for a CAS process.

Solution pH plays a significant role in the removal of MPs as the highly acidic or highly basic solutions affect the solubility of the MPs and also hinder growth of the microbial community [68]. As listed in **Table 1**, MPs exhibit a wide range of pKa values. At pH range of 6–8, as found in most wastewater, many antibiotics and other MPs with pKa values in this range will be ionized. For example, about 40% of pharmaceuticals contain at least one functional group with pKa values in the range of 5–10 [69]. The degree of speciation of such ionizable compounds and their subsequent adsorption and biotransformation will be affected by pH.

The microbial growth and activity, as well as solubility and other physicochemical properties of MPs, are significantly affected by temperature. Temperature variability has been related to deterioration in bulk water quality and system instability; it has also been linked to sludge deflocculation and decreased sludge metabolic activity [70]. Vieno et al. [71] reported that the removal of ibuprofen, diclofenac, benzafibrate, ketoprofen, and naproxen increased during the summer (average temperature 17°C) and decreased in the winter (average temperature 7°C). However, Lesjean et al. [72] reported that in a conventional WWTP, temperature variation between 12 and 25°C brought about little or no change to the degradation process of MPs whereas for a MBR the removal rates were greatly affected by the seasonal changes. Hai et al. [70] reported that the removal of most hydrophobic compounds (log *K*ow > 3.2) in a MBR was stable in the temperature range of 10–35°C, while for less hydrophobic compounds, significant variation occurred in the lower temperature regimes (10–35°C). Lower and more variable removal efficiency at 10°C was observed for certain hydrophilic compounds, which have been reported to be moderately recalcitrant in MBR treatment.

No quantitative relationship between structure and activity can be found for the biological transformation. Overall, it can be concluded that for compounds with a sorption coefficient (*Kd*) below 300 L/kg, sorption onto secondary sludge is not relevant, and their transformation can consequently be assessed simply by comparing influent and effluent concentrations.

At low dissolved concentration, the kinetics of biodegradation/biotransformation of MPs follow first order as

$$\text{rate} = K\_{\text{bio}} C\_{\text{ss}} C\_{\text{dis}} \tag{8}$$

where bio is the biodegradation rate constant, ss is the suspended solids concentration, and dis is the dissolved concentration of MPsss. Typically, complex aromatic structure with more than one benzene ring and/or with chlorine and nitro groups are not efficiently biodegraded [32, 73]. The aerobic biodegradation constants of 20 aromatic species using activated sludge were reported, and the kinetic constants were correlated to the structure of the molecules [73]. The normalized first-order rate constants bio (L/gss/h) using ss (g/L) were 0.003, 0.02, and 3.80 for 3, 5 dinitrobenzoic acid, 2, 6 dichlorophenol, and benzoic acid, respectively. Pomiesa et al. [32] summarized a list of both aerobic and anaerobic rate constants for 20 pharmaceuticals including antibiotics, and other compounds such as bisphenol A and nonylphenol, and the aerobic bio (L/gss/h) varied from 0.0025 to 7.08 with carbamazepine being the lowest, and galaxolide (a synthetic fragrance) being the highest biodegradable compound. The difference in rate constants for aerobic and anaerobic conditions is less than 15% for some substances (e.g., celestolide and galaxolide) or can be much higher in some other cases (e.g., >50% for estradiol and roxithromycin). Compounds with *k*bio < 0.0042 L/gss/ h are not removed significantly (<20%), whereas compounds with *k*bio > 0.4 L/gss/h can be transformed by >90%. Therefore, with the existing biological treatment schemes in municipal wastewater, 90% of the MPs are not removed or biotransformed. Many of the plant data do not distinguish between adsorption and biotransformation due to challenging chemical analyses. In most cases, overall removal is estimated based on the influent and effluent concentrations, and information about the intermediate steps is either missing or not reliable [74]. Other challenges are the fate of metabolites, transformation products of pharmaceuticals, and complex chemistry involving these compounds with background water quality, which are all unknown at this point.

Tertiary treatment of wastewater using various combinations of membrane processes, activated carbon adsorption, and advanced oxidation are being performed or characterized in various jurisdictions with stringent water quality requirements. Above technologies all work well for the removal of trace concentration of organics in lab studies and will be described below.

#### *3.2.3. Activated carbon adsorption*

between 12 and 25°C brought about little or no change to the degradation process of MPs whereas for a MBR the removal rates were greatly affected by the seasonal changes. Hai et al. [70] reported that the removal of most hydrophobic compounds (log *K*ow > 3.2) in a MBR was stable in the temperature range of 10–35°C, while for less hydrophobic compounds, significant variation occurred in the lower temperature regimes (10–35°C). Lower and more variable removal efficiency at 10°C was observed for certain hydrophilic compounds, which have been

No quantitative relationship between structure and activity can be found for the biological transformation. Overall, it can be concluded that for compounds with a sorption coefficient (*Kd*) below 300 L/kg, sorption onto secondary sludge is not relevant, and their transformation can consequently be assessed simply by comparing influent and effluent concentrations.

At low dissolved concentration, the kinetics of biodegradation/biotransformation of MPs

where bio is the biodegradation rate constant, ss is the suspended solids concentration, and dis is the dissolved concentration of MPsss. Typically, complex aromatic structure with more than one benzene ring and/or with chlorine and nitro groups are not efficiently biodegraded [32, 73]. The aerobic biodegradation constants of 20 aromatic species using activated sludge were reported, and the kinetic constants were correlated to the structure of the molecules [73]. The normalized first-order rate constants bio (L/gss/h) using ss (g/L) were 0.003, 0.02, and 3.80 for 3, 5 dinitrobenzoic acid, 2, 6 dichlorophenol, and benzoic acid, respectively. Pomiesa et al. [32] summarized a list of both aerobic and anaerobic rate constants for 20 pharmaceuticals including antibiotics, and other compounds such as bisphenol A and nonylphenol, and the aerobic bio (L/gss/h) varied from 0.0025 to 7.08 with carbamazepine being the lowest, and galaxolide (a synthetic fragrance) being the highest biodegradable compound. The difference in rate constants for aerobic and anaerobic conditions is less than 15% for some substances (e.g., celestolide and galaxolide) or can be much higher in some other cases (e.g., >50% for estradiol and roxithromycin). Compounds with *k*bio < 0.0042 L/gss/ h are not removed significantly (<20%), whereas compounds with *k*bio > 0.4 L/gss/h can be transformed by >90%. Therefore, with the existing biological treatment schemes in municipal wastewater, 90% of the MPs are not removed or biotransformed. Many of the plant data do not distinguish between adsorption and biotransformation due to challenging chemical analyses. In most cases, overall removal is estimated based on the influent and effluent concentrations, and information about the intermediate steps is either missing or not reliable [74]. Other challenges are the fate of metabolites, transformation products of pharmaceuticals, and complex chemistry involving these compounds with background water quality, which

bio ss dis rate = *K CC* (8)

reported to be moderately recalcitrant in MBR treatment.

94 Physico-Chemical Wastewater Treatment and Resource Recovery

follow first order as

are all unknown at this point.

Adsorption as a unit operation using either granular‐ or powder‐activated carbon (GAC and PAC) to remove organics from water metrics is well established. The mechanism of adsorption, relevant parameters, and adsorption models discussed in the section of adsorption on sludge are applicable for GAC and PAC adsorption. In absence of experimental data on adsorption isotherm, a correlation developed by Crittenden et al. [75] combining Polanyi potential theory and linear solvation energy relationships (LSERs) can be used.

Activated carbon adsorption for the removal of MPs has been applied in both secondary and tertiary treatment units. Simultaneous adsorption of sulfamethoxazole and carbamazepine to powdered‐activated carbon (PAC) in a membrane bioreactor (MBR) was reported at PAC dosage of 0.1–1 g/L [76–78]. Altmann et al. [77] compared the performance of PAC and ozonation for seven MPs from four different wastewater plants. Typical dosages were about 20 mg/L of PAC and 5–7 mg/L of ozone, respectively, and the performances of both technologies were very much dependent on the type of pollutants. Hydrophobic compounds with log *K*ow > 5 have much better removal potential by adsorption than polar compounds, with the exceptions of some protonated bases and deprotonated acids. Empty bed contact time (EBCT) for a biological‐activated carbon filter for the removal of numerous MPs for three full‐scale reclamation plants varied from 9 to 45 min.

#### *3.2.4. Membrane processes*

Membrane‐based process systems can be classified as direct membrane‐based, integrated membrane‐based, and combined direct and integrated membrane system. Pressure‐driven membrane filtration processes, such as nanofiltration (NF), ultrafiltration (UF), microfiltration (MF), and reverse osmosis (RO), are routinely used for various effluent treatments. While MF and UF are low‐pressure processes, NF and RO are high‐pressure processes. In tertiary treatment of wastewater for MPs, UF and NF can be effectively used. The removal of MPs by membrane depends on many different factors including characteristics of membrane, MP, aqueous media/solute characteristics, operating conditions, and membrane fouling. The fundamental mechanism of membrane filtration is size exclusion, although adsorption due to hydrophobic interactions, electrostatic repulsion, and adsorption on fouling layer all can play a part [79–82]. Size exclusion mechanism is mostly applicable to noncharged MPs, however, shape of the molecule should also be taken into consideration. Hydrophobic interaction and hydrogen bonding contribute to the adsorption of MPs on the membrane surface. Membrane fouling and the presence of dissolved organic carbon could also increase adsorption by changing the membrane surface characteristics and pore size. For charged MP, electrostatic interaction between the compound and membrane surface gives rise to electrostatic exclusion for membrane surfaces with like charges. **Figure 3** shows the four mechanisms of MP removal by membrane processes. Membrane‐based processes have several advantages such as good adaptability, high removal rate, robustness, and no harmful intermediates are formed. An overview of research at laboratory, pilot and full‐scale applications of MPs removal is pre‐ sented in **Table 5**.

**Figure 3.** Micropollutants removal mechanism in polymeric membranes. (a) size exclusion, (b) adsorption (hydropho‐ bic interaction), (c) electrostatic repulsion, and (d) adsorption (fouling layer interaction) (concept adopted from Ojajuni et al. [83]).



**Table 5.** Membrane systems for micropollutants removal in different scales.

for membrane surfaces with like charges. **Figure 3** shows the four mechanisms of MP removal by membrane processes. Membrane‐based processes have several advantages such as good adaptability, high removal rate, robustness, and no harmful intermediates are formed. An overview of research at laboratory, pilot and full‐scale applications of MPs removal is pre‐

**Figure 3.** Micropollutants removal mechanism in polymeric membranes. (a) size exclusion, (b) adsorption (hydropho‐ bic interaction), (c) electrostatic repulsion, and (d) adsorption (fouling layer interaction) (concept adopted from Ojajuni

**MPs % Removal Remarks Reference**

∼40–50% removal for

85±/4% for estradiol, 65±/3% for estrone, 91±/1% for salicine

metals

>70% UF and NF; laboratory scale;

45– 94% Full scale UF; raw sewage of WWTP

50–85% NF; laboratory scale; spiked

secondary effluent

synthetic solution

NF; laboratory scale; spiked synthetic solution

UF; full scale; secondary clarified effluent

Acero et al. [84]

Battistoni et al.

Sahar et al. [86]

Nghiem et al. [87]

Braeken and Van der Bruggen [88]

[85]

sented in **Table 5**.

96 Physico-Chemical Wastewater Treatment and Resource Recovery

et al. [83]).

pesticides)

12.17 μg/L

(TMP)

sulfonamides, and

and Ibuprofen (500 μg/L)

concentration of 1 mg/L

11 MPs 500 μg/L, (pharmaceuticals and

Macrolides, roxithromycin (ROX), clarythromycin (CLA), erythromycin (ERY),

trimethoprim:sulfamethazine (SMZ), sulfamethoxazole (SMX), and trimethoprim

Pharmaceutically active contaminants (PhACs): sulfamethoxazole, carbamazepine,

EDCs–estrone, estradiol, and salicin at initial

80 MPs; Metals 18–265 μg/L, VOC 0.65–7.10 μg/L, PAH 0.23–0.67 μg/L, and HVOC 1.45–

#### *3.2.5. Advanced oxidation processes*

Advanced oxidation processes (AOPs) using hydroxyl radicals (OH•) are increasingly used for tertiary treatment of municipal wastewater and for water recycling. These processes are fast, nonselective, and effective for recalcitrant compounds. Among numerous combinations of AOPs, UV‐, hydrogen peroxide‐, and ozone‐based processes are easy to implement for tertiary treatment of WWTP effluent. In a comprehensive research, removal efficiency of 220 MPs with postozonation was studied at full scale for a WWTP [1]. Compounds with activated aromatic moieties, amine functions, or double bonds such as sulfamethoxazole, diclofenac, or carbamazepine had second‐order rate constants for ozonation >104 /M/s at pH 7 (fast reacting) were eliminated to concentrations below the detection limit for an ozone dose of 0.47 g O3/g DOC. Higher ozone dosage of 0.6 g O3/g DOC was needed for more recalcitrant compounds such as atenolol and benzotriazole for >85%. Rahman et al. [99] summarized the second‐order ozone and OH• oxidation constants for commonly found EDCs and pharmaceuticals in pure water, which varied from 0.8 to 7 × 109 and 1.2 × 109 to 9.8 ×109 /MS, respectively. In wastewater, rates will be somewhat lower due to the competition of background organics, suspended particulates, and radical scavengers. However, the effect of background organics competition was found to be minimal for estrone degradation in wastewater by Sarkar et al. [100]. The overall cost of ozonation was found to be lower than that of UV/H2O2 process for estrone degradation, although electrical energy per order was lower for UV/H2O2. AOPs are effective in a wide range of pH (i.e., 4–11) depending on the type of target compounds; although ozonation is more effective in alkaline pH. In some cases, transformation products that form due to AOPs may be even more toxic compared to parent compounds. For example, inter‐ mediates of UV/H2O2 oxidation of bisphenol A exhibited different estrogenic activity depend‐ ing on the treatment conditions [101]. Whole effluent analysis methods are better for assessing the toxicity of resulting water instead of time‐ and labor‐intensive chemical analyses.

#### **4. Conclusion**

Fate and removal processes of micropollutants (MPs) in wastewater treatment are complex, and difficult to assess due to tedious and cost‐intensive analyses. However, these processes can be somewhat estimated based on their physical properties such as log *K*ow, pKa, and solubility. Adsorption on colloidal and suspended particles and subsequent removal in sludge may occur for compounds with log *K*ow > 4.0. Majority of the MPs are not removed in conven‐ tional‐activated sludge process, although better removal for some cases occurs in membrane bioreactors due to greater diversity and adaptability of microorganisms. Compounds with biological degradation constant <0.0042 L/gss/h are not removed significantly (<20%), whereas compounds with rate constants >0.4 L/gss/h can be transformed by >90%. Tertiary treatment of wastewater effluent using activated carbon adsorption, membrane filtration, and advanced oxidation processes are capable of removing MPs with varying degrees of success, although both lab and pilot‐scale studies are required to establish their rates of removal. In the case of intermediates or transformation, products are produced during a treatment, whole effluent analysis using a bioassay is a better method to evaluate the quality of effluent instead of conducting compounds specific chemical analyses.

## **Author details**

*3.2.5. Advanced oxidation processes*

98 Physico-Chemical Wastewater Treatment and Resource Recovery

water, which varied from 0.8 to 7 × 109

**4. Conclusion**

Advanced oxidation processes (AOPs) using hydroxyl radicals (OH•) are increasingly used for tertiary treatment of municipal wastewater and for water recycling. These processes are fast, nonselective, and effective for recalcitrant compounds. Among numerous combinations of AOPs, UV‐, hydrogen peroxide‐, and ozone‐based processes are easy to implement for tertiary treatment of WWTP effluent. In a comprehensive research, removal efficiency of 220 MPs with postozonation was studied at full scale for a WWTP [1]. Compounds with activated aromatic moieties, amine functions, or double bonds such as sulfamethoxazole, diclofenac, or

were eliminated to concentrations below the detection limit for an ozone dose of 0.47 g O3/g DOC. Higher ozone dosage of 0.6 g O3/g DOC was needed for more recalcitrant compounds such as atenolol and benzotriazole for >85%. Rahman et al. [99] summarized the second‐order ozone and OH• oxidation constants for commonly found EDCs and pharmaceuticals in pure

rates will be somewhat lower due to the competition of background organics, suspended particulates, and radical scavengers. However, the effect of background organics competition was found to be minimal for estrone degradation in wastewater by Sarkar et al. [100]. The overall cost of ozonation was found to be lower than that of UV/H2O2 process for estrone degradation, although electrical energy per order was lower for UV/H2O2. AOPs are effective in a wide range of pH (i.e., 4–11) depending on the type of target compounds; although ozonation is more effective in alkaline pH. In some cases, transformation products that form due to AOPs may be even more toxic compared to parent compounds. For example, inter‐ mediates of UV/H2O2 oxidation of bisphenol A exhibited different estrogenic activity depend‐ ing on the treatment conditions [101]. Whole effluent analysis methods are better for assessing

to 9.8 ×109

and 1.2 × 109

the toxicity of resulting water instead of time‐ and labor‐intensive chemical analyses.

Fate and removal processes of micropollutants (MPs) in wastewater treatment are complex, and difficult to assess due to tedious and cost‐intensive analyses. However, these processes can be somewhat estimated based on their physical properties such as log *K*ow, pKa, and solubility. Adsorption on colloidal and suspended particles and subsequent removal in sludge may occur for compounds with log *K*ow > 4.0. Majority of the MPs are not removed in conven‐ tional‐activated sludge process, although better removal for some cases occurs in membrane bioreactors due to greater diversity and adaptability of microorganisms. Compounds with biological degradation constant <0.0042 L/gss/h are not removed significantly (<20%), whereas compounds with rate constants >0.4 L/gss/h can be transformed by >90%. Tertiary treatment of wastewater effluent using activated carbon adsorption, membrane filtration, and advanced oxidation processes are capable of removing MPs with varying degrees of success, although both lab and pilot‐scale studies are required to establish their rates of removal. In the case of intermediates or transformation, products are produced during a treatment, whole effluent

/M/s at pH 7 (fast reacting)

In wastewater,

/MS, respectively.

carbamazepine had second‐order rate constants for ozonation >104

Sreejon Das, Nillohit Mitra Ray, Jing Wan, Adnan Khan, Tulip Chakraborty and Madhumita B. Ray\*

\*Address all correspondence to: mray@eng.uwo.ca

Department of Chemical and Biochemical Engineering, University of Western Ontario, London, Ontario, Canada


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#### **PCDDs/PCDFs and PCBs in Wastewater and Sewage Sludge PCDDs/PCDFs and PCBs in Wastewater and Sewage Sludge**

Magdalena Urbaniak and Anna Wyrwicka Magdalena Urbaniak and Anna Wyrwicka

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/66204

#### **Abstract**

The chapter includes the information concerning the wastewater treatment plants (WWTPs) functioning in respect to polychlorinated dibenzo‐*p*‐dioxins (PCDDs)/ polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs). In particular, the chapter describes the occurrence and fate of PCDDs/PCDFs and PCBs in WWTPs, at different treatment stages, including the tertiary wastewater treatment (e.g. constructed, wetlands biofilters) and factors affecting the removal of these micropollu‐ tants during treatment process. Considering the production of growing amounts of sewage sludge as an end product of the wastewater treatment process, the chapter describes also the occurrence and fate of above‐mentioned compounds in sewage sludge and the ways of their utilization with the special emphasis on agricultural uses, bioremediation and phytoremediation processes. With regard to the agricultural use of sewage sludge, the impact of sludge‐born PCDDs/PCDFs and PCBs on plant growth and plant metabolism is described, together with the current state of knowledge on the accumulation and translocation of the studied compounds in plant tissues.

**Keywords:** PCDDs/PCDFs, PCBs, wastewater, sewage sludge, phytoremediation, plant growth, plant metabolism

## **1. Introduction**

Rapid growth in global population has been observed from approximately 5.3 billion in 1992 [1] to about 6.97 billion in 2011 [2]. United Nation predicted that in 2030, the global population reach over 8 billion, whereas in 2050 exceed 9 billion. The growing population affect the consumption of water and consequential production of wastewater. The projections

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

concerning influent wastewater flow in USA estimate its rise from 100,000,000 m3 /day in 1996 to 170,000,000 m3 /day in 2025 [3–5].

An increased usage of water around the world led to an increased concern about the outgoing wastewater quality from municipal wastewater treatment plants (WWTPs) [6]. Usually, quantification of wastewater quality is based on monitoring of traditional parameters which can be analysed in easy and inexpensive way and are regulated by the European Urban Wastewater Directive (91/271/EEC). These parameters include biochemical oxygen demand (BOD), chemical oxygen demand (COD), nitrates, phosphates and total suspended solids [7]. Nevertheless, these routine chemical analyses cannot give a complete overview of the threat to the water environment posed by other substances released through the WWTPs effluents such as polychlorinated dibenzo‐*p*‐dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs), which are toxic, carcinogenic and known endocrine disrupters posing a serious risk for living organisms [6]. According the Directive of the European Parliament and the Council 2013/39/EC of 12 August 2013 amending Directive 2000/60/EC and 2008/105/EC in respect of priority substances in the field of water policy, PCDDs/PCDFs and PCBs have been identified as priority hazardous substances, which need to be eliminated from the water environment.

Considering the above, the present chapter reviews the available data concerning the occur‐ rence and fate of PCDDs/PCDFs and PCBs in wastewater (point 2) and sewage sludge (point 3) with the special emphasis of the ways of sewage sludge utilization and impact of sludge‐ born PCDDs/PCDFs and PCBs on the plant growth and plant metabolism.

## **2. The occurrence and fate of PCDDs/PCDFs and PCBs in WWTPs**

WWTPs represent an obligatory and final step prior to the release of wastewater into the environment. Hence, an emerging task for WWTPs would be to act as a barrier for micropol‐ lutants, preventing the emission of potentially harmful substances into the aqueous environ‐ ment. WWTPs use different kinds of methods including biological, physical and chemical processes, to fulfil the regulatory standards regarding the quality of the effluent discharges. Regardless of the methods used at any particular WWTP, all the treatment processes can be generally divided into three categories: (1) primary, (2) secondary and (3) advanced tertiary treatment [5]. The primary treatment removes large objects from incoming wastewater through floatation, settling and screening mechanisms and the smaller objects such as sand are removed in grit chambers and sedimentations tanks. The secondary treatment is designed to substan‐ tially degrade organic matter and dissolved nutrients using trickling filters and activated sludge. The purpose of the tertiary treatment is to further improve the effluent quality before it is discharged to the receiving environment and include filtration, chlorination and UV radiation.

The occurrence of PCDDs/PCDFs and PCBs in the untreated wastewater and sewage sludge has been studied very intensively during recent decades and revealed their very high concen‐ trations with a predominance of highly chlorinated congeners [8–15].

The available literature data indicate that conventional wastewater treatment systems are not able to sufficiently remove hydrophobic contaminants, which have adverse effects on the receiving water ecosystem [5, 7, 13]. Thus, organic compounds are detected in the river water worldwide [13, 16–22]. This is due to the fact that for many years, quantification of wastewater effluents and receiving river water pollution were restricted to monitor biochemical oxygen demand (BOD), chemical oxygen demand (COD), nitrogen and phosphorus concentrations and total suspended solids [7]. However, as shown in the work of Urbaniak et al. [13] and Urbaniak and Kiedrzyńska [14], significant concentrations of PCDDs/PCDFs and PCBs may be present in treated wastewater, with the highest values in the smallest WWTPs. All WWTPs studied by Urbaniak et al. [13] were found to discharge toxic PCDD/PCDF and PCB compounds into their receiving rivers. This is the effect of insufficient regulation of the discharge of toxic congeners of PCDDs/PCDFs by municipal WWTPs: the existing regulations only apply to municipal WWTPs with a population equivalent (p.e.) of 100,000. In consequence, the release of PCDDs/PCDFs in treated wastewater from the studied WWTPs is not regulated, as the plants are below this p.e. This, together with the increasing number of municipal WWTPs, and the results presented by Sztamberek‐Gola et al. [23] and Oleszek‐Kudlak et al. [24], which demonstrate increases in the concentrations of the lower chlorinated, and hence, more toxic, PCDDs/PCDFs in WWTP outlet water, may result in lower quality of the receiving waters. Data presented by Sztamberek‐Gola et al. [23] and Oleszek‐Kudlak et al. [24] obtained on the basis of three WWTP analyses, revealed total and toxic equivalency (TEQ) concentrations within the range of 107.26–219.19 pg/m3 for total PCDDs, from 201.75 to 736.50 pg/m3 for total PCDFs and from 14.70 to 116.40 pg I‐TEQ/m3 for TEQ. Moreover, the authors observed increased PCDD and PCDF concentrations to be related to increased daily wastewater flow: the lowest values were noted in effluents from the smallest WTP, with a daily flow of 20,000 m3 , whereas samples coming from WWTPs with twice the flow (40,000 and 45,000 m3 ) were found to have concentrations about two times higher. Considering the above results, the authors note that wastewater treatment affects the fate of PCDDs/PCDFs, with increased amounts of lower chlorinated, and thus more toxic, congeners in the outlet effluents. As a consequence, the International‐TEQ (I‐TEQ) concentrations are more than five times higher in the outgoing treated effluent than the incoming wastewater. Moreover, the authors observe a predominance of PCDFs over PCDDs in the outgoing effluents. Also other studies confirm the presence of PCDDs/PCDFs in wastewater effluents. The study of Rappe et al. [10] showed that the TEQ and PCDD/PCDF concentrations in wastewater effluents from publicly owned treatment works ranged between 0.264 and 3.84 pg TEQ/L. Urbaniak et al. [13] examined 17 outflows of treated wastewater from municipal wastewater treatment plants. Sewage treatment plants were divided into three classes based on their p.e. size, that is: class I (0–1999 p.e.), class II (2000–9999 p.e.), class III (10,000–14,999 p.e.) and class IV (15,000– 99,999 p.e.). The analysis of the treated wastewater collected at the sewage outlets revealed that toxic PCDDs/PCDFs and dl‐PCBs were present at a range of concentrations from 32.30 to 732.79 pg/L. The mean values at high water flow and during stable hydrological conditions were respectively 81.96 and 216.92 pg/L for class I wastewater treatment plants, 80.47 and 74.30 pg/L for class II, and 69.82 and 137.06 pg/L for class IV. These results indicate

concerning influent wastewater flow in USA estimate its rise from 100,000,000 m3

An increased usage of water around the world led to an increased concern about the outgoing wastewater quality from municipal wastewater treatment plants (WWTPs) [6]. Usually, quantification of wastewater quality is based on monitoring of traditional parameters which can be analysed in easy and inexpensive way and are regulated by the European Urban Wastewater Directive (91/271/EEC). These parameters include biochemical oxygen demand (BOD), chemical oxygen demand (COD), nitrates, phosphates and total suspended solids [7]. Nevertheless, these routine chemical analyses cannot give a complete overview of the threat to the water environment posed by other substances released through the WWTPs effluents such as polychlorinated dibenzo‐*p*‐dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs), which are toxic, carcinogenic and known endocrine disrupters posing a serious risk for living organisms [6]. According the Directive of the European Parliament and the Council 2013/39/EC of 12 August 2013 amending Directive 2000/60/EC and 2008/105/EC in respect of priority substances in the field of water policy, PCDDs/PCDFs and PCBs have been identified as priority hazardous substances, which need

Considering the above, the present chapter reviews the available data concerning the occur‐ rence and fate of PCDDs/PCDFs and PCBs in wastewater (point 2) and sewage sludge (point 3) with the special emphasis of the ways of sewage sludge utilization and impact of sludge‐

WWTPs represent an obligatory and final step prior to the release of wastewater into the environment. Hence, an emerging task for WWTPs would be to act as a barrier for micropol‐ lutants, preventing the emission of potentially harmful substances into the aqueous environ‐ ment. WWTPs use different kinds of methods including biological, physical and chemical processes, to fulfil the regulatory standards regarding the quality of the effluent discharges. Regardless of the methods used at any particular WWTP, all the treatment processes can be generally divided into three categories: (1) primary, (2) secondary and (3) advanced tertiary treatment [5]. The primary treatment removes large objects from incoming wastewater through floatation, settling and screening mechanisms and the smaller objects such as sand are removed in grit chambers and sedimentations tanks. The secondary treatment is designed to substan‐ tially degrade organic matter and dissolved nutrients using trickling filters and activated sludge. The purpose of the tertiary treatment is to further improve the effluent quality before it is discharged to the receiving environment and include filtration, chlorination and UV

The occurrence of PCDDs/PCDFs and PCBs in the untreated wastewater and sewage sludge has been studied very intensively during recent decades and revealed their very high concen‐

trations with a predominance of highly chlorinated congeners [8–15].

born PCDDs/PCDFs and PCBs on the plant growth and plant metabolism.

**2. The occurrence and fate of PCDDs/PCDFs and PCBs in WWTPs**

/day in 2025 [3–5].

110 Physico-Chemical Wastewater Treatment and Resource Recovery

to be eliminated from the water environment.

to 170,000,000 m3

radiation.

/day in 1996

that small wastewater treatment plants had higher concentrations of the studied compounds than the larger ones. In the case of the concentrations measured as TEQs, the obtained values were less diverse, amounting to 4.38 and 3.81 pg TEQ/L for high and stable flow in class I wastewater treatment plants, 4.72 and 3.97 pg TEQ/L for class II, and 3.94 and 3.15 pg  TEQ/L for class IV.

With reference to the PCBs, only a few publications refer to their concentrations in the WWTPs effluents. Katsoyiannis and Samara [25, 26] demonstrated the occurrence of the sum of indicator PCBs (∑7PCBs) in the raw urban wastewater and wastewater after primary and secondary treatment steps. The authors showed decreasing mean concentrations of ∑7PCBs from 1,000,000 through 631,000 to 250,000 pg/L in raw wastewater and effluents from primary and secondary treatment stage, respectively. Another research conducted by Blanchard et al. [27] in the outflow from the Montreal WTP (Canada) showed much lower concentrations (measured as sum of 13 PCB congeners) ranged from 20 to 860 pg/L with the mean value of 310 pg/L, whereas Pham and Prolux [28] found a concentration of ∑13PCBs in the treated wastewater from the same WTP of 1400 pg/L. The study of Bergqvist et al. [6] conducted in two WTPs in Umea (Sweden) and in Siauliai (Lithuania) showed higher ∑7PCBs ranged from 1000 to 6000 pg/L. The authors also demonstrated a rapid increase of the ∑7PCBs during the treatment process (ranged from 300 to 1000 pg/L and from 1000 to 6000 pg/L in the case of Umea and Siauliai WTP, respectively). However, other authors suggest that treatment proc‐ esses such as sorption to the sludge remove up to 70% of PCBs, whereas volatilization led to eliminate of about 50% of the Aroclor 1254 [29, 30]. According to Pham and Prolux, 1997, the removal rates ranged from 33% (for PCB: 101) up to 100% (for PCB: 194) with the average value for the∑13PCBs of 67%. Despite the above, Urbaniak and Kiedrzyńska [14] note that the treated wastewater effluent of the smallest wastewater treatment plants, class I, is characterized by dl‐ PCB values more than double those of medium and large wastewater plants. This phenomenon was not noted for TEQ values, which were found in the narrow range of 0.31–0.37 pg TEQ/L. The study of Urbaniak and Kiedrzyńska [14] demonstrates a significant problem with the maintenance of the proper purification efficiency in all the studied WWTPs and in this way effluents quality which have the potential to affect the quality of river water.

In order to enhance the removal of PCDDs/PCDFs and PCBs from wastewater effluents and the receiving river waters, the land‐water ecotones constructed in a river valley with different kinds of plants and micro‐organisms may be applied. Such structures may partially purify the inflowing surface water and groundwater contaminated by PCDDs/PCDFs and PCBs through their capturing, immobilization and/or degradation [31–33]. Wetlands are another promising solution towards wastewater purification due to their intrinsic function to transform and store organic matter and nutrients [34, 35] and associated micropollutants such as PCDDs/PCDFs and PCBs. Due to these properties, wetlands have been used for water quality improvement worldwide [36]. Constructed wetlands were first used for treatment of wastewater in the 1950s, while in last years, they are also used for treatment of runoff water from city areas and agriculture. Constructed wetlands exploit natural processes to remove pollutants in a sustain‐ able cost and in an energy effective way with minimal operation and maintenance cost. Moreover, their usage as tools in the treatment of polluted waters has been gaining popularity as an ecological engineering alternative over conventional, chemical‐based methods.

that small wastewater treatment plants had higher concentrations of the studied compounds than the larger ones. In the case of the concentrations measured as TEQs, the obtained values were less diverse, amounting to 4.38 and 3.81 pg TEQ/L for high and stable flow in class I wastewater treatment plants, 4.72 and 3.97 pg TEQ/L for class II, and 3.94 and 3.15 pg 

With reference to the PCBs, only a few publications refer to their concentrations in the WWTPs effluents. Katsoyiannis and Samara [25, 26] demonstrated the occurrence of the sum of indicator PCBs (∑7PCBs) in the raw urban wastewater and wastewater after primary and secondary treatment steps. The authors showed decreasing mean concentrations of ∑7PCBs from 1,000,000 through 631,000 to 250,000 pg/L in raw wastewater and effluents from primary and secondary treatment stage, respectively. Another research conducted by Blanchard et al. [27] in the outflow from the Montreal WTP (Canada) showed much lower concentrations (measured as sum of 13 PCB congeners) ranged from 20 to 860 pg/L with the mean value of 310 pg/L, whereas Pham and Prolux [28] found a concentration of ∑13PCBs in the treated wastewater from the same WTP of 1400 pg/L. The study of Bergqvist et al. [6] conducted in two WTPs in Umea (Sweden) and in Siauliai (Lithuania) showed higher ∑7PCBs ranged from 1000 to 6000 pg/L. The authors also demonstrated a rapid increase of the ∑7PCBs during the treatment process (ranged from 300 to 1000 pg/L and from 1000 to 6000 pg/L in the case of Umea and Siauliai WTP, respectively). However, other authors suggest that treatment proc‐ esses such as sorption to the sludge remove up to 70% of PCBs, whereas volatilization led to eliminate of about 50% of the Aroclor 1254 [29, 30]. According to Pham and Prolux, 1997, the removal rates ranged from 33% (for PCB: 101) up to 100% (for PCB: 194) with the average value for the∑13PCBs of 67%. Despite the above, Urbaniak and Kiedrzyńska [14] note that the treated wastewater effluent of the smallest wastewater treatment plants, class I, is characterized by dl‐ PCB values more than double those of medium and large wastewater plants. This phenomenon was not noted for TEQ values, which were found in the narrow range of 0.31–0.37 pg TEQ/L. The study of Urbaniak and Kiedrzyńska [14] demonstrates a significant problem with the maintenance of the proper purification efficiency in all the studied WWTPs and in this way

effluents quality which have the potential to affect the quality of river water.

In order to enhance the removal of PCDDs/PCDFs and PCBs from wastewater effluents and the receiving river waters, the land‐water ecotones constructed in a river valley with different kinds of plants and micro‐organisms may be applied. Such structures may partially purify the inflowing surface water and groundwater contaminated by PCDDs/PCDFs and PCBs through their capturing, immobilization and/or degradation [31–33]. Wetlands are another promising solution towards wastewater purification due to their intrinsic function to transform and store organic matter and nutrients [34, 35] and associated micropollutants such as PCDDs/PCDFs and PCBs. Due to these properties, wetlands have been used for water quality improvement worldwide [36]. Constructed wetlands were first used for treatment of wastewater in the 1950s, while in last years, they are also used for treatment of runoff water from city areas and agriculture. Constructed wetlands exploit natural processes to remove pollutants in a sustain‐ able cost and in an energy effective way with minimal operation and maintenance cost.

TEQ/L for class IV.

112 Physico-Chemical Wastewater Treatment and Resource Recovery

The promising solution provides ecohydrology [37], through the use of cascade system of biofilters for the purification of wastewater, runoff water, leachate, etc. The biofilters which consist with zone of intensive sedimentation, which facilitate the deposition of matter, nutrients and micropollutants and their further biodegradation by existing microbiota and macrophyte zone where an intensive phytodegradation processes occur, are considered to be one of the most effective solutions for pollutant removal. Our earlier results obtained on the basis of such systems functioning in the urban area and receiving the untreated sewage and storm water (Sokołówka River, Poland) showed the removal efficiency reaching 95% for mineral matter, 86% for organic matter, 81% for total nitrogen and 86% for total phosphorus [38]. At the same time, removal efficiency of biofilter located in rural area (Asella lake, Ethiopia) was 67%, 36%, 76% and 93% for mineral matter, organic matter, total nitrogen and total phosphorus, respectively [38]. Moreover, results from the biofilter located in Asella (Central Ethiopia) demonstrated a 70% reduction of the lake sediment TEQ after one year of biofilter implementation (data not published). The implementation of such biofiltration system enabled a reduction in the input of micropollutants into the river recipients through sedimentation and acceleration of biodegradation and phytodegradation processes and in this way indicates the positive role of such systems in the quality of water ecosystems and in consequence of human health.

## **3. The occurrence and fate of PCDDs/PCDFs and PCBs in sewage sludge and sewage sludge amended soil**

The occurrence of PCDDs/PCDFs and PCBs in inflowing wastewater causes considerable problems for the WWTPs because conventional biological and chemical processes are insufficient for removing them. What is more is scarce data exist to explain how wastewater treatment affects the behaviour and fate of PCDDs/PCDFs and PCBs. Since they have a very high sorption potential [39], they are expected to partition into the sewage sludge part of the wastewater during treatment processes. In addition, the majority of treatment processes are very conducive to volatilization; hence, low volatilization potentials of PCDDs/PCDFs reduce their loss [24].

Various studies confirm that sewage sludge contains a very high level of PCDDs/PCDFs and PCBs ranging between 2.26 and 1270 ng I‐TEQ/kg in the United States [10], from 19 to 225 ng I‐TEQ/kg in UK [40], from 7 to 160 ng I‐TEQ/kg in Spain [41], and between 16.85 and 74.56  ng I‐TEQ/kg in Poland [24, 42]. Our study from the Lodz Wastewater Treatment plant showed the concentration of 17 toxic congeners PCDDs/PCDF in sewage sludge equal to 3270.07  ng/kg and the TEQ concentration equal to 29.71 ng TEQ/kg.

These findings confirm that the majority of PCDDs, PCDFs and PCBs are deposited in sludge. This in turn implicates problems with the further use of such contaminated sludge as a fertilizer especially because PCDDs/PCDFs and PCBs toxicity is further enhanced by their accumulation in soil, and bioaccumulation and biomagnification within food chains.

Concerning the above, the further part of the chapter is focused on the methods dedicated to safe disposal and utilization of sewage sludge and the fate of sludge‐born PCDDs/PCDFs and PCBs in the environment.

## **4. Sewage sludge utilization**

The main methods of sewage sludge utilization include storage, natural resources and agricultural land use, and burning. At present, most often the sewage sludge is stored on sludge lagoons. This practice became insufficient because (1) the storage has a limited capacity and (2) sludge could be used as a potential recyclable material; however, this method requires drying of sewage sludge to the content of 58–95% dry weight (d.w.), which is high energy‐ consuming.

The use of thermal processes removes organic compounds associated with the sewage sludge but leave contaminated fly ash. Moreover, this kind of sewage sludge utilization led to air pollution and airborne diseases among human population due to smoke production which may contain toxic compounds like heavy metals. The use of efficient equipment led to the reduction of emissions of harmful elements to the atmosphere but at the same time move the problem of pollutant emissions to the captured ashes. Additionally, during the incineration process as an end by‐product, the hydrogen (H2), methane (CH4), carbon monoxide (CO) and carbon dioxide (CO2) are produced. This led to increased production of greenhouse gases, which are the main concern of the Kyoto Protocol regarding climate change.

The alternative method of sewage sludge utilization is their use as a soil and plant fertilizer. This way of their utilization is possible thanks to high organic matter content and high levels of nitrogen and phosphorus which are required for plant growth [43, 44]. Moreover, the organic matter increases the water capacity influences by this way positive on the structure, texture and microbial activity of the soil. The use of sewage sludge as a fertilizer is widespread [9]. The amount of sludge used for agricultural purpose is 25% in Germany and up to 90% in Sweden [9].

The use of sewage sludge as plant fertilizer is not only the method of sewage sludge manage‐ ment but also the method for implementation of the Renewable Energy Sources Directive when using energetic crops (2001/77/EC) and the Kyoto Protocol (OJ L 203 of 2005, p.1684). Moreover, crops may be used for the reduction of enhanced pollutants levels in soil after sludge application.

Following application, the sludge is present in relatively thin film on the soil surface. Nevertheless, it should be stated that due to high persistence of the PCDDs/PCDFs and PCBs, the addition of these compounds to the soil through the application of sludge must lead to an increase in soil contamination. This is important because Regulation of the Ministry of Environment of Poland from 16 April 2002 recommended to not exceed the 20 ng PCB/g d.w. in agricultural soil (PCB: 28, 52, 101, 118, 138, 153 and 180) and 2000 ng PCB/g d.w. in industrial soil (OJ 2002, 63 item 634). Nevertheless, there is no law regulation concerning the concentrations of 17 toxic congeners of PCDD, PCDF and 12 toxic dl‐PCB in soil. In contrast, in Germany, the limit for PCB is 0.4, 2.0, 0.8 and 40.0 mg/kg d.w. for playgrounds, parks, residential and industrial areas, respectively [45]. The limit for PCDD/PCDF is the following: 100, 1000, 1000 and 10,000 ng TEQ/kg d.w. for playgrounds, parks, residential and industrial areas, respectively [46].

Concerning the above, the further part of the chapter is focused on the methods dedicated to safe disposal and utilization of sewage sludge and the fate of sludge‐born PCDDs/PCDFs and

The main methods of sewage sludge utilization include storage, natural resources and agricultural land use, and burning. At present, most often the sewage sludge is stored on sludge lagoons. This practice became insufficient because (1) the storage has a limited capacity and (2) sludge could be used as a potential recyclable material; however, this method requires drying of sewage sludge to the content of 58–95% dry weight (d.w.), which is high energy‐

The use of thermal processes removes organic compounds associated with the sewage sludge but leave contaminated fly ash. Moreover, this kind of sewage sludge utilization led to air pollution and airborne diseases among human population due to smoke production which may contain toxic compounds like heavy metals. The use of efficient equipment led to the reduction of emissions of harmful elements to the atmosphere but at the same time move the problem of pollutant emissions to the captured ashes. Additionally, during the incineration process as an end by‐product, the hydrogen (H2), methane (CH4), carbon monoxide (CO) and carbon dioxide (CO2) are produced. This led to increased production of greenhouse gases,

The alternative method of sewage sludge utilization is their use as a soil and plant fertilizer. This way of their utilization is possible thanks to high organic matter content and high levels of nitrogen and phosphorus which are required for plant growth [43, 44]. Moreover, the organic matter increases the water capacity influences by this way positive on the structure, texture and microbial activity of the soil. The use of sewage sludge as a fertilizer is widespread [9]. The amount of sludge used for agricultural purpose is 25% in Germany and up to 90% in

The use of sewage sludge as plant fertilizer is not only the method of sewage sludge manage‐ ment but also the method for implementation of the Renewable Energy Sources Directive when using energetic crops (2001/77/EC) and the Kyoto Protocol (OJ L 203 of 2005, p.1684). Moreover, crops may be used for the reduction of enhanced pollutants levels in soil after sludge

Following application, the sludge is present in relatively thin film on the soil surface. Nevertheless, it should be stated that due to high persistence of the PCDDs/PCDFs and PCBs, the addition of these compounds to the soil through the application of sludge must lead to an increase in soil contamination. This is important because Regulation of the Ministry of Environment of Poland from 16 April 2002 recommended to not exceed the 20 ng PCB/g d.w. in agricultural soil (PCB: 28, 52, 101, 118, 138, 153 and 180) and 2000 ng PCB/g d.w. in industrial soil (OJ 2002, 63 item 634). Nevertheless, there is no law regulation concerning the

which are the main concern of the Kyoto Protocol regarding climate change.

PCBs in the environment.

consuming.

Sweden [9].

application.

**4. Sewage sludge utilization**

114 Physico-Chemical Wastewater Treatment and Resource Recovery

In case of sewage sludge, the Directive 86/278/EEC on the protection of the environment, and in particular of the soil when sewage sludge is used in agriculture, does not provide any limit values or requirements for organic compounds in sewage sludge. Thus, several national regulations on the use of sludge have added specifications on organic compounds. This is the case in particular of Austria, France and Germany which have all included limit values for some organic compounds in the relevant regulation for the use of sludge, for example, in Austria, the limit values for PCDDs/PCDFs in sewage sludge are 100 ng TEQ/kg d.w. in Lower and Upper Austria and Burgenland and 50 ng TEQ/kg d.w. in Carinthia [47]. The limit of 100  ng TEQ/kg d.w. is also valid in Germany [48]. The limits for PCBs in sewage sludge are the following: 0.2 mg/kg d.w. in Lower and Upper Austria and Burgenland and 1.0 mg/kg d.w. in Burgenland. In France, the limit for sum of seven principal PCBs (PCB 28, 52, 101, 118, 138, 153, 180) is 0.8 mg/kg d.w.; in Germany is 0.2 mg/kg d.w. for each of the six PCB congeners; and in Sweden is 0.4 mg/kg d.w. In Poland, according to Ministerial Decree (OJ 2009, 27 item 169), PCBs should be completely removed from the sewage during their treatment; neverthe‐ less, there is no limits of the aforementioned compounds in sewage sludge.

Despite the above national regulation, European Union proposed some limit values for concentrations of organic compounds and PCDDs/PCDFs in sludge for use on land. The mentioned proposed limit values are following: 0.8 mg/kg of dry matter for PCBs (sum of PCBs 28, 52, 101, 118, 138, 153 and 180) and 100 ng TEQ/kg of dry matter for PCDDs/PCDFs [49]. Also the US EPA proposed the limit of 300 ng TEQ/kg of dry matter for 17 toxic PCDDs/PCDFs and 12 coplanar PCBs [50].

## **5. Bioremediation and phytoremediation of sludge originated PCDDs/ PCDFs and PCBs in soil**

Reports on the biodegradation of chlorinated dioxins in the soil are contradicting. On the one hand, there are studies that indicate that chlorinated PCDDs/PCDFs are persistent. One of such studies considered chlorinated PCDDs/PCDFs that were introduced into soil through land application of sewage sludge [51]. According to this study, the PCDDs/PCDFs concentrations did not change significantly after 260 days of monitoring. On the other hand, the evidence obtained in other experiments suggests that PCDDs/PCDFs are degraded in soil, for example, the concentration of 2,3,7,8‐TCDD was monitored over 10 years in the soil and was shown to be significantly decrease. Biodegradation was also observed in the soil spiked with one to 100 ppm of 2,3,7,8‐TCDD. Between 37 and 44% of added 2,3,7,8‐TCDD was eliminated during 1 year [52].

The above results are connected with the soil microbial transformation of micropollutants. Many literature data suggest that microbial biodegradation is the critical event determining the fate and persistence of PCDDs/PCDFs in the soil [53].

Biodegradation is dependent on micro‐organism enzymes which modify toxic compounds into less toxic forms. Biodegradation can be carried as two processes: mineralization, when organic compound uses a sole source of carbon and energy by micro‐organisms, and co‐ metabolism where transformation of given pollutant depends on the presence of other substrate. Products of this process can be further mineralized; otherwise, incomplete degra‐ dation occurs, leading to a formation and accumulation of metabolites more toxic than parent substrates.

In this place, there is a need to underline the role of humic acids—major components of soil organic matter which consist of complex polymers of hydroxyphenols, hydroxybenzoic and methoxybenzoic acids and other aromatic structures with linked peptides, amino sugar compounds, fatty acids and possibly other constituents. Hydroquinone/quinone‐type couples are perceived to affect the redox properties of humic acid and to act either as electron transfer mediators or as direct donors of electrons. Thus, the amount of humic acids may determine the microbial dechlorination of PCDDs/PCDFs [54].

The use of anaerobic and aerobic micro‐organisms is the only known process of PCDDs/PCDFs degradation in soil and aquatic systems, leading to a removal of chlorine atoms from the biphenyl molecule and theoretically releasing CO2, chlorine and water. Highly chlorinated congeners have been found to be reductively dechlorinated under anaerobic conditions through a preferential meta‐ and para‐chlorine removal and production of less chlorinated congeners, which can then be used in aerobic transformations. Thus, complete degradation of PCDDs/PCDFs can be achieved by a sequential exposure to anaerobic and aerobic micro‐ organisms [55].

The fungi, similarly to bacteria, are also capable to degrade PCDDs/PCDFs in the presence of oxygen using both processes: mineralization and co‐metabolism. The fungi use specific enzymes named lignin peroxidase or manganese peroxidase which enable to oxidize the pollutant molecule. The fungal aerobic biodegradation was first reported by Bumpus et al. [56]. The authors documented the mineralization of [14C] 2,3,7,8‐TCDD to 14CO2 by *Phanerochaete chrysosporium* within 30 days. *P. chrysosporium* also successfully been used to degrade 2,7‐ DCDD. It should also be mentioned that the biodegradation activity of fungi is not limited to less chlorinated congeners, for example, *P. chrysosporium* is able to remove 34 and 48% of a mixture of PCDD/PCDF congeners containing from 5 to 8 chlorine atoms in the molecule during 7 to 14 days [57].

It was estimated that the highest rate of microbial degradation of pollutants occurs in the plant rhizosphere [58, 59]. Rhizodegradation of organic micropollutants is one of the most effective remediation processes due to existing interactions in the rhizosphere between plant roots, plant exudates, soil and micro‐organisms. Moreover, plants are able to store in their rhizo‐ sphere up to 40% of aminoacids, carbohydrates and other photosynthesis products. This influences on the availability of carbon used by micro‐organisms in the co‐metabolism process. Whipps [60] demonstrated that 1 g of planted soil contains 1012 higher amount of micro‐ organisms in comparison with non‐planted one. Rhizosphere microbiota plays also an intrinsic role in the protection of plants against pathogens and stress caused by too high concentration of pollutants and eases the uptake of biogenic substances by a given plant [61]. The effective‐ ness of rhizosphere biodegradation depends on the ability of micro‐organisms to adapt to a given pollution concentration and effectiveness of colonization of roots [61]. The study of Kuiper et al. [62] demonstrated that naturally occurred rhizosphere biodegradation may be enhanced by an addition of micro‐organisms to the rhizosphere.

The above results are connected with the soil microbial transformation of micropollutants. Many literature data suggest that microbial biodegradation is the critical event determining

Biodegradation is dependent on micro‐organism enzymes which modify toxic compounds into less toxic forms. Biodegradation can be carried as two processes: mineralization, when organic compound uses a sole source of carbon and energy by micro‐organisms, and co‐ metabolism where transformation of given pollutant depends on the presence of other substrate. Products of this process can be further mineralized; otherwise, incomplete degra‐ dation occurs, leading to a formation and accumulation of metabolites more toxic than parent

In this place, there is a need to underline the role of humic acids—major components of soil organic matter which consist of complex polymers of hydroxyphenols, hydroxybenzoic and methoxybenzoic acids and other aromatic structures with linked peptides, amino sugar compounds, fatty acids and possibly other constituents. Hydroquinone/quinone‐type couples are perceived to affect the redox properties of humic acid and to act either as electron transfer mediators or as direct donors of electrons. Thus, the amount of humic acids may determine

The use of anaerobic and aerobic micro‐organisms is the only known process of PCDDs/PCDFs degradation in soil and aquatic systems, leading to a removal of chlorine atoms from the biphenyl molecule and theoretically releasing CO2, chlorine and water. Highly chlorinated congeners have been found to be reductively dechlorinated under anaerobic conditions through a preferential meta‐ and para‐chlorine removal and production of less chlorinated congeners, which can then be used in aerobic transformations. Thus, complete degradation of PCDDs/PCDFs can be achieved by a sequential exposure to anaerobic and aerobic micro‐

The fungi, similarly to bacteria, are also capable to degrade PCDDs/PCDFs in the presence of oxygen using both processes: mineralization and co‐metabolism. The fungi use specific enzymes named lignin peroxidase or manganese peroxidase which enable to oxidize the pollutant molecule. The fungal aerobic biodegradation was first reported by Bumpus et al. [56]. The authors documented the mineralization of [14C] 2,3,7,8‐TCDD to 14CO2 by *Phanerochaete chrysosporium* within 30 days. *P. chrysosporium* also successfully been used to degrade 2,7‐ DCDD. It should also be mentioned that the biodegradation activity of fungi is not limited to less chlorinated congeners, for example, *P. chrysosporium* is able to remove 34 and 48% of a mixture of PCDD/PCDF congeners containing from 5 to 8 chlorine atoms in the molecule

It was estimated that the highest rate of microbial degradation of pollutants occurs in the plant rhizosphere [58, 59]. Rhizodegradation of organic micropollutants is one of the most effective remediation processes due to existing interactions in the rhizosphere between plant roots, plant exudates, soil and micro‐organisms. Moreover, plants are able to store in their rhizo‐ sphere up to 40% of aminoacids, carbohydrates and other photosynthesis products. This influences on the availability of carbon used by micro‐organisms in the co‐metabolism process.

the fate and persistence of PCDDs/PCDFs in the soil [53].

116 Physico-Chemical Wastewater Treatment and Resource Recovery

the microbial dechlorination of PCDDs/PCDFs [54].

substrates.

organisms [55].

during 7 to 14 days [57].

The study with application of plants for phytoremediation/rhizoremediation of soil contami‐ nated with organic compounds showed the decline in the concentration of organochlorine compounds of 30% during 2 years of plant cultivation. At the same time, the unplanted soil demonstrated the reduction of about two times lower [63]. On the basis of 21‐month study, Nedunuri et al. [64] showed the decrease of aromatic compounds concentrations of about 42 and 50% in soil cultivated with fibre flax (*Lolium annual*) and St. Augustine grass (*Stenotaphrum secundatum),* respectively. Other examples showed phytoremediation of soil contaminated with crude oil using combination of grass and fertilizers [64–66]. Despite grasses, the shrubs and trees can be also used as effective phytoremediation tools. The example can be the study of Vervaeke et al. [67] who reported 57% reduction of aromatic compounds and mineral oils during 1.5 years of willow (*Salix viminalis*) cultivation.

With respect to the removal of sludge‐born PCBs, Wyrwicka et al. [68] demonstrated that the use of cucumber (*Cucumis sativus* L. var. Cezar) resulted in a decrease in PCB concentrations by an average of 38.63%. However, the efficiency of PCB removal decreased as the dose of sludge increased in sludge‐treated soil (41.28, 38.39 and 36.22% PCB reduction at doses of 3, 9 and 18 tonnes/ha). Urbaniak et al. [43] demonstrated that the use of other plant from the *Cucurbitaceae* family—*Cucurbita pepo* L. cv Atena Polka—reduced total PCDDs/PCDFs and TEQ concentration by 37 and 68%, respectively, in soil amended with sewage sludge. The comparative study of the use of *Cucurbita pepo* L. cv Atena Polka (zucchini) and *Cucumis sativus* L. var. Cezar (cucumber) showed that zucchini was more efficient in sludge‐born PCDDs/ PCDFs removal, while cucumber demonstrated higher efficiency in soil phytotoxicity allevi‐ ation [44]. Presented studies demonstrate that cultivation of the plants from the *Cucurbitaceae* family plays a positive role in reducing the PCDDs/PCDFS in soil amended with sewage sludge.

The above data confirm the positive role of plant‐bacteria systems in the removal of PCDDs/ PCDFs and PCBs from soil contaminated through agricultural utilization of sewage sludge.

## **6. Impact of PCDDs/PCDFs on plant growth and plant metabolism**

There is limited data on the impact of PCDDs/PCDFs on the plant growth and biomass production. The literature on this issue mainly comes from studies on the effects of sewage sludge on plant growth and metabolism [69–71]. Application of sewage sludge as soil organic amendment and as a source of macronutrients and micronutrients can contribute not only to restore the soil cover and vegetation on devastated land [72] but also can be used in the organization and maintenance of green areas in cities and recreational facilities. An important aspect of the use of sewage sludge is to improve soil fertility, of low quality class, which can be used, for example, for energy crops (biomass extraction). The addition of sewage sludge may have beneficial effects on plants and soil expressing itself by improving the physico‐ chemical properties of the soil, increased nutrient content for plants, increased production of plant biomass, and increased activity of soil enzymes and soil micro‐organisms [73]. However, the presence of pollutants in sewage sludge may have a negative impact on the growth and development of plants. The content of heavy metals, toxic organic compounds including PCDDs/PCDFs and microbiological contaminants may contribute to the occurrence of secondary oxidative stress [74–76]. The occurrence of environmental stresses can lead to an imbalance in cellular redox state and predominance of oxidation reaction over reduction reactions. The reactive oxygen species (ROS) are highly reactive and toxic and can damage important from the biological point of view molecules such as nucleic acids, proteins and lipids [77]. It is well known that oxidative stress is a common plant reaction to numerous biotic [78] and abiotic stresses including drought [79], high salinity [80], temperature extremes [81, 82], anoxia [83], mineral nutrients' deficiencies and metal toxicity [84], increased UV‐B radiation [85], gaseous pollutants [86], acid rain [87] and PCDD/PCDFs [88].

The enzymatic and non‐enzymatic antioxidant systems present in the plant tissues prevent the accumulation of ROS caused by stress factors. The enzymatic free radical scavengers include, among others, superoxide dismutase (SOD), catalase (CAT) and peroxidases; ascorbate peroxidase (APx), glutathione peroxidase (GSH‐Px), phenolic peroxidase (POx). Non‐ enzymatic, low molecular weight antioxidants mainly include ascorbic acid, glutathione, carotenoids, flavonoids, α‐tocopherol and the phenolic compounds [89, 90].

Currently, there is little literature concerning the impact of PCDDs/PCDFs and PCBs on plant antioxidative system, and usually, information are related to multistress associated with the presence of organic pollutants and heavy metals. The plants belonging to the *Cucurbitaceae* family are known to accumulate high levels of PCDDs/PCDFs and PCBs compared with other plant species. However, the studies showed also that plant belonging to cucurbits: zucchini and cucumber, activate the antioxidative system and detoxification mechanisms as an effect of application of sewage sludge containing high levels of POPs including PCDDs/PCDFs and PCB [68, 91]. Obtained results indicate that signs of sewage sludge toxicity were greater in zucchini than in cucumber plants. Visible symptoms of leaf blade damage after sewage sludge application occurred only on the zucchini plants. Activity of peroxidases such as ascorbate peroxidase (APx) and guaiacol peroxidase (POx) increased in zucchini plants significantly with increasing of sewage sludge dose, but they decreased in cucumber plants. Moreover, both in zucchini and cucumber plants, the relationship between peroxidases activity and catalase (CAT) activity was inverse. Activity of detoxifying enzyme—glutathione S‐transferase (GST) —increased progressively with the sludge concentration in both the zucchini and cucumber leaves. Moreover, the increase in GST activity was greater in zucchini plants and was visible at the lowest dose used. Concentration of α‐tocopherol, a lipophilic antioxidant, increased with sewage sludge dose in both investigated species.

Other research focused on the influence of light soil fertilization using sewage sludge on soil toxicity showed its negative impact on growth and development of three plant species *Lepidium sativum*, *Sorgo saccharatum* and *Sinapis alba* [92].

restore the soil cover and vegetation on devastated land [72] but also can be used in the organization and maintenance of green areas in cities and recreational facilities. An important aspect of the use of sewage sludge is to improve soil fertility, of low quality class, which can be used, for example, for energy crops (biomass extraction). The addition of sewage sludge may have beneficial effects on plants and soil expressing itself by improving the physico‐ chemical properties of the soil, increased nutrient content for plants, increased production of plant biomass, and increased activity of soil enzymes and soil micro‐organisms [73]. However, the presence of pollutants in sewage sludge may have a negative impact on the growth and development of plants. The content of heavy metals, toxic organic compounds including PCDDs/PCDFs and microbiological contaminants may contribute to the occurrence of secondary oxidative stress [74–76]. The occurrence of environmental stresses can lead to an imbalance in cellular redox state and predominance of oxidation reaction over reduction reactions. The reactive oxygen species (ROS) are highly reactive and toxic and can damage important from the biological point of view molecules such as nucleic acids, proteins and lipids [77]. It is well known that oxidative stress is a common plant reaction to numerous biotic [78] and abiotic stresses including drought [79], high salinity [80], temperature extremes [81, 82], anoxia [83], mineral nutrients' deficiencies and metal toxicity [84], increased UV‐B

118 Physico-Chemical Wastewater Treatment and Resource Recovery

radiation [85], gaseous pollutants [86], acid rain [87] and PCDD/PCDFs [88].

carotenoids, flavonoids, α‐tocopherol and the phenolic compounds [89, 90].

sewage sludge dose in both investigated species.

The enzymatic and non‐enzymatic antioxidant systems present in the plant tissues prevent the accumulation of ROS caused by stress factors. The enzymatic free radical scavengers include, among others, superoxide dismutase (SOD), catalase (CAT) and peroxidases; ascorbate peroxidase (APx), glutathione peroxidase (GSH‐Px), phenolic peroxidase (POx). Non‐ enzymatic, low molecular weight antioxidants mainly include ascorbic acid, glutathione,

Currently, there is little literature concerning the impact of PCDDs/PCDFs and PCBs on plant antioxidative system, and usually, information are related to multistress associated with the presence of organic pollutants and heavy metals. The plants belonging to the *Cucurbitaceae* family are known to accumulate high levels of PCDDs/PCDFs and PCBs compared with other plant species. However, the studies showed also that plant belonging to cucurbits: zucchini and cucumber, activate the antioxidative system and detoxification mechanisms as an effect of application of sewage sludge containing high levels of POPs including PCDDs/PCDFs and PCB [68, 91]. Obtained results indicate that signs of sewage sludge toxicity were greater in zucchini than in cucumber plants. Visible symptoms of leaf blade damage after sewage sludge application occurred only on the zucchini plants. Activity of peroxidases such as ascorbate peroxidase (APx) and guaiacol peroxidase (POx) increased in zucchini plants significantly with increasing of sewage sludge dose, but they decreased in cucumber plants. Moreover, both in zucchini and cucumber plants, the relationship between peroxidases activity and catalase (CAT) activity was inverse. Activity of detoxifying enzyme—glutathione S‐transferase (GST) —increased progressively with the sludge concentration in both the zucchini and cucumber leaves. Moreover, the increase in GST activity was greater in zucchini plants and was visible at the lowest dose used. Concentration of α‐tocopherol, a lipophilic antioxidant, increased with

## **7. Accumulation and translocation of PCDDs/PCDFs and PCBs in plant tissue**

Plants are the organisms, which are the first stage in the food chain. Widely distributed at low concentration in the environment, extraordinary toxic PCDDs/PCDFs and PCBs have the ability to bioaccumulate in the food chain. For these reasons, accumulation of these compounds by plants is an important step for the transfer of PCDDs/PCDFs and PCBs into the higher trophic levels and biomagnification. Understanding the mechanisms of uptake and transloca‐ tion of PCDDs/PCDFs and PCBs allows to control the risk of unexpected contamination of important vegetative plants. On the other hand, this knowledge can be used as a tool for selecting plants that have high phytoremediative potential [59, 93].

The accumulation of persistent pollutants such as chlorinated pesticides, chlorobenzenes, PCBs and PAH as well as PCDDs/PCDFs in vegetation has been demonstrated in several investigations carried out in the 1980s of the twentieth century [94–96]. In recent years, our understanding of the uptake of PCDDs/PCDFs by plants increased considerably [97, 98] but the pathway by which above‐mentioned organic pollutants enter to the plant tissues still remain under discussion. Early evidence suggested that organic compounds were unlikely to be taken up from soil and translocated within the plants due to their hydrophobicity [99]. Nevertheless, according to many publications, the absorption from soil vapour may be the major pathway by which PCDDs/PCDFs from soil enter into the aerial plant tissues [100–102]. Other studies also evidence that dry gaseous deposition is the dominant pathway of PCDDs/ PCDFs in plant tissue, such as lettuce, potato, apple, pear, rice, pea and oilseed rape [103– 108]. However, more recent studies have shown that some species of plant have the ability to mobilize and accumulate significant concentrations of several organic compounds from soil. Generally, it is estimated that there are several pathways of PCDDs/PCDFs and PCBs accu‐ mulation in plants: (1) adsorption to the root surface, (2) root uptake trough absorption from soil vapour or water phases of soil and translocation to upper plant organs, (3) contamination of the foliage and fruits by soil particles which develop in contact with or in close proximity to the ground, (4) absorption of volatilized from soil PCDDs/PCDFs and PCBs by aerial plants parts, (5) atmospheric deposition of airborne PCDDs/PCDFs and PCBs for both gas and particle phase [97, 109].

Uptake and distribution of PCDDs/PCDFs and PCBs are a function of chemical and physical properties of particular pollutant, such as hydrophobicity, water solubility and vapour pressure, as well as environmental conditions, such as temperature, pH, organic carbon content of the soil and plant species [110].

The most important property which determines the possibility of absorption of various compounds from the soil by roots is hydrophobicity. Usually, it is expressed as the 1‐octanol/ water partition coefficients (Kow) and extends over a wide range for different organic com‐ pounds [111]. Kow values vary over several orders of magnitude and are expressed as log Kow. More hydrophobic substances that having a higher log Kow value are sorbed more strongly to soil organic particles. In consequences, if the log Kow value of the compound is around 2, the compound could be easily absorbed by plants, whereas if the log Kow value is over 5, the compound is hardly absorbed [112, 113]. The value closely related to the hydrophobicity is solubility. Water solubility describes the amount of a chemical which can dissolve in a known quantity of water. The solubility of the chemical compound is dependent on temperature and is pH‐dependent. Another feature describing the tendency of the substance to move from the aqueous phase to the gas phase is Henry's constant (Hc). This parameter can be useful in predicting the ability of chemical to volatilize from soil, water or plant surfaces into the atmosphere. Although chemical properties are important predictors of uptake, the physiology and composition of the plant root itself are also a significant influence. One explanation for such difference in uptake potential is the varying types and amounts of lipids in roots cells.

The uptake of organic chemicals by plants is also influenced by soil properties. Transfer of organic pollutants from soil to plant roots might be carried out by the uptake of soil pore water during plant transpiration. Non‐ionized organic pollutants, which usually are lipophilic, are principally sorbed or bound to several components in soil including clays, iron oxides and onto the organic fraction of the soil's solid phase. The latter usually exerts the strongest influence on the organic chemicals pore water concentration [114]. Similarly, compounds with a high log Kow associate with particulates in the wastewater and become incorporated into sewage sludge during sedimentation, and thus, substances with a log Kow of <2.0 appear less frequently in sewage sludge. It is considered that with the increase in the organic matter content of a soil, the proportion of the chemical in the pore water decreases, and consequently, its uptake by plant also decreases. Moreover, it should be noted that the increase of the amount of organic carbon fraction reduces the optimum of Kow for uptake into plants.

Deposition of non‐ionic organic compounds on leaves and its sorption at the leaf surface or rapid movement into the leaf depends on diffusion through the cuticle or stomata. The concentration of all these compounds on leaves is mainly due to adsorption from the gaseous phase. Accumulation of PCDDs/PCDFs in above‐ground plant parts mainly results from atmospheric deposition in the gaseous state alone. The contribution of particle‐bound depo‐ sition may be, despite areas of extreme particle loading, of secondary importance [94].

The pathways of PCDDs/PCDFs accumulation in rice plants were carefully examined by Uegaki et al., [105] who estimated that dioxins were not absorbed from the soil by growing plants, but its uptake from atmosphere has the greatest importance. They reported that dioxin levels in rice plants were strongly influenced by soil adhesion but only at the early growth stage of brown rice plants grown in three different soils: dioxin‐contaminated soil, paddy soil and upland soil. In the later stage of growth, over the time of experiment, predominant influence on dioxin level in rice leaf and steam was attributed to concentrations of these compounds in atmospheric gas phase. This remarks remain in agree with results other investigations which indicate that approximately 70% of 2,3,7,8‐TCDD added to the growth solution, but only 3% of 2‐chlorobiphenyl was adsorbed by the roots of soybean and corn [108], and the most important mechanism of foliar contamination is connected with volatility of 2,3,7,8‐TCDD from the growth solution.

Taking above relations into consideration, the hydrophobic nature of PCDDs/PCDFs and PCBs (log Kow values between 4.8 and 10.5) and their consequent strong adsorption to soil particles renders them to largely immobile and generally unavailable to plants [115, 116]. The majority of available evidence nevertheless suggests that the adsorption or absorption of PCDDs/PCDFs and PCBs into plant roots and their subsequent translocation into other parts of the plant structure is minimal. However, the notable exceptions are several plants of the genus *Cucur‐ bita*, which readily take up PCDDs/PCDFs from soil and translocate them to leaves and fruits [97, 117]. It was also found that *Cucurbita* plants (e.g. zucchini, pumpkin and squash) can phytoextract polychlorinated biphenyls (PCBs) [118, 119], p,p′ DDE [120, 121] and chlordane [122, 123] from soil and translocate some quantities to aerial tissues, as well as it was found that there is remarkable diversity in the uptake and transportation of persistent organic pollutants (POPs) among subspecies [119, 121, 123]. In case of willow, study of Oleszczuk and Baran [124] demonstrated the uptake of 16 Polycyclic aromatic hydrocarbons (PAHs) by willow from the soil amendment with the contaminated sewage sludge. The authors showed that soil total (PAHs) content decreased significantly within the first half year, followed by minimal changes over the subsequent 3 years of treatment. The authors showed that the total content of (PAHs) in control ranged between 3.6–7.3 μg/kg in shoots and 13–27 μg/kg in leaves, whereas treated plant demonstrated higher concentrations ranged from 5.5 to 17.6 and 13.5 to 33.8 μg/kg in shoots and leaves, respectively.

## **8. Conclusions**

water partition coefficients (Kow) and extends over a wide range for different organic com‐ pounds [111]. Kow values vary over several orders of magnitude and are expressed as log Kow. More hydrophobic substances that having a higher log Kow value are sorbed more strongly to soil organic particles. In consequences, if the log Kow value of the compound is around 2, the compound could be easily absorbed by plants, whereas if the log Kow value is over 5, the compound is hardly absorbed [112, 113]. The value closely related to the hydrophobicity is solubility. Water solubility describes the amount of a chemical which can dissolve in a known quantity of water. The solubility of the chemical compound is dependent on temperature and is pH‐dependent. Another feature describing the tendency of the substance to move from the aqueous phase to the gas phase is Henry's constant (Hc). This parameter can be useful in predicting the ability of chemical to volatilize from soil, water or plant surfaces into the atmosphere. Although chemical properties are important predictors of uptake, the physiology and composition of the plant root itself are also a significant influence. One explanation for such difference in uptake potential is the varying types and amounts of lipids in roots cells.

120 Physico-Chemical Wastewater Treatment and Resource Recovery

The uptake of organic chemicals by plants is also influenced by soil properties. Transfer of organic pollutants from soil to plant roots might be carried out by the uptake of soil pore water during plant transpiration. Non‐ionized organic pollutants, which usually are lipophilic, are principally sorbed or bound to several components in soil including clays, iron oxides and onto the organic fraction of the soil's solid phase. The latter usually exerts the strongest influence on the organic chemicals pore water concentration [114]. Similarly, compounds with a high log Kow associate with particulates in the wastewater and become incorporated into sewage sludge during sedimentation, and thus, substances with a log Kow of <2.0 appear less frequently in sewage sludge. It is considered that with the increase in the organic matter content of a soil, the proportion of the chemical in the pore water decreases, and consequently, its uptake by plant also decreases. Moreover, it should be noted that the increase of the amount

Deposition of non‐ionic organic compounds on leaves and its sorption at the leaf surface or rapid movement into the leaf depends on diffusion through the cuticle or stomata. The concentration of all these compounds on leaves is mainly due to adsorption from the gaseous phase. Accumulation of PCDDs/PCDFs in above‐ground plant parts mainly results from atmospheric deposition in the gaseous state alone. The contribution of particle‐bound depo‐ sition may be, despite areas of extreme particle loading, of secondary importance [94].

The pathways of PCDDs/PCDFs accumulation in rice plants were carefully examined by Uegaki et al., [105] who estimated that dioxins were not absorbed from the soil by growing plants, but its uptake from atmosphere has the greatest importance. They reported that dioxin levels in rice plants were strongly influenced by soil adhesion but only at the early growth stage of brown rice plants grown in three different soils: dioxin‐contaminated soil, paddy soil and upland soil. In the later stage of growth, over the time of experiment, predominant influence on dioxin level in rice leaf and steam was attributed to concentrations of these compounds in atmospheric gas phase. This remarks remain in agree with results other investigations which indicate that approximately 70% of 2,3,7,8‐TCDD added to the growth solution, but only 3% of 2‐chlorobiphenyl was adsorbed by the roots of soybean and corn [108],

of organic carbon fraction reduces the optimum of Kow for uptake into plants.

PCDDs/PCDFs and PCBs pose one of the most challenging problems in environmental science and technology. Their discharges via insufficiently treated wastewater are responsible for their occurrence in river ecosystems, both water and bottom sediments. The administration of sewage sludge, as end products of purification processes, additionally generates problems with the occurrence of PCDDs/PCDFs and PCBs in the environment and creates risk for ecosystem functioning and human well‐being. Despite the above, the available literature data concerning PCDDs/PCDFs and PCBs removal from the environment using range of bio‐ and phytoremediation technologies demonstrate a promising tool towards safe and effective elimination of the compounds and in this way improvement of ecosystems quality.

## **Acknowledgements**

The research is financed within the project 'Impact of sludge originated PCDDs/PCDFs on soil contamination and Salix sp. Metabolism' (UMO—2013/09/D/ST10/04043) and the project funded under the Ministry of Science and Higher Education programme under the name 'Iuventus Plus' for the years 2015–2017: project no. IP2014 049273.

## **Author details**

Magdalena Urbaniak1,2\* and Anna Wyrwicka3

\*Address all correspondence to: m.urbaniak@unesco.lodz.pl

1 European Regional Centre for Ecohydrology of the Polish Academy of Sciences, Łódz, Poland

2 Department of Applied Ecology, Faculty of Biology and Environmental Protection, Univer‐ sity of Łódz, Łódz, Poland

3 Department of Plant Physiology and Biochemistry, Faculty of Biology and Environmental Protection, University of Łódz, Łódz, Poland


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**Author details**

sity of Łódz, Łódz, Poland

Poland

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**Resource Recovery and Management**

**Provisional chapter**

## **Phosphorus Recovery by Struvite Crystallization from Livestock Wastewater and Reuse as Fertilizer: A Review Livestock Wastewater and Reuse as Fertilizer: A Review**

**Phosphorus Recovery by Struvite Crystallization from** 

Tao Zhang, Rongfeng Jiang and Yaxin Deng Tao Zhang, Rongfeng Jiang and Yaxin Deng Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65692

#### **Abstract**

In China, the intensive livestock farming produces massive livestock wastewater with high concentration of phosphorus. Discharge of these compounds to surface water not only causes water eutrophication but also wastes phosphorus resources for plant growth. Therefore, it's necessary combining the removal of phosphorus from livestock wastewater with its recovery and reuse as fertilizer. As a valuable slow-release mineral fertilizer, struvite crystallization has become a focus in phosphorus recovery. In this chapter, struvite crystallization mechanism, reaction factors, crystallizers, and the applications of struvite as fertilizer are discussed. Two steps of nucleation and crystal growth for struvite crystallization from generation to growth are introduced. The reaction factors, including molar ratio of magnesium and phosphate, solution pH, coexisting substances, and seeding assist, of struvite crystallization are summarized. Several innovate types of crystallizer, which relate to the shape and size of harvest struvite to realize the phosphorus recycling, are demonstrated. Due to the influence of toxic or harmful impurities in struvite on its reuse as fertilizer, the environmental risk evaluation of struvite application is introduced. In conclusion, struvite crystallization is a promising tool for recovering phosphorus from livestock wastewater.

**Keywords:** phosphorus, struvite, livestock wastewater, fertilizer, review

## **1. Introduction**

Phosphorus is a key factor causing water eutrophication, on the other hand, it is also a nonrecyclable, nonrenewable, and quite valuable resource. According to the Mineral Commodity Summaries 2015 [1] from the United States Geological Survey (USGS), the reserve of phosphate

rock in China is 3.7 billion tons in 2014, which is in second place in the world. However, with a total of 43–48% of the world's phosphate rock production over the last 3 years [2], the phosphate rock might run out in less than 40 years. So phosphate rock has been one of the 20 minerals that could not meet the demand of the national economy development after 2010 as reported by the Ministry of Land and Resources in China.

On the other side, the intensive livestock farming is a pillar industry in agricultural economy and an important way to increase rural incomes in China [3]. However, it usually produces large amount of livestock wastewater containing high concentration of phosphorus [4]. If this wastewater was not treated reasonably, it would not only lead to the pollution of water eutrophication, but also waste nonrenewable resources and would become one of the major contributors to phosphorus loss [5]. According to the first national sources of pollution survey [6] in China in 2008, nonpoint source pollution in agriculture is a major cause of eutrophication. It accounts for 34.24% of the total phosphorus emission amount, including the livestock and poultry industry. Therefore, it is quite valuable to combine nutrient recycling with environmental pollution control to recover losing phosphorus from livestock wastewater [7].

Numerous phosphorus recovery technologies have been developed, such as biological phosphorus removal, chemical precipitation, electrolysis, adsorption, and crystallization. Biological phosphorus removal utilizes polyphosphate-accumulating organisms to capture phosphorus in their cells. However, this method is limited by the lack of carbon sources and the difficulty of culturing pure bacteria [8]. Chemical precipitation process may consume expensive chemicals and produce large amounts of chemical sludge [9]. Electrolysis is restricted by the small capacity of handling wastewater and the frequent renewal of electrodes [10, 11]. Recovering phosphorus from wastewater using chemical adsorbents is expensive, so cheaper and more efficient adsorbents are necessary for research [12].

Recovering phosphorus by crystallization, by contrast, is a more economical and efficient way. As long as the crystallization conditions are suitable, the struvite crystal would be generated just by adding magnesium (Mg2+) in the raw wastewater which has high concentrations of HnPO4 n−3 and NH4 + -N. This technology can remove nitrogen at the same time and its production can be used as fertilizer. So it had been studied in many kinds of wastewater, such as multiple wastewater [13], industrial wastewater [14, 15], municipal landfill leachate [16], biogas slurry [17], and effluent of sewage sludge [18], and livestock wastewater is no exception.

## **2. Struvite characteristic**

Magnesium ammonium phosphate, also known as struvite, is a white crystal generated in neural or mild alkali condition, for which the chemical formula is Mg(NH4 )PO4 ⋅6H2 O. Struvite consists of one molecule of magnesium(Mg2+), one molecule of ammonium (NH4 + ), one molecule of phosphate (PO4 3−), and six molecules of water (H2 O), whose relative molar mass is 245.43 g/mol. It is only slightly soluble in water but soluble in acid solution [19]. Struvite is a light crystal with low relative density of 1.65–1.7. It is not easy to be rush off by rainfall [20]. Pure struvite belongs to orthorhombic crystals which consists of regular PO4 3− octahedron, distorted Mg(H2 O)6 2+ octahedron, and groups of NH4 + connected by hydrogen bonding [21], but shows rod-like structure [22] or irregular structure [23] sometimes (**Figure 1**). And struvite of rod-like structure is of low purity, because of the coprecipitation with foreign ions.

rock in China is 3.7 billion tons in 2014, which is in second place in the world. However, with a total of 43–48% of the world's phosphate rock production over the last 3 years [2], the phosphate rock might run out in less than 40 years. So phosphate rock has been one of the 20 minerals that could not meet the demand of the national economy development after 2010 as

On the other side, the intensive livestock farming is a pillar industry in agricultural economy and an important way to increase rural incomes in China [3]. However, it usually produces large amount of livestock wastewater containing high concentration of phosphorus [4]. If this wastewater was not treated reasonably, it would not only lead to the pollution of water eutrophication, but also waste nonrenewable resources and would become one of the major contributors to phosphorus loss [5]. According to the first national sources of pollution survey [6] in China in 2008, nonpoint source pollution in agriculture is a major cause of eutrophication. It accounts for 34.24% of the total phosphorus emission amount, including the livestock and poultry industry. Therefore, it is quite valuable to combine nutrient recycling with environmental pollution control to recover losing phosphorus from livestock

Numerous phosphorus recovery technologies have been developed, such as biological phosphorus removal, chemical precipitation, electrolysis, adsorption, and crystallization. Biological phosphorus removal utilizes polyphosphate-accumulating organisms to capture phosphorus in their cells. However, this method is limited by the lack of carbon sources and the difficulty of culturing pure bacteria [8]. Chemical precipitation process may consume expensive chemicals and produce large amounts of chemical sludge [9]. Electrolysis is restricted by the small capacity of handling wastewater and the frequent renewal of electrodes [10, 11]. Recovering phosphorus from wastewater using chemical adsorbents is expensive, so cheaper and more

Recovering phosphorus by crystallization, by contrast, is a more economical and efficient way. As long as the crystallization conditions are suitable, the struvite crystal would be generated just by adding magnesium (Mg2+) in the raw wastewater which has high concen-

its production can be used as fertilizer. So it had been studied in many kinds of wastewater, such as multiple wastewater [13], industrial wastewater [14, 15], municipal landfill leachate [16], biogas slurry [17], and effluent of sewage sludge [18], and livestock wastewater is no

Magnesium ammonium phosphate, also known as struvite, is a white crystal generated

Struvite consists of one molecule of magnesium(Mg2+), one molecule of ammonium (NH4

in neural or mild alkali condition, for which the chemical formula is Mg(NH4


)PO4 ⋅6H2 O.

> + ),

reported by the Ministry of Land and Resources in China.

136 Physico-Chemical Wastewater Treatment and Resource Recovery

efficient adsorbents are necessary for research [12].

n−3 and NH4

+

wastewater [7].

trations of HnPO4

**2. Struvite characteristic**

exception.

Actually, struvite had been widely studied as early as 1937 for the congestion in the pipes of the sludge anaerobic digester [24]. The general struvite forming reaction equation is shown below:

$$\rm NH\_4^\* + Mg^{2+} + PO\_4^{3-} + 6\, H\_2O \to MgNH\_4PO\_4 \cdot 6\, H\_2O \tag{1}$$

$$\mathrm{NH\_4^+ + Mg^{2+} + HPO\_4^{2-} + 6H\_2O \to MgNH\_4PO\_4 \cdot 6H\_2O + H^+} \tag{2}$$

$$\mathrm{NH\_4^+ + Mg^{2+} + H\_2PO\_4^- + 6H\_2O \to \cdot MgNH\_4PO\_4 \cdot 6H\_2O + 2H^+} \tag{3}$$

When Mg2+, NH4 + , and HnPO4 n−3 (*n* = 0, 1, or 2) exist in the solution and the product of their concentrations are bigger than the solubility product constant (Ksp) of struvite, the crystal would be generated spontaneously. And the calculation formula of struvite's Ksp is shown below:

$$\mathbf{K}\_{sp} = \left[ \mathbf{M} \mathbf{g}^{2\*} \right] \cdot \left[ \mathbf{N} \mathbf{H}\_4^\* \right] \cdot \left[ \mathbf{P} \mathbf{O}\_4^{3\*} \right] \tag{4}$$

**Figure 1.** The SEM figure of magnesium ammonium phosphate crystal.

where [Mg2+], [NH4 + ], and [PO4 3−] are concentrations of Mg2+, NH4 + , and PO4 3− in the solution, respectively. As the molar ratio of Mg2+, NH4 + , and PO4 3− is 1:1:1 in struvite, so C\* is used to present the same concentration of these three ions, which means C\* = [Mg2+] = [NH4 + ] = [PO4 3−]. So the calculation formula of struvite's Ksp also can be shown as:

$$\mathbf{K}\_{sp} = \text{ (C\text{\textquotedblleft})}{}^{3} \tag{5}$$

Snoeyink et al. [25] got the Ksp of struvite is 10−12.6 as early as 1980. Ohlinger et al. [26] corrected it to 10−13.26 in 1999. And then Bhuiyan et al. [27] corrected it again to 10−13.36 in 2007, which is widely used now. However, Ksp of struvite is hard to get in the real wastewater for the negative impact of the soluble coexisting ions. Therefore, in the estimation of the saturability of the real wastewater, ionic activity coefficient (Kso) is more widely useful than Ksp. Considering the impact of ionic strength (I) and the ionic activity (Ai) in estimating the Kso, the value of Kso is bigger than Ksp. And the calculation formula of struvite's Kso is shown below:

$$\mathbf{K}\_{\rm so} = \boldsymbol{\alpha}\_{\rm Mg^{\circ}} \times \boldsymbol{\alpha}\_{\rm NH\_{i}^{+}} \times \boldsymbol{\alpha}\_{\rm PO\_{i}^{\circ}} \tag{6}$$

$$a\_i = \gamma\_i \{ \mathbb{C} \} \tag{7}$$

where α*<sup>i</sup>* presents the ionic activity (Ai), γ*<sup>i</sup>* presents the activity coefficient of the ionic strength (I), and [*Ci* ] presents the concentration of the ion. Only when the value of γ*<sup>i</sup>* is 1, Ksp is able to represent Kso. Therefore, it is necessary to eliminate the interruptions of the soluble coexisting ions (like Ca2+, CO3 2−, and SO4 2−) and clear of the ionic activities of Mg2+, NH4 + , and PO4 3− in the specific pH condition. **Table 1** shows the ionization equations and pKa value in magnesium ammonium phosphate solution at 25°C, which is helpful to estimate the distribution of these ions and predict the probability to generate struvite under such environment of solution.


**Table 1.** The ionization equations and pKa value in magnesium ammonium phosphate solution at 25° [30, 31].

## **3. Mechanism of struvite crystallization**

where [Mg2+], [NH4

below:

where α*<sup>i</sup>*

of solution.

PO4

strength (I), and [*Ci*

1 NH4

2 H3

3 H2

4 HPO4

5 MgNH4

7 MgH2

9 MgPO4

8 MgHPO4 ⇌ HPO4

coexisting ions (like Ca2+, CO3

+

], and [PO4

*Kso* = *α*Mg 2+ × *α*NH 4

*α<sup>i</sup>* = *γ<sup>i</sup>* [ *Ci*

<sup>+</sup> ⇌ NH3

PO4 −

<sup>−</sup> ⇌ HPO4

2− ⇌ PO4

PO4 ⋅6H2

PO4 <sup>+</sup> ⇌ H2 PO4 −

<sup>−</sup> ⇌ PO4

PO4 ⇌ H2

PO4

presents the ionic activity (Ai), γ*<sup>i</sup>*

2−, and SO4

**No. Ionization equation pKa**

O ⇌ Mg2+ + NH4

6 MgOH+ ⇌ Mg2+ + OH<sup>−</sup> 2.56

So the calculation formula of struvite's Ksp also can be shown as:

respectively. As the molar ratio of Mg2+, NH4

138 Physico-Chemical Wastewater Treatment and Resource Recovery

3−] are concentrations of Mg2+, NH4

, and PO4

+ × *α*PO 4

] presents the concentration of the ion. Only when the value of γ*<sup>i</sup>*

(aq) + H+ 9.26

+ H+ 2.12

2− + H+ 7.20

3− + H+ 12.36

+ Mg2+ 0.45

2− + Mg2+ 2.91

3− + Mg2+ 4.80

3− + 6H2

<sup>+</sup> + PO4

**Table 1.** The ionization equations and pKa value in magnesium ammonium phosphate solution at 25° [30, 31].

is able to represent Kso. Therefore, it is necessary to eliminate the interruptions of the soluble

3− in the specific pH condition. **Table 1** shows the ionization equations and pKa value in magnesium ammonium phosphate solution at 25°C, which is helpful to estimate the distribution of these ions and predict the probability to generate struvite under such environment

+

K sp = ( C \* ) 3 (5)

Snoeyink et al. [25] got the Ksp of struvite is 10−12.6 as early as 1980. Ohlinger et al. [26] corrected it to 10−13.26 in 1999. And then Bhuiyan et al. [27] corrected it again to 10−13.36 in 2007, which is widely used now. However, Ksp of struvite is hard to get in the real wastewater for the negative impact of the soluble coexisting ions. Therefore, in the estimation of the saturability of the real wastewater, ionic activity coefficient (Kso) is more widely useful than Ksp. Considering the impact of ionic strength (I) and the ionic activity (Ai) in estimating the Kso, the value of Kso is bigger than Ksp. And the calculation formula of struvite's Kso is shown

present the same concentration of these three ions, which means C\* = [Mg2+] = [NH4

+

, and PO4

3− is 1:1:1 in struvite, so C\* is used to

<sup>3</sup><sup>−</sup> (6)

] (7)

presents the activity coefficient of the ionic

2−) and clear of the ionic activities of Mg2+, NH4

O 12.70

3− in the solution,

3−].

is 1, Ksp

+ , and

+ ] = [PO4 Nucleation and crystal growth are two classical steps in the process of struvite crystallization from generation to development. As shown in **Figure 2**, nucleation is the first step of the struvite crystallization. When Mg2+, NH4 + , and PO4 3− meet under the proper pH value, nucleation occurs. And the nucleation time is the time required to form a saturated solution to the beginning of the nucleation. It is mainly influenced by the pH of solution, mixing energy, coexisting ions, and saturation index (SI). The ion activity affected by the value of pH significantly leads to differentiation of combine speed of free ions [28]. Weak ion activity means slow combine speed and longer nucleation time indirectly. When the rate of struvite nucleation and growth is greater than or equal to the rate of mixing magnesium to the solution, there needs additional mixing energy. Kim et al. [15] emphasized that mixing energy could influence the quantity and size of struvite strongly. However, the greatest impaction of struvite nucleation is saturation index (SI) of solution which decides the development of crystal to homogeneous or heterogeneous directly [26]. SI is used to describe the saturation state of the reaction system of struvite. And the SI calculation of struvite is shown as follows:

$$SI = \log \frac{IAP}{K\_p} \tag{8}$$

where *IAP* and *K*sp represent ionic activity product and the thermodynamic solubility product of struvite, respectively [29]. The homogeneous crystallization that we want happens on metastable region in the solution. In this region, nucleation is not spontaneous, which differentiated between the process of crystallization and precipitation, and avoids the occurrence of undesirable spontaneous nucleation to a great extent [28]. However, mestastable state of solution is very difficult to control. Therefore, SI, as the indicator for metastable state, is very important. Bonurophoulos et al. [30] found that the threshold between homogeneous and heterogeneous precipitation is the condition where SI ≈ 2.0 and the nucleation rate of 1 nucleus/ (cm3 ⋅s). When the SI is less than 1.716, the struvite crystals are in heterogeneous precipitation and vice versa. Bhuiyan et al. [31] and Mehta et al. [32] also got the threshold at SI = 1.83 and SI = 1.7 at the special nucleation rate, respectively. In addition, Durrant et al. [33] emphasized the great influence of SI on the shape of struvite as early as 1999. And it also has a SI threshold between rhombic structure and rod-like structure of struvite.

**Figure 2. The crystal nucleation, growth and aggregation mechanism of magnesium ammonium phosphate.**

After the crystal nucleus generates, the ions in the solution used to form the crystal begin to deposit on the crystal nucleus and the nucleus grow to the settling particles. During that time, there are two trends for the development of particles. One is orientation growth, which means the ions sequence in the crystal is arranged according to a certain lattice. The other one is nonorientation growth, which means these ions are too late to arrange in order. It is the fast growth rate that causes disorder. And two types of the crystal growth mechanisms lead to different trends. One is the integration mechanism, and the other is the mass transfer mechanism. The former is the integration of solute molecules into the surface; the latter is the transfer (by diffusion or convection) of solute molecules from the bulk solution to the crystal surface. When the effect of mass transfer is greater than the effect of integration, the crystal growth mainly depends on the diffusion effect and the growth of crystal would be orientable. However, if the effect of integration is greater, the integration on the surface of solute decides the crystal growth. And the relative sizes of the nucleation rate, aggregation rate, and directional array rate also decide the trend of crystal growth, which can be changed by precipitation conditions [31]. Abe et al. [34] showed that the growth rate of struvite was very slow. In the high concentration of phosphate (greater than 200 mg/L), the daily growth rate of struvite was 0.173 mm. In the low concentration of phosphate (30–100 mg/L), the daily growth rate of struvite was 0.061 mm. Therefore, increasing crystal growth rate and crystal size of struvite is not only beneficial to further removing phosphorus from livestock wastewater, but also to recycling phosphorus with a bigger size struvite. There is a metastable zone in industrial crystallization to make the crystal bigger and more even. The metastable zone is defined as a region bounded by the solubility curve in which the solution is supersaturated but the spontaneous nucleation cannot occur in such a short time [35]. In the metastable zone, the solute condenses on the nucleus as constantly as possible. As we known, the process of struvite constant growth is also the further recovery of phosphorus from livestock wastewater. So it is meaningful to study the metastable zone of struvite for the industrialized application.

#### **4. The factors influencing struvite crystallization**

#### **4.1. Molar ratio of P and Mg**

Generally speaking, livestock wastewater is rich in ammonium and phosphorus. So it is needed to add extra magnesium to form struvite. Therefore, the addition amount of magnesium affects the solubility product constant (Ksp) directly, which further affects the quantity of struvite crystal and the recovery rate of phosphorus in livestock wastewater. So the molar ratio of phosphate and magnesium is the key factor to control the yield of phosphorus recovery. The molar ratio of phosphate and magnesium is 1:1 in theory. However, the real molar of the added magnesium is larger than the total amount of phosphorus in the real livestock wastewater. As shown in **Table 2**, for a higher phosphorus removal rate, the molar ratio of phosphate and magnesium is about 1:1**–**1.2 from livestock wastewater, 1:1.4 from synthetic livestock wastewater, and 1:1–1.4 from anaerobic digesters of livestock wastewater. It is mainly based on the effect of coexisting ions in the livestock wastewater. The coexisting ions, such as OH<sup>−</sup> and CO3 2 −, are apt to coprecipitate with Mg2+, which prevent the Mg2+ from touching Phosphorus Recovery by Struvite Crystallization from Livestock Wastewater and Reuse as Fertilizer: A Review http://dx.doi.org/10.5772/65692 141


**Table 2.** The summary of parameters on magnesium ammonium phosphate crystallization.

with the NH4 + and PO4 3−, so more magnesium is needed. Marit et al. [36] indicated that there would be many other kinds of magnesium phosphate precipitates except for struvite at different value of pH, such as Mg3 (PO4 ) 2 , MgHPO4 , and Mg(H2 PO4 ) 2 . So it does not mean that more magnesium means more struvite. Excessive amounts of magnesium would increase the pH value of the solution as well as the degree of saturation of the magnesium salts, resulting in the formation of other kinds of magnesium phosphate precipitates mentioned.

#### **4.2. The wastewater pH value**

After the crystal nucleus generates, the ions in the solution used to form the crystal begin to deposit on the crystal nucleus and the nucleus grow to the settling particles. During that time, there are two trends for the development of particles. One is orientation growth, which means the ions sequence in the crystal is arranged according to a certain lattice. The other one is nonorientation growth, which means these ions are too late to arrange in order. It is the fast growth rate that causes disorder. And two types of the crystal growth mechanisms lead to different trends. One is the integration mechanism, and the other is the mass transfer mechanism. The former is the integration of solute molecules into the surface; the latter is the transfer (by diffusion or convection) of solute molecules from the bulk solution to the crystal surface. When the effect of mass transfer is greater than the effect of integration, the crystal growth mainly depends on the diffusion effect and the growth of crystal would be orientable. However, if the effect of integration is greater, the integration on the surface of solute decides the crystal growth. And the relative sizes of the nucleation rate, aggregation rate, and directional array rate also decide the trend of crystal growth, which can be changed by precipitation conditions [31]. Abe et al. [34] showed that the growth rate of struvite was very slow. In the high concentration of phosphate (greater than 200 mg/L), the daily growth rate of struvite was 0.173 mm. In the low concentration of phosphate (30–100 mg/L), the daily growth rate of struvite was 0.061 mm. Therefore, increasing crystal growth rate and crystal size of struvite is not only beneficial to further removing phosphorus from livestock wastewater, but also to recycling phosphorus with a bigger size struvite. There is a metastable zone in industrial crystallization to make the crystal bigger and more even. The metastable zone is defined as a region bounded by the solubility curve in which the solution is supersaturated but the spontaneous nucleation cannot occur in such a short time [35]. In the metastable zone, the solute condenses on the nucleus as constantly as possible. As we known, the process of struvite constant growth is also the further recovery of phosphorus from livestock wastewater. So it is

140 Physico-Chemical Wastewater Treatment and Resource Recovery

meaningful to study the metastable zone of struvite for the industrialized application.

Generally speaking, livestock wastewater is rich in ammonium and phosphorus. So it is needed to add extra magnesium to form struvite. Therefore, the addition amount of magnesium affects the solubility product constant (Ksp) directly, which further affects the quantity of struvite crystal and the recovery rate of phosphorus in livestock wastewater. So the molar ratio of phosphate and magnesium is the key factor to control the yield of phosphorus recovery. The molar ratio of phosphate and magnesium is 1:1 in theory. However, the real molar of the added magnesium is larger than the total amount of phosphorus in the real livestock wastewater. As shown in **Table 2**, for a higher phosphorus removal rate, the molar ratio of phosphate and magnesium is about 1:1**–**1.2 from livestock wastewater, 1:1.4 from synthetic livestock wastewater, and 1:1–1.4 from anaerobic digesters of livestock wastewater. It is mainly based on the effect of coexisting ions in the livestock wastewater. The coexisting ions, such

−, are apt to coprecipitate with Mg2+, which prevent the Mg2+ from touching

**4. The factors influencing struvite crystallization**

**4.1. Molar ratio of P and Mg**

as OH<sup>−</sup>

and CO3

2

The value of pH of livestock wastewater is an important parameter to the formation of struvite. It affects the quality and the purity of struvite at the same time. As shown in **Table 2**, the best value of pH to form struvite is between 8 and 10, while 8.0–9.0 is the best for livestock wastewater. However, Hao et al. [37] indicated that struvite could get the highest purity at pH = 7.0, and the purity seemed to have fallen with the increasing of the pH value of wastewater. When the value of pH is higher than 10, the formed precipitate mainly consists of Mg3 (PO4 )2 (Ksp = 9.8 × 10−25). Song et al. [38] also found that the precipitate of Mg(OH)2 would form at the pH of 11. It does not mean that it is better to form struvite at a lower value of pH. However, considering the phosphorous recovery, as long as the productions of phosphorus salts are harmless and nontoxic, the aim of recovering phosphorus from livestock wastewater is reached. Anyway, the pH value of livestock wastewater is generally between 7.5 and 8.5, which is more convenient to recover phosphorus without the need for adjusting the pH value. It is helpful to simplify the technology and reduce the cost of livestock wastewater treatment.

#### **4.3. The coexisting ions**

There are many kinds of coexisting ions interfering with the crystallization of struvite in livestock wastewater, such as calcium ion (Ca2+), carbonate ion (CO3 2−), suspended solids (SS), and heavy metal ions (HMI). Moerman et al. [39] found that Ca2+ could enhance the phosphorus removal with forming the precipitate of Ca3 (PO4 ) 2 . However, it reduces the size of struvite. Meanwhile, lots of Ca3 (PO4 ) 2 powder flows out with effluent easily, declining the effluent water quality. Le Corre et al. [40] also declared that Ca2+ would compete with Mg2+ and form the precipitates of Ca3 (PO4 ) 2 (Ksp = 2.1 × 10−33) and CaHPO4 (Ksp = 1.8 × 10−7) at the pH value of 9. By performing batch experiments, Zhang et al. [41] found that the degree of the supersaturation would decrease with the increase of the concentration of CO3 2−. The CO3 2−, easily combining with Mg2+, increases the ion saturation in the solution and decreases the concentration of Mg2+ forming struvite. Suzuki et al. [42] showed that negatively charged SS adsorbed NH4 + and Mg2+ easily in the alkaline environment, which retarded the struvite crystalline rate. And Muryanto et al. [43] studied on the influence of copper ions (Cu2+) and zinc ions (Zn2+) in struvite crystallization and showed that the existence of Cu2+ and Zn2+ would delay the nucleation rate and the growth rate of struvite. Although they had little impact on the crystal shape, the crystal would have some cracks on the surface.

All in all, in the process of recovering phosphorus from livestock wastewater, some pretreatments are necessary to implement for removing these coexisting ions before forming the struvite. Laridi et al. [44] tried to reduce the negative impacts of organics and SS by adding ferric chloride and flocculants into the livestock wastewater, and it worked with a higher phosphorus removal rate at the same time. Suzuki et al. [42] tried to separate the struvite from suspended solids containing heavy metals by the differences of their settlement characteristics. It improved the purity of struvite and reduced the negative impact of SS and heavy metal ions.

#### **4.4. Seed crystal**

Seed crystals have positive influence in the struvite growth. Adding seed crystals can reduce the saturation of struvite crystallization in need, shorten the nucleation time, and increase the rate of crystal growth. What is more, struvite crystallizes on the surface of seed crystals, which enhances the separation of crystals and water, prevents the tiny crystals from flowing out with the effluent, and improves the phosphorus removal efficiency. Ariyanto et al. [45] showed that the smaller the added crystal nucleus is, the faster is the rate of crystal growth. Kim et al. [18] emphasized that the excessive amount of seed crystals added could not improve the phosphorus removal efficiency, and the pH value of wastewater also influenced the efficiency at the same dosage of seed crystals. The phosphorus removal efficiency is more significant at the pH value of 9. So only adding proper amount of seed crystals with a proper average size can the phosphorus removal efficiency be higher.

## **5. Crystallizer of struvite**

It is important to realize the phosphorus recycling in crystallizer, as the struvite crystallization equipment. The pros and cons of its design decide the shape and size of the struvite and the phosphorus removal efficiency from livestock wastewater. A series of struvite crystallizers had been developed and put into production successfully abroad previously, which had obtained environmental and economic benefits simultaneously. According to the mode of agitation, these crystallizers can be divided into air agitation type, water agitation type, and mechanical agitation type.

**4.3. The coexisting ions**

Meanwhile, lots of Ca3

the precipitates of Ca3

**4.4. Seed crystal**

There are many kinds of coexisting ions interfering with the crystallization of struvite in live-

heavy metal ions (HMI). Moerman et al. [39] found that Ca2+ could enhance the phosphorus

(PO4 ) 2

water quality. Le Corre et al. [40] also declared that Ca2+ would compete with Mg2+ and form

By performing batch experiments, Zhang et al. [41] found that the degree of the supersatura-

ing with Mg2+, increases the ion saturation in the solution and decreases the concentration of Mg2+ forming struvite. Suzuki et al. [42] showed that negatively charged SS adsorbed NH4

and Mg2+ easily in the alkaline environment, which retarded the struvite crystalline rate. And Muryanto et al. [43] studied on the influence of copper ions (Cu2+) and zinc ions (Zn2+) in struvite crystallization and showed that the existence of Cu2+ and Zn2+ would delay the nucleation rate and the growth rate of struvite. Although they had little impact on the crystal shape, the

All in all, in the process of recovering phosphorus from livestock wastewater, some pretreatments are necessary to implement for removing these coexisting ions before forming the struvite. Laridi et al. [44] tried to reduce the negative impacts of organics and SS by adding ferric chloride and flocculants into the livestock wastewater, and it worked with a higher phosphorus removal rate at the same time. Suzuki et al. [42] tried to separate the struvite from suspended solids containing heavy metals by the differences of their settlement characteristics. It improved the purity of struvite and reduced the negative impact of SS and heavy metal ions.

Seed crystals have positive influence in the struvite growth. Adding seed crystals can reduce the saturation of struvite crystallization in need, shorten the nucleation time, and increase the rate of crystal growth. What is more, struvite crystallizes on the surface of seed crystals, which enhances the separation of crystals and water, prevents the tiny crystals from flowing out with the effluent, and improves the phosphorus removal efficiency. Ariyanto et al. [45] showed that the smaller the added crystal nucleus is, the faster is the rate of crystal growth. Kim et al. [18] emphasized that the excessive amount of seed crystals added could not improve the phosphorus removal efficiency, and the pH value of wastewater also influenced the efficiency at the same dosage of seed crystals. The phosphorus removal efficiency is more significant at the pH value of 9. So only adding proper amount of seed crystals with a proper average size can the

It is important to realize the phosphorus recycling in crystallizer, as the struvite crystallization equipment. The pros and cons of its design decide the shape and size of the struvite and

(Ksp = 2.1 × 10−33) and CaHPO4

2−), suspended solids (SS), and

. However, it reduces the size of struvite.

2−. The CO3

(Ksp = 1.8 × 10−7) at the pH value of 9.

2−, easily combin-

+

powder flows out with effluent easily, declining the effluent

stock wastewater, such as calcium ion (Ca2+), carbonate ion (CO3

tion would decrease with the increase of the concentration of CO3

removal with forming the precipitate of Ca3

142 Physico-Chemical Wastewater Treatment and Resource Recovery

(PO4 )2

(PO4 )2

crystal would have some cracks on the surface.

phosphorus removal efficiency be higher.

**5. Crystallizer of struvite**

The air agitation type crystallizer is a kind of crystallizer that is studied widely. The special aerate system can not only mix the solution more efficiently, improving the collision chance of crystal forming ions, but also vent the gas carbon dioxide and insoluble ammonium from the solution, increasing the pH value of the solution, the ammonium removal efficiency, and the effluent water quality at the same time. The crystallizer used by the British slough sewage treatment plant (**Figure 3a**) reached the soluble phosphorus removal rate of 94% and the total phosphorus removal rate of 87.5% with drumming into air under the inner reaction zone [46]. Le Corre et al. designed two concentric meshes made of stainless steel as a substrate to grow struvite in the crystallizer, which can trap and then accumulate the struvite in the reactor as an adhesive (**Figure 3b**). With the help of crystallizers, the phosphorus removal rate can increase from 81 to 86%. However, because of the limitations of volume and growth time, the struvite crystal cannot grow large enough in the air agitation type crystallizer, which causes the loss of phosphorus recovery. Moreover, some kinds of air agitation type crystallizer have the problem of replacing padding or membranes frequently. What is more, the congestion problem becomes serious once the crystallizer broke down for some reasons, and it is hard to restart.

Phosphorus Recovery by Struvite Crystallization from Livestock Wastewater and Reuse as Fertilizer: A Review http://dx.doi.org/10.5772/65692 145

144 Physico-Chemical Wastewater Treatment and Resource Recovery

**Figure 3.** Different kinds of crystallizers forming struvite. (a) The air agitation crystallizer from the British slough sewage treatment plant. (b) The air agitation crystallizer from Le Corre et al. (c) The water agitation crystallizer from Guadia et al. (d) The water agitation crystallizer from Rahaman et al. (e) The MSMPR type crystallizer with 1—internal circulation of suspension, 2—thermostat, 3—computer, 4—rural wastewater (including aqueous solution of MgCl2 ), 5—pump, 6 alkalinity agent tank: aqueous solution of NaOH, 7 and 8—pump, 9—storage tank of a product crystal suspension, 10, 11, and 12—electronic balances, M—stirrer speed control, T—temperature control, and pH—acid/alkaline reaction control.

The water agitation type crystallizer realizes uniform mixing by changing the solution flow direction, speed, or gravity changed the flow rate by increasing the diameters of the equipment from the bottom to the top, inserted cone-shape structures at an angle of 45° between every diameter-changed parts to reduce unwanted crystal loss at each junction, and recycled finer particles with the effluent through the external recycler (**Figure 3c**). This crystallizer can remove 92% phosphate, and the purity of struvite goes up to 99%. Rahaman et al. [47] designed four distinct zones at the same principle, added a settling zone (also called seed  hopper) at the top, getting the phosphate removal rate of up to 90% and the size of particles up to 3.5 mm. Seed crystals are added into the crystallizer from the seed hopper and allowed the finer crystals to continue to grow up in the upper supersaturated solution.

The mechanical agitation type crystallizer, mixing solution by impellers, is simple in design and easy to operate. However, it causes greater energy consumption and uneven size of crystals distribution. Recently, a new crystallizer called mixed suspension, mixed product removal crystallizer (MSMPR for short) can solve the problem of uneven crystal size distribution (**Figure 3e**). With the mechanical agitation centered and the water agitation assisted, MSMPR can uniform the suspension density and particle size of the crystals, and remove productions evenly by controlling the speed and time of mixing. Hutnik et al. [48] and Kozik et al. [49] both got the phosphate removal rate up to 99% from industrial wastewater and wastewater with low concentration of phosphorus. And both of them confirmed that MSMPR could increase the size of crystals, improve the crystallization rate of the struvite, and enhance the phosphorus recovery rate.

## **6. Application of struvite as a fertilizer**

It is reckoned that 100 m3 wastewater could form 1 kg of struvite. If all the wastewater in the world is treated by struvite crystallization, 63,000 tons of P2 O5 could be recovered, equaling to 16% of the phosphate rock production of the world [50]. And 171 g struvite can be recovered from livestock wastewater per square meter at most and the purity as high as 95% without washing. Therefore, recovery of struvite returning to the farmland is a developmental trend of struvite crystallization technology. Struvite, as a slightly soluble crystal, for containing the equal molar concentrations of magnesium (Mg), ammonium, and phosphate, has been successfully used on herbages [51], vegetables [14, 52], and grain crops [53] as a fertilizer, especially on the magnesium-fond crops, like sugar beet [54]. Especially, the presence of Mg in struvite makes it more attractive as an alternative to contemporary fertilizers for a few crops, which require magnesium [55]. Ryu et al. [52] found that the struvite source provided the essential crop nutrients of N, P, K, Ca, and Mg for Chinese cabbage as much as other commercial fertilizers. Moreover, it has a lasting positive function to roots and does not burn the seeding or roots due to its slow release characteristics. Besides, compared with other highly soluble fertilizers, struvite is more suitable for use in the vast areas of forest. Since the area of forest is too large to fertilize frequently, the use of struvite can decrease the frequency of fertilization and reduce the loss of nutrients [54]. However, as livestock wastewater is full of impurities, especially the heavy metal ions, the struvite recovered from livestock wastewater still contain more or less heavy metal ions. From livestock-based struvite, toxic substances may diffuse into the aquatic environment or accumulate in soils and have an adverse effect on the human health and environment [56, 57]. Although currently no specific threshold values are available for micropollutants in fertilizers, the introduction of potential hazardous substances into the environment should be avoided. The accumulation of heavy metal ions will be a serious concern for sustainability [58]. Ryu et al. [52] made a security evaluation for struvite as a fertilizer used in the soil. They affirmed the fertilizer efficiency of struvite and emphasized the negative effect of higher concentrations of copper and cadmium in struvite at the same time. Because copper and cadmium were tested in the cabbage fertilized by the struvite used as fertilizer, the struvite, especially recovered from livestock wastewater, needs to tested for the amount of toxic or harmful substances, followed by the security evaluation as a fertilizer.

## **7. Summary and outlook**

 hopper) at the top, getting the phosphate removal rate of up to 90% and the size of particles up to 3.5 mm. Seed crystals are added into the crystallizer from the seed hopper and allowed

The mechanical agitation type crystallizer, mixing solution by impellers, is simple in design and easy to operate. However, it causes greater energy consumption and uneven size of crystals distribution. Recently, a new crystallizer called mixed suspension, mixed product removal crystallizer (MSMPR for short) can solve the problem of uneven crystal size distribution (**Figure 3e**). With the mechanical agitation centered and the water agitation assisted, MSMPR can uniform the suspension density and particle size of the crystals, and remove productions evenly by controlling the speed and time of mixing. Hutnik et al. [48] and Kozik et al. [49] both got the phosphate removal rate up to 99% from industrial wastewater and wastewater with low concentration of phosphorus. And both of them confirmed that MSMPR could increase the size of crystals, improve the crystallization rate of the struvite, and enhance

16% of the phosphate rock production of the world [50]. And 171 g struvite can be recovered from livestock wastewater per square meter at most and the purity as high as 95% without washing. Therefore, recovery of struvite returning to the farmland is a developmental trend of struvite crystallization technology. Struvite, as a slightly soluble crystal, for containing the equal molar concentrations of magnesium (Mg), ammonium, and phosphate, has been successfully used on herbages [51], vegetables [14, 52], and grain crops [53] as a fertilizer, especially on the magnesium-fond crops, like sugar beet [54]. Especially, the presence of Mg in struvite makes it more attractive as an alternative to contemporary fertilizers for a few crops, which require magnesium [55]. Ryu et al. [52] found that the struvite source provided the essential crop nutrients of N, P, K, Ca, and Mg for Chinese cabbage as much as other commercial fertilizers. Moreover, it has a lasting positive function to roots and does not burn the seeding or roots due to its slow release characteristics. Besides, compared with other highly soluble fertilizers, struvite is more suitable for use in the vast areas of forest. Since the area of forest is too large to fertilize frequently, the use of struvite can decrease the frequency of fertilization and reduce the loss of nutrients [54]. However, as livestock wastewater is full of impurities, especially the heavy metal ions, the struvite recovered from livestock wastewater still contain more or less heavy metal ions. From livestock-based struvite, toxic substances may diffuse into the aquatic environment or accumulate in soils and have an adverse effect on the human health and environment [56, 57]. Although currently no specific threshold values are available for micropollutants in fertilizers, the introduction of potential hazardous substances into the environment should be avoided. The accumulation of heavy metal ions will be a serious concern for sustainability [58]. Ryu et al. [52] made a security evaluation for

wastewater could form 1 kg of struvite. If all the wastewater in the

O5

could be recovered, equaling to

the finer crystals to continue to grow up in the upper supersaturated solution.

the phosphorus recovery rate.

It is reckoned that 100 m3

**6. Application of struvite as a fertilizer**

146 Physico-Chemical Wastewater Treatment and Resource Recovery

world is treated by struvite crystallization, 63,000 tons of P2

Struvite crystallization represents a promising tool for recovering phosphorus from livestock wastewater. Based on this study, the conclusions are as follows.

Struvite is a white crystal which is formed in neural or mild alkali conditions. Nucleation and crystal growth are two steps for struvite crystallization from generation to growth. The molar ratio of magnesium and phosphate, and solution pH are the key factors to control. The coexisting substances, such as calcium, carbonate, suspended solids, and heavy metals, interfere the crystallization of struvite. Seed crystals have positive influence on struvite growth. Adding seed crystals can reduce the saturation of struvite crystallization, shorten the nucleation time, and improve the rate of crystal growth.

Crystallizer, its design decides the shape and size of struvite, is important to realize the phosphorus recycling. According to the agitation mode, it can be divided into air agitation, water agitation, and mechanical agitation.

The recovered struvite can be used on herbages, vegetables, and grain crops as a fertilizer, especially on the magnesium-fond crops, like sugar beet.

However, there are still some problems. Livestock wastewater belongs to the organic wastewater with high concentrations of ammonium, phosphorus, organics, and suspend solids. And the existing forms are complex, such as simple monoester phosphorus, phytate-like phosphorus, and polynucleotide-like phosphoric. So it is necessary to use some physical or chemical measures to transform different kinds of phosphorus to phosphate, as many as possible, before removing phosphorus from livestock wastewater.

The design for struvite crystallizer is still a key to struvite crystallization technology. Although enhancing the phosphorus removal rate has got a big breakthrough by the current crystallizer, the crystals are still too small to recover and block the crystallizer easily. It is the influence of negative zeta potential on the surface of crystals that makes further aggregating hard for small crystals. So finding a way to change the zeta potential on the surface of the crystals, enhancing the aggregation capability, and increasing the size of the crystals is required.

Struvite, a fertilizer with high concentrations of nutrients, might be difficult in application for the influence of other toxic or harmful impurities. Therefore, to reduce the environmental risk at source, it is necessary that estimating the potential effects of struvite on the ecosystem before use.

## **Acknowledgements**

The work was supported by a grant from the National Natural Science Foundation of China (No. 31401944), the National key Research and Development Program of China (No. 2016YFD0501404), the Beijing Municipal Natural Science Foundation (No. 6144026), China Agricultural University Education Foundation "Da Bei Nong Group Education Foundation" (No. 1031-2415005), and the Specialized Research Fund for the Doctoral Program of Higher Education of China (No. 20120008120013).

## **Author details**

Tao Zhang\*, Rongfeng Jiang and Yaxin Deng

\*Address all correspondence to: taozhang@cau.edu.cn

Key Laboratory of Plant-Soil Interactions of Ministry of Education, College of Resources and Environmental Sciences, China Agricultural University, Beijing, China


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#### **Slaughterhouse Wastewater: Treatment, Management and Resource Recovery Slaughterhouse Wastewater: Treatment, Management and Resource Recovery**

Ciro Bustillo-Lecompte and Mehrab Mehrvar Ciro Bustillo-Lecompte and Mehrab Mehrvar

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/65499

#### **Abstract**

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152 Physico-Chemical Wastewater Treatment and Resource Recovery

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The meat processing industry is one of the largest consumers of total freshwater used in the agricultural and livestock industry worldwide. Meat processing plants (MPPs) produce large amounts of slaughterhouse wastewater (SWW) because of the slaugh‐ tering process and cleaning of facilities. SWWs need significant treatment for a sustainable and safe discharge to the environment due to the high content of organics and nutrients. Therefore, the treatment and final disposal of SWW are a public health necessity. In this chapter, the regulatory frameworks relevant to the SWW management, environmental impacts, health effects, and treatment methods are discussed. Although physical, chemical, and biological treatment can be used for SWW degradation, each treatment process has different advantages and drawbacks depending on the SWW characteristics, best available technology, jurisdictions, and regulations. SWWs are typically assessed using bulk parameters because of the various pollutant loads derived from the type and the number of animals slaughtered that fluctuate amid the meat industry. Thus, an on‐site treatment using combined processes would be the best option to treat and disinfect the slaughterhouse effluents to be safely discharged into receiving waters.

**Keywords:** Anaerobic digestion, Activated sludge, Advanced oxidation processes, Combined processes, Slaughterhouse wastewater

## **1. Introduction**

The meat processing industry consumes 29% of the total freshwater used by the agricultural sector worldwide [1, 2]. Moreover, the global production of beef, pork, and poultry meat has been doubled in the past decade and is projected to grow steadily until 2050. Thus, the number

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

of slaughterhouse facilities is increasing, which results in an expected higher volume of slaughterhouse wastewater (SWW) to be treated [3]. SWWs are classified as one of the most detrimental industrial wastewaters to the environment by the United States Environmental Protection Agency (US EPA) because the inadequate disposal of SWW is one of the reasons for river deoxygenation and groundwater pollution [4]. Thus, SWWs require significant treatment for a safe and sustainable release to the environment, and the treatment and disposal of wastewater from slaughterhouses are an economic and public health necessity [5, 6].

The organic matter concentration in meat processing plant (MPP) effluents is usually high, and the residues are moderately solubilized, leading to a polluting effect due to the high levels of organics and pathogens present in SWW along with detergents used for cleaning purposes. SWWs are typically assessed using bulk parameters because of the various pollutant loads derived from the type and the number of animals slaughtered that fluctuate amid the meat industry [7].

Anaerobic treatment is the preferred biological treatment because of its effectiveness in treating high‐strength wastewater such as SWW with less complex equipment requirements [8]. Although anaerobic treatment is efficient, anaerobically treated effluents require posttreatment to comply with required discharge limits where the complete stabilization of the organic matter is not possible by anaerobic treatment alone. Anaerobically treated effluents contain solubi‐ lized organic matters, which are more suited for treatment using aerobic processes. Therefore, aerobic treatment systems are more frequently used in wastewater treatment systems since they operate at higher rates than conventional anaerobic treatment methods. Taking into account that oxygen requirements and treatment time are directly proportional to an increase in wastewater strength, aerobic treatment is frequently applied as posttreatment of anaerobic effluents as well as for nutrient removal [9].

Nevertheless, biological processes alone do not produce effluents that comply with current effluent discharge limits when treating high‐organic‐strength wastewaters. The use of combined anaerobic and aerobic processes is beneficial for its potential resource recovery and high treatment efficiency [10].

On the other hand, some slaughterhouse effluents contain toxic, bioresistant, recalcitrant, and nonbiodegradable substances. Thus, advanced oxidation processes (AOPs) could be used to improve the biodegradability of SWW and inactivate pathogenic microorganisms and viruses, left after biological treatment of the wastewater. Consequently, AOPs are an attractive alter‐ native and a complementary treatment method to biological processes for the treatment of slaughterhouse effluents, especially as a posttreatment method [5–7]. Adopting combined biological treatment and AOPs for the treatment of slaughterhouse effluents is considered operationally and economically advantageous. Combined processes incorporate advantages of diverse technologies to achieve high‐quality effluents from industrial and high‐strength wastewaters for water reuse and resource recovery purposes [9, 10].

In this chapter, the regulatory frameworks relevant to the SWW management, environmental impacts, and health effects are discussed along with common practices for SWW treatment. Significant progress in the combination of biological treatment and AOPs is emphasized. A case study for the treatment of an actual SWW by integrated anaerobic baffled reactor (ABR)‐ aerobic activated sludge (AS)‐UV/H2 O2 processes is presented. The overall treatment efficiency of organics and nutrients, the potential energy recovery from CH₄ production, and the H2 O2 residual are discussed. A cost‐effectiveness analysis is used to minimize the treatment time as well as the overall incurred treatment costs required for the efficient treatment of slaughter‐ house effluents. Finally, the chapter ends with a discussion on the need for an adequate SWW management, resource recovery, and the improvement actions required.

## **2. Characteristics of slaughterhouse wastewater**

of slaughterhouse facilities is increasing, which results in an expected higher volume of slaughterhouse wastewater (SWW) to be treated [3]. SWWs are classified as one of the most detrimental industrial wastewaters to the environment by the United States Environmental Protection Agency (US EPA) because the inadequate disposal of SWW is one of the reasons for river deoxygenation and groundwater pollution [4]. Thus, SWWs require significant treatment for a safe and sustainable release to the environment, and the treatment and disposal of wastewater from slaughterhouses are an economic and public health necessity [5, 6].

The organic matter concentration in meat processing plant (MPP) effluents is usually high, and the residues are moderately solubilized, leading to a polluting effect due to the high levels of organics and pathogens present in SWW along with detergents used for cleaning purposes. SWWs are typically assessed using bulk parameters because of the various pollutant loads derived from the type and the number of animals slaughtered that fluctuate amid the meat

Anaerobic treatment is the preferred biological treatment because of its effectiveness in treating high‐strength wastewater such as SWW with less complex equipment requirements [8]. Although anaerobic treatment is efficient, anaerobically treated effluents require posttreatment to comply with required discharge limits where the complete stabilization of the organic matter is not possible by anaerobic treatment alone. Anaerobically treated effluents contain solubi‐ lized organic matters, which are more suited for treatment using aerobic processes. Therefore, aerobic treatment systems are more frequently used in wastewater treatment systems since they operate at higher rates than conventional anaerobic treatment methods. Taking into account that oxygen requirements and treatment time are directly proportional to an increase in wastewater strength, aerobic treatment is frequently applied as posttreatment of anaerobic

Nevertheless, biological processes alone do not produce effluents that comply with current effluent discharge limits when treating high‐organic‐strength wastewaters. The use of combined anaerobic and aerobic processes is beneficial for its potential resource recovery and

On the other hand, some slaughterhouse effluents contain toxic, bioresistant, recalcitrant, and nonbiodegradable substances. Thus, advanced oxidation processes (AOPs) could be used to improve the biodegradability of SWW and inactivate pathogenic microorganisms and viruses, left after biological treatment of the wastewater. Consequently, AOPs are an attractive alter‐ native and a complementary treatment method to biological processes for the treatment of slaughterhouse effluents, especially as a posttreatment method [5–7]. Adopting combined biological treatment and AOPs for the treatment of slaughterhouse effluents is considered operationally and economically advantageous. Combined processes incorporate advantages of diverse technologies to achieve high‐quality effluents from industrial and high‐strength

In this chapter, the regulatory frameworks relevant to the SWW management, environmental impacts, and health effects are discussed along with common practices for SWW treatment. Significant progress in the combination of biological treatment and AOPs is emphasized. A

wastewaters for water reuse and resource recovery purposes [9, 10].

industry [7].

effluents as well as for nutrient removal [9].

154 Physico-Chemical Wastewater Treatment and Resource Recovery

high treatment efficiency [10].

Meat processing effluents are considered harmful worldwide due to the SWW complex composition of fats, proteins, fibers, high organic content, pathogens, and pharmaceuticals for veterinary purposes. Slaughterhouse effluents are typically evaluated using bulk parameters because of the broad range of SWW and pollutant loads. SWW contains large amounts of biochemical oxygen demand (BOD), chemical oxygen demand (COD), total organic carbon (TOC), total nitrogen (TN), total phosphorus (TP), and total suspended solids (TSS) [7]. Typical characteristics of an actual SWW are summarized in **Table 1**.


**Table 1.** Typical characteristics of the slaughterhouse wastewater.

As a result, due to the diverse characteristics of the SWW, it is appropriate to classify and minimize wastewater production at its source. Meat processing effluents are becoming one of the major agribusiness concerns due to the vast amount of water used during slaughtering, processing, and cleaning of the slaughtering facilities.

## **3. Regulations for slaughterhouse wastewater management**

Regulations are necessary to mitigate the environmental impact of slaughterhouses, and the treatment methods are used as the main regulatory requirement [11]. The compliance with current environmental legislation and the state‐of‐the‐art technologies may also provide some economic relief via resource recovery from biogas generation using high‐rate anaerobic treatment.

**Table 2** describes current regulations and discharge limits for organics and nutrients in SWW for an adequate release to the environment in different jurisdictions worldwide, including the World Bank Group [12], the Council of the European Communities [13], the US EPA [14], the Environment Canada [15, 16], the Colombian Ministry of Environment and Sustainable Development Colombia [17], the People's Republic of China Ministry of Environmental Protection [18], the Indian Central Pollution Control Board [19], and the Australian and New Zealand Environment and Conservation Council [20, 21].


**Table 2.** Comparison of standard limits for slaughterhouse wastewater discharge in different jurisdictions worldwide.

Although it can be seen that Canadian standards are stricter than other international regula‐ tions such as those in the European Union (EU), Australia and New Zealand, or the USA, Canada does not have a specific regulation for the meat processing industry. Moreover, Australia and New Zealand and the USA have been incorporating an integrated approach to the regulation of the MPPs, where industry and regulatory sectors are working together to achieve a common goal of reducing the threats caused by the hazardous and high‐strength wastewaters produced in slaughterhouses. Finally, emerging economies such as India, China, and Colombia have less strict standards, but their legislation is focused on specific industries to attain certain levels of treatment depending on the wastewater strength. Therefore, the selection of a specific treatment method depends on the characteristics of the SWW being treated, the best available technology economically achievable (BAT), and the compliance with regulations in different political jurisdictions.

## **4. Environmental impact and health effects of slaughterhouse wastewater**

**3. Regulations for slaughterhouse wastewater management**

156 Physico-Chemical Wastewater Treatment and Resource Recovery

Zealand Environment and Conservation Council [20, 21].

regulations in different political jurisdictions.

treatment.

Regulations are necessary to mitigate the environmental impact of slaughterhouses, and the treatment methods are used as the main regulatory requirement [11]. The compliance with current environmental legislation and the state‐of‐the‐art technologies may also provide some economic relief via resource recovery from biogas generation using high‐rate anaerobic

**Table 2** describes current regulations and discharge limits for organics and nutrients in SWW for an adequate release to the environment in different jurisdictions worldwide, including the World Bank Group [12], the Council of the European Communities [13], the US EPA [14], the Environment Canada [15, 16], the Colombian Ministry of Environment and Sustainable Development Colombia [17], the People's Republic of China Ministry of Environmental Protection [18], the Indian Central Pollution Control Board [19], and the Australian and New

**Parameter World Bank EU USA Canada Colombia China India Australia** BOD (mg/L) 30 25 16–26 5–30 50 20–100 30–100 5–20 COD (mg/L) 125 125 n.a. n.a. 150 100–300 250 40 TN (mg/L) 10 10–15 4–8 1.25 10 15–20 10–50 10–20 TOC (mg/L) n.a. n.a. n.a. n.a. n.a. 20–60 n.a. 10 TP (mg/L) 2 1–2 n.a. 1.00 n.a. 0.1–1.0 5 2 TSS (mg/L) 50 35–60 20–30 5–30 50 20–30 100 5–20 pH 6–9 n.a. 6–9 6–9 6–9 6–9 5.5–9.0 5–9 Temperature (°C change) n.a. n.a. n.a. <1 °C n.a. n.a. <5 °C <2 °C

**Table 2.** Comparison of standard limits for slaughterhouse wastewater discharge in different jurisdictions worldwide.

Although it can be seen that Canadian standards are stricter than other international regula‐ tions such as those in the European Union (EU), Australia and New Zealand, or the USA, Canada does not have a specific regulation for the meat processing industry. Moreover, Australia and New Zealand and the USA have been incorporating an integrated approach to the regulation of the MPPs, where industry and regulatory sectors are working together to achieve a common goal of reducing the threats caused by the hazardous and high‐strength wastewaters produced in slaughterhouses. Finally, emerging economies such as India, China, and Colombia have less strict standards, but their legislation is focused on specific industries to attain certain levels of treatment depending on the wastewater strength. Therefore, the selection of a specific treatment method depends on the characteristics of the SWW being treated, the best available technology economically achievable (BAT), and the compliance with The commercialization of animal products for consumption leads to the production of a large volume of SWW. Although the environment can handle a certain amount of pollutants through natural degradation processes, as the SWW concentration increases, these mechanisms come to be overburdened, where contamination problems commence [22].

The discharge of raw SWW to water bodies affects the quality of water particularly by causing a reduction of dissolved oxygen (DO), which may lead to the death of aquatic life [23]. Moreover, macronutrients, such as nitrogen and phosphorus, may cause eutrophication events. The discharge of these nutrients triggers an excessive algae growth and subsequent decay. Thus, the mineralization of the algae may lead to the deterioration of aquatic life due to depletion of DO levels. Finally, SWW may contain compounds, such as chromium and un‐ ionized ammonia, which are directly toxic to aquatic life [24].

Another source of contamination of the meat processing industry is the addition of surfactants as a result of the cleaning process. Surfactants, major components in detergents, may enter the aquatic environment due to an inadequate SWW treatment, causing short‐term and long‐term changes in the ecosystem that affect humans, fish, and vegetation [25].

The environmental impact of SWW is not only characterized by pollution via surfactants, nitrate, and chloric anions but also pathogens, which persist in the soil and reproduce continuously. Pathogens from SWW can also be transmitted to humans who are exposed to the water body, making those areas nonsuitable for drinking, swimming, or irrigation pur‐ poses [5, 26].

The general public health effects of the meat processing industry are related to the direct interaction of human communities with the slaughterhouse activities and indirect interactions with the environment, which can be previously affected by the inadequate management of the liquid effluents, solid waste, and obnoxious odors [27]. According to Um et al. [28], conven‐ tional treatment processes have no major impact on the reduction of antibiotic‐resistant *Escherichia coli* strains present in SWW, highlighting the public health risks associated with inadequately treated slaughterhouse effluents concerning the propagation of antibiotic‐ resistant and pathogenic bacteria into the environment.

The unsanitary conditions in some slaughterhouses allow the proliferation of pathogens to the final meat product to be consumed. People from developing countries in Africa, Asia, and South America have experienced serious gastrointestinal diseases, bloody diarrhea, liver malfunctions, and, in some cases, death associated with the presence of viruses, protozoa, helminthic eggs, and bacteria in SWW [5, 27]. Furthermore, the presence of hepatitis A and E viruses has been reported in the sewage of animal origin in Spain. Therefore, SWW must be treated efficiently before discharge into water bodies to avoid environmental pollution and human health effects [29].

## **5. Treatment methods for slaughterhouse wastewater**

The freshwater consumption substantially varies in the meat processing sector, and a typical MPP generates a large amount of wastewater from the slaughtering process and cleaning of the facilities. Therefore, the water reuse and the recovery of valuable by‐products from the meat processing effluents are the main focus in the agribusiness toward a cleaner production focused on high‐quality effluents, biogas production and exploitation, and recovery of nutrients and fertilizers [7].

Treatment methods for SWW are comparable to those used in municipal wastewater treatment and include primary, secondary, and tertiary treatment. However, this does not eliminate the need for primary treatment. There are numerous SWW treatment methods after preliminary treatment, which can be divided into four main categories: physicochemical treatment, biological treatment, AOPs, and combined processes [2, 7]. Each method has advantages and disadvantages, which are discussed below.

#### **5.1. Preliminary treatment**

The purpose of the preliminary treatment is to separate solids and large particles from the liquid portion in SWW and remove up to 30% of the BOD. The most common unit operations for preliminary treatment of SWW include screeners, sieves, and strainers. Thus, large solids with a 10–30 mm diameter are retained while the SWW passes through. Other preliminary treatment methods include homogenization and equalization and flotation, among other systems such as catch basins and settlers [30].

#### **5.2. Physicochemical treatment**

After preliminary treatment, the effluent should be further treated using primary and secon‐ dary treatment. One of the most practical methods of primary treatment for SWW is dissolved air floatation (DAF) for the reduction of fat, oil, grease, TSS and BOD [31]. The most commonly used physicochemical treatment methods are presented below.

#### *5.2.1. Coagulation-flocculation and sedimentation*

In the coagulation process, colloidal particles in the SWW are grouped into larger particles, called flocs. The colloidal particles in SWW are nearly negatively charged which make them stable and resistant to aggregation. For this reason, coagulants with positively charged ions are added to destabilize the colloidal particles to form flocs and facilitate the sedimentation process. Various coagulant types can be found in the market, and the most widely used are inorganic metal based‐coagulants such as aluminum sulfate, aluminum chlorohydrate, ferric chloride, ferric sulfate, and poly‐aluminum chloride with removal efficiencies of up to 80% for BOD, COD, and TSS [32].

#### *5.2.2. Dissolved air flotation*

**5. Treatment methods for slaughterhouse wastewater**

nutrients and fertilizers [7].

**5.1. Preliminary treatment**

disadvantages, which are discussed below.

158 Physico-Chemical Wastewater Treatment and Resource Recovery

systems such as catch basins and settlers [30].

*5.2.1. Coagulation-flocculation and sedimentation*

used physicochemical treatment methods are presented below.

**5.2. Physicochemical treatment**

BOD, COD, and TSS [32].

The freshwater consumption substantially varies in the meat processing sector, and a typical MPP generates a large amount of wastewater from the slaughtering process and cleaning of the facilities. Therefore, the water reuse and the recovery of valuable by‐products from the meat processing effluents are the main focus in the agribusiness toward a cleaner production focused on high‐quality effluents, biogas production and exploitation, and recovery of

Treatment methods for SWW are comparable to those used in municipal wastewater treatment and include primary, secondary, and tertiary treatment. However, this does not eliminate the need for primary treatment. There are numerous SWW treatment methods after preliminary treatment, which can be divided into four main categories: physicochemical treatment, biological treatment, AOPs, and combined processes [2, 7]. Each method has advantages and

The purpose of the preliminary treatment is to separate solids and large particles from the liquid portion in SWW and remove up to 30% of the BOD. The most common unit operations for preliminary treatment of SWW include screeners, sieves, and strainers. Thus, large solids with a 10–30 mm diameter are retained while the SWW passes through. Other preliminary treatment methods include homogenization and equalization and flotation, among other

After preliminary treatment, the effluent should be further treated using primary and secon‐ dary treatment. One of the most practical methods of primary treatment for SWW is dissolved air floatation (DAF) for the reduction of fat, oil, grease, TSS and BOD [31]. The most commonly

In the coagulation process, colloidal particles in the SWW are grouped into larger particles, called flocs. The colloidal particles in SWW are nearly negatively charged which make them stable and resistant to aggregation. For this reason, coagulants with positively charged ions are added to destabilize the colloidal particles to form flocs and facilitate the sedimentation process. Various coagulant types can be found in the market, and the most widely used are inorganic metal based‐coagulants such as aluminum sulfate, aluminum chlorohydrate, ferric chloride, ferric sulfate, and poly‐aluminum chloride with removal efficiencies of up to 80% for The DAF technology refers to the method of liquid‐solid separation by air introduction. The fat and grease along with light solids are moved to the surface creating a sludge blanket. Thus, it can be continuously removed via scum scraping. Furthermore, flocculants and blood coagulants can be added to enhance the effectiveness of the DAF treatment for COD and BOD removals of up to 75%. Nevertheless, common DAF disadvantages include occasional mal‐ functioning, poor TSS elimination, and moderate nutrient removal [33].

## *5.2.3. Electrocoagulation*

The electrocoagulation (EC) process has been employed as a cost‐effective technology for the removal of organics, heavy metals, and pathogens from slaughterhouse effluents by inducing an electric current without chemical addition. The EC process generates M3+ ions, mainly Fe3+ and Al3+, using different electrode materials. Other electrode types including Pt, SnO2 , and TiO2 can interact with H+ or OH− ions in acidic or alkaline conditions, respectively. Thus, removal efficiencies of up to 80, 81, 84, 85, and 96% can be achieved for BOD, TSS, TN, COD, and color, respectively [34, 35].

#### *5.2.4. Membrane processes*

Membrane processes are becoming an alternative treatment method for meat processing effluents. Different membrane processes, including microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO), have been used for SWW treatment to remove particulates, colloids, macromolecules, organic matter, and pathogens with overall efficiencies of up to 90%. However, membrane processes are required to be coupled with conventional processes for nutrient removal in SWW. Another drawback of membrane processes refers to the membrane fouling when treating high‐strength wastewater because of the formation of biofouling layers on the membranes, restricting the permeation rate [36].

#### **5.3. Biological treatment**

Primary treatment and physicochemical processes typically do not treat SWW completely, to a degree of satisfaction set by regulations. Thus, secondary treatment is used for the removal of the remaining soluble organic compounds from primary treatment. Biological processes include lagoons with anaerobic, aerobic, or facultative microorganisms, trickling filters, activated sludge (AS) bioreactors, and constructed wetlands (CWs) for organic and nutrient removal efficiencies of up to 90% [7].

#### *5.3.1. Anaerobic treatment*

Anaerobic digestion is the preferred method for SWW treatment due to its effectiveness in treating highly concentrated industrial effluents since organic compounds are degraded by anaerobic bacteria in the absence of oxygen into CO2 and CH₄ . Anaerobic systems have the advantage of achieving low sludge production, minimum energy requirements with potential resource recovery, and high COD removal. Typical anaerobic processes for the treatment of meat processing effluents comprise anaerobic baffled reactor (ABR), anaerobic digester (AD), anaerobic filter (AF), anaerobic lagoon (AL), septic tanks (ST), and up-flow anaerobic sludge blanket (UASB) [30].

Nevertheless, anaerobic treatment barely complies with current discharge limits. Complete stabilization of the organic compounds is difficult due to the high organic strength of SWW. Therefore, an additional treatment stage is recommended to remove the organics, nutrients, and pathogens that remain after anaerobic treatment. On the other hand, anaerobic treatment requires a higher space and a higher residence time to achieve high overall treatment efficiency, affecting the economic viability of anaerobic treatment alone. Accordingly, the combination of anaerobic and aerobic processes is necessary to achieve a maximum efficiency for the treatment of SWW [37].

#### *5.3.2. Aerobic treatment*

Aerobic processes are frequently employed for nutrient removal and further treatment after primary treatment. The required oxygen and treatment time are directly related to the strength of the SWW, which makes it inadequate as primary treatment of SWW but adequate after anaerobic treatment [38].

There are many advantages of using aerobic wastewater treatment processes, including low odor production, fast biological growth rate, and rapid adjustments to the temperature and loading rate changes. Conversely, the operating costs of aerobic systems are higher than those for anaerobic systems due to the maintenance and energy requirements for artificial oxygenation. There are different aerobic unit operations for SWW treatment, such as aerobic AS, rotating biological contactors (RBCs), and sequencing batch reactors (SBRs) [39].

#### *5.3.3. Constructed wetlands*

Constructed wetlands (CWs) emulate the degradation mechanisms of natural wetlands for water decontamination, integrating biological and physicochemical processes from the interaction of vegetation, soil, microorganisms, and atmosphere for the adsorption, biodegradation, filtration, photooxidation, and sedimentation of organics and nutrients.

The performance of CW systems for the treatment of SWW has been evaluated using both horizontal and vertical subsurface flow CWs. Results have shown a wide range of organic and nutrient removal for different vegetation with encouraging maximum removals of 99, 97, 85, and 78% for BOD, COD, TSS, and TN, respectively [40]. As a result, CWs are simple methods with low operation and maintenance costs and few negative impacts on the environment, which make them an attractive alternative to conventional treatment [41].

#### **5.4. Advanced oxidation processes**

AOPs are an interesting complementary treatment option for primary or secondary treatment of SWW, showing excellent overall treatment efficiencies for water reuse. AOPs are diverse and include gamma radiation, ozonation, ultrasound technology (UST), UV/H2 O2 , UV/O₃ , and photocatalysis, among others, for the oxidation and degradation of organic matter. The disinfection is another benefit of AOPs, which can inactivate pathogens without adding additional chemicals in comparison to other disinfection methods, such as chlorination, preventing the formation of hazardous by‐products [5]. Another main advantage of the AOPs is the high reaction rates as well as very low treatment time.

Photocatalysis using photo‐Fenton–based processes and photooxidation using UV/H2 O2 are the most commonly used AOPs for SWW treatment. Although these processes are usually expensive if applied alone, removal efficiencies of over 90% can be achieved for SWW secon‐ dary effluents in terms of TOC and COD as a posttreatment method. Thus, the combination of biological processes and AOPs is recommended for SWW treatment [42, 43].

#### **5.5. Combined processes**

meat processing effluents comprise anaerobic baffled reactor (ABR), anaerobic digester (AD), anaerobic filter (AF), anaerobic lagoon (AL), septic tanks (ST), and up-flow anaerobic sludge

Nevertheless, anaerobic treatment barely complies with current discharge limits. Complete stabilization of the organic compounds is difficult due to the high organic strength of SWW. Therefore, an additional treatment stage is recommended to remove the organics, nutrients, and pathogens that remain after anaerobic treatment. On the other hand, anaerobic treatment requires a higher space and a higher residence time to achieve high overall treatment efficiency, affecting the economic viability of anaerobic treatment alone. Accordingly, the combination of anaerobic and aerobic processes is necessary to achieve a maximum efficiency for the treat-

Aerobic processes are frequently employed for nutrient removal and further treatment after primary treatment. The required oxygen and treatment time are directly related to the strength of the SWW, which makes it inadequate as primary treatment of SWW but adequate after

There are many advantages of using aerobic wastewater treatment processes, including low odor production, fast biological growth rate, and rapid adjustments to the temperature and loading rate changes. Conversely, the operating costs of aerobic systems are higher than those for anaerobic systems due to the maintenance and energy requirements for artificial oxygenation. There are different aerobic unit operations for SWW treatment, such as aerobic AS,

Constructed wetlands (CWs) emulate the degradation mechanisms of natural wetlands for water decontamination, integrating biological and physicochemical processes from the interaction of vegetation, soil, microorganisms, and atmosphere for the adsorption, biodegra-

The performance of CW systems for the treatment of SWW has been evaluated using both horizontal and vertical subsurface flow CWs. Results have shown a wide range of organic and nutrient removal for different vegetation with encouraging maximum removals of 99, 97, 85, and 78% for BOD, COD, TSS, and TN, respectively [40]. As a result, CWs are simple methods with low operation and maintenance costs and few negative impacts on the environment,

AOPs are an interesting complementary treatment option for primary or secondary treatment of SWW, showing excellent overall treatment efficiencies for water reuse. AOPs are diverse

O2

, UV/O₃

, and

rotating biological contactors (RBCs), and sequencing batch reactors (SBRs) [39].

dation, filtration, photooxidation, and sedimentation of organics and nutrients.

which make them an attractive alternative to conventional treatment [41].

and include gamma radiation, ozonation, ultrasound technology (UST), UV/H2

blanket (UASB) [30].

160 Physico-Chemical Wastewater Treatment and Resource Recovery

ment of SWW [37].

*5.3.2. Aerobic treatment*

anaerobic treatment [38].

*5.3.3. Constructed wetlands*

**5.4. Advanced oxidation processes**

The implementation of combined processes is operationally and economically beneficial for SWW treatment since it couples the advantages of different technologies to treat high‐strength industrial wastewaters. The combined ABR‐AS‐UV/H2 O2 system is recognized as a cost‐ effective solution for SWW treatment with removal efficiencies of over 95% for organics and nutrients at optimum operating conditions [6, 9, 10].

An overview of the state‐of‐the‐art technologies for SWW treatment, during the last two years, is presented in **Table 3**. Particular attention is given to organic and nutrient removal, in terms of bulk parameters such as BOD, COD, TOC, TN, and TP. As shown in **Table 3**, SWW treatment efficiencies vary extensively and depend on the SWW characteristics, the treatment time, and the influent concentration, as well as the type of treatment and BAT to comply with current regulations [7].



**Table 3.** Comparison of different slaughterhouse wastewater treatment methods.

#### **6. Case study**

Actual SWW samples with average concentrations of 1950, 1400, 850, 750, 200, and 40 mg/L for COD, BOD, TOC, TSS, TN, and TP, respectively, were taken from selected licensed MPPs in Ontario, Canada [62]. Anaerobic and aerobic sludge inocula in concentrations of 40,000 and 3000 mg/L, respectively, were obtained from the Ashbridges Bay Municipal Wastewater Treatment Plant in Toronto, Canada. The inocula were acclimatized in a period of 60 days.

The combined ABR‐AS‐UV/H2 O2 system consisted of a 36‐L ABR with five equal‐volume chambers and individual biogas collection, a 12.65‐L aerobic AS reactor with controlled air flow to maintain DO concentrations of 2 mg/L, and a 1.35‐L UV‐C photoreactor with recycle, output power of 6 W, and uniform light distribution (**Figure 1**).

**Figure 1.** Schematic diagram of the combined ABR‐AS‐UV/H2 O2 system for SWW treatment.

Bulk parameters including BOD, COD, TOC, TN, TP, and TSS were analyzed as the main parameters for the treatment of an actual SWW. **Figure 2** shows the obtained maximum removal values of more than 99% for COD (**Figure 2a**), BOD (**Figure 2b**), TOC (**Figure 2c**), TSS (**Figure 2d**), TN (**Figure 2e**), and TP (**Figure 2f**) from the SWW by the combined ABR‐AS‐UV/ H2 O2 processes, operated in continuous mode.

**Method HRT**

**6. Case study**

The combined ABR‐AS‐UV/H2

**(h)**

**BODin (mg/L)**

162 Physico-Chemical Wastewater Treatment and Resource Recovery

**CODin (mg/L)**

**Table 3.** Comparison of different slaughterhouse wastewater treatment methods.

O2

output power of 6 W, and uniform light distribution (**Figure 1**).

**Figure 1.** Schematic diagram of the combined ABR‐AS‐UV/H2

**TOCin (mg/L)** **TNin (mg/L)**

SBR 12 4240 6057 1436 576 – 93 – 93 [56] AOP 1 – 406 – – – 84 – – [57] SBR 3 – 8604 – 1493 – 80 – 88 [58] UST‐AF‐UF 343 – 5200 – 74 – 96 – – [59] AOP 2 340 – 94 55 – – 81 – [42] UF‐RO 160 – 7970 – – – 80 – – [60] AD‐AOP 723 658 1494 513 181 98 98 97 31 [43] ABR‐AS‐AOP 55 1635 2000 1200 841 100 99 100 85 [6] ABR‐AS‐AOP 10 1831 2043 1691 866 100 99 100 90 [61]

Actual SWW samples with average concentrations of 1950, 1400, 850, 750, 200, and 40 mg/L for COD, BOD, TOC, TSS, TN, and TP, respectively, were taken from selected licensed MPPs in Ontario, Canada [62]. Anaerobic and aerobic sludge inocula in concentrations of 40,000 and 3000 mg/L, respectively, were obtained from the Ashbridges Bay Municipal Wastewater Treatment Plant in Toronto, Canada. The inocula were acclimatized in a period of 60 days.

chambers and individual biogas collection, a 12.65‐L aerobic AS reactor with controlled air flow to maintain DO concentrations of 2 mg/L, and a 1.35‐L UV‐C photoreactor with recycle,

O2

system for SWW treatment.

**BODrem (%)**

**CODrem (%)**

system consisted of a 36‐L ABR with five equal‐volume

**TOCrem (%)**

**TNrem (%)**

**Ref.**

**Figure 2.** Maximum removal values of (a) COD, (b) BOD, (c) TOC, (d) TN, (e) TP, and (f) TSS from an actual slaughter‐ house wastewater using combined ABR‐AS‐UV/H2 O2 processes.

The ABR process alone achieved high TSS and TN removals, providing an effluent that complies with most of the current standards worldwide (**Table 2**), with concentrations of 15 and 8 mg/L, respectively. A further treatment with the aerobic AS bioreactor was required to achieve high BOD and TP removals reaching concentrations of 14 and 0.04 mg/L, respectively. However, the COD and TOC concentrations, which are not included broadly as standard parameters, remain with considerable concentrations of 132 and 128 mg/L, respectively. These concentrations are more related to nonbiodegradable organics that can be mineralized using AOPs as a posttreatment process. Thus, after the treatment by the UV/H2 O2 process, the effluent concentrations for COD and TOC reached values of less than 0.4 and 0.1 mg/L, respectively, which could be used for water reuse.

The effects of the influent TOC concentration, flow rate, and pH on the TOC and TN remov‐ als, H2 O2 residual, and CH₄ yield in the combined ABR‐AS‐UV/H2 O2 system were also evalu‐ ated. **Figure 3** shows that the influent TOC concentration and the flow rate are inversely proportional to both TOC and TN removals. On the other hand, results indicate that an optimum TOC concentration with no pH adjustment and low flow rate are required to achieve a minimum H2 O2 residual in the effluent. Finally, results also demonstrate that a high influent TOC concentration is needed to achieve a maximum CH₄ yield with an optimum flow rate and no pH adjustments.

**Figure 3.** Individual effect of the (a) influent concentration of TOC, (b) flow rate, and (c) pH on the TOC removal, TN removal, H2 O2 residual, and CH₄ yield. Dashed lines represent model predicted values, whereas marker points repre‐ sent experimental values. The y‐axis is in percentage (%) units, except H2 O2 residual, which is expressed in per mil (%) for scaling purposes.

As a final point, the treatment costs per volume for the individual ABR, AS, and UV/H2 O2 processes were compared with those of the combined ABR‐AS‐UV/H2 O2 system for the treatment of an actual SWW and plotted versus the TOC removal for each configuration (**Figure 4**). Consequently, a minimum overall treatment cost of 0.12 \$/m3 for a maximum TOC removal of more than 90 % can be achieved at optimum operating conditions in the combined ABR‐AS‐UV/H2 O2 system for the treatment of an actual SWW.

The effects of the influent TOC concentration, flow rate, and pH on the TOC and TN remov‐

yield in the combined ABR‐AS‐UV/H2

ated. **Figure 3** shows that the influent TOC concentration and the flow rate are inversely proportional to both TOC and TN removals. On the other hand, results indicate that an optimum TOC concentration with no pH adjustment and low flow rate are required to achieve

**Figure 3.** Individual effect of the (a) influent concentration of TOC, (b) flow rate, and (c) pH on the TOC removal, TN

As a final point, the treatment costs per volume for the individual ABR, AS, and UV/H2

treatment of an actual SWW and plotted versus the TOC removal for each configuration

removal of more than 90 % can be achieved at optimum operating conditions in the combined

processes were compared with those of the combined ABR‐AS‐UV/H2

system for the treatment of an actual SWW.

(**Figure 4**). Consequently, a minimum overall treatment cost of 0.12 \$/m3

yield. Dashed lines represent model predicted values, whereas marker points repre‐

residual, which is expressed in per mil (%)

O2

O2

system for the

for a maximum TOC

O2

residual in the effluent. Finally, results also demonstrate that a high influent

O2

system were also evalu‐

yield with an optimum flow rate and

als, H2 O2

a minimum H2

removal, H2

O2

for scaling purposes.

ABR‐AS‐UV/H2

residual, and CH₄

O2

sent experimental values. The y‐axis is in percentage (%) units, except H2

no pH adjustments.

residual, and CH₄

164 Physico-Chemical Wastewater Treatment and Resource Recovery

O2

TOC concentration is needed to achieve a maximum CH₄

**Figure 4.** Comparison of operational costs per cubic meter of treated actual slaughterhouse wastewater using individu‐ al ABR, AS, and UV/H2 O2 processes and the combined ABR‐AS‐UV/H2 O2 system.

## **7. Slaughterhouse wastewater management and resource recovery**

The meat processing industry needs to incorporate both waste minimization and resource recovery into SWW management strategies considering the portion of the industry's waste and by‐products that have a potential of recovery for direct reuse, including nutrients and methane as biofuel. **Figure 5** presents a schematic illustration of the ideal operation of a meat processing plant and supply chain from the animal farming and raw materials to the final product, waste disposal, and recoverable resources [27, 63].

**Figure 5.** Schematic illustration of the ideal operation and supply chain of a meat processing plant.

A cleaner production should be the focus of meat processing plants due to the increasing interest in environmental initiatives and demands for green practices. Thus, it is appropriate to classify and minimize waste generation at the source, and on‐site treatment is the preferred option for water reuse and potential energy recovery. As a result, there are some considerations to be made for the adequate treatment of SWW effluents. **Figure 6** presents a proposed layout of the pretreatment, treatment, and disinfection of slaughterhouse wastes for a typical meat processing plant, as well as the potential resource recovery for water reuse and products recycling [63, 64].

**Figure 6.** Proposed layout of the pretreatment, treatment, and disinfection of slaughterhouse wastes for a typical meat processing plant.

#### **8. Conclusions**

Meat processing effluents are usually pretreated using screeners, settlers, and blood collection systems, followed by physicochemical treatment methods, such as coagulation, flocculation, sedimentation, DAF, or secondary biological treatment. Although biological treatment is able to provide high organic and nutrient removal efficiencies, further treatment by AOPs, or other BAT, is required for a high‐quality effluent.

The presented case study provided an example of the application of a combined ABR‐AS‐ UV/H2 O2 system for the treatment of an actual SWW. Maximum organic and nutrient removal reached over 90% in terms of TOC and TN, respectively. Moreover, a potential resource recovery achieved a maximum CH₄ yield of up to 56%, and minimization of residual by‐ products from disinfection was attained in terms of H2 O2 residual of less than 2% at the effluent. Finally, the cost‐effectiveness analysis found a minimum overall treatment cost of 0.12 \$/m3 for the treatment of an actual SWW using the combined ABR‐AS‐UV/H2 O2 system at optimum operating conditions.

All types of waste, liquid, solid, or gaseous, must be treated prior to their release into the environment. Whereas the use of recoverable resources is recommended as a feasible and practical alternative to conventional energy sources in the long term, costs associated to the application of these technologies will be offset by the reduction in local electricity consumption and by‐product recycling and reuse. Thus, the potential of biogas production as an energy source, the use of fertilizers from nutrient recovery, and the SWW high‐quality treated effluents for water reuse are to be considered toward a sustainable and cleaner production in the meat processing industry.

Consequently, the use of combined processes as an alternative to conventional methods has become a cost‐effective approach for the treatment of meat processing effluents to comply with applicable current regulations worldwide.

## **Acknowledgements**

The financial support of the Natural Sciences and Engineering Research Council of Canada (NSERC), the Ontario Trillium Scholarship (OTS) program, Colciencias (Colombian Research Funding Agency), and Ryerson University is greatly appreciated.

#### **Nomenclature**

A cleaner production should be the focus of meat processing plants due to the increasing interest in environmental initiatives and demands for green practices. Thus, it is appropriate to classify and minimize waste generation at the source, and on‐site treatment is the preferred option for water reuse and potential energy recovery. As a result, there are some considerations to be made for the adequate treatment of SWW effluents. **Figure 6** presents a proposed layout of the pretreatment, treatment, and disinfection of slaughterhouse wastes for a typical meat processing plant, as well as the potential resource recovery for water reuse and products

**Figure 6.** Proposed layout of the pretreatment, treatment, and disinfection of slaughterhouse wastes for a typical meat

Meat processing effluents are usually pretreated using screeners, settlers, and blood collection systems, followed by physicochemical treatment methods, such as coagulation, flocculation, sedimentation, DAF, or secondary biological treatment. Although biological treatment is able to provide high organic and nutrient removal efficiencies, further treatment by AOPs, or other

recycling [63, 64].

166 Physico-Chemical Wastewater Treatment and Resource Recovery

processing plant.

**8. Conclusions**

BAT, is required for a high‐quality effluent.



## **Author details**

Ciro Bustillo-Lecompte and Mehrab Mehrvar\*

\*Address all correspondence to: mmehrvar@ryerson.ca

Ryerson University, Toronto, Canada


[3] Valta K., Kosanovic T., Malamis D., Moustakas K., Loizidou M. Overview of water usage and wastewater management in the food and beverage industry. Desalination and Water Treatment. 2015;53(12):3347. DOI: 10.1080/19443994.2014.934100

DAF dissolved air floatation DO dissolved oxygen EC electrocoagulation EU European Union MF microfiltration

168 Physico-Chemical Wastewater Treatment and Resource Recovery

MPP meat processing plant

SBR sequencing batch reactor

SWW slaughterhouse wastewater

UASB up-flow anaerobic sludge blanket

UST ultrasound technology

US EPA United States Environmental Protection Agency

Ciro Bustillo-Lecompte and Mehrab Mehrvar\*

Ryerson University, Toronto, Canada

\*Address all correspondence to: mmehrvar@ryerson.ca

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TOC total organic carbon TN total nitrogen TP total phosphorus TSS total suspended solids

UF ultrafiltration

**Author details**

**References**

NF nanofiltration RO reverse osmosis

ST septic tank


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## **Treatment of Antibiotics in Wastewater Using Advanced Oxidation Processes (AOPs)** Provisional chapter

Treatment of Antibiotics in Wastewater Using Advanced

Ayse Kurt, Berna Kiril Mert, Nihan Özengin,

Oxidation Processes (AOPs)

Özge Sivrioğlu and Taner Yonar Ayse Kurt, Berna Kiril Mert, Nihan Özengin,

Additional information is available at the end of the chapter Özge Sivrioğlu and Taner Yonar

http://dx.doi.org/10.5772/67538 Additional information is available at the end of the chapter

#### Abstract

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174 Physico-Chemical Wastewater Treatment and Resource Recovery

651-666. DOI: 10.1016/j.jenvman.2016.07.044

j.ijpe.2016.07.022 (in press).

Antibiotics are nonbiodegradable, can survive at aquatic environments for long periods and they have a big potential bio-accumulation in the environment. They are extensively metabolized by humans, animals and plants. After metabolization, antibiotics or their metabolites are excreted into the aquatic environment. Removal of these compounds from the aquatic environment is feasible by different processes. But antibiotics are not treated in conventional wastewater treatment plants efficiently. During the last years studies with advanced oxidation processes (AOPs) for removal of these pharmaceuticals from waters has shown that they can be useful for removing them fully. Advanced oxidation processes (AOPs) can work as alternatives or complementary method in traditional wastewater treatment, and highly reactive free radicals, especially hydroxyl radicals (OH) generated via chemical (O3/H2O2, O3/OH- ), photochemical (UV/O3, O3/H2O2) reactions, serve as the main oxidant. This study presents an overview of the literature on antibiotics and their removal from water by advanced oxidation processes. It includes almost all types of antibiotics which are consumed by human and veterinary processes. It was found that most of the investigated advanced oxidation treatment processes for the oxidation of antibiotics in water are direct and indirect photolysis with the combinations of H2O2, TiO2, ozone and Fenton's reagent.

Keywords: wastewater, antibiotics, endocrine disrupter, advanced oxidation

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

© The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons

Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and eproduction in any medium, provided the original work is properly cited.

## 1. Introduction

The "antibiotic" term qua generic is used to specify any class of organic molecule that blocks or ravage microbes by specific interactions with bacterial marks, without considering any compound or class [1]. Antibiotics are designed to act very effectively even at low doses and, in case of intracorporal administration, to be completely excreted from the body after a short time of residence [2]. They are nonbiodegradable and can survive in aquatic environments for long periods [3]. The entrance of these compounds into the environment owing to anthropogenic sources can result in a potential risk for organisms. Although antibiotics exist at residual levels, they can cause resistance in bacterial populations, making them inactive in the treatment of several diseases in the near future [4, 5]. And they cause endocrine-disrupting effects when they are consumed by living organisms. They interfere with the synthesis, secretion, transport, binding, action, and elimination of hormones in the human body [6].

The annual usages of antibiotics are determined between 100,000 and 200,000 t globally [7]. Traditionally, these compounds were not accepted as environmental contaminants, but their existence in the aquatic ecosystems has become an apprehension as biological impacts and potential threat to the environment [8–10]. Furthermore, it has been shown up that residual antibiotics are able to support the election of genetic variants of microorganisms concluding in the existence of antibiotic-resistant pathogens [11, 12].

Removal of these compounds from the aquatic environment is feasible by different processes. This can be carried out using biotic (biodegradation) or nonbiotic (chemical oxidation and advanced oxidation) ways. But antibiotics are not treated in conventional wastewater treatment plants efficiently. During the last years, studies with advanced oxidation processes (AOPs) for removal of these pharmaceuticals from waters have shown that they can be useful for removing them completely. In this chapter, we aim to introduce a review of literature on antibiotics and their removal from water by advanced oxidation processes. An effort to include as many studies as possible was made in order to highlight important findings and present the knowledge currently available on the removal efficiency of antibiotics from wastewater.

## 2. General description of antibiotics

Antibiotic as a word is reproduced from the Greek anti (¼against) and biotikos (¼living).

Most of the living organisms are able to compose matters that can influence other organisms' capacity for growth, endurance, and reproduction. Microorganisms have a versatile ability to inhibit the growth and purpose of other microorganisms and produce and release biologically effective substances at the appropriate moment. We denominate substances of this kind as antibiotics [13].

In addition, organisms' ability to compose antibiotics has been of great importance for the development of different life forms and their capability to accommodate to new circumambient. Nowadays, antibiotics are important components for the functions of various biological systems.

With the development of synthetic antibiotics, a large number of substances with specific areas of application have been given to access. Today, the level of usage of synthetic antibiotics and their effects to the environment are at a critical rate. Since antibiotics are bioavailable, they can show long-term biological effects in the environment.

1. Introduction

176 Physico-Chemical Wastewater Treatment and Resource Recovery

The "antibiotic" term qua generic is used to specify any class of organic molecule that blocks or ravage microbes by specific interactions with bacterial marks, without considering any compound or class [1]. Antibiotics are designed to act very effectively even at low doses and, in case of intracorporal administration, to be completely excreted from the body after a short time of residence [2]. They are nonbiodegradable and can survive in aquatic environments for long periods [3]. The entrance of these compounds into the environment owing to anthropogenic sources can result in a potential risk for organisms. Although antibiotics exist at residual levels, they can cause resistance in bacterial populations, making them inactive in the treatment of several diseases in the near future [4, 5]. And they cause endocrine-disrupting effects when they are consumed by living organisms. They interfere with the synthesis, secretion, transport,

The annual usages of antibiotics are determined between 100,000 and 200,000 t globally [7]. Traditionally, these compounds were not accepted as environmental contaminants, but their existence in the aquatic ecosystems has become an apprehension as biological impacts and potential threat to the environment [8–10]. Furthermore, it has been shown up that residual antibiotics are able to support the election of genetic variants of microorganisms concluding in

Removal of these compounds from the aquatic environment is feasible by different processes. This can be carried out using biotic (biodegradation) or nonbiotic (chemical oxidation and advanced oxidation) ways. But antibiotics are not treated in conventional wastewater treatment plants efficiently. During the last years, studies with advanced oxidation processes (AOPs) for removal of these pharmaceuticals from waters have shown that they can be useful for removing them completely. In this chapter, we aim to introduce a review of literature on antibiotics and their removal from water by advanced oxidation processes. An effort to include as many studies as possible was made in order to highlight important findings and present the knowledge currently available on the removal efficiency of antibiotics from wastewater.

Antibiotic as a word is reproduced from the Greek anti (¼against) and biotikos (¼living).

Most of the living organisms are able to compose matters that can influence other organisms' capacity for growth, endurance, and reproduction. Microorganisms have a versatile ability to inhibit the growth and purpose of other microorganisms and produce and release biologically effective substances at the appropriate moment. We denominate substances of this kind as

In addition, organisms' ability to compose antibiotics has been of great importance for the development of different life forms and their capability to accommodate to new circumambient. Nowadays, antibiotics are important components for the functions of various biological

binding, action, and elimination of hormones in the human body [6].

the existence of antibiotic-resistant pathogens [11, 12].

2. General description of antibiotics

antibiotics [13].

systems.

Antibiotics can be grouped according to their chemical structure or mechanism of action. There are various groups of chemicals that can be arranged to different subgroups, such as ß-lactams, quinolones, tetracyclines, macrolides, sulphonamides, and others. They are complicated molecules, which may have different functionalities within the same molecule. Consequently, they act as neutral, cationic, anionic, or zwitterionic under different pH conditions. Owing to different functionalities in a single molecule, their physico-chemical and biological properties (like octanol-water partition coefficients (log Pow), sorption behavior, photoreactivity and antibiotic activity, and toxicity) may change with pH [14].

b-Lactam antibiotics contain cefradine, amoxicillin ceftriaxone, sultamicillin, and penicillins G and VK. Actually, these antibiotics have been insulated from molds and have been adapted to obtain different physicochemical and pharmacological properties [15]. They suppress bacterial cell wall synthesis.

Sulfonamides are synthetic antibiotics, and they inhibit generation of bacteria by behaving as competitive inhibitors of p-aminobenzoic acid in the folic acid metabolism cycle [15]. A diversity of sulfonamides have been developed, consumed, and finally detected in the environment, and some of them have been studied for their degradation by ozonation and AOPs. These compounds include sulfadiazine, sulfadimethoxine, sulfachlopyridazine, sulfamethazine, sulfamethizole, sulfamethoxazole, sulfisoxazole, sulfapyridine, sulfathiazole, sulfamoxole, and sulfamerazine.

Most common quinolone antibiotics are enrofloxacin and ofloxacin. They have been examined in terms of their degradation by ozonation and AOPs. More particularly, both are fluoroquinolones. While enrofloxacin is utilized as a veterinary antibiotic, despite that ofloxacin is designed for human uptake. These compounds have a benefit to suppress the activity of bacterial DNA gyrase. It is known that the quinolones are metabolized in the liver and eliminated in the urine [15, 16].

When we examine other antibiotics, clarithromycin, azithromycin, erythromycin, and roxithromycin are macrolide antibiotics, and lincomycin is a lincosamide antibiotic. These antibiotics are described by a property to inhibit bacterial protein synthesis. These antibiotics are mostly eliminated in the bile [15, 16].

Antibiotic is a chemotherapeutic agent that inhibits the growth of microorganisms (bacteria, fungi, protozoa, or viruses) even at very low concentrations. They are nonbiodegradable and can survive in aquatic environments for long periods. So they can bio-accumulate in the environment [3]. Also antibiotics in the environment may contribute to the emergence of antibiotic-resistant bacteria [17]. And they cause endocrine-disrupting effects when they are consumed by living organisms. They interfere with the synthesis, secretion, transport, binding, action, and elimination of hormones in the human body [6].

Releasing of antibiotics into the aquatic environment by human beings and animals depends mainly on the consumption rates of antibiotics [14, 18]. According to the investigations, some antibiotics have toxic effects on humans, animals, and also microorganisms even at low concentrations. At the same time, they are nonbiodegradable and can survive in the environment even in the conventional wastewater treatments. So they cause bio-accumulation [3]. Therefore, the presence of antibiotics in the environment can cause the occurrence of antibiotic-resistant bacteria [17]. They are not treated in the conventional wastewater treatment plants completely. According to the recent studies, advanced oxidation processes (AOPs) are useful to remove these toxic compounds completely from waters.

#### 2.1. Sources of antibiotics in the environment

In these last years, the use of antibiotics in veterinary and human medicine has been widespread, and consequently, the possibility of water contamination with such compounds has been increased [19]. These pollutants are continually discharged into the natural environment as parent compounds, metabolites/degradation products, or both forms by a diversity of input sources as shown in Figure 1 [5].

Fertilizers present in the fields can contaminate soil and consequently surface and groundwater through runoff or filtration [20]. Likewise, human antibiotics which are present into the environment through discharge, entering in the sewage and reaching the Waste-water treatment plants (WWTP). Despite most of WWTPs are not projected to remove highly polar micropollutants [19], they can be transferred to surface waters and reach groundwater after leaching.

The sludge produced in WWTPs is utilized as soil manure and can cause problems when used as a fertilizer. Some other significant pollution source is the direct delivery of veterinary antibiotics through the implementation in aquaculture. Inappropriate elimination of unused/expired

Figure 1. Pathways of antibiotics in the environment [5, 16].

drugs can also be considered as significant points of contamination. These are derived directly from sewage discharge or landfills deposition, waste effluents from manufacture, or accidental spills during manufacturing or distribution [5].

#### 2.2. Occurrence

antibiotics have toxic effects on humans, animals, and also microorganisms even at low concentrations. At the same time, they are nonbiodegradable and can survive in the environment even in the conventional wastewater treatments. So they cause bio-accumulation [3]. Therefore, the presence of antibiotics in the environment can cause the occurrence of antibiotic-resistant bacteria [17]. They are not treated in the conventional wastewater treatment plants completely. According to the recent studies, advanced oxidation processes (AOPs) are

In these last years, the use of antibiotics in veterinary and human medicine has been widespread, and consequently, the possibility of water contamination with such compounds has been increased [19]. These pollutants are continually discharged into the natural environment as parent compounds, metabolites/degradation products, or both forms by a diversity of input

Fertilizers present in the fields can contaminate soil and consequently surface and groundwater through runoff or filtration [20]. Likewise, human antibiotics which are present into the environment through discharge, entering in the sewage and reaching the Waste-water treatment plants (WWTP). Despite most of WWTPs are not projected to remove highly polar micropollutants [19],

The sludge produced in WWTPs is utilized as soil manure and can cause problems when used as a fertilizer. Some other significant pollution source is the direct delivery of veterinary antibiotics through the implementation in aquaculture. Inappropriate elimination of unused/expired

they can be transferred to surface waters and reach groundwater after leaching.

useful to remove these toxic compounds completely from waters.

2.1. Sources of antibiotics in the environment

178 Physico-Chemical Wastewater Treatment and Resource Recovery

Figure 1. Pathways of antibiotics in the environment [5, 16].

sources as shown in Figure 1 [5].

Research has quite extensively studied the presence of antibiotics in the environment (for a short overview, see Table 1). With respect to other pharmaceuticals, the concentrations of antibiotics measured in different countries were found in the same range of concentrations in the different compartments [14, 21–23]. The antibiotic groups that have been analyzed up to now include a number of different important classes of antibiotics. They include primarily macrolides, aminoglycosides, tetracyclines, sulfonamides, sulfanilamides, and quinolones to name just a few [14, 23–29].




\*Directly impacted by surface water.

Antibiotic Sewage treatment plant

180 Physico-Chemical Wastewater Treatment and Resource Recovery

Chinolones

Sulfonamides

Tetracyclines

specified)

Tetracycline (no more

effluent (ng L<sup>1</sup>

)

Surface water (ng L<sup>1</sup> )

Fluorchinolone up to 100 up to 5 [32] Fluorchinolone up to 106 up to 19 [34, 35] Ciprofloxacin 9 [33]

Norfloxacin up to 120 [38] Ofloxacin up to 82 [30]

Sulfamethoxazole up to 370 [30]

Sulfamethazin up to 160 [8]

Sulfamethizole up to 130 [38] Sulfadiazine up to 17 [36] Sulfadimidine up to 23 [36]

Tetracycline up to 110 [38] Chlortetracycline up to 690 [38]

Oxytetracycline up to 340 [38]

up to 2000 up to 480 up to 470 [8]

up to 20 up to 1 [32]

up to 1000 up to 560 [8]

Ground water\*/bank filtrate (ng L<sup>1</sup>

up to 14 [33] up to 180 [38] up to 350 [31]

up to 30 [38] up to 26.2 [37] up to 1300 [31] up to 26 [39]

20 [33]

up to 52 [33] up to 1900 [38] up to 2000 [31]

up to 220 [38]

up to 7 [33]

up to 600 [31] up to 100 [38]

up to 19.2 [37]

)

References

Table 1. Examples of measured concentrations of antibiotics in the aquatic environment [14, 18, 30, 31].

## 3. Advanced oxidation processes (AOPs)

During the oxidation of organic contaminants, the ultimate goal is to produce simple, relatively harmless inorganic molecules [40]. Advanced oxidation processes are characterized by their production of the hydroxyl radical (�OH), a very strong oxidant, in sufficiently high concentrations to affect water quality. The symbol "�"represents the radical center, a single unpaired electron [41].

At optimum operation conditions, for instance sufficient contact time, it is possible to mineralize the target contaminant to CO2 and H2O, the most stable end products of chemical oxidation. For this reason, the extraordinary definition of AOPs on chemical processes is that they are completely described as "environmentally friendly" [42].

The basic treatment of AOPs can be explained in two steps: one is the generation of hydroxyl radicals and the other is the oxidative reaction of these radicals with molecules [43]. The dissolved organic pollutants can be converted into CO2 and H2O by AOPs. The generation of hydroxyl radical might be by the use of UV, UV/H2O2, UV/O3, TiO2/H2O2, Feþ<sup>2</sup> /H2O2 and one or two processes [44].

AOPs can be classified in two groups: (1) nonphotochemical AOPs and (2) photochemical AOPs. Nonphotochemical AOPs include cavitation, ozonation, Fenton and Fenton-like processes, wet air oxidation, ozone/hydrogen peroxide, etc. Photochemical oxidation processes include homogeneous and heterogeneous processes [45].

#### 3.1. Nonphotochemical oxidation processes

Nonphotochemical oxidation processes can be classified as follows: ozonation, peroxide, Fenton process, ozone/hydrogen, supercritical water oxidation, electrochemical oxidation, cavitation, gamma-ray, X-ray, electrical discharge-based nonthermal plasma, and electron beam.

#### 3.1.1. Ozonation

Ozone is a powerful oxidizer and has been increasingly used for the treatment of wastewater [46]. High pH values (>11.0) causes high efficiency and ozone behaves randomly with all organic and inorganic compositions present in the reacting medium [45]. Ozone reacts with substances in two different ways: indirect and direct. These two reaction pathways are managed by different type of kinetics and lead to different oxidation products [47].

Simplified reaction mechanism of ozone at high pH is given in below:

OH� 3O3 þ H2O ! 2OH• þ 4O2 ð1Þ

#### 3.1.2. Ozone/hydrogen peroxide (peroxone) process (O3/H2O2)

The principle of peroxonation is based on the coupling between ozone (O3) and H2O2, resulting in the generation of oxidizing radicals. As pointed out by Zaviska et al. [47], the peroxonation mechanism could be more productive than ozonation alone, and H2O2 impacts on increasing the decomposition percentage of O3 in water, which generates a larger number of very reactive •OH radicals [49]. Because of the high cost of ozone generation, this combination makes the process economically feasible [50]. Several factors limit the usefulness of the peroxonation process such as important energetic consumption, low water solubility of ozone, and its sensitivity to several factors [51]. A general mechanism of peroxon process is given below:

$$\rm H\_2O\_2 + 2O\_3 \to 2OH^\cdot + 3O\_2 \tag{2}$$

Solution pH is critical as well for the process output like other AOPs. Higher production rates of hydroxyl radicals will be obtained by the addition of hydrogen peroxide to the aqueous O3 solution at high pH conditions. Independence of peroxone process from any light source or UV delivers a certain benefits to this operation [44].

#### 3.1.3. Fenton process

Fenton's reaction is known as the dark reaction of ferrous iron (Fe(II)) with H2O2 (Eq.(15)) [6]. •OH radical is generated through the agency of reaction between H2O2 and Feþ<sup>2</sup> salts as described below.

$$\text{Fe}^{+2} + \text{H}\_2\text{O}\_2 \rightarrow \text{Fe}^{+3} + \text{OH}^- + \text{\textdegree OH} \tag{3}$$

Thus, composed hydroxyl radical can react with Fe(II) to develop ferric ion (Fe(III)) (Eq. (16))

$$\cdot \text{OH} + \text{Fe}^{+2} \rightarrow \text{Fe}^{+3} + \text{OH}^{-} \tag{4}$$

As an alternative, hydroxyl radicals are able to react with organic pollutants and start oxidation in a waste stream,

$$\text{RH} + \cdot\\\text{OH} \rightarrow \text{R} \cdot + \text{H}\_2\text{O} \tag{5}$$

Reactions can result into the degradation of Feþ<sup>3</sup> to Feþ<sup>2</sup> at a value of pH between 2.7 and 2.8.

$$\text{Fe}^{2+} + \text{H}\_2\text{O}\_2 \rightarrow \text{Fe}^{3+} + \text{\text{\textdegree OH}} + \text{OH}^- \tag{6}$$

$$\text{Fe}^{3+} + \text{H}\_2\text{O}\_2 \rightarrow \text{Fe}^{2+} + \text{\textdegreeCOOH} + \text{H}^+ \tag{7}$$

In these circumstances, iron can be considered as a true catalyst [156].

Process efficiency is closely related to the solution pH whose optimal values are between 2 and 4 as well as the COD:H2O2:catalyst ratio in the feed [52].

Basically, the Fenton process possesses several important advantages for water/wastewater treatment [48, 53]:


AOPs can be classified in two groups: (1) nonphotochemical AOPs and (2) photochemical AOPs. Nonphotochemical AOPs include cavitation, ozonation, Fenton and Fenton-like processes, wet air oxidation, ozone/hydrogen peroxide, etc. Photochemical oxidation processes

Nonphotochemical oxidation processes can be classified as follows: ozonation, peroxide, Fenton process, ozone/hydrogen, supercritical water oxidation, electrochemical oxidation, cavitation, gamma-ray, X-ray, electrical discharge-based nonthermal plasma, and electron beam.

Ozone is a powerful oxidizer and has been increasingly used for the treatment of wastewater [46]. High pH values (>11.0) causes high efficiency and ozone behaves randomly with all organic and inorganic compositions present in the reacting medium [45]. Ozone reacts with substances in two different ways: indirect and direct. These two reaction pathways are man-

3O3 þ H2O ! 2OH• þ 4O2

The principle of peroxonation is based on the coupling between ozone (O3) and H2O2, resulting in the generation of oxidizing radicals. As pointed out by Zaviska et al. [47], the peroxonation mechanism could be more productive than ozonation alone, and H2O2 impacts on increasing the decomposition percentage of O3 in water, which generates a larger number of very reactive •OH radicals [49]. Because of the high cost of ozone generation, this combination makes the process economically feasible [50]. Several factors limit the usefulness of the peroxonation process such as important energetic consumption, low water solubility of ozone, and its sensitivity to several factors [51]. A general mechanism of peroxon process is given below:

Solution pH is critical as well for the process output like other AOPs. Higher production rates of hydroxyl radicals will be obtained by the addition of hydrogen peroxide to the aqueous O3 solution at high pH conditions. Independence of peroxone process from any light source or UV

Fenton's reaction is known as the dark reaction of ferrous iron (Fe(II)) with H2O2 (Eq.(15)) [6]. •OH radical is generated through the agency of reaction between H2O2 and Feþ<sup>2</sup> salts as

H2O2 þ 2O3 ! 2OH• þ 3O2 ð2Þ

ð1Þ

aged by different type of kinetics and lead to different oxidation products [47].

OH�

Simplified reaction mechanism of ozone at high pH is given in below:

3.1.2. Ozone/hydrogen peroxide (peroxone) process (O3/H2O2)

delivers a certain benefits to this operation [44].

3.1.3. Fenton process

described below.

include homogeneous and heterogeneous processes [45].

3.1. Nonphotochemical oxidation processes

182 Physico-Chemical Wastewater Treatment and Resource Recovery

3.1.1. Ozonation

#### 3.2. Photochemical oxidation processes

#### 3.2.1. Homogeneous photochemical oxidation processes

#### 3.2.1.1. Vacuum UV (VUV) photolysis

The vacuum ultraviolet (UV) is absorbed by all the materials from water to air, therefore can only be transferred in a vacuum. The absorption of a VUV photon causes breaking of one or more bond. As an example, water is decomposed by

$$\mathrm{H\_2O} + h\nu(<190\,\mathrm{nm}) \to \mathrm{H^{\bullet}} + \mathrm{HO^{\bullet}}\tag{8}$$

$$\mathrm{H\_2O} + h\nu(<190\,\mathrm{nm}) \rightarrow \mathrm{H} + + \mathrm{e} - + \mathrm{HO}\bullet \tag{9}$$

VUV photolysis has a high feasibility for the oxidative degradation of organic pollutants in water. In spectral domain (approx. 140–200 nm), it produces hydrogen atoms and hydroxyl radicals. Due to the high absorption cross-section of water and quantum yields of water homolysis of 0.45–0.3 at stimulation wavelengths (between 140 and 185 nm) provide productive local concentrations of hydroxyl and hydrogen radicals. VUV photolysis is a new technique for water treatment and suggests the benefit to generate unusually high local concentrations of oxidative reactive intermediates without the addition of supplementary oxidant [44, 54].

#### 3.2.1.2. Hydrogen peroxide/UV (H2O2/UV) process

Hydrogen peroxide can be photolyzed by UV radiations by producing the homolytic scission of the O–O bond of the H2O2 and resulting the formation of •OH radicals which can also be supplied to the decomposition of H2O2 by secondary reactions [48]. The main reaction is given below:

$$\text{H}\_2\text{O}\_2 + \text{hv} \to 2\text{HO}\bullet\tag{10}$$

UV/H2O2 process is effective in mineralizing organic pollutants. As an disadvantage the process cannot use solar light as the source of UV light owing to the fact that the required UV energy is not available in the solar spectrum [55]. Over and above, H2O2 has poor UV absorption characteristics. At last, special reactors designed for UV illumination are required [56].

The major factors influencing the process are the amount of H2O2 used, presence of bicarbonate, wastewater pH, the initial concentration of the object compound, and reaction time [57].

#### 3.2.1.3. Ozone/UV (O3/UV) process

The advanced oxidation process with ozone and UV radiation is initiated by the photolysis of ozone. Hydroxyl radicals can be composed by those in hydrogen peroxide under UV and/or ozone. The equations are given below:

$$\rm O\_3 + hv + H\_2O \to H\_2O\_2 + O\_2 \tag{11}$$

$$\text{H}\_2\text{O}\_2 + hv \rightarrow 2\cdot\text{OH} \tag{12}$$

$$\text{2O}\_3 + \text{H}\_2\text{O}\_2 \to \text{2} \cdot \text{OH} + \text{3O}\_2 \tag{13}$$

All kinds of UV light origins can be utilized for this process, especially low-pressure mercury vapor lamps. The O3/UV process does not have the same limitations as that of H2O2/UV process. Many variables (temperature, pH, UV intensity, tubidity, lamp spectral characteristics, and pollutant type, etc.) affect the performance of the system [42, 44].

#### 3.2.1.4. Ozone/hydrogen peroxide/UV (O3/H2O2/UV) process

This method is considered to be the most effective and powerful method, which provides a fast and complete mineralization of pollutants [42, 50]. The addition of H2O2 to the O3/UV process accelerates the decomposition of ozone, which results in an increased rate of ·OH generation.

The main short mechanism of the O3/H2O2/UV process is given below:

$$2\text{O}\_3 + \text{H}\_2\text{O}\_2 \xrightarrow{\text{UV}} 2\text{ HO}\cdot + 3\text{O}\_2\tag{14}$$

The capital and operating costs for the system vary widely depending on the wastewater flow rate, types, and concentrations of contaminants present and the degree of removal required [58].

#### 3.2.1.5. Photo-Fenton process

treatment and suggests the benefit to generate unusually high local concentrations of oxidative

Hydrogen peroxide can be photolyzed by UV radiations by producing the homolytic scission of the O–O bond of the H2O2 and resulting the formation of •OH radicals which can also be supplied to the decomposition of H2O2 by secondary reactions [48]. The main reaction is given below:

UV/H2O2 process is effective in mineralizing organic pollutants. As an disadvantage the process cannot use solar light as the source of UV light owing to the fact that the required UV energy is not available in the solar spectrum [55]. Over and above, H2O2 has poor UV absorption characteristics. At last, special reactors designed for UV illumination are required [56].

The major factors influencing the process are the amount of H2O2 used, presence of bicarbonate, wastewater pH, the initial concentration of the object compound, and reaction time [57].

The advanced oxidation process with ozone and UV radiation is initiated by the photolysis of ozone. Hydroxyl radicals can be composed by those in hydrogen peroxide under UV and/or

All kinds of UV light origins can be utilized for this process, especially low-pressure mercury vapor lamps. The O3/UV process does not have the same limitations as that of H2O2/UV process. Many variables (temperature, pH, UV intensity, tubidity, lamp spectral characteristics,

This method is considered to be the most effective and powerful method, which provides a fast and complete mineralization of pollutants [42, 50]. The addition of H2O2 to the O3/UV process accelerates the decomposition of ozone, which results in an increased rate of ·OH generation.

The capital and operating costs for the system vary widely depending on the wastewater flow rate, types, and concentrations of contaminants present and the degree of removal required [58].

and pollutant type, etc.) affect the performance of the system [42, 44].

The main short mechanism of the O3/H2O2/UV process is given below:

2O3 þ H2O2 !

3.2.1.4. Ozone/hydrogen peroxide/UV (O3/H2O2/UV) process

H2O2 þ hν ! 2HO• ð10Þ

O3 þ hv þ H2O ! H2O2 þ O2 ð11Þ

2O3 þ H2O2 ! 2 �OH þ 3O2 ð13Þ

H2O2 þ hv ! 2 �OH ð12Þ

UV 2 HO � þ 3O2 <sup>ð</sup>14<sup>Þ</sup>

reactive intermediates without the addition of supplementary oxidant [44, 54].

3.2.1.2. Hydrogen peroxide/UV (H2O2/UV) process

184 Physico-Chemical Wastewater Treatment and Resource Recovery

3.2.1.3. Ozone/UV (O3/UV) process

ozone. The equations are given below:

The Photo-Fenton process occurs by the combination of H2O2 and UV radiation with Fe(II) or Fe(III). The main factor of the mechanism is that iron salts act as photocatalysts and H2O2 as an oxidizing agent. It offers a productive and cheap method for wastewater treatment and produces hydroxyl radicals to a greater extent [59].

The reaction is given below:

$$\text{Fe(OH)}^{+2} + h\upsilon \rightarrow \text{Fe}^{+3} + \text{HO}\cdot \text{} \tag{15}$$

A highly low reaction time is required for the photo-Fenton process, and depending on the operating pH value, the concentrations of H2O2 and iron are added.

#### 3.2.2. Heterogeneous photochemical oxidation processes

Widely investigated and applied heterogeneous photochemical oxidation processes are semiconductor-sensitized photochemical oxidation processes.

A semiconductor consists of two energy bands: one is high energy conduction and the other is low energy valence band. This kind of photolytic chemical oxidation is used for the generation of OH radical in heterogeneous processes. Zinc oxide, strontium titanium trioxide, and TiO2 have been used for commercial implementation. Valance and conduction bands of a semiconductor material are distinguished by energy gap/band gap [60].

Moreover, the photocatalyst TiO2 is a wide band gap semiconductor (3.2 eV) and is successfully used as a photocatalyst for the treatment of organic pollutants [61, 62]. To summarize, in the TiO2 process, the photon energy given to achieve the band gap energy and to induce an electron into the transmission band from the valence band can be fed with a wavelength shorter than 387.5 nm. Clarified reaction mechanisms of TiO2/UV process are given below [Eq. (16)–(19)].

$$\text{TiO}\_2 + h\upsilon \to \text{e}^-\text{C}\_\text{CB} + h^+\text{v}\_\text{VB} \tag{16}$$

$$\mathrm{H\_2O} + h^+ \_{\mathrm{VB}} \rightarrow \mathrm{OH} \bullet + \mathrm{H}^+ \tag{17}$$

$$\text{O}\_2 + \text{e}^-\text{сB} \rightarrow \text{O}\_2\text{•}^-\tag{18}$$

$$\text{O}\_2\text{O}^- + \text{H}\_2\text{O} \rightarrow \text{OH}\bullet + \text{OH}^- + \text{O}\_2 + \text{HO}\_2^- \tag{19}$$

The basic reason of this reversal is the production of photons. The reversal mechanism importantly decreases the photocatalytic efficiency of a semiconductor. Main benefit of TiO2/UV process is low energy consumption thus sunlight can be utilized as a light source [44].

AOPs have been examined in terms of limitations and summarized below.

As an example, UV oxidation process with H2O2 is just effective at low wavelengths (especially under 200 nm).The treated aqueous flux must supply good transmission of UV light. Scavengers and high doses of chemical subscriptions may limit the process. Insoluble oil and grease, heavy metal ions, insoluble oil and grease, carbonates, and high alkalinity may cause clogging of the UV quartz handle. Air emission problems with O3 may arise. The cost of the AOPs is expensive when compared to rival technologies [42].

## 3.3. Assessment of AOPs performance for antibiotic removal

Ozonation and AOPs are required for efficient degradation of antibiotics in water and wastewater. These treatment processes have an advantage of elimination of such pollutants through mineralization or conversion to the products that are less harmful to human health and the aquatic environment.

Various studies have published the effective AOPs treatment for the removal of antibiotics from wastewater [63–67]. High-quality and effective publications relevant with the AOP studies on the mechanisms and applications of water and wastewater treatments have been pronounced for last two decades. From the theoretical, environmental, and economical point of view, they demonstrate a great and increasing interest. As shown in Table 2, several studies have been conducted on the applicability of AOPs on different antibiotic classes.

Ozone is a potent oxidant and has been progressively applied for the treatment of wastewater. Ozone and/or hydroxyl radicals passivate bactericidal characteristics of antibiotics by disrupting or modulating their pharmaceutically active functional groups, such as N-etheroxime and dimethylamino groups of macrolides [68, 69], aniline moieties of sulfonamides [70], thioether groups of penicillins, unsaturated bonds of cephalosporin, and the phenol ring of trimethoprim [69]. High removal rates (>90%) were achieved by ozonation of the compounds with electron-rich aromatic systems, such as hydroxyl, amino (e.g., sulfamethoxazole), acylamino, alkoxy, and alkyl aromatic compounds, as well as those compounds with deprotonated amine (e.g., erythromycin, ofloxacin, and trimethoprim) and nonaromatic alkene groups, since these structural moieties are highly amendable to oxidative attack [1]. Ozonation process was found to be effective for the removal of b-lactams, macrolides, sulfonamides, trimethoprim, quinolones, tetracyclines, and lincosamides [5].

The performance of ozone treatment can be improved providing ozone is combined with UV irradiation, hydrogen peroxide, or catalysts such as iron or copper complexes [52]. Regardless, optimum process and operational circumstances have still been determined for the different water and wastewater types together with various types of antibiotics [152].

In general, Fenton process has been widely used successfully for the oxidation of many groups of antibiotics, including b-lactams, quinolones, trimethoprim, and tetracyclines. Fenton's oxidation is a homogeneous oxidation process and considered to be a metal-catalyzed oxidation reaction, in which iron acts as a catalyst [65, 153]. The main handicap of the process is the low pH value. It is required to avoid iron precipitation that takes place at high pH [154, 155].

Heterogeneous photocatalysis with TiO2 semiconductor is generally accomplished by the illumination of a suspension of TiO2 in aqueous solution with light energy which is greater than its bandgap energy. This causes the generation of high energy electron-hole pairs (e�/hþ), which may migrate to the surface of the catalyst and may either reunite producing thermal


heavy metal ions, insoluble oil and grease, carbonates, and high alkalinity may cause clogging of the UV quartz handle. Air emission problems with O3 may arise. The cost of the AOPs is

Ozonation and AOPs are required for efficient degradation of antibiotics in water and wastewater. These treatment processes have an advantage of elimination of such pollutants through mineralization or conversion to the products that are less harmful to human health and the

Various studies have published the effective AOPs treatment for the removal of antibiotics from wastewater [63–67]. High-quality and effective publications relevant with the AOP studies on the mechanisms and applications of water and wastewater treatments have been pronounced for last two decades. From the theoretical, environmental, and economical point of view, they demonstrate a great and increasing interest. As shown in Table 2, several studies

Ozone is a potent oxidant and has been progressively applied for the treatment of wastewater. Ozone and/or hydroxyl radicals passivate bactericidal characteristics of antibiotics by disrupting or modulating their pharmaceutically active functional groups, such as N-etheroxime and dimethylamino groups of macrolides [68, 69], aniline moieties of sulfonamides [70], thioether groups of penicillins, unsaturated bonds of cephalosporin, and the phenol ring of trimethoprim [69]. High removal rates (>90%) were achieved by ozonation of the compounds with electron-rich aromatic systems, such as hydroxyl, amino (e.g., sulfamethoxazole), acylamino, alkoxy, and alkyl aromatic compounds, as well as those compounds with deprotonated amine (e.g., erythromycin, ofloxacin, and trimethoprim) and nonaromatic alkene groups, since these structural moieties are highly amendable to oxidative attack [1]. Ozonation process was found to be effective for the removal of b-lactams, macrolides, sulfonamides,

The performance of ozone treatment can be improved providing ozone is combined with UV irradiation, hydrogen peroxide, or catalysts such as iron or copper complexes [52]. Regardless, optimum process and operational circumstances have still been determined for the different

In general, Fenton process has been widely used successfully for the oxidation of many groups of antibiotics, including b-lactams, quinolones, trimethoprim, and tetracyclines. Fenton's oxidation is a homogeneous oxidation process and considered to be a metal-catalyzed oxidation reaction, in which iron acts as a catalyst [65, 153]. The main handicap of the process is the low pH value. It is required to avoid iron precipitation that takes place at high pH [154, 155].

Heterogeneous photocatalysis with TiO2 semiconductor is generally accomplished by the illumination of a suspension of TiO2 in aqueous solution with light energy which is greater than its bandgap energy. This causes the generation of high energy electron-hole pairs (e�/hþ), which may migrate to the surface of the catalyst and may either reunite producing thermal

have been conducted on the applicability of AOPs on different antibiotic classes.

expensive when compared to rival technologies [42].

186 Physico-Chemical Wastewater Treatment and Resource Recovery

aquatic environment.

3.3. Assessment of AOPs performance for antibiotic removal

trimethoprim, quinolones, tetracyclines, and lincosamides [5].

water and wastewater types together with various types of antibiotics [152].



Amoxicillin

UV, O3, Fenton, Fenton-like,

Fenton, UV/O3/H

Fe(III)

O2 2, TiO2, Fe(II), and

 photo-

1 μM

 name

AOP

Concentration

 Reaction conditions

oxygen flow rate: 600 sccm, power: 2 W, pH 8,

synthetic wastewater

Temperature:

intensity: 1.81 micro Einstein s1, optical path: 5.09

cm, O was generated from O

3

mgh1photo-Fenton

other for Fenton and

concentration:10

water, municipal WWTP

15-W black-light

 fluorescent

pharmaceutical

pH 2.5, ferric nitrate or FeOx conc.: 0.2 mmol L1

H

O2 2 conc.: 1.0–10.0 mmol L1, sewage treatment

plant effluent

H

pH 3.5,

were performed with a modified version of the

domestic electric oven: power of 1200 W, of 2450

MHz, synthetic wastewater

microwave-assisted

 oxidation reactions

O2 2 conc.: 2 g L1, FeSO4.7H O conc.: 0.2 g L

2

1

,

[83]

 solution flow rate: 80 mL min1

,

[82]

,

 lamp (365 nm),

Amoxicillin Amoxicillin Amoxicillin Amoxicillin Amoxicillin Amoxicillin

Sulfate radicals under ultrasound

0.095 mmolL1

irradiation

UV and UV/H

O2 2

00 μM

Photo-Fenton

50 mg L1

Ozonation

5.0 104 M

 Ozonation were performed in a

stirred gas–liquid reactor. 25C, flow rate: 361 h1

synthetic wastewater

Solar simulator: 1100-W xenon arc lamp (290 nm),

minimum intensity (250 W m2) pH 6.2, TOC: 26.3

mg C L1, FeSO4.7H O or FeO

H

O2 2 2.8, synthetic wastewater

Ultrasonic generator: 20 kHz, Ti probe, synthetic

[86]

wastewater

Low-pressure

Photon fluence rate: 8 107 Einstein L1 s

effective light path: 5.5 cm, T: 20

 Hg arc-UV (254 nm) 1

[87]

,

 2C pH 7

concentration

 used was 120 mg L1, pH 2.5–

2

x conc.: 0.05 mM,

semicontinuous

[84]

,

[85]

Microwave

 assisted Fenton

450 μg L1

Photo-Fenton

0.1 mM

groundwater,

 secondary effluents from

 μM, ultra-pure water, reservoir

experiments

 were carried out at natural pH

photo-Fenton

 Fe(II) and H

O2 2

188 Physico-Chemical Wastewater Treatment and Resource Recovery

 and

photo-Fenton:

 pH 3 and

2.flow rate of O3: 16

 20C, 15 W Hg lamp (254 nm), light

[81]



 name Chlortetracycline,

oxytetracycline

Chlortetracycline,

Chlortetracycline

Chlortetracycline

Ciprofloxacin

Ciprofloxacin,

Ciprofloxacin Ciprofloxacin

Ciprofloxacin Ciprofloxacin

Photo-Fenton

0.15 mM

T: 298 K, 125-W

(9 104 μEs m2 s

high-pressure

1), synthetic wastewater

 lamp, photonic flux

[100]

UV, TiO2/UV, O3 and H

O2 2

200 μgL1

O3/H

O2 2

Electron ionization

moxifloxacin

UV and TiO2/UV

 and

sulfamethoxazole

 persulfate

Photocatalytic

 ozonation

0.5 mM

Photocatalytic

 ozonation

0.15 mM

sulfamethoxazole

 UV, electron beam, ozone

doxycycline,

Ozone

AOP

Concentration

5 106 M

30 mgL1

pH 4.63 and 4.33,

T: 22

was produced from O2, electron accelerator

and 40 kW), synthetic wastewater

Ozone generated from O2 (air pressure: 5 bar, air

flow rate: 1200 L h–1), ozone input: 20 g m3 and

flow rate of the ozone/air: 20 L h1, T: 25C, 15-W

UV

low-pressure

wastewater

Ozone generated from pure oxygen. flow rate: 20

mg min1, T: 20

lamp (260, 275, 290, 302, 307, 315, 336, 366, 406, and

434 nm), TiO2: Degussa P25 and 0.1 g L1

synthetic wastewater

Initial pH 6 and decreased to 3–4.K2S O2 8 and Fe

(II)/Fe(II)-chelate:

Photocatalyst:

1K, synthetic wastewater

10 MeV, 10-kW electron ionizing energy unit,

synthetic wastewater

 TiO2-P25, TiO2: 0.5 g L1

,T: 298

[103]

[104]

 4.8 and 4.8 mM, river water

0.15 mM CIP:45.3 μM and

MOX: 37.4 μM

100 mg L1

45.27 μM

A bubble reactor was used for ozonation. generated from O2. T: 6.0–62C, ozone conc.: 2500

ppm, gas flow rate: 120 mL min1

2–990 μmol L1, synthetic wastewater

Photocatalyst:

Hg lamp, pH 3 (UV and TiO2/UV), TiO2 conc.: 571

ppm, ozone generated from O2,flowrate:

pH 9 (O3 and H

mgL1, hospital wastewater

O2 2), H

O2 2 conc.: 500 and 1000

 TiO2-P25, 125W

medium-pressure

[106]

 8 Lmin1

,

, T: 27.5C, H

O2 2:

 Ozone

[105]

 2C, 125-W

high-pressure

 UV

[101]

,

[102]

 lamp (254 nm), synthetic

 2C, 6-W single UV-C lamp (254 nm), ozone

atmospheric

 pressure

[99]

 (1 MeV

[100]

190 Physico-Chemical Wastewater Treatment and Resource Recovery

 Reaction conditions

Flow rate: 80 cm3/min, T: 20–21C, synthetic

wastewater

References

[98]



Flumequine,

 ofloxacin,

Modified

photo-Fenton

100 μgL�1

5 mgL�1 Fe, 35 mgL�1, 50 mgL�1 H

initial pH ≈ 7, λ < 400 nm

Solar UV power: 30 Wm�2, municipal wastewater

treatment plant effluent was taken

the MWTP secondary biological treatment in El

Ejido (province of Almería, Spain). pilot compound

parabolic collector (CPC) was used for photo-

Fenton experiment

Low-pressure

irradiance:

L�1, Fe(II): 0.25–1.0 mmol L�1. NaHSO4/H

Ozone flow rate: 3.3 gh�1, oxygen was used as a

feed gas, commercial

as catalyst, T: 17�C, pH 6.5, synthetic wastewater

T: 20�C, synthetic wastewater

T: 25�C, pH: 3.0, 7.0 and 10.0, stirring speed: 13.2

rps, reactor volume: 200 mL, catalyst loading: 1.0 g

L�1, air flow: 60 mLmin�1, phosphate buffer conc.:

10 mM, light intensity UV-A 104 mW, synthetic

wastewater

 TiO2 Degussa P25 was used

 8.3 mW cm�2, H

O2 2: 0.5–10.0 mmol

O2 2: 1

[113]

[114]

[115]

 mercury lamp, 15 W, λmax

¼ 254 nm,

[97]

downstream

 of

192 Physico-Chemical Wastewater Treatment and Resource Recovery

O2 2, oxalic acid,

[112]

sulfamethoxazole

Flumequine Levofloxacin Metronidazole

Moxifloxacin Moxifloxacin

TiO2/UV

12.5, 24.9, 37.4, 49.9,

T: 5, 15, 25, 35, 45, and 65�C, pH 7, stirring speed

2.3, 7.9 and 13.2 rps, reactor volume 200 mL,

catalyst loading 0.25, 0.5, 1, 3, 5, and 8 g L�1

oxygen, air, nitrogen flow: 60 mL min�1, buffer

concentration

485 W cm�2, ISO

3740, 37.4 � 103, 74.8 � 103, and 18.7 � 104, KI

concentration

L�1, synthetic wastewater

> Ofloxacin and

trimethoprim

 Solar

photo-Fenton

 process

100 μg L�1

T: 25�C, UV power: 30 W m�2, secondary treated

[117]

domestic effluents

 3.74, 37.4, 374, 3740, and 7480 mol

 10 mM, light intensity UV-A at 3 cm,

concentration μmol L�1 37.4, 374,

,

[116]

62.3

and 124.6 μM

Electro-Fenton

TiO2/UV

80 mg L�1

37.4 and 124.6 μM

OzonationandTiO2/UV

20 mgL�1

Fenton and

photo-Fenton

500 μg L�1

 name

AOP

Concentration

 Reaction conditions

wastewaters

(Madrid)

 were taken from Alcala de Henares




 name Sulfamethoxazole,

erythromycin,

sulfathiazole

Sulfamethoxazole,

sulfadiazine,

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

Sulfamethoxazole

TiO2/UV

2.5–30 mgL�1

9W lamp (350–400 nm), photon flux: 2.81 � 10�4

Einstein min�1

, T: 25�C, synthetic wastewater

TiO2/UV

Solar

photo-Fenton

10 mg L�1

100 mg L�1

Photo-Fenton

200 mg L�1

Ozone

200 μg L�1

Ozone

0.150 mM

Photoelectro-Fenton

200–300 mg L�1

T: 20�C, anode: RuO2/Ti, cathode: RuO2/Ti, UV

Lamp SLUV-8, 254/365 nm, energy input: 1407

Wcm�2, current: 0.36 A, synthetic wastewater

Ozone generated from pure oxygen, pH: 2 and 8,

H

O2 2: 0.013 M, flow: 3.0 ml min�1

flow: 8.5 gNm�3, synthetic wastewater

Ozone generated from pure oxygen, 15-W black

light lamps (365 nm), flux of radiation: 7.05

� 10�5 Einstein min�1, primary wastewater

effluent

Black-light

400 nm), photon flow: 6.85–5.67 Einstein s�1

� 0.8�C; TOC

L�1, synthetic wastewater

1100-W xenon arc lamp (below 290 nm), intensity:

250 Wm2

, T: 25�C, synthetic wastewater,

Catalyst: TiO2 Degussa P25, T: 25�C, xenon lamp

(1000 W), wavelength:

wastewater

 below 290 nm, synthetic

 seawater [135]

[136]

[134]

¼ 94.5 mgL�1 and COD

. T: 25

[133]

¼ 290 mgO2

 blue lamps with power of 8W (350 and

� 0.05

[132]

, T: 25�C, gas

[131]

Anodic oxidation and

electro-Fenton

 1.3 mM

trimethoprim

sulfamethazine,

UV and UV/H

O2 2

4 μM

ciprofloxacin

 and

roxithromycin,

UV

1

μgmL�1

Mercury vapor lamp (UV 254 nm) or black light

[127]

phosphor bulb (UV 350 nm), xenon lamp (750

Wcm�2, 250 Wcm�2

pH: 5.5–8.1, T: 20�C, synthetic wastewater

Low-pressure

H

O2 2 dose

surface water, wastewater

Catalyst: 0.2 mM Fe2þ and/or 0.2 mM

Cu2þ, pH 3.0 and T: 23

synthetic wastewater

� 2�C, current: 30–450 mA,

[129]

[130]

 treatment plant effluent

¼ 6 mg L�1, synthetic wastewater,

 UV lamps, fluence

¼ 540 mJ cm�2

,

[128]

194 Physico-Chemical Wastewater Treatment and Resource Recovery

,

AOP

Concentration

 Reaction conditions

lamp has

wastewater

polychromatic

 emission, hospital


Table 2. Summary of reaction conditions for antibiotic removal from water by AOP.

energy or participate in redox reactions with the compounds that are adsorbed on the catalyst's surface [1].

Due to some disadvantages of the heterogeneous photocatalysis (e.g., rather small quantum efficiency of the process; comparatively narrow light-response reach of TiO2; the requirement of postseparation and recovery of the catalyst particles from the reaction mixture in aqueous slurry systems), TiO2 appear to have some interesting properties, such as high chemical stability in a wide pH range, strong resistance to chemical breakdown and photocorrosion, and high efficiency. The catalyst is also inexpensive and can be reprocessed [134, 156]. The characteristics of antibiotics to be treated like pKa and molecular structure will identify not only the performance of their photocatalytic breakdown but also the mechanisms of the oxidation products formation.

Ultraviolet (UV) disinfection is progressively discovering practices in UWTPs. Photolytic breakdown can be either direct or indirect. In direct photolysis, the target pollutant assimilates a solar photon, which causes to a breakup of the molecule. In an indirect photolysis mechanism, as a matter of course occurring molecules in the system such as dissolved organic matter (DOM) behave as sensitizing species, which creates strong reactive agents such as hydroxyl radicals, singlet oxygen, and hydrate electrons under solar radiation [1, 117].

Ultraviolet irradiation has been greatly used for the treatment of waters and wastewaters worldwide. Different studies have stated the effective treatment of UV irradiation for the removal of antibiotics in wastewater effluents [63]. It has been lately stated that at high UV doses of almost 11,000–30,000 mJ cm2 , a nearly complete removal of tetracyclines and ciprofloxacin was obtained. Kim et al. [99] noticed that sulfonamides and quinolones demonstrate high removal efficiency in the reach of 86–100% throughout the UV process [1].

Many of the antibiotics have aromatic rings, structural moieties (such as phenol and nitro groups) heteroatoms, and other functional chromophore groups that can either absorb solar radiation or react with photogenerated transient species in natural. The organic material, UV dosage, contact time, and the chemical construction of the compound are significant agents ruling the removal performance of antibiotics throughout direct photolysis. This technique is only practicable to wastewater-containing photosensitive compounds and waters with low COD concentrations [5].

Most traditional operation performed in WWTPs and DWTPs (such as coagulation, flocculation, sedimentation, and filtration) were ineffective in the removal of these compounds [63], taking the improvement of new effective methodologies. Owing to the recalcitrant nature of the effluents including antibiotics residues, the implementation of the AOPs arises as an alternate. In fact, ozonation and Fenton's oxidation are the most tried methodologies. Although ozonation has the benefit of being used to fluctuate flow rates and compositions, the high cost of material and the energy required to provide the process constitute a primary disadvantage. Oxyhydroxides produce precipitate (if the pH range is not controlled well) when a homogeneous process is used and the necessity of recovering dissolved catalyst is a disadvantage. This is another process that is applied often to the group of beta-lactam antibiotics, combined with UV irradiation (photo-Fenton).

Compound

Tetracycline Tetracycline Tetracycline Tetracycline Tetracycline

Tinidazole Tinidazole Trimethoprim Trimethoprim

Table 2.

Summary of reaction conditions

 for antibiotic removal from water by AOP.

Anodic oxidation, photoelectro-Fenton,

Fenton

BDD,

electrochemical

 oxidation

 1.72 � 10� 4 mol L�1 T: 25�C, synthetic wastewater

 solar

photoelectro-

electro-Fenton,

20.0 mg L�1

Photocatalysis

Ozone Sonolysis

67 mgL�1

30 mgL�1 45, 80, and 100 ppm

 pH: 3, 5, 7, 9; H

mML�1, frequency: 40, 80, 120, and 160 kHz, input

power:750 W,

Iran), synthetic wastewater

6-W fluorescent synthetic wastewater,

 blacklight blue lamp (350–410 nm),

[151]

[120]

 wastewater

pharmaceutical

 wastewater

 (Tehran,

O2 2 conc.: 83, 167, 250, 333, and 417

T: 25�C, synthetic wastewater,

sewage treatment plant effluent

 surface water and a

Ozonation Photo-Fenton

20–100 mgL�1

TOC: 13 mg L�1

15-W black-light Wm�2, flow rate: 80 mL min�1, synthetic

wastewater,

plant effluent

Medium mercury lamp, synthetic wastewater

 [148] [149]

[150]

 surface water and a sewage treatment

 lamp (365 nm), irradiance:

 19

[147]

O3 was generated from oxygen. T: 25�C, synthetic

[146]

wastewater

Electrochemical

Fenton

 oxidation, electro

100 mg L�1

 name

AOP Electrochemical

 oxidation

Concentration

TC

¼ 200 mg L�1

 Reaction conditions The process was performed using a DSA (mixed

metal oxide, cathode. synthetic wastewater

Three electrodes as anode: commercial

boron-doped

on a niobium substrate), and commercial

(mixed metal oxide

tridimensional,

were used. T: 23�C, synthetic wastewater

 carbon-felt

 electrodes as cathode

Ti/RuO2–IrO2),

 and a

196 Physico-Chemical Wastewater Treatment and Resource Recovery

 diamond (BDD, thin-film deposited

 pure Pt,

[145]

 DSA

Ti/RuO2-IrO2)

 anode carbon-felt

 from

[144]

## 4. Conclusions

The consumption of antibiotics worldwide by human and veterinary uses has been increasing significantly. This is an important public concern because they have endocrine-disrupting properties even in trace concentrations and can cause microorganism resistance in aquatic environments. According to the researches made in recent years, advanced oxidation processes are promising treatment methods for the removal of the antibiotic compounds from water.

In the event of the photochemical technologies, we can determine that the photochemical AOPs are usually easy, clean, comparatively inexpensive, and productive against the classical, chemical AOPs. Four basic types of photochemical AOPs (H2O2/UV, O3/UV, H2O2/Fe2<sup>þ</sup>/UV, and TiO2/UV) have been enforced to reduce and/or mineralize organic pollutants. We have defined that, within these photochemical processes, the photocatalytic ones had mainly a better performance.

Furthermore, it is significant to point that heterogeneous photocatalysis has been the aim of an enormous improvement in the last decade. In fact, TiO2 is a semiconducter approach that exists, for example, as a chemically very stable, biologically inefficient, very easy to manufacture, cheaper than the photocatalytic viewpoint, active and several important photocalysts with an energy vacancy comparable to solar photons.

The economic robust of AOPs for full-scale wastewater treatment needs to be extensively investigated. These technologies should be modified to achieve both technical efficiency and cost effectiveness so that water industries could afford the adaptation of such technologies.

## Author details

Ayse Kurt, Berna Kiril Mert, Nihan Özengin, Özge Sivrioğlu and Taner Yonar\*

\*Address all correspondence to: yonar@uludag.edu.tr

Department of Environmental Engineering, Faculty of Engineering, Uludağ University, Görükle Campus, Bursa, Turkey


[3] Alaton I, Dogruel S. Pre-treatment of penicillin formulation effluent by advanced oxidation processes. J. Hazard. Mater. 2004;B112:105–113. DOI:10.1016/j.jhazmat.2004.04.009

4. Conclusions

198 Physico-Chemical Wastewater Treatment and Resource Recovery

better performance.

Author details

References

11.027

Görükle Campus, Bursa, Turkey

with an energy vacancy comparable to solar photons.

\*Address all correspondence to: yonar@uludag.edu.tr

The consumption of antibiotics worldwide by human and veterinary uses has been increasing significantly. This is an important public concern because they have endocrine-disrupting properties even in trace concentrations and can cause microorganism resistance in aquatic environments. According to the researches made in recent years, advanced oxidation processes are promising treatment methods for the removal of the antibiotic compounds from water.

In the event of the photochemical technologies, we can determine that the photochemical AOPs are usually easy, clean, comparatively inexpensive, and productive against the classical, chemical AOPs. Four basic types of photochemical AOPs (H2O2/UV, O3/UV, H2O2/Fe2<sup>þ</sup>/UV, and TiO2/UV) have been enforced to reduce and/or mineralize organic pollutants. We have defined that, within these photochemical processes, the photocatalytic ones had mainly a

Furthermore, it is significant to point that heterogeneous photocatalysis has been the aim of an enormous improvement in the last decade. In fact, TiO2 is a semiconducter approach that exists, for example, as a chemically very stable, biologically inefficient, very easy to manufacture, cheaper than the photocatalytic viewpoint, active and several important photocalysts

The economic robust of AOPs for full-scale wastewater treatment needs to be extensively investigated. These technologies should be modified to achieve both technical efficiency and cost effectiveness so that water industries could afford the adaptation of such technologies.

Ayse Kurt, Berna Kiril Mert, Nihan Özengin, Özge Sivrioğlu and Taner Yonar\*

Department of Environmental Engineering, Faculty of Engineering, Uludağ University,

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210 Physico-Chemical Wastewater Treatment and Resource Recovery

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**Provisional chapter**

## **Wastewater Treatment through Low Cost Adsorption Technologies Technologies**

**Wastewater Treatment through Low Cost Adsorption** 

Gonzalo Montes-Atenas and Fernando Valenzuela Fernando Valenzuela

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/67097

Gonzalo Montes-Atenas and

#### **Abstract**

This chapter addresses the wastewater treatment of mining residues through adsorption methodologies. It preferentially focuses its attention on (but not limited to) the removal of heavy metals. It begins with a brief description of the most used wastewater treatment pathways highlighting both their advantages and disadvantages and focusing on adsorption industrial practice. Classic models of adsorption thermodynamics and kinetics are presented. It finalises with a more detailed description of two methodologies of low cost sorbents: (i) inorganic nanostructured silicates and (ii) organicbased sorbents—pine bark.

**Keywords:** adsorption, adsorbent, wastewater treatment, sorbent, mathematical modelling, mining

## **1. Wastewaters containing heavy metals and other contaminants: mining applications**

#### **1.1. Introduction**

One of the two major environmental problems concerning the management of liquids in mining and mining-related activities, particularly relevant to a country like Chile, is the natural generation of acidic mine drainages (AMD) or acidic rock drainages (ARD) produced by chemical and/or microbial oxidation of sulphides in the presence of air and water [1, 2]. Obviously, the other problem is the impact of industrial residual aqueous solutions originated in the metallurgical and mining processing plants. All these solutions exhibit an important amount of chemical contaminants, either dissolved or suspended, at concentrations that normally surpass the limit fixed by the national discharge regulations and their

Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

© 2016 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons

discard in an acceptable manner into surface and groundwater bodies is an imperious need [3]. The complexity behind the latter issue is even higher considering that the water requirements in the mining industry are becoming a real problem, especially in regions such as the Atacama Desert or The Andes Mountains, where this vital liquid is scarce. Mining activities must then share the few water resources with the needs of the local communities for human life and agriculture. Therefore, it is absolutely necessary to optimise the use of all the available water resources, by controlling the waters entering mining and metallurgical plants and by treating the liquid residues that the processes such as leaching, solvent extraction and froth flotation plants produce.

For instance, AMD occurs when sulphide ores, mainly iron- and copper-bearing sulphides minerals, are exposed to water and air, resulting in the formation of sulphuric acid and metal hydroxides that release toxic heavy metal ions, and acidic hydrogen into surface and ground waters [4, 5]. They are also generated at waste disposal sites and around abandoned mine sites. Its formation is primarily a function of the local geology and hydrology. The AMD usually consists of an aqueous solution that exhibits a variable acidity and contains many and variable toxic or valuable dissolved metals, with a high content of sulphate (SO4 2− ) and other anions such as phosphate (PO<sup>4</sup> 3-), nitrate (NO<sup>3</sup> - ), molybdate (MoO<sup>3</sup> 2-) and chloride (Cl- ), and a high amount of colloidal-type suspended fine solids difficult to settle, becoming a water pollution problem difficult to remediate. The metals remain in solution until the pH increases to a level where precipitation occurs. However, precipitation presents many problems, such as redissolving of precipitates, the need of a large amount of chemicals and the generation of large volumes of sludge whose disposal is quite complex. Just as an exemplification, some chemical reactions that represent the chemistry of AMD formation are as follows [6, 7]:

$$4\,\text{FeS}\_{\text{2(s)}} + 15\,\text{O}\_{\text{2(g)}} + 14\,\text{H}\_2\text{O}\_{\text{(l)}} \to 4\,\text{Fe(OH)}\_{\text{3(s)}} + 8\,\text{H}\_2\text{SO}\_{\text{4(aq)}}\tag{1}$$

$$4\,\mathrm{Fe^{2+}\_{(aq)}} + \mathrm{O\_{2(g)}} + 4\,\mathrm{H^{+}\_{(aq)}} \leftrightarrow 4\,\mathrm{Fe^{3+}\_{(aq)}} + 2\,\mathrm{H\_2O\_{(l)}}\tag{2}$$

$$4\text{ Fe}^{3+}\_{\text{(aq)}} + 12\text{ H}\_{2}\text{O}\_{\text{(l)}} \leftrightarrow 4\text{Fe}\text{(OH)}\_{\text{3(s)}} + 12\text{ H}^{+}\_{\text{(aq)}}\tag{3}$$

$$\mathrm{FeS}\_{2(s)} + 14\,\mathrm{Fe^{3+}\_{(aq)}} + 8\,\mathrm{H}\_{2}\mathrm{O}\_{(l)} \to 15\,\mathrm{Fe^{2+}\_{(aq)}} + 2\,\mathrm{SO^{2-}\_{4(aq)}} + 16\,\mathrm{H^{+}\_{(aq)}}\tag{4}$$

Equation (1) represents the oxidation of pyrite by oxygen. This reaction generates two moles of acid per mole of oxidised pyrite. Reaction (2) involves the conversion of Fe(II) into Fe(III), reaction that is enhanced by the bacteria *Acidithiobacillus ferrooxidans* (whenever these microorganisms are present in the system). Reaction (3) represents the hydrolysis of iron with the formation of the corresponding hydroxide and the production of new acid molecules. Finally, reaction (4) represents the oxidation of additional pyrite by ferric iron generated in the previous reactions. These reactions are based on the dissolution of pyrite (FeS2 ); nevertheless, it is expected that they also occur simultaneously with other metal sulphides present in the ore. An overall chemical reaction cycle takes place very rapidly and continues until either the ferric iron or the metal sulphide is depleted. The latter implies that a variety of metals and anions may dissolve in the process both conferring a significant toxicity to the streams containing them and turning AMD into a potential source of many valuable and, sometimes, scarce metals.

#### **1.2. Wastewater treatment paths and their comparison to adsorption operations**

discard in an acceptable manner into surface and groundwater bodies is an imperious need [3]. The complexity behind the latter issue is even higher considering that the water requirements in the mining industry are becoming a real problem, especially in regions such as the Atacama Desert or The Andes Mountains, where this vital liquid is scarce. Mining activities must then share the few water resources with the needs of the local communities for human life and agriculture. Therefore, it is absolutely necessary to optimise the use of all the available water resources, by controlling the waters entering mining and metallurgical plants and by treating the liquid residues that the processes such as leaching, solvent

For instance, AMD occurs when sulphide ores, mainly iron- and copper-bearing sulphides minerals, are exposed to water and air, resulting in the formation of sulphuric acid and metal hydroxides that release toxic heavy metal ions, and acidic hydrogen into surface and ground waters [4, 5]. They are also generated at waste disposal sites and around abandoned mine sites. Its formation is primarily a function of the local geology and hydrology. The AMD usually consists of an aqueous solution that exhibits a variable acidity and contains many and variable

), molybdate (MoO<sup>3</sup>

of colloidal-type suspended fine solids difficult to settle, becoming a water pollution problem difficult to remediate. The metals remain in solution until the pH increases to a level where precipitation occurs. However, precipitation presents many problems, such as redissolving of precipitates, the need of a large amount of chemicals and the generation of large volumes of sludge whose disposal is quite complex. Just as an exemplification, some chemical reactions

+ 14 H2 O (l) → 4Fe (OH)

↔ 4 Fe (aq)

 + 4 H(aq) +

3+ + 12 H2 O (l) ↔ 4Fe (OH)

3+ + 8 H2 O (l) → 15 Fe(aq)

Equation (1) represents the oxidation of pyrite by oxygen. This reaction generates two moles of acid per mole of oxidised pyrite. Reaction (2) involves the conversion of Fe(II) into Fe(III), reaction that is enhanced by the bacteria *Acidithiobacillus ferrooxidans* (whenever these microorganisms are present in the system). Reaction (3) represents the hydrolysis of iron with the formation of the corresponding hydroxide and the production of new acid molecules. Finally, reaction (4) represents the oxidation of additional pyrite by ferric iron generated in the previous reactions. These reactions are based on the dissolution of pyrite

); nevertheless, it is expected that they also occur simultaneously with other metal sulphides present in the ore. An overall chemical reaction cycle takes place very rapidly and continues until either the ferric iron or the metal sulphide is depleted. The latter implies that a variety of metals and anions may dissolve in the process both conferring a significant toxicity to the streams containing them and turning AMD into a potential source of many

2−

2-) and chloride (Cl-

3 (s )

 + 12 H(aq) +

> + 16 H(aq) +

3(s)

2+ + 2 SO 4(aq) 2−

) and other anions such

+ 8 H2 SO 4(aq) (1)

(3)

(4)

3+ + 2 H2 O (l) (2)

), and a high amount

toxic or valuable dissolved metals, with a high content of sulphate (SO4


that represent the chemistry of AMD formation are as follows [6, 7]:

2+ + O 2(g)

+ 15 O 2 (g )

+ 14 Fe (aq)

extraction and froth flotation plants produce.

214 Physico-Chemical Wastewater Treatment and Resource Recovery

3-), nitrate (NO<sup>3</sup>

as phosphate (PO<sup>4</sup>

4 FeS 2(s)

FeS 2(s)

(FeS2

4 Fe (aq)

4 Fe (aq)

valuable and, sometimes, scarce metals.

*Solubility driven treatments:* Precipitation operations widely used in the removal of heavy metals are implemented based on pH or the addition of counter-ions leading the formation of sparingly soluble compounds (Eqs. (5) and (6), respectively)

$$\mathrm{M}^{\ast \ast}\_{\mathrm{(aq)}} + \mathrm{\boldsymbol{\eta}}\,\mathrm{OH}^{-}\_{\mathrm{(aq)}} \leftrightarrow \,\mathrm{M}\,\mathrm{(OH)}\_{\mathrm{(s)}}\tag{5}$$

$$m\,\mathrm{M}\_{\mathrm{(aq)}}^{n\star} + n\,\mathrm{X}\_{\mathrm{(aq)}}^{n\star} \leftrightarrow \,\mathrm{M}\_{\mathrm{m}}\,\mathrm{X}\_{\mathrm{n\cdot(s)}}\tag{6}$$

In these cases, the solubility product constants and the solubility of the compounds are written as in Eqs. (7) and (8).

$$K\_{ps} = \left. a\_{M^{s-}} a\_{\text{OH}^{s-}}^{u} \right|\_{S^{s}} = \sqrt[m+1]{\frac{|\overline{K\_{ps}}|}{M}} \tag{7}$$

$$K\_{ps} = \left. a\_{\chi^{\circ}}^{u} \right|\_{\chi^{\circ}} a\_{\chi^{\circ}}^{u} \quad\_{S} = \begin{array}{c} \stackrel{m \mapsto n}{\sqrt{\frac{K\_{ps}}{m+n}}}\\ \end{array} \tag{8}$$

where *Kps* and *ai* represent the solubility product constant and the activity or real (effective) concentration of the ion *i* in the aqueous solution.

The remaining amount of the pollutant in the solution depends on the species solubility which is not only a function of the solubility product constant but also of the number and equilibrium constants associated with the formation of anionic and others species related to counterions forming other molecular complexes [8]. The formation of metal hydroxides may leave large amounts of pollutants in aqueous solutions. In this case, counterions with lower *Kps* values are used, for instance sulphide ions (S 2<sup>−</sup> ) are used to improve the wastewater treatment of waters containing mercury or cadmium. Most of the treatments employed in our country only consider a alkaline chemical treatment by contacting the acidic stream with CaCO3 or CaO or Ca (OH) 2 raising the pH up to a point where metals are partially precipitated.

*Cementation and other electrochemically driven treatments:* Cementation operations correspond to spontaneous electrochemical reactions largely displaced towards the formation of the products. They have identical problems to those found in classic precipitation operations and, additionally, the process have to deal with the release of the reduction agent (*R*) to the aqueous phase (Eq. (9)).

$$M\_{\text{(aq)}}^{n+} + n \ R\_{\text{(s)}}^{0} \leftrightarrow M\_{\text{(s)}}^{0} + n \ R\_{\text{(aq)}}^{+} \tag{9}$$

Cementation processes are commonly implemented in acidic conditions to avoid metal hydrolysis and consider oxidation reactions which are located outside the water stability domain. The latter introduces in these systems parallel competitive reactions associated with the hydrogen evolution and others decreasing the efficiency of the cementation process. Other non-spontaneous electrochemical treatment is used such as electrodeposition, a clean technology claimed to avoid the generation of residues or electrocoagulation. The latter having an anode made of a polyvalent metal able to promote coagulation such as iron or aluminium. However, according to the elements to be removed, appropriate aqueous solution conditions must be reached. Relatively high cost associated with the application of voltage differences and mass transport limitations due to low conductivity and/or low element concentration make this wastewater treatment strategy expensive compared to other spontaneous processes.

*Membrane-based separation treatments:* This path includes ultrafiltration, nanofiltration and reverse osmosis. Ultrafiltration allows the removal of dissolved and colloidal materials. Isolated metal ions are difficult to remove directly due to their small sizes so the use of micelles and other complexing agents enhances its separation capabilities. Reverse osmosis removes a wide range of dissolved species from aqueous solutions; however, it has high power pumping requirements to restore the membranes. Membrane separation techniques have gone through significant progresses developing liquid membrane processes based on the formation of emulsions; however, there is still room for improvement with regard to their instability in salty and acidic conditions and fouling produced by other species present in the wastewater. Electrodialysis does not require the addition of external species to the wastewater, exhibits high selectivity and does not produce sludges, which makes it a promising technology. However, energy consumption and anodic/cationic membrane fouling is still a challenge.

*Ion-exchange treatment*: Ion-exchange resins either natural or synthetic are selected when high treatment capacity, high efficiency and fast kinetics are required. The stronger resins are built on sulphonic or carboxylic groups as in Eqs. (10) and (11). In both cases, the acidification of the aqueous media is commonly observed as it occurs in the case of solvent extraction using kerosene and implemented in countercurrent when using oximes. The only difference between ion exchange and solvent extraction would be that the organic phase would be in the dissolved phase in the organic phase rather than in the solid phase.

$$n\left(\mathrm{RSO}\_{3}\mathrm{H}\right)\_{\mathrm{(s)}} + M\_{\mathrm{(aq)}}^{\*\*} \leftrightarrow \left(\mathrm{RSO}\_{3}^{-}\right)\_{n}M\_{\mathrm{(s)}}^{\*\*} + n\,\mathrm{H}\_{\mathrm{(aq)}}^{\*}\tag{10}$$

$$n\text{ (RCOOH)}\_{\text{(s)}} + M\_{\text{(aq)}}^{\*\*} \leftrightarrow \text{ (RCOO}^{-}\text{)}\_{\text{u}}M\_{\text{(s)}}^{\*\*} + n\text{ H}\_{\text{(aq)}}^{\*}\tag{11}$$

*Froth flotation*: This process can be used to treat wastewaters contaminated with heavy metals and can be implemented following different technologies such as dissolved air flotation (DAF), electroflotation, ion flotation or precipitation flotation. DAF processes are based on the gas oversaturation and decompression producing microbubbles that are able to separate small particles or agglomerates. The electroflotation process is associated with the water electrolysis producing the smallest bubble sizes of hydrogen and oxygen known at the industrial level. Ion flotation is based on ionic complex formation with surfactant molecules and subsequent frother-aided flotation. The precipitation flotation is a mixed technology between the precipitation followed by flotation operations (sulphide precipitation are commonly implemented).

*Adsorption separation treatments:* In heterogeneous systems, whenever two immiscible states of the matter, namely gas and liquid, gas and solid or liquid and solid, are set in contact they are separated by a surface layer having properties different to those of the two states forming it [9]. When one or more components present in one of the two phases (or in both) tend towards increasing its concentration in the surface layer, it is said that the adsorption process is taking place (**Figure 1**).

In **Figure 1**, the adsorbat is represented by the orange circles. The adsorbent is represented by the continuous marble-like colour. The adsorption process can be either physical or chemical in nature. The physical adsorption or physisorption requires energies in the order of some kcal/mol and is mainly associated with condensation processes of the species at the interface. Its spontaneity is enhanced by reducing the temperature of the system (Eq. (12)).

$$\boldsymbol{d}\,\mathbf{G}\_{\mathrm{ads}} = \boldsymbol{d}\,\boldsymbol{H}\_{\mathrm{ads}} - \mathrm{Td}\,\mathbf{S}\_{\mathrm{ads}} \tag{12}$$

where *<sup>d</sup> <sup>G</sup>*ads, *<sup>d</sup> <sup>H</sup>*ads, *<sup>d</sup> <sup>S</sup>*ads represent the adsorption Gibbs free energy, the adsorption enthalpy and the adsorption entropy, respectively, while *T* is the temperature of the system. The term *<sup>d</sup> <sup>H</sup>*ads <sup>&</sup>lt; <sup>0</sup>(heat of condensation). In consequence, the spontaneity *<sup>d</sup> <sup>G</sup>*ads <sup>&</sup>lt; <sup>0</sup>would be favoured by lower temperature conditions. In this case, the rate of the adsorption process is rapid, reversible between the adsorbed and desorbed states, and it can be either monomolecular or multimolecular with respect to the number of adsorbat molecules associated

**Figure 1.** Schematics of an adsorption process.

*Membrane-based separation treatments:* This path includes ultrafiltration, nanofiltration and reverse osmosis. Ultrafiltration allows the removal of dissolved and colloidal materials. Isolated metal ions are difficult to remove directly due to their small sizes so the use of micelles and other complexing agents enhances its separation capabilities. Reverse osmosis removes a wide range of dissolved species from aqueous solutions; however, it has high power pumping requirements to restore the membranes. Membrane separation techniques have gone through significant progresses developing liquid membrane processes based on the formation of emulsions; however, there is still room for improvement with regard to their instability in salty and acidic conditions and fouling produced by other species present in the wastewater. Electrodialysis does not require the addition of external species to the wastewater, exhibits high selectivity and does not produce sludges, which makes it a promising technology. However, energy consumption and anodic/cationic membrane fouling is still a challenge.

*Ion-exchange treatment*: Ion-exchange resins either natural or synthetic are selected when high treatment capacity, high efficiency and fast kinetics are required. The stronger resins are built on sulphonic or carboxylic groups as in Eqs. (10) and (11). In both cases, the acidification of the aqueous media is commonly observed as it occurs in the case of solvent extraction using kerosene and implemented in countercurrent when using oximes. The only difference between ion exchange and solvent extraction would be that the organic phase would be in the

↔ (RSO<sup>3</sup>

↔ (RCOO−)

*Froth flotation*: This process can be used to treat wastewaters contaminated with heavy metals and can be implemented following different technologies such as dissolved air flotation (DAF), electroflotation, ion flotation or precipitation flotation. DAF processes are based on the gas oversaturation and decompression producing microbubbles that are able to separate small particles or agglomerates. The electroflotation process is associated with the water electrolysis producing the smallest bubble sizes of hydrogen and oxygen known at the industrial level. Ion flotation is based on ionic complex formation with surfactant molecules and subsequent frother-aided flotation. The precipitation flotation is a mixed technology between the precipitation followed by flotation operations (sulphide precipitation are commonly implemented).

*Adsorption separation treatments:* In heterogeneous systems, whenever two immiscible states of the matter, namely gas and liquid, gas and solid or liquid and solid, are set in contact they are separated by a surface layer having properties different to those of the two states forming it [9]. When one or more components present in one of the two phases (or in both) tend towards increasing its concentration in the surface layer, it is said that the adsorption process is taking place (**Figure 1**).

In **Figure 1**, the adsorbat is represented by the orange circles. The adsorbent is represented by the continuous marble-like colour. The adsorption process can be either physical or chemical in nature. The physical adsorption or physisorption requires energies in the order of some kcal/mol and is mainly associated with condensation processes of the species at the interface. Its spontaneity is enhanced by reducing the temperature of the system (Eq. (12)).

− )*<sup>n</sup> M*(s) *n*+

> *<sup>n</sup> M*(s) *n*+

 + *n* H(aq) +

> + *n* H(aq) +

(10)

(11)

dissolved phase in the organic phase rather than in the solid phase.

 H)(s)

(s) + *M*(aq) *n*+

 + *M*(aq) *n*+

*n* (RSO<sup>3</sup>

*n* (RCOOH)

216 Physico-Chemical Wastewater Treatment and Resource Recovery

with the number of available adsorption sites. On the other hand, chemical adsorption or chemisorption requires energies in the order of dozens of kcal/mol and involves the generation of bonding between the adsorbed specie and the interface. In general terms, its spontaneity is enhanced by increasing the temperature as *<sup>d</sup> <sup>H</sup>*ads <sup>&</sup>gt; <sup>0</sup>(activation energy needs to be reached). The adsorption process is slower and is highly dependent on the adsorbent nature and generally irreversible due to the strong forces between the adsorbat and the adsorbent taking place in the adsorption process. Adsorption operations have grown as a feasible way to treat wastewaters. The process is versatile in terms of the materials that can be used ranging from natural substrates or biosorbents such as stems, bark, leaves, among other; to synthetic sorbents like activated carbon, metal oxides, etc. The maximum adsorption capacity of the sorbent is commonly studied in thermodynamic experiments which are represented in terms of the adsorption isotherms. A list of the most used adsorption isotherms and the systems is presented in **Tables 1** and **2**. The \_\_\_\_ *<sup>V</sup> V*max ratio represents the adsorbed volume of adsorbat over the maximum volume of adsorbat able to be adsorbed. Otherwise named as a degree of coverage or *qe* ⁄ *<sup>q</sup>*max where *qi* has units of mg of adsorbat / g of sorbent defined as θ. Please, note that *q*max is referred to as the maximum adsorption associated with the monolayer. In all mathematical expressions, the variable *P* refers to the pressure of the adsorbat when a gas phase is set in contact with a condensed phase, namely, liquid or solid. In more general terms, the pressure *P* can be replaced by the activity or concentration of the adsorbat in aqueous phase.



**Table 1.** Adsorption isotherms and their use.

**Model No.**

(1) Langmuir [10] *<sup>θ</sup>* <sup>=</sup> *<sup>k</sup>*<sup>1</sup> *<sup>P</sup>* \_\_\_\_\_

1 + *k*<sup>1</sup> *<sup>P</sup>* (3 )

Or *<sup>k</sup>*<sup>1</sup> *<sup>P</sup>* <sup>=</sup> \_\_\_\_ *<sup>θ</sup>* <sup>1</sup><sup>−</sup> *<sup>θ</sup>* (3')

218 Physico-Chemical Wastewater Treatment and Resource Recovery

 \_\_1 *qe* = \_\_\_\_ <sup>1</sup> *qmax* + \_\_\_\_\_\_ <sup>1</sup> *qmax <sup>k</sup>*<sup>1</sup> *<sup>P</sup>* (3.5 )

 \_\_*P qe* = \_\_\_\_\_ <sup>1</sup> *k*<sup>1</sup> *qmax* + \_\_\_\_ *<sup>P</sup> qmax* (3.6 )

 *q* \_\_*e*

 \_\_1 *<sup>P</sup>*<sup>=</sup> *<sup>k</sup>*<sup>1</sup> *<sup>q</sup>* \_\_\_\_\_*max qe*

(2) Freundlich (also known by the name Halsey and Taylor) (Freundlich, 1906) [12]

Graphically:

*qe* = *qmax* −

*qe* <sup>=</sup> *<sup>k</sup>*<sup>1</sup> *<sup>P</sup>*\_\_1

Graphically: log *qe* = log *k*<sup>1</sup>

*<sup>n</sup>* (11 )

+ \_\_1

*<sup>n</sup>* log*P* (11.3 )

*<sup>q</sup>*\_\_\_*<sup>e</sup> <sup>k</sup>*<sup>1</sup> *<sup>P</sup>* (3.7 )

*<sup>P</sup>* = *k*<sup>1</sup> *qmax* − *k*<sup>1</sup> *qe* (3.8 )

− *k*<sup>1</sup> (3.9 )

**Name Mathematical model Specific features of the model and when to use it…**

as follows:

surface site

surface sites

The conditions are

concentrations

If \_\_1

occurs. If \_\_1

cooperative adsorption

*H <sup>a</sup>* = − *Δ H* <sup>0</sup> Log*θ* (11.1 ) .

This two-parameter model is widely used in cases where ideal localised monolayer is obtained by either chemical or physical adsorption. This is perhaps the most relevant model in the field of adsorption and has been derived from kinetic premises, thermodynamics conceptualisation and also statistically [11] The particular conditions under which the model is valid are

(1) Once adsorbed the adsorbat is fixed onto one surface site and there is no migration throughout the sorbent surface (or interface). Therefore, the adsorbent exhibits a limited capacity for adsorption (2) One adsorbate molecule is associated with one

(3) The surface energy is identical for every surface site and therefore the same applies to the adsorption energy. In other words, the following equation holds *Δ G a* = *Δ H a* − *TΔ S a* = *cte*( 3.3 ) , where the subscript

(4) A homogeneous surface without lateral interaction and steric hindrance between adsorbed species is required even if they are adsorbed in adjacent sorbent

The strength of intermolecular attractive forces decreases dramatically with the distance to the surface sorbent

This two-parameter isotherm is applicable to chemical and physical adsorption without lateral interactions between adsorbed molecules. It is an empirical two parameter model which describes a multilayer adsorption process where the surface sites do not follow any uniform distribution in adsorption heat of adsorption or affinities between the sorbent and adsorbent. It is based on the fact that the concentration of the adsorbat on the surface increases with the concentration in the other phase

1. The heat of adsorption can be represented by − *Δ*

2. The model cannot be used at high and low degree of coverage. At very low adsorbat concentration, the Freundlich equation does not provide Henry's law; however, it is valid for ion adsorption at low solute

*n* is close to zero, the surface heterogeneity of the sorbent is more significant. Such heterogeneity is characterised by the exponential decaying sorption site distribution (or the adsorption enthalpy changes logarithmically with the degree of coverage). Particularly, in this case, when n varies from 1 to 10, the adsorption is highly favourable. If , then, the partition between the two phases is independent of the concentration of the adsorbate and linear adsorption

*<sup>n</sup>*is higher than 1, there is a symptom of

*a* refers to the adsorption process


**Table 2.** Adsorption kinetics models.

## **2. Inorganic sorbents case study: Silica based sorbents - Nanostructured silicates**

## **2.1. Introduction**

The synthesis, characterisation and performance of sorbents based on nanostructured calcium silicates without modification or chemically modified with other atoms with the purpose to improve the adsorption of some specific contaminant species are reviewed. The nanostructured calcium silicates possess a very high capacity to remove contaminants from water even when these are present in higher concentrations, and also have the capability of a collective and simultaneous removal of cationic and anionic species from acidic aqueous solutions. Effectively, there is a strong need for adsorbent materials that have the capacity for removing simultaneously both anions and cations from industrial and mining solutions. The synthesized nanostructured calcium silicates exhibit a specific surface area, averaging values over 100–400 m2 /g, showing a good sorption capacities for removing high contents of different ionic species. The mechanisms governing the sorption process are related to the presence of silanol, calcium and hydroxyl groups in its structure, which would act as binding or nucleation sites on its surface [16]. Recent efforts for modifying nanostructured calcium silicates by introducing in their structures other metals that generate an improved sorbent addressed to enhance the sorption of other species present in solutions like the AMD, by using simple and low-cost processes, have been achieved. The efficiencies accomplished with nanostructured calcium silicates modified with iron, magnesium and aluminium introduced in its structure opens the possibility of tailoring a mix of modified silicates to suit the composition of various wastewaters requiring decontamination treatment. 'The more insoluble is the hydroxide precipitate or the salt formed during the reaction of the calcium silicate with the ionic species to be removed, the more efficient is the sorption process'.

The following modifications have been obtained successfully:


This monography addresses the base case and these four case studies.

#### **2.2. Experimental procedures – Nanostructured silicate synthesis procedure**

**2. Inorganic sorbents case study: Silica based sorbents - Nanostructured** 

The synthesis, characterisation and performance of sorbents based on nanostructured calcium silicates without modification or chemically modified with other atoms with the purpose to improve the adsorption of some specific contaminant species are reviewed. The nanostructured calcium silicates possess a very high capacity to remove contaminants from water even when these are present in higher concentrations, and also have the capability of a collective and simultaneous removal of cationic and anionic species from acidic aqueous solutions. Effectively, there is a strong need for adsorbent materials that have the capacity for removing simultaneously both anions and cations from industrial and mining solutions. The synthesized nanostructured calcium silicates exhibit a specific surface area, averaging values over

ionic species. The mechanisms governing the sorption process are related to the presence of silanol, calcium and hydroxyl groups in its structure, which would act as binding or nucleation sites on its surface [16]. Recent efforts for modifying nanostructured calcium silicates by introducing in their structures other metals that generate an improved sorbent addressed to enhance the sorption of other species present in solutions like the AMD, by using simple and low-cost processes, have been achieved. The efficiencies accomplished with nanostructured calcium silicates modified with iron, magnesium and aluminium introduced in its structure opens the possibility of tailoring a mix of modified silicates to suit the composition of various wastewaters requiring decontamination treatment. 'The more insoluble is the hydroxide precipitate or the salt formed during the reaction of the calcium silicate with the ionic species

(a) Nanostructured calcium silicate derivatives containing Fe atoms have been prepared to selectively promote and enhance the adsorption of arsenic species through the formation

(b) The introduction of Mg atoms of *n*-calcium silicate by partial substitution by Caatoms have permitted the preparation of a sorbent material that benefits the removal of phosphate and ammonium ions, frequently found in mining wastewaters, due the formation

of a very insoluble calcium-containing double phosphate of Mg and ammonium.

(c) The partial replacement of Caatoms in *n*-calcium silicate by Al atoms generates a stronger sorbent that allow and improve the sulphate removal by forming a high insoluble alu-

(d) Nanostructured calcium silicate by forming a composite with magnetite in order to provide to this nanosorbent of magnetic properties and facilitate its separation from the re-

to be removed, the more efficient is the sorption process'.

minium and calcium double basic salt.

sulting solution.

The following modifications have been obtained successfully:

of highly insoluble and stable double-salt of calcium and iron.

This monography addresses the base case and these four case studies.

/g, showing a good sorption capacities for removing high contents of different

**silicates**

100–400 m2

**2.1. Introduction**

220 Physico-Chemical Wastewater Treatment and Resource Recovery

The synthesis of nanostructured calcium silicate can be conducted using several routes of preparation. In this study, calcium silicate was synthesized by precipitation as a result of the room temperature chemical reaction between liquid Na2 SiO 3 and Ca (OH) 2 at a pH value of 12.3, following a process in two steps. Firstly, a suspension of Ca (OH) 2 in water is treated with HCl 33% w/w to give a slurry with a pH value of 12–12.5, varying the stirring velocity in a range between 425 and 1000 min−1. Secondly, a sodium silicate solution containing 28.5% w/w SiO 2 is diluted with water resulting in a 0.32% w/w SiO 2 solution. Then, this solution is vigorously mixed with the previously prepared Ca (OH) 2 suspension in a reactor, varying the stirring velocity between 1000 and 6000 min−1, immediately forming a precipitate, nanostructured calcium silicates containing an average value of 35 g SiO 2 /kg of formed solid. The formed slurry is aged for 20 min, allowed to settle for 13 h, before the obtained solid is recovered by vacuum filtration. Afterwards, the filter cake is washed with water and ethanol in order to disperse the nanostructured particles and lower the surface tension of the still wet solid. The solid is dried at 383 K for 2 days. Treatment with ethanol is necessary because during the drying of silicate, which presents a high surface area, is generated a great liquid-air interface that causes a high surface tension collapsing the structure of the material reducing its pore volume and surface area. The synthesised calcium silicate is an amorphous substance which does not present a defined chemical structure as is shown in **Figure 2**.

**Figure 2.** Probable structure of nanostructured calcium silicate.

**Figure 2** shows the presence of OH− groups and adsorbed water on the surface of the material. **Figure 3** shows scanning electron microscope (SEM) micrographs of the synthesised nanostructured calcium silicates. Particles exhibit a mean particle size between 0.2 and 1.0 µm. The particles appear in the SEM pictures forming larger agglomerates. This structure would correspond to a silicate backbone with plates of a thickness around of 10 nm and would consist of a tetrahedral silicate with Ca(II) ions and silanol groups on the surface forming a wollastonite-like structure, CaSiO3 , which contains SiO 3 2− species in which Ca(II) ions and the silanol groups would act as probable binding sites for the species being removed. The modifications done to the adsorbent by introducing other atoms that replace part of the calcium in the silicate, have not change significantly the mentioned structure, making possible the preparation of strong sorbent composites. Because of their low-cost and their disposable character, the regeneration of the modified calcium silicate is not considered. However, given their silicatebase composition similar to cement, later uses can be outlined.

**Figure 3.** SEM micrographs of the non-modified nanostructured silicate adsorbent.

#### **2.3. Modifications to the nanostructured calcium silicate**

Variants of this methodology to generate sorbent modifications are described below:


In all cases, there is a maximum element to calcium replacement ratio without compromising the nanostructure of silicate. Over certain replacement proportions, the nanostructure suffers a deleterious effect.

#### **2.4. Characterization of nanostructured calcium silicate**

The characterisation consisted of:


#### **2.5. Adsorption experiments**

done to the adsorbent by introducing other atoms that replace part of the calcium in the silicate, have not change significantly the mentioned structure, making possible the preparation of strong sorbent composites. Because of their low-cost and their disposable character, the regeneration of the modified calcium silicate is not considered. However, given their silicate-

base composition similar to cement, later uses can be outlined.

222 Physico-Chemical Wastewater Treatment and Resource Recovery

**2.3. Modifications to the nanostructured calcium silicate**

**Figure 3.** SEM micrographs of the non-modified nanostructured silicate adsorbent.

**2.4. Characterization of nanostructured calcium silicate**

mole % magnesium is added as Mg (OH)

Featoms using as iron source FeCl 3

a deleterious effect.

The characterisation consisted of:

Variants of this methodology to generate sorbent modifications are described below:

(a) Mg -substituted calcium silicate: add Mg during the synthesis of calcium silicate, a variable

(b) Fe -calcium silicate: are prepared by replacing part of the calcium atoms of the silicate by

(d) Magnetic properties of the calcium silicate that allows its easy separation from the treated waters, during the synthesis are added a suitable proportion of magnetite (Fe 3 O 4

In all cases, there is a maximum element to calcium replacement ratio without compromising the nanostructure of silicate. Over certain replacement proportions, the nanostructure suffers

• Measuring the mean particle size using a Malvern Mastersizer Hydro 2000 MU apparatus.

3

to replace part of the Ca atoms by Mg.

, Al (OH) 3

2 . or poly-alu-

), what

2

or Fe (OH)

(c) Al -substituted calcium silicate, aluminium can be added as NaAlO<sup>2</sup>

minium chloride (PAC) at the start of the synthesis with the Ca (OH)

avoid an important degrading of the accessible surface area of the silicate.

• Observing by scanning electron microscopy (SEM) JEOL JSM-25SII instrument.

Batch equilibrium sorption tests are conducted at 303 K in a batch-type reactor by mixing a variable amount of nanostructured calcium silicate and different volumes of a copper mine aqueous solution having the following main composition:


Batch experiments are conducted over sufficient time to reach equilibrium conditions. During the experiments, samples of the solution are collected at defined intervals and filtered using a 0.45 µm nitrocellulose Millipore membrane. The latter was performed before measuring the pH value and the concentration of metallic ions by atomic absorption spectrophotometry on a Perkin Elmer PinAAcle 900F instrument. The quantity of metal adsorbed is determined by the difference between the concentrations of metal in the initial aqueous feed phase and that in the raffinate solutions. At the end of the tests, the nanostructured calcium silicates were separated from the resulting aqueous solution by filtration. The SO4 <sup>2</sup><sup>−</sup> ion content is determined using a standard barium sulphate method [17] and PO<sup>4</sup> 3− ion concentration is measured using the vanadate-molybdate-phosphoric UV-spectroscopy method [18].

#### **2.6. Results**

#### *2.6.1. Synthesis and characterization*

The chemical reaction between Ca(OH) 2 and Na<sup>2</sup> SiO 3 is fast generating a thixotropic precipitate that produces particle agglomeration. The reaction of Ca(II) ions with soluble silicate as Na<sup>2</sup> SiO 3 resulted in the precipitation of calcium silicate, an insoluble solid, where Ca(II) appeared to be strongly bound to a silicate backbone. The extent of the reaction and the particle size of the solid formed depend strongly on the pH of the reaction mixture, the proportions and concentration of calcium and silicate ions and the intensity of the stirring employed during the process. Intensive stirring in basic medium produces a quite homogeneous but colloidal and amorphous insoluble calcium silicate. The higher the stirring velocity, the smaller the particle size of the resultant calcium silicate and the larger and more accessible the substrate surface area comprising of micro-pores and meso-pores.

The modification of the adsorbent consists of a first step where Ca(OH) 2 and the source of the replacing elements (Mg (OH) 2 or FeCl 3 or the Al compound) are mixed with the HCl solution. Afterwards, the resulting phase is mixed with the Na<sup>2</sup> SiO3 solution at high-stirring velocity, ideally over 2000 min−1. It is not easy to establish a single stoichiometry of the chemical reaction due to the wide variety of silicate species possible to be formed and to the proportion of hydroxyl and silanol groups that silicates would contain. Even the synthesized calcium silicate as an amorphous substance does not present a defined chemical structure as shown in the scheme of **Figure 1**. Notwithstanding, in all cases, the prepared substrate corresponds effectively to an amorphous material without a defined structure. Particles present a mean particle size averaging 0.5–1.0 µm forming larger agglomerates. Details of nanostructured calcium silicate have been described in former communications by Cairns et al. using 29Si-NMR spectroscopy and X-ray photoelectron spectroscopy (XPS) analysis [19]. They suggest the structure consists of a silicate backbone with plates of thickness around 10 nm. It would comprise silicate tetrahedral sites with Ca(II) ions and silanol groups on the surface forming a wollastonite-like structure, CaSiO 3 , where Ca(II) ions and silanol groups would act as probable binding sites. Cairns et al. propose that the calcium silicate would contain 1.5% of hydroxyl groups meaning that approximately 8% of silicon atoms in the structure are silanol groups allowing metallic ions such as Cu(II) and Zn(II) to form the corresponding hydroxides on the surface of the silicates acting as nucleation sites [20]. Porosimetry analyses using the N<sup>2</sup> sorptometer, indicate a variable surface area ranging between 80 and 300 m<sup>2</sup> /g, a value much higher than the surface area shown by other sorbents [21]. The mean pore diameter varies between 11 and 25 nm and the pore volume between 0.200 and 0.400 cm<sup>3</sup> /g for calcium silicates varying with the stirring velocity used during the synthesis.

X-ray diffraction analysis confirms that the prepared compounds are basically amorphous or at least polycrystalline. However, elements of patterns associated with wollastonite, CaSiO 3 , and larnyta-syn, Ca 2 SiO 4 , were observed confirming the synthesis of a calcium silicate. Although natural silicates are crystalline, like natural wollastonite, the solids prepared in this study appeared amorphous, probably because they were prepared by precipitation from aqueous solutions. Soluble silicates found in solutions comprise silicate ions of different size, most of them polymerized, which are not able to organise themselves in a crystal form resulting in a colloidal solid.

#### *2.6.2. Adsorption experiments*

**2.6. Results**

as Na<sup>2</sup>

*2.6.1. Synthesis and characterization*

the replacing elements (Mg (OH)

tonite-like structure, CaSiO 3

larnyta-syn, Ca 2 SiO 4

The chemical reaction between Ca(OH)

224 Physico-Chemical Wastewater Treatment and Resource Recovery

2

strate surface area comprising of micro-pores and meso-pores.

2 or FeCl 3

Afterwards, the resulting phase is mixed with the Na<sup>2</sup>

The modification of the adsorbent consists of a first step where Ca(OH)

and Na<sup>2</sup>

tate that produces particle agglomeration. The reaction of Ca(II) ions with soluble silicate

ideally over 2000 min−1. It is not easy to establish a single stoichiometry of the chemical reaction due to the wide variety of silicate species possible to be formed and to the proportion of hydroxyl and silanol groups that silicates would contain. Even the synthesized calcium silicate as an amorphous substance does not present a defined chemical structure as shown in the scheme of **Figure 1**. Notwithstanding, in all cases, the prepared substrate corresponds effectively to an amorphous material without a defined structure. Particles present a mean particle size averaging 0.5–1.0 µm forming larger agglomerates. Details of nanostructured calcium silicate have been described in former communications by Cairns et al. using 29Si-NMR spectroscopy and X-ray photoelectron spectroscopy (XPS) analysis [19]. They suggest the structure consists of a silicate backbone with plates of thickness around 10 nm. It would comprise silicate tetrahedral sites with Ca(II) ions and silanol groups on the surface forming a wollas-

ing sites. Cairns et al. propose that the calcium silicate would contain 1.5% of hydroxyl groups meaning that approximately 8% of silicon atoms in the structure are silanol groups allowing metallic ions such as Cu(II) and Zn(II) to form the corresponding hydroxides on the surface

the surface area shown by other sorbents [21]. The mean pore diameter varies between 11 and

X-ray diffraction analysis confirms that the prepared compounds are basically amorphous or at least polycrystalline. However, elements of patterns associated with wollastonite, CaSiO 3

ural silicates are crystalline, like natural wollastonite, the solids prepared in this study appeared amorphous, probably because they were prepared by precipitation from aqueous solutions. Soluble silicates found in solutions comprise silicate ions of different size, most of them polymerized, which are not able to organise themselves in a crystal form resulting in a colloidal solid.

, were observed confirming the synthesis of a calcium silicate. Although nat-

of the silicates acting as nucleation sites [20]. Porosimetry analyses using the N<sup>2</sup>

indicate a variable surface area ranging between 80 and 300 m<sup>2</sup>

25 nm and the pore volume between 0.200 and 0.400 cm<sup>3</sup>

the stirring velocity used during the synthesis.

SiO 3

 SiO 3 resulted in the precipitation of calcium silicate, an insoluble solid, where Ca(II) appeared to be strongly bound to a silicate backbone. The extent of the reaction and the particle size of the solid formed depend strongly on the pH of the reaction mixture, the proportions and concentration of calcium and silicate ions and the intensity of the stirring employed during the process. Intensive stirring in basic medium produces a quite homogeneous but colloidal and amorphous insoluble calcium silicate. The higher the stirring velocity, the smaller the particle size of the resultant calcium silicate and the larger and more accessible the sub-

is fast generating a thixotropic precipi-

2

SiO3 solution at high-stirring velocity,

or the Al compound) are mixed with the HCl solution.

, where Ca(II) ions and silanol groups would act as probable bind-

and the source of

sorptometer,

, and

/g, a value much higher than

/g for calcium silicates varying with

Sorption experiments are carried out to measure the capability of these nanostructured calcium silicates to uptake diverse metallic ions and some anions found in many residual industrial and mining wastewaters. Most of the experiments were conducted using the aqueous solutions whose initial metals concentrations have been described before. However, in order to establish the maximum metal sorption capacity of the nanostructured calcium silicates, some experiments were carried out using a feed-aqueous solution with higher content of metals (up to 5 g/L) (**Figure 4**).

**Figure 4.** Adsorption of some metallic ions onto nanostructured calcium silicate.

Metallic ions such as Cu(II), Zn(II) and Cd(II) readily form insoluble hydroxides under basic conditions. However, their removal by precipitation only with lime is far from an ideal solution when their content in solution is high enough to consider them pollutants. In the presence of nanostructured calcium silicate the extent of precipitation reached is significantly higher than that observed with only lime being used. The degree of removal of Cu(II), Zn(II) and Cd(II) ions is higher and it produces more stable structures using the nanostructured calcium silicates prepared in this study. In fact, it is reasonable to think that silicates kept the pH value in a more basic region acting as a buffer. Thereby, they ensure a good precipitation of metallic ions probably onto the surface of the formed silicate, rather than in the bulk solution. This way, the precipitates generated in this case are of more granular nature, meaning that they could settle or get filtrated more easily than the precipitates obtained using lime or NaOH. This fact suggested that the metal removal from aqueous solutions would be a cation exchange between the metals to remove and the calcium ions associated with the silicate structure.

**Figure 5** presents the results observed for the sorption of the anions PO<sup>4</sup> 3− , SO 4 2− and CrO 4 2− introduced in the test solutions employing as a sorbent in the same sample of nanostructured calcium silicate without modification. The adsorption of PO<sup>4</sup> 3- is fairly significant, SO<sup>4</sup> = is also adsorbed, until a maximum uptake point. Cr(VI) is not adsorbed at all probably because this ion is present in aqueous solution as anionic species such CrO 4 2− , HCrO<sup>4</sup> − or Cr 2 O 7 2− that do not tend to form insoluble salts with Ca(II) ions. The adsorption of PO<sup>4</sup> 3− was higher than that of SO 4 2− in agreement with the lower solubility of Ca 3 (PO4)<sup>2</sup> compared to CaSO 4 . The combined use of the nanostructured calcium silicate with Al(III) using poly-aluminium chloride improves the sulphate adsorption, even starting from initial solutions that contain over 2–5 g/L, obtaining a raffinate solution containing only a few mg/L accomplishing this way with the Chilean environmental regulations of discharge. The formation of ettringite, a double calcium and aluminium basic salt, during the adsorption would explain these results. Then, it is possible to achieve an almost complete sulphate removal.

**Figure 5.** Adsorption of some anionic species onto nanostructured calcium silicate.

The dissolved species diffuse from the solution to the surface of the adsorbent, and then to the internal structure. The rate of adsorption is usually limited by mass transfer and depends on the properties of the sorbate and sorbent. The equilibrium adsorption results of ionic species adsorption have been explained using conventional equilibrium isotherms including the Langmuir model, the empirical Freundlich model and the Redlich-Peterson isotherm, which is a hybrid sorption model employed to analyse experimental data that do not fit well with other mentioned models. Normally, Langmuir isotherm only can explain the experimental results when synthetic and quite pure and ideal aqueous solutions are used. Freundlich and Redlich-Peterson used to fit equilibrium experimental results when real mining or more complex chemical matrices are used as aqueous solutions. With respect to kinetics experiments, normally experimental results are satisfactorily well explained by applying a pseudo-secondorder kinetics model which is based on the sorption capacity of the sorbent.

#### **2.7. Conclusions**

calcium silicate without modification. The adsorption of PO<sup>4</sup>

226 Physico-Chemical Wastewater Treatment and Resource Recovery

in agreement with the lower solubility of Ca 3 (PO4)<sup>2</sup>

to achieve an almost complete sulphate removal.

SO 4 2−

ion is present in aqueous solution as anionic species such CrO 4

tend to form insoluble salts with Ca(II) ions. The adsorption of PO<sup>4</sup>

adsorbed, until a maximum uptake point. Cr(VI) is not adsorbed at all probably because this

of the nanostructured calcium silicate with Al(III) using poly-aluminium chloride improves the sulphate adsorption, even starting from initial solutions that contain over 2–5 g/L, obtaining a raffinate solution containing only a few mg/L accomplishing this way with the Chilean environmental regulations of discharge. The formation of ettringite, a double calcium and aluminium basic salt, during the adsorption would explain these results. Then, it is possible

The dissolved species diffuse from the solution to the surface of the adsorbent, and then to the internal structure. The rate of adsorption is usually limited by mass transfer and depends on the properties of the sorbate and sorbent. The equilibrium adsorption results of ionic species adsorption have been explained using conventional equilibrium isotherms including the Langmuir model, the empirical Freundlich model and the Redlich-Peterson isotherm, which is a hybrid sorption model employed to analyse experimental data that do not fit well with other mentioned models. Normally, Langmuir isotherm only can explain the experimental results when synthetic and quite pure and ideal aqueous solutions are used. Freundlich and Redlich-Peterson used to fit equilibrium experimental results when real mining or more complex chemical matrices are used as aqueous solutions. With respect to kinetics experiments, normally experimental results are satisfactorily well explained by applying a pseudo-second-

order kinetics model which is based on the sorption capacity of the sorbent.

**Figure 5.** Adsorption of some anionic species onto nanostructured calcium silicate.

3- is fairly significant, SO<sup>4</sup>

2− , HCrO<sup>4</sup> − or Cr 2 O 7 2−

compared to CaSO 4

3−

= is also

that do not

was higher than that of

. The combined use

In general, the modification with other atoms like Fe, Al and Mg of the nanostructured calcium silicate hydrate does not affect its high potential as adsorbent for treating polluted acidic solutions. The modified materials are superior to the unmodified material in terms of stability and sorption capacity by forming highly insoluble and stable double salts. The silicates would act as seeding material of insoluble hydroxides and salt species.

## **3. Natural sorbents case study: biosorbents - pine bark**

#### **3.1. Introduction**

Several natural and low-cost products have been assessed as sorbents or as removal agents for heavy metal ions and other pollutants present in industry-discarded aqueous solutions [22–25]. Some examples of biosorbents are sugar maple [26]; mulch [27]; papaya wood [28]; tea industry wastes [29]; sawdust [30] and pine bark [31]. Particularly, the primary use of bark (and other by-products of the wood industry) is mainly associated with fuel with little added value. Therefore, their use as a sorbent is expected to produce an increase of the overall industry sustainability. Moreover, its use as biosorbent may become especially interesting when the pollutant loaded material can undertake further steps leading to its reutilisation as a sorbent (elution, pyrolysis, etc.) or carry on with other uses such as the production of activated carbon.

In this case, the adsorption efficiency, evaluated in terms of the adsorption capacity, varies from only a few micrograms up to 200 mg per grams of dry bark. The efficiency of the overall wastewater treatment process depends on the type of natural sorbent, the nature of the metal or pollutant to be adsorbed, the initial pollutant concentration, pH, temperature, pulp density, the cell design (batch, column, etc.) and the contacting time. In fact, many scientific reports have revealed that this material can adsorb and act simultaneously as biofilter for several types of pollutants, such as heavy metals and organic products.

This chapter focuses on the treatment of wastewaters containing heavy metals commonly observed at large scale operations related to industrial activities such as mining and metallurgical processes [32, 33]. The use of adsorption techniques overcomes this problem leaving aqueous solutions with heavy metal concentrations typically in the order of parts per billion [28].

It has been shown that the governing mechanism for the adsorption of cationic heavy metals using biosorbents is based on ionic exchange reactions [34], which occurs throughout the removal of protons present in the different bark molecular structures (Eq. (13)),

$$R-OH + M^{n+} \rightarrow \ R-O \, M^{(n+1)+} + H^{\*} \tag{13}$$

where *R* represents the organic structure of the sorbent, *M<sup>n</sup>*<sup>+</sup> is the dissolved heavy metal ion in solution. As a consequence of the adsorption, the acidity of the aqueous phase increases. The adsorption may coordinate one or more adsorption sites and the charge of the metallic ion can be totally or partially compensated by the surface charge of the biosorbent.

#### **3.2. Substrate pre-treatment**

One problem associated with the use of pine bark in the removal of heavy metals is its content of natural organic complexation agents such as tannins, which can be released in significant concentrations and stabilise contaminants in solution rather than removing them. There are a number of biosorbent pre-treatments reported in the literature aiming at removing or fixing such soluble compounds and simultaneously improving the subsequent pollutant adsorption efficiency/ capacity. For instance, it has been shown that best adsorption results are obtained throughout carrying out a prior 'bark chemical activation'. This treatment consists of immersing the pine bark into diluted acid media at slightly high temperatures such as 50°C–70°C [34–36] before performing the adsorption tests. The acidic solution needs to be diluted to secure the integrity of the sorbent during the pre-treatment. The pH range in which the impact of chemical activation is the highest is narrow and it depends on the heavy metal nature and on the pulp density used [37]. Another pre-treatment consists of just washing the pine bark with water facilitating the removal of tannins and other soluble species which, if not removed, could again, irreversibly pollute even more the wastewater.

#### **3.3. Experimental procedures**

This experimental part gathers a number of findings concerning the bark structure, its response to chemical activation, and experiments generating adsorption equilibrium and kinetic data.

#### *3.3.1. Pine bark characterisation and preparation*

Pine bark samples are reduced in size to particle size distributions below 1 mm diameter (16# Tyler). The chemical activation is commonly performed using sulphuric acid 0.1–0.2 M and up to 1 M for 2 h at a temperature ranging from 20°C to 50°C with a solid to liquid ratio equal to 1:10.

Pine bark washing procedures are carried out with distilled water in a Soxhlet apparatus. Scanning electron microscopy (FEI Quanta 250 FEG SEM) is used to analyse the pine bark unloaded and loaded with the pollutant. Semi-quantitative elemental analysis was performed by EDX analysis of fluorescence intensities. FT-IR spectroscopy experiments are carried out on an IFS 55 spectrometer (Bruker) using diffuse reflectance mode (Harrick Attachment). The detector was of an MCT type and cooled at liquid nitrogen temperature (77 K).

XPS analysis is performed using a Kratos Axis Ultra instrument with a monochromatic Al Kα source. Samples are fixed on the sample holder using double-sided tape.

#### *3.3.2. Adsorption studies*

In all cases, there is an initial concentration of the pollutant (heavy metal) commonly in the dissolved state in the aqueous phase and introduced as a sulphate or a nitrate salt. The pH condition is close to the point where the hydrolysis starts producing insoluble or neutral species. All tests are carried out at 20°C and the pulp density used varies from 1 to 20 g sorbent/L aqueous solution unless stated otherwise. The adsorbed amount of pollutant is obtained by computing the difference between the initial amount of pollutant and the residual concentration after adsorption. The metal concentration in solution is obtained using atomic absorption spectrophotometry (AAS) Perkin Elmer Model 2380.

Equilibrium data are obtained at least after 48 h contact between the sorbent and the polluted aqueous phase. Kinetic data are generated following similar experimental protocols

In all cases, high purity reagents were used. Further details or changes in this procedure are pointed out when appropriate.

### **3.4. Results**

**3.2. Substrate pre-treatment**

228 Physico-Chemical Wastewater Treatment and Resource Recovery

more the wastewater.

equal to 1:10.

*3.3.2. Adsorption studies*

**3.3. Experimental procedures**

*3.3.1. Pine bark characterisation and preparation*

One problem associated with the use of pine bark in the removal of heavy metals is its content of natural organic complexation agents such as tannins, which can be released in significant concentrations and stabilise contaminants in solution rather than removing them. There are a number of biosorbent pre-treatments reported in the literature aiming at removing or fixing such soluble compounds and simultaneously improving the subsequent pollutant adsorption efficiency/ capacity. For instance, it has been shown that best adsorption results are obtained throughout carrying out a prior 'bark chemical activation'. This treatment consists of immersing the pine bark into diluted acid media at slightly high temperatures such as 50°C–70°C [34–36] before performing the adsorption tests. The acidic solution needs to be diluted to secure the integrity of the sorbent during the pre-treatment. The pH range in which the impact of chemical activation is the highest is narrow and it depends on the heavy metal nature and on the pulp density used [37]. Another pre-treatment consists of just washing the pine bark with water facilitating the removal of tannins and other soluble species which, if not removed, could again, irreversibly pollute even

This experimental part gathers a number of findings concerning the bark structure, its response to chemical activation, and experiments generating adsorption equilibrium and kinetic data.

Pine bark samples are reduced in size to particle size distributions below 1 mm diameter (16# Tyler). The chemical activation is commonly performed using sulphuric acid 0.1–0.2 M and up to 1 M for 2 h at a temperature ranging from 20°C to 50°C with a solid to liquid ratio

Pine bark washing procedures are carried out with distilled water in a Soxhlet apparatus. Scanning electron microscopy (FEI Quanta 250 FEG SEM) is used to analyse the pine bark unloaded and loaded with the pollutant. Semi-quantitative elemental analysis was performed by EDX analysis of fluorescence intensities. FT-IR spectroscopy experiments are carried out on an IFS 55 spectrometer (Bruker) using diffuse reflectance mode (Harrick Attachment). The

XPS analysis is performed using a Kratos Axis Ultra instrument with a monochromatic Al Kα

In all cases, there is an initial concentration of the pollutant (heavy metal) commonly in the dissolved state in the aqueous phase and introduced as a sulphate or a nitrate salt. The pH condition is close to the point where the hydrolysis starts producing insoluble or neutral species. All tests are carried out at 20°C and the pulp density used varies from 1 to 20 g sorbent/L

detector was of an MCT type and cooled at liquid nitrogen temperature (77 K).

source. Samples are fixed on the sample holder using double-sided tape.

#### *3.4.1. Characterisation and modelling of the substrate*

**Table 3** shows the chemical composition of both raw and chemically (acid) activated bark resulting from the EDX elemental analysis of original washed bark and after the sulphuric acid treatment. As expected, most of the soluble species decrease their content due to the activation. The acid treatment reduces the concentrations of K , Mg , Mn and Ca significantly, whereas other metal concentrations ( Na , Al , Si , Ti , Fe , Cu ) remain almost constant.


**Table 3.** Semi-quantitative EDX analysis of pine bark (values are in mg per g of dry bark).

**Figure 6** shows the DRIFT analysis obtained for a section of the finger print region. A broad band is observed from 3600 to 3000 cm–1, indicating the presence of O − H and C − H stretching vibrations. Between 2000 and 1500 cm–1 there is a strong band at 1610 cm–1 representing the carbonyl C = O stretch. When Pb(II) is adsorbed onto bark, a broadening of the band occurs and a simultaneous modifications at wavenumbers of approximately 1512 cm−1 related to aromatic skeletal vibrations. These changes are in agreement with the structures which are easier to ionise (**Figure 7**).

A high concentration of the metallic ions in solution may then activate the adsorption sites having lower ionization constant ultimately increasing the specific adsorption.

**Figure 8** shows the SEM images after sorption using low pulp densities conditions. EDX elemental analysis at the spots vmarked in the figures revealed a high heterogeneity in the Pb distribution across the bark. The Pb/O weight ratio varied from 2:1 to 3:1 regardless the pulp density value. Similar observations of varying metal content were previously seen for untreated bark [39].

**Figure 6.** DRIFTS spectra for activated bark (a) and bark charged with solutions containing 1 (b), 1.5 (c) and 3 (d) g Pb L−1. Pulp density: 1.5 g/L.

**Figure 7.** Schematics of the organic groups, its relative proportion in the structure of bark, and radicals through which the ionization could lead to heavy metals adsorption onto bark. *Ka* represents the acidity constant, C : carbon, O : oxygen, Ar : aromatic structure (from Ref. [38]).

**Figure 8.** SEM images of activated pine bark loaded with adsorbed Pb(II). Left: pulp density 1.0 g/L. EDX analysis indicates Pb/O weight ratios of 2:1 and 3:1 in spots (a) and (b), respectively. Right: pulp density 1.5 g/L. EDX indicates Pb/O weight ratios decreasing from (a) to (c) ranging from approximately 2:1 to 3:1.

#### *3.4.2. Adsorption tests: thermodynamics and kinetics*

**Figure 6.** DRIFTS spectra for activated bark (a) and bark charged with solutions containing 1 (b), 1.5 (c) and 3 (d) g Pb

**Figure 7.** Schematics of the organic groups, its relative proportion in the structure of bark, and radicals through which the ionization could lead to heavy metals adsorption onto bark. *Ka* represents the acidity constant, C : carbon, O : oxygen,

L−1. Pulp density: 1.5 g/L.

230 Physico-Chemical Wastewater Treatment and Resource Recovery

Ar : aromatic structure (from Ref. [38]).

As an example, **Figure 9** shows the adsorption percentage of Cu(II) from batch experiments. Langmuir and Freundlich isotherms fit well with the experimental data. The adsorption capacity varies from one chemical element to another and as a function of the experimental conditions used. Pb(II) ions reach similar adsorption capacity at 10 g/L pulp density but it can reach 90 mg/g at 1 g/L. Concerning adsorption competitive reactions, when comparing Cu(II) and Zn(II) adsorption, Cu(II) reaches higher adsorption capacities than Zn(II). Binary adsorption reveals both metallic ions which do not interfere in the adsorption of each other reaching similar adsorption capacities when mixed and separated; however, this is not the case for all metals. Al-Asheh and Duvnjak (1997) have proved that Ni2+ and Cu2+ interfere slightly whereas the pairs (Cu2+, Cd2+) and (Cd2+, Ni2+) show a more significant interaction [22].

**Figure 9.** Adsorption isotherm of copper (II) ions onto washed pine bark at pH 5, 10 g/L pulp density [40].

If the heavy metal needs to be recovered, elution stages may allow removing and concentrating at least 90 and 71% or Cu(II) and Zn(II), respectively. The resulting solutions are 20 and 9 times more concentrated than the original polluted solution [41]. In consequence, the zinc surface complexes are more stable than the ones of copper. **Figure 10** shows the adsorption percentage and the specific adsorption results of copper obtained at different pulp densities with washed and activated pine bark.

**Figure 10.** (a) Kinetics of adsorption percentage (%) and (b) specific adsorption (mg copper (II) ions/g dry bark) of copper (II) ions onto washed and activated bark [42].

The higher the pulp density (number of sorbent surface sites available), the higher the amount of metal removed from solution. Depending on the pulp density, the concentration of the heavy metal in solution and the contact time of the solution with the bark, the adsorption percentage may reach a *plateau* or maximum value. Such maximum value is usually referred to as 'saturation' which might be misleading. It can also be observed that the higher the pulp density, the lower the specific adsorption. At constant initial metal concentration, the increase of pulp density favours the adsorption process on bark surface sites that are easier to reach and to be replaced by a metal ion (high energy surface sites). In other words, the material is not being used at its maximum capacity.

Moving from the high to low pulp density values, the specific adsorption increases. This indicates that higher concentrations of copper (II) ions activate sites. This is expected, given that the increase of copper (II) ions will shift the adsorption equilibrium constants of secondary reactions towards increasing the metal concentration in its adsorbed state. Indeed, the ion exchange mechanism (the main responsible for metal adsorption) occurs throughout many organic radicals with different acidity constants.

#### *3.4.3. Surface analysis of the sorbent*

If the heavy metal needs to be recovered, elution stages may allow removing and concentrating at least 90 and 71% or Cu(II) and Zn(II), respectively. The resulting solutions are 20 and 9 times more concentrated than the original polluted solution [41]. In consequence, the zinc surface complexes are more stable than the ones of copper. **Figure 10** shows the adsorption percentage and the specific adsorption results of copper obtained at different pulp densities

The higher the pulp density (number of sorbent surface sites available), the higher the amount of metal removed from solution. Depending on the pulp density, the concentration of the heavy metal in solution and the contact time of the solution with the bark, the adsorption percentage may reach a *plateau* or maximum value. Such maximum value is usually referred to as 'saturation' which might be misleading. It can also be observed that the higher the pulp

**Figure 10.** (a) Kinetics of adsorption percentage (%) and (b) specific adsorption (mg copper (II) ions/g dry bark) of copper

with washed and activated pine bark.

232 Physico-Chemical Wastewater Treatment and Resource Recovery

(II) ions onto washed and activated bark [42].

Several scientific reports have been devoted only to analyse and to interpret the results from X-ray photoelectron spectra of wood and compounds derived from it [43, 44]. **Figure 11** shows the O1s emission lines obtained for the untreated sample of pine bark and charged with Pb(II) ions. The energy scale was fixed and referred to 285 eV, in agreement with most studies reported [44].

**Figure 11.** XPS spectra of pine bark samples. Sample 1: **Original** pine bark; sample 2 loaded for 24 h with Pb(II) in an acid aqueous solution (pH 5) at 5 g/L pulp density [45].

The O1s line is strongly perturbed by Pb(II) adsorption. The peak at the lowest position shifts to lower binding energy values, confirming a reductive environment due to Pb(II) adsorption. The peak associated with O-C-O is similarly shifted towards lower binding energies but in smaller magnitude compared to that representing C-O- type structures. Additionally, the intensity of C-O peak increases relative to the other peak. This suggests that not all sites present similar active sites. Most of them have C-O sites for adsorption while the amount of phenolic sites varies locally through bark structure. The mechanism, therefore, should take place in well distributed C-O groups and as a second type of sites in significance it would be the phenolic groups. The latter is in agreement with DRIFT results.

#### **3.5. Conclusions**

Results have shown that the adsorption efficiency depends not only on the pollutant nature but also on the presence of other species which may interact with the pollutant interfering positively or negatively on the adsorption performance. There are cases where the simultaneous presence of different pollutants results in influencing the overall adsorption capacity (for example Cd(II) and other divalent ions) and other cases where pollutants adsorb regardless the presence of others (for example Cu(II) and Zn(II)). The elution process carried out with 1N nitric acid solution proved to be promising to concentrate both ions. Cu(II) ions were more likely to be destabilised than Zn(II). The tests performed with a pulp density of 6 mg/g dry bark resulted in solutions 20 and 9 times more concentrated in the case of Cu(II) and Zn(II), respectively. The chemical activation of pine bark material increases the rate at which copper (II) ions are adsorbed and it simultaneously reduces the maximum achievable adsorption equilibrium. The chemical activation is useful primarily at high pulp density values (above 10 g/L). The adsorption capacity for Pb(II) at very low pulp density is about 93.7 mg Pb(II)/g dry bark. Mono-hydroxylated species and free lead ions were presented as the major responsible for adsorption. DRIFTS analysis revealed that the adsorption mechanism is complex, mainly driven by bark surface sites involving C-O groups and this was confirmed by XPS analysis. Phenolic and cellulose oxygen sites are also relevant.

## **Acknowledgements**

The authors wish to thank The National Fund for Development of Science and Technology of Chile, FONDECYT, for all the financial support throughout the Grant N° 1140331. The authors would like to dedicate this manuscript to Professor Sergio Montes Sotomayor for all his contributions to the field of wastewater treatment. We thank Mrs. Ana Maria Rojo MSc. for all her comments that greatly improved the manuscript.

## **Author details**

Gonzalo Montes-Atenas1,\* and Fernando Valenzuela<sup>2</sup>

\*Address all correspondence to: gmontes@ing.uchile.cl

1 Minerals and Metals Characterisation and Separation (M<sup>2</sup> CS) Research Group, Departamento de Ingeniería de Minas, Facultad de Ciencias Físicas y Matemáticas, Universidad de Chile, Santiago, Chile

2 Laboratorio de Operaciones Unitarias e Hidrometalurgia, Facultad de Ciencias Químicas y Farmacéuticas, Universidad de Chile, Independencia, Santiago, Chile

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place in well distributed C-O groups and as a second type of sites in significance it would be

Results have shown that the adsorption efficiency depends not only on the pollutant nature but also on the presence of other species which may interact with the pollutant interfering positively or negatively on the adsorption performance. There are cases where the simultaneous presence of different pollutants results in influencing the overall adsorption capacity (for example Cd(II) and other divalent ions) and other cases where pollutants adsorb regardless the presence of others (for example Cu(II) and Zn(II)). The elution process carried out with 1N nitric acid solution proved to be promising to concentrate both ions. Cu(II) ions were more likely to be destabilised than Zn(II). The tests performed with a pulp density of 6 mg/g dry bark resulted in solutions 20 and 9 times more concentrated in the case of Cu(II) and Zn(II), respectively. The chemical activation of pine bark material increases the rate at which copper (II) ions are adsorbed and it simultaneously reduces the maximum achievable adsorption equilibrium. The chemical activation is useful primarily at high pulp density values (above 10 g/L). The adsorption capacity for Pb(II) at very low pulp density is about 93.7 mg Pb(II)/g dry bark. Mono-hydroxylated species and free lead ions were presented as the major responsible for adsorption. DRIFTS analysis revealed that the adsorption mechanism is complex, mainly driven by bark surface sites involving C-O groups and this was confirmed by XPS analysis. Phenolic and cellulose oxygen sites

The authors wish to thank The National Fund for Development of Science and Technology of Chile, FONDECYT, for all the financial support throughout the Grant N° 1140331. The authors would like to dedicate this manuscript to Professor Sergio Montes Sotomayor for all his contributions to the field of wastewater treatment. We thank Mrs. Ana Maria Rojo MSc. for

de Ingeniería de Minas, Facultad de Ciencias Físicas y Matemáticas, Universidad de Chile,

2 Laboratorio de Operaciones Unitarias e Hidrometalurgia, Facultad de Ciencias Químicas y

CS) Research Group, Departamento

the phenolic groups. The latter is in agreement with DRIFT results.

234 Physico-Chemical Wastewater Treatment and Resource Recovery

**3.5. Conclusions**

are also relevant.

**Author details**

Santiago, Chile

**Acknowledgements**

all her comments that greatly improved the manuscript.

Gonzalo Montes-Atenas1,\* and Fernando Valenzuela<sup>2</sup> \*Address all correspondence to: gmontes@ing.uchile.cl

1 Minerals and Metals Characterisation and Separation (M<sup>2</sup>

Farmacéuticas, Universidad de Chile, Independencia, Santiago, Chile


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#### **Integral use of** *Nejayote***: Characterization, New Strategies for Physicochemical Treatment and Recovery of Valuable By-Products Integral use of** *Nejayote***: Characterization, New Strategies for Physicochemical Treatment and Recovery of Valuable By-Products**

Eduardo Alberto López-Maldonado, Mercedes Teresita Oropeza-Guzmán and Karla Alejandra Suárez-Meraz Eduardo Alberto López-Maldonado, Mercedes Teresita Oropeza-Guzmán and Karla Alejandra Suárez-Meraz

Additional information is available at the end of the chapter Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/66223

#### **Abstract**

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[43] Nzokou P, Kamdem DP, You have free access to this content X-ray photoelectron spectroscopy study of red oak-(*Quercus rubra*), black cherry- (Prunus serotina) and red pine- (*Pinus resinosa*) extracted wood surfaces, *Surface and Interface Analysis*, 2005; 37: 689–694.

[44] Shchukarev A, Sundberg B, Mellerowicz E, Persson P, XPS study of living tree, *Surface* 

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1312 p. ISBN-10: 084005498X

238 Physico-Chemical Wastewater Treatment and Resource Recovery

10.1260/0263-6174.27.5.513

10.1007/s12665-014-3066-3

DOI: 10.1002/sia.2064

In this research, an innovative physicochemical strategy is presented to address the problem of *nejayote*, from two perspectives: the first focusing on sanitation and reuse of *nejayote* using waste from shrimp shells, thereby adding value to the recovered solids of *nejayote*. Zeta potential measurements are a proactive electrochemical tool to define the strategy to allow integral use of *nejayote* in the industry nixtamalization. The treated water can be discharged from the municipal sewer system using a process of coagulation-flocculation, with an optimal dose of 1250 mg/L chitosan at pH 5, achieving removal of up to 80% of total suspended solids and turbidity. Moreover, zeta potential measurements show that the anionic biopolyelectrolyte obtained from *nejayote* has potential to be applied in the area of water treatment as a green chelating agent.

**Keywords:** nejayote, nixtamalization, biopolyelectrolytes, zeta potential, coagulationflocculation

## **1. Introduction**

Nixtamalized products such as maize tortillas originated in Mexico, are the main sources of energy, protein, calcium and other important nutrients and are considered the national breads and consumed with other fillings such as beans, meats, eggs and vegetables [1–3].

© 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. © 2017 The Author(s). Licensee InTech. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

The ancient, laborious or traditional process (nixtamalization) to obtain tortillas is a process widely used by indigenous people in Mexico (41%), the Southern United States, Central America, Asia and parts of Europe, that consumes significant amounts of water, energy and time [2]. Traditional maize is lime-cooked in clay pots over a fire, followed by steeping for 8– 16 h (generally overnight), the supernatant called maize wastewater or commonly known as "*nejayote*", derived from the Nahuatl word meaning "lime broth ashes" is discarded and then the nixtamal is hand-washed. Nixtamal is ground into a fine masa with a stone grinder called metate and then hand-molded, patted or pressed into disks, which are baked on both sides on a hot griddle [4–10].

Nixtamalization causes a loss of about 5% by weight dry basis of corn; 3% is suspended and the remaining 2% is dissolved. The suspended matter can be separated easily and inexpensively by sedimentation and the dissolved substance should "precipitate" to separate solids which is also done by sedimentation [2, 11–16].

A typical maize nixtamalization facility, processing 50 kg of maize everyday, uses over 75 L of water per day and generates nearly the equivalent amount of alkaline wastewater on a daily basis [4]. The estimated monthly volume of *nejayote* generated in Mexico is about 1.2 m3 [17].

*Nejayote* is considered an environmental pollutant because it is an alkaline wastewater, with high chemical and biological oxygen demand [2, 9]. Due to the presence of lime in the process, the pH of the wastewater is very high (12–14), with a high temperature between 40 and 70°C), containing suspended solids (corn husks and broken grains) and a very high portion of dissolved material from the alkaline hydrolysis of corn components [14]. The *nejayote* with these physicochemical characteristics is thrown, often without treatment, into drainage systems and even directly to the soil and groundwater. Thus, alternatives for sanitation of *nejayote* and utilization are needed [14]. Among the solutions that have been reported, they are from biological treatment processes [9], membrane filtration, nixtamalization methods that minimize water use and the use of *nejayote* as a supplement in animal feed [17].

In this research, an innovative physicochemical strategy is presented to address the problem of *nejayote* from two perspectives: the first focused on remediation of *nejayote* and the second is on water reuse using biopolyelectrolyte (BPE) from waste shrimp shells. With the use of effluents generated by 20 tons of corn nixtamalized equivalent to one ton of corn or sorghum protein is obtained [2, 16]. Another benefit, both economic and social, which could have *nejayote* recovery is that wastewater could be recycled, either in nixtamalization industry itself or for any other use. The second is based on the use of *nejayote* for obtaining anionic BPE (maize gum) for treating wastewater from electroplating industry. In both cases, zeta potential (*ζ*) measurements as electrochemical parameters were used to develop the process of sanitation and water reuse and for the extraction and application of anionic BPE in the separation of heavy metals.

## **2. Experimental**

The ancient, laborious or traditional process (nixtamalization) to obtain tortillas is a process widely used by indigenous people in Mexico (41%), the Southern United States, Central America, Asia and parts of Europe, that consumes significant amounts of water, energy and time [2]. Traditional maize is lime-cooked in clay pots over a fire, followed by steeping for 8– 16 h (generally overnight), the supernatant called maize wastewater or commonly known as "*nejayote*", derived from the Nahuatl word meaning "lime broth ashes" is discarded and then the nixtamal is hand-washed. Nixtamal is ground into a fine masa with a stone grinder called metate and then hand-molded, patted or pressed into disks, which are baked on both sides on

Nixtamalization causes a loss of about 5% by weight dry basis of corn; 3% is suspended and the remaining 2% is dissolved. The suspended matter can be separated easily and inexpensively by sedimentation and the dissolved substance should "precipitate" to separate solids

A typical maize nixtamalization facility, processing 50 kg of maize everyday, uses over 75 L of water per day and generates nearly the equivalent amount of alkaline wastewater on a daily basis [4]. The estimated monthly volume of *nejayote* generated in Mexico is about

*Nejayote* is considered an environmental pollutant because it is an alkaline wastewater, with high chemical and biological oxygen demand [2, 9]. Due to the presence of lime in the process, the pH of the wastewater is very high (12–14), with a high temperature between 40 and 70°C), containing suspended solids (corn husks and broken grains) and a very high portion of dissolved material from the alkaline hydrolysis of corn components [14]. The *nejayote* with these physicochemical characteristics is thrown, often without treatment, into drainage systems and even directly to the soil and groundwater. Thus, alternatives for sanitation of *nejayote* and utilization are needed [14]. Among the solutions that have been reported, they are from biological treatment processes [9], membrane filtration, nixtamalization methods that minimize water use and the use of *nejayote* as a supplement

In this research, an innovative physicochemical strategy is presented to address the problem of *nejayote* from two perspectives: the first focused on remediation of *nejayote* and the second is on water reuse using biopolyelectrolyte (BPE) from waste shrimp shells. With the use of effluents generated by 20 tons of corn nixtamalized equivalent to one ton of corn or sorghum protein is obtained [2, 16]. Another benefit, both economic and social, which could have *nejayote* recovery is that wastewater could be recycled, either in nixtamalization industry itself or for any other use. The second is based on the use of *nejayote* for obtaining anionic BPE (maize gum) for treating wastewater from electroplating industry. In both cases, zeta potential (*ζ*) measurements as electrochemical parameters were used to develop the process of sanitation and water reuse and for the extraction and application of anionic BPE in the separation of heavy

a hot griddle [4–10].

1.2 m3 [17].

in animal feed [17].

metals.

which is also done by sedimentation [2, 11–16].

240 Physico-Chemical Wastewater Treatment and Resource Recovery

It is shown in **Figure 1** that zeta potential measurements were used to interconnect the physicochemical characteristics of *nejayote* and chitosan flocculant capacity to achieve sanitation and water reuse in nixtamalization industry. In the first stage, plots of *ζ* vs pH of *ne‐ jayote*, chitosan and maize gum were constructed to determine the behavior of surface charge and isoelectric point (IEP). Then the strategic dosage of chitosan was done in the process of coagulation-flocculation of *nejayote*. The coagulation-flocculation window was constructed by measuring the water-quality parameters of environmental interest (turbidity and total suspended solids) and zeta potential. Moreover, zeta potential measurements were used to to exploring the interaction capacity of maize gum obtained from *nejayote* with metal ions, frequently contained in wastewater from the electroplating industry.

**Figure 1.** Using zeta potential measurements for *nejayote* sanitation and water reuse, and its use for obtaining a green flocculant for the separation of heavy metals.

#### **2.1. Materials**

Commercial testing water-quality reagents from HACH® were used. Milli-Q grade water was used in all the experiments. All other reagents were of analytical grade and were used without further purification.

#### *2.1.1. Wastewater sampling in the nixtamalization industry*

*Nejayote* was provided by a local tortilla-making industry. The wastewater sampling protocol was followed as recommended by Mexican sampling standard (NMX-AA-003-1980).

#### *2.1.2. Chitosan extraction from waste shrimp shells*

Chitosan is obtained from waste shrimp shells using the method proposed by the authors Goycoolea et al. [15].

### *2.1.3. Maize gum extraction of nejayote*

Maize gum was obtained by fractional separation, using hexane, ethanol and hydrochloric acid, isopropanol, acetone, methanol formed by the steps of desalmidonado, deproteinization, delipidation, delignification which are proposed by the authors of [8, 18, 19].

## **2.2. Methods**

## *2.2.1. Physicochemical characterization of nejayote*

The main parameters of quality wastewater used in this research were performed following the Mexican standard procedures to determine the biochemical oxygen demand (BOD5), chemical oxygen demand (COD), total nitrogen (TN), the solids content, total organic carbon (TOC), total phosphorus (TP), and other parameter fields such as pH, electrical conductivity (EC) and temperature were carried out based on the Hach methods.

#### *2.2.2. Profiles of ζ = f(pH) of nejayote, maize gum and chitosan*

The charge density, isoelectric point and chitosan-dosing strategy for treating *nejayote* were determined in a *ζ* = *f*(pH) plot. The zeta potential measurement was performed using the SZ-100 of Horiba Scientific equipment based on studies by López-Maldonado et al. [20, 21].

#### *2.2.3. Nejayote treatability tests by coagulation-flocculation using chitosan*

A sample of 20 mL of *nejayote* was taken in a vial and the pH was adjusted to 5. The chitosan dosage tests were performed in 20-mL-vials. Progressive additions of 0.1 g/L chitosan solution were done and after each one, the vials were shaken for 2 min at 250 rpm and 5 min at 50 rpm and allowed to settle for 5 more min. Finally the supernatant to a height of 2 cm from the vial was suctioned to determine the parameters of water quality in the supernatant.

#### *2.2.4. Evaluation of the capacity of polyelectrolyte maize gum for decontaminating wastewater*

The anionic BPE obtained from *nejayote* is characterized by Fourier Transform Infra-Red spectroscopy (FTIR), scanning electron microscopy (SEM) and measurements of zeta potential (*ζ*). FTIR spectra of maize gum were recorded using a Nicolet FT-IR spectrometer. The samples of maize gum were analyzed by SEM and X-ray microanalysis. The analysis was performed on SEM (ZEISS EVO-MA15), equipped with an EDS (energy dispersive spectroscopy) BRUKER detector microscope to observe the composition. The zeta potential measurement was performed using the SZ-100 of Horiba Scientific equipment based on studies by López-Maldonado et al. [20]. This was developed with the maize gum dispersion in a 0.1% solution, which took different levels of acidity and alkalinity in the range of 2–12 and injected into a cell with electrode graphite.

## **3. Results and discussion**

*2.1.3. Maize gum extraction of nejayote*

242 Physico-Chemical Wastewater Treatment and Resource Recovery

*2.2.1. Physicochemical characterization of nejayote*

**2.2. Methods**

al. [20, 21].

in the supernatant.

into a cell with electrode graphite.

Maize gum was obtained by fractional separation, using hexane, ethanol and hydrochloric acid, isopropanol, acetone, methanol formed by the steps of desalmidonado, deproteinization,

The main parameters of quality wastewater used in this research were performed following the Mexican standard procedures to determine the biochemical oxygen demand (BOD5), chemical oxygen demand (COD), total nitrogen (TN), the solids content, total organic carbon (TOC), total phosphorus (TP), and other parameter fields such as pH, electrical conductivity

The charge density, isoelectric point and chitosan-dosing strategy for treating *nejayote* were determined in a *ζ* = *f*(pH) plot. The zeta potential measurement was performed using the SZ-100 of Horiba Scientific equipment based on studies by López-Maldonado et

A sample of 20 mL of *nejayote* was taken in a vial and the pH was adjusted to 5. The chitosan dosage tests were performed in 20-mL-vials. Progressive additions of 0.1 g/L chitosan solution were done and after each one, the vials were shaken for 2 min at 250 rpm and 5 min at 50 rpm and allowed to settle for 5 more min. Finally the supernatant to a height of 2 cm from the vial was suctioned to determine the parameters of water quality

*2.2.4. Evaluation of the capacity of polyelectrolyte maize gum for decontaminating wastewater*

The anionic BPE obtained from *nejayote* is characterized by Fourier Transform Infra-Red spectroscopy (FTIR), scanning electron microscopy (SEM) and measurements of zeta potential (*ζ*). FTIR spectra of maize gum were recorded using a Nicolet FT-IR spectrometer. The samples of maize gum were analyzed by SEM and X-ray microanalysis. The analysis was performed on SEM (ZEISS EVO-MA15), equipped with an EDS (energy dispersive spectroscopy) BRUKER detector microscope to observe the composition. The zeta potential measurement was performed using the SZ-100 of Horiba Scientific equipment based on studies by López-Maldonado et al. [20]. This was developed with the maize gum dispersion in a 0.1% solution, which took different levels of acidity and alkalinity in the range of 2–12 and injected

delipidation, delignification which are proposed by the authors of [8, 18, 19].

(EC) and temperature were carried out based on the Hach methods.

*2.2.3. Nejayote treatability tests by coagulation-flocculation using chitosan*

*2.2.2. Profiles of ζ = f(pH) of nejayote, maize gum and chitosan*

#### **3.1. Physicochemical characterization of** *nejayote*

In this investigation the *nejayote* generated by a tortilla factory in Mexico was taken as the object of study. A typical maize nixtamalization facility, processing 500 kg of maize every day, uses over 750 L of water per day and generates nearly the equivalent amount of alkaline wastewater on a daily basis. **Figure 2** shows the stages of the nixtamalization process used for the manufacture of nixtamal mass and the generation of *nejayote*.

**Figure 2.** Diagram of the nixtamalization process and the point of generation *nejayote*.

As shown in **Table 1**, the physicochemical characteristics of *nejayote* concerning the content of organic matter determined by the parameters COD, TP, BOD5 and TOC normed indicate that najeyote exceeds the maximum permissible limits of NOM-002-SEMARNAT-1996. The *nejayote* has a pH of 11.6 as already well known, is a wastewater alkaline by the use of lime in the nixtamalization.



**Table 1.** Maize industry wastewater physicochemical composition.

For this research the measurement of other nonregulatory parameters was performed, and they are key to evaluate the performance of coagulation-flocculation process and determine the best operating conditions. *ζ* = −10 mV (pH = 12) and particle size of the dissolved part of *nejayote* (100–600 nm), which indicates containing dispersed particles very stable. Considering the surface charge of the *nejayote* colloids and particle size to be separated by coagulationflocculation, it requires the addition of a cationic BPE.

#### **3.2. Profiles of** *ζ* **=** *f***(pH) of** *nejayote* **and chitosan**

The zeta potential is a parameter by electrochemical nature that allows to study and predict the interactions occurring at the molecular level between the colloidal particles *nejayote* and the different ionic species of the medium, also it indicates the degree of stability of dispersion in an aqueous medium from the point electrically. The aim is to employ *ζ* measurements to know and understand the behavior of the BPE type chitosan in this kind of wastewater treatment (see **Figure 3**).

Surface charge of chitosan and *nejayote* colloids are pH-dependent and their behavior has great influence on coagulation-flocculation performance [22]. In addition, *ζ* measurements are required to characterize the colloidal system to understand repulsion and aggregation between colloidal particles.

In **Figure 4**, chitosan shows an amphoteric behavior, in the region of pH = 2–5.5 has a positive surface charge (*ζ* = 51.1 mV) due to protonation of amine groups, at pH = 6–10 its surface charge remains neutral, this is due to the insolubilization phenomenon occuring at pH > IEP (pH = 5– 6) of chitosan and increases their hydrophobicity.

Integral use of *Nejayote*: Characterization, New Strategies for Physicochemical Treatment... http://dx.doi.org/10.5772/66223 245

**Parameter** *Nejayote* **Maximum permissible limit**

For this research the measurement of other nonregulatory parameters was performed, and they are key to evaluate the performance of coagulation-flocculation process and determine the best operating conditions. *ζ* = −10 mV (pH = 12) and particle size of the dissolved part of *nejayote* (100–600 nm), which indicates containing dispersed particles very stable. Considering the surface charge of the *nejayote* colloids and particle size to be separated by coagulation-

The zeta potential is a parameter by electrochemical nature that allows to study and predict the interactions occurring at the molecular level between the colloidal particles *nejayote* and the different ionic species of the medium, also it indicates the degree of stability of dispersion in an aqueous medium from the point electrically. The aim is to employ *ζ* measurements to know and understand the behavior of the BPE type chitosan in this kind of wastewater

Surface charge of chitosan and *nejayote* colloids are pH-dependent and their behavior has great influence on coagulation-flocculation performance [22]. In addition, *ζ* measurements are required to characterize the colloidal system to understand repulsion and aggregation between

In **Figure 4**, chitosan shows an amphoteric behavior, in the region of pH = 2–5.5 has a positive surface charge (*ζ* = 51.1 mV) due to protonation of amine groups, at pH = 6–10 its surface charge remains neutral, this is due to the insolubilization phenomenon occuring at pH > IEP (pH = 5–

Temperature (°C) 30–39 40b Color (Pt-Co) 5653–8580 NI pH 11.61–12.1 5.5–10a Chemical Oxygen Demand, COD (mg O2/L) 9800–28,450 NI Total Organic Carbon, TOC (mg C/L) 7337–9836 NI Inorganic Carbon, IC (mg/L) 23–28 NI Total Carbon, TC (mg C/L) 7360–9864 NI Biochemical Oxygen Demand, BOD5 (mg O2/L) 2700 200b Total Phosphorus, TP (mg P/L) 905–1321 30b Total Nitrogen, TN (mg N/L) 303–418 60b Biodegradability (BOD5/COD) 0.27 NI

244 Physico-Chemical Wastewater Treatment and Resource Recovery

NI= Not included in the standard.

**Table 1.** Maize industry wastewater physicochemical composition.

flocculation, it requires the addition of a cationic BPE.

**3.2. Profiles of** *ζ* **=** *f***(pH) of** *nejayote* **and chitosan**

6) of chitosan and increases their hydrophobicity.

NOM-001-SEMARNAT-1996.

NOM-002-SEMARNAT-1996.

treatment (see **Figure 3**).

colloidal particles.

a

b

**Figure 3.** Model of the electrical double layer and zeta potential concept adopted for sanitation of *nejayote*.

**Figure 4.** Electrokinetic properties of *nejayote* and biopolyelectrolyte type chitosan.

Moreover, the *nejayote* (*ζ* = 0) has a negative surface charge throughout the pH range. From the electrical viewpoint, at pH = 5, the interaction between oppositely charged species chitosan*nejayote* is ensured and therefore the strategic dosage cationic BPE was performed.

#### **3.3.** *Nejayote* **treatability tests by coagulation-flocculation using chitosan**

Dosing strategy for chitosan was determined by *ζ* of *nejayote* colloids and chitosan, and also by observing critical pH value of the IEP. In this study, charged chitosan purpose is to reduce the repulsion forces between particles by neutralizing the negatively charged molecules. In general, the electroneutrality zone for chitosan-*nejayote* system is below pH = 6 (see **Figure 4**), this has a practical application since higher charge density with less BPE concentration can be achieved.

Since the best wastewater clarification was at pH = 5.5 for chitosan, a turbidity-dosage profile was performed near the same pH to determine the optimal quantity of chitosan needed to flocculate *nejayote* colloids.

**Figure 5.** Turbidity and *ζ* of the supernatant in the coagulation-flocculation of nejayote at pH= 5 with chitosan: a) Nejayote and b) Nejayote visual appearance treated with chitosan.

The coagulation-flocculation window of *nejayote* using chitosan at pH = 5 was constructed based on the methodology reported by López Maldonado et al. [23].

**Figure 5** shows the behavior of the zeta potential and turbidity with respect to the concentration of chitosan. In the region of low doses (250 and 750 mg/L), a decrease in turbidity (1590– 500 FAU) is achieved, and the variation of zeta potential ζ = −10 mV to more positive values (ζ = −5.4 mV) shows that the mechanism of destabilization of *nejayote* colloids occurs by charge neutralization [24]. At a dose of 1250 mg/L chitosan, point of zero charge was reached and the better quality of treated water (turbidity = 22 FAU, color = 315 Pt-Co, TSS = 12 mg/L) was obtained. At higher concentration (>1250 mg/L) the best dose, the restabilization processes occur due to excess chitosan adsorbed on the colloids of *nejayote*. In this region of overdose, the addition of chitosan had an adverse effect on the quality of wastewater, increasing turbidity and stability of the dispersed particles (*ζ* = 15 mV and turbidity = 450 FAU).

Moreover, the *nejayote* (*ζ* = 0) has a negative surface charge throughout the pH range. From the electrical viewpoint, at pH = 5, the interaction between oppositely charged species chitosan-

Dosing strategy for chitosan was determined by *ζ* of *nejayote* colloids and chitosan, and also by observing critical pH value of the IEP. In this study, charged chitosan purpose is to reduce the repulsion forces between particles by neutralizing the negatively charged molecules. In general, the electroneutrality zone for chitosan-*nejayote* system is below pH = 6 (see **Figure 4**), this has a practical application since higher charge density with less BPE concentration can be

Since the best wastewater clarification was at pH = 5.5 for chitosan, a turbidity-dosage profile was performed near the same pH to determine the optimal quantity of chitosan needed to

**Figure 5.** Turbidity and *ζ* of the supernatant in the coagulation-flocculation of nejayote at pH= 5 with chitosan: a) Ne-

The coagulation-flocculation window of *nejayote* using chitosan at pH = 5 was constructed

jayote and b) Nejayote visual appearance treated with chitosan.

based on the methodology reported by López Maldonado et al. [23].

*nejayote* is ensured and therefore the strategic dosage cationic BPE was performed.

**3.3.** *Nejayote* **treatability tests by coagulation-flocculation using chitosan**

246 Physico-Chemical Wastewater Treatment and Resource Recovery

achieved.

flocculate *nejayote* colloids.

The coagulation-flocculation window was obtained from 1000 to 1500 mg/L chitosan with optimal dosage of 1250 mg/L chitosan, obtaining with this removal turbidity and suspended solids of about 80% (see **Figure 6**). At this dose, the surface charges of *nejayote* colloids were neutralized by chitosan molecules, resulting in a ζ value very close to zero.

**Figure 6.** Diagram of the engineering for the sanitation process of *nejayote* using chitosan.

#### **3.4. Evaluation of the polyelectrolyte capacity of maize gum for decontaminating wastewater**

The behavior of zeta potential vs pH of anionic BPE obtained from *nejayote* is shown in **Figure 7**, which has a high negative charge density (−35 mV) in the pH range 6–12, having the isoelectric point close to pH = 2.

This negative surface charge is very interesting for the treatment of wastewater containing high concentration of heavy metal. In the FTIR spectrum (see **Figure 8**) shows that the BPE has the characteristic functional groups of a polysaccharide (3400 cm−1 corresponding to stretching of the OH groups and 2900 cm−1 corresponding to the CH2 groups) which give the negative surface charge and that can interact with oppositely charged species, such as heavy metal ions [25].

**Figure 7.** Zeta potential vs pH profiles of anionic BPE.

**Figure 8.** FT-IR spectrum of anionic BPE.

**Figure 9** shows the morphology of anionic BPE and analysis of chemical composition, indicating that its content is primarily carbon, oxygen and calcium, because lime is used in the nixtamalization.

**Figure 9.** SEM micrograph and EDS spectrum of the anionic BPE: Inset Table shows the composition analysis.

## **4. Conclusion**

the OH groups and 2900 cm−1 corresponding to the CH2 groups) which give the negative surface charge and that can interact with oppositely charged species, such as heavy metal ions [25].

**Figure 7.** Zeta potential vs pH profiles of anionic BPE.

248 Physico-Chemical Wastewater Treatment and Resource Recovery

**Figure 8.** FT-IR spectrum of anionic BPE.

Zeta potential measurements are a proactive electrochemical tool to define the strategy of chitosan dosage that allows sanitation and water reuse industry nixtamalization. The use of chitosan allows the use and reuse of byproducts recovered from *nejayote* and it serves as a source of protein for animal feed. The treated water can be discharged into the municipal sewer system using an optimal dose of 1250 mg/L chitosan at pH = 5, achieving removal of up to 80% in the removal of total suspended solids and turbidity. This work evidenced the potential use of *nejayote* as a raw material for obtaining anionic biopolyelectrolyte in the treatment of wastewater with heavy metals of the electroplating industry.

## **Acknowledgements**

The authors gratefully acknowledge support from Consejo Nacional de Ciencia y Tecnologia, México (CONACyT) Project Ciencia Básica 2015 No. 237032, Project Problemas Nacionales No. 247236.

## **Author details**

Eduardo Alberto López-Maldonado\* , Mercedes Teresita Oropeza-Guzmán and Karla Alejandra Suárez-Meraz

\*Address all correspondence to: elopez@cideteq.mx

Faculty of Chemical Sciences and Engineering, Autonomous University of Baja California, Tijuana, Baja California, Mexico


gelling capability and cross-linking content. Food Chemistry 2009; 115: 1286–1290. DOI: 10.1016/j.foodchem.2009.01.046.

**Author details**

**References**

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250 Physico-Chemical Wastewater Treatment and Resource Recovery

\*Address all correspondence to: elopez@cideteq.mx

132. DOI: 10.1016/j.jfoodeng.2009.12.018.

Karla Alejandra Suárez-Meraz

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[2] Rosentrater KA. Review of corn masa processing residues: generation, properties, and potential utilization. Waste Management 2006; 26: 284–292. DOI: 10.1016/j.wasman.

[3] Valderrama C, Rojas A, Gutiérrez E, Rojas I, Oaxaca A, Rosa E, Rodríguez ME. Mechanism of calcium uptake in corn kernels during the traditional nixtamalization process: diffusion, accumulation and percolation. Journal of Food Engineering 2010; 98: 126–

[4] Fernandez JL, Acosta AA, Gruintal MA, Zelaya O. Kinetics of water diffusion in corn grain during the alkaline cooking at different temperatures and calcium hydroxide

[5] Gutiérrez E, Rojas I, Rojas A, Arjona JL, Cornejo MA, Zepeda Y, Velázquez R, Ibarra C, Rodríguez ME. Microstructural changes in the maize kernel pericarp during cooking stage in nixtamalization process. Journal of Cereal Science 2010; 51: 81–88. DOI: 10.1016/

[6] Fernández JL, Martín ES, Díaz JAI, Calderon A, Alvarado A, Ortiz H, Leal M. Steeping time and cooking temperature dependence of calcium ion diffusion during microwave nixtamalization of corn. Journal of Food Engineering 2006; 76: 568–572. DOI: 10.1016/

[7] Osorio P, Agama E, Bello LA, Islas JJ, Gomez NO, Paredes O. Effect of endosperm type on texture and in vitro starch digestibility of maize tortillas, LWT. Food Science and

[8] Niño-Medina G, Carvajal-Millán E, Lizardi J, Rascon-Chu A, Marquez-Escalante J, Gardea A, Martínez-López A, Guerrero V. Maize processing wastewater arabinoxylans:

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**Provisional chapter**

## **Social Perspectives on the Effective Management of Wastewater Wastewater**

**Social Perspectives on the Effective Management of** 

Dalia Saad, Deirdre Byrne and Pay Drechsel

[20] López-Maldonado EA, Oropeza-Guzman MT, Jurado-Baizaval JL, Ochoa-Terán A. Coagulation–flocculation mechanisms in wastewater treatment plants through zeta potential measurements. Journal of Hazardous Materials 2014; 279: 1–10. DOI: 10.1016/

[21] López-Maldonado EA, Ochoa-Terán A, Oropeza-Guzmán MT. A multiparameter colloidal titrations for the determination of cationic polyelectrolytes. Journal of

[22] Suarez Meraz KA, Ponce Vargas SM, Lopez Maldonado JT, Cornejo Bravo JM, Oropeza-Guzmán MT, López-Maldonado EA. Eco-friendly innovation for nejayote coagulation– flocculation process using chitosan: evaluation through zeta potential measurements. Chemical Engineering Journal 2015; 284: 536–542. DOI: 10.1016/j.cej.2015.09.026. [23] López Maldonado EA, Oropeza Guzmán MT, Ochoa Terán A. Improving the efficiency of a coagulation-flocculation wastewater treatment of the semiconductor industry through zeta potential measurements. In: Monsalvo, VM, editor. Ecological Technologies for Industrial. Wastewater Management. Oakville, Canada: Apple Academic Press

[24] Bolto B, Gregory J. Organic polyelectrolytes in water treatment. Water Research 2007;

[25] Egüés I, Stepan AM, Eceiza A, Toriz G, Gatenholm P, Labidi J. Corncob arabinoxylan for new materials. Carbohydrate Polymers 2014; 102: 12–20. DOI: 10.1016/j.carbpol.

Enviromental Protection. 2012; 3: 1559–1570. DOI: 10.4236/jep.2012.311172.

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Inc.; 2015. p. 167–195.

2013.11.011.

41: 2301–2324. DOI: 10.1016/j.watres.2007.03.012.

Dalia Saad, Deirdre Byrne and Pay Drechsel Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/67312

#### **Abstract**

The chapter discusses how adopting a holistic methodology that acknowledges socio‐ logical factors, including community participation, public involvement, social perception, attitudes, gender roles and public acceptance, would lead to improvements in wastewater management practice. It highlights the social dimension as a tool, a lens through which wastewater management and reuse can take on new dimensions. In this way, this chapter aims to shift the focus from perceiving wastewater as a nuisance that needs disposal, toward a resource not to be wasted, which can contribute to food security, human and environmental health, access to energy as well as water security.

**Keywords:** wastewater reuse, social dimension, community participation, public involvement, public acceptance, gender

## **1. Introduction**

The global water crisis, the shortage of fresh water, contamination of water and increasing volumes of wastewater being produced have eventually necessitated the use of wastewater. A paradigm shift is therefore required not only to prevent further damage to the ecosystems, but also to emphasize that wastewater is a resource whose effective management is essential for future water security [1, 2].

Wastewater can be recycled and reused for a variety of water demanding activities such as agriculture, firefighting, flushing of toilets, industrial cooling, park watering, formation of wetlands for wildlife habitats, etc. [3]. Treated wastewater reuse can be seen as a sustainable way of addressing long‐term imbalances between water dem and and supply, which makes sound economic sense also in view of increased imbalances due to climate change [4]. The

focus of most wastewater research has been on the technical aspects and improvements in terms of water quality and on minimizing the environmental and health impacts, without paying sufficient attention to their basic social and sustainability dimensions. Recent research has shown that ignoring broader social issues that impact the adoption of sustainable solu‐ tions prolongs global environmental problems as well as unjust public health and social con‐ ditions [5]. Thus, more attention is needed to the social aspects of wastewater management strategies.

## **2. The Global Demand for Water**

The development of human societies is heavily dependent on the availability of water of suitable quality in adequate quantities. However, the demand for water is ever increasing due to population growth, technological advancement, industrial expansion, pollution and urbanization, which put great stress on the natural water cycle [1]. These demands were met by constructing ever‐larger dams, which in turn affect both water quality and quantity [6]. Moreover, the available freshwater supplies are not evenly distributed, and there is a growing competition for water from different sectors, including industry, agriculture, power genera‐ tion, domestic use, etc. As a result, one‐third of the world's population is currently experienc‐ ing water scarcity. In water‐scarce regions and countries, inequity in access to water resources is increasing because of competition for limited resources, and this particularly affects poor people [7]. However, the focus on freshwater without enough attention to its end products (wastewater) will exacerbate the water quality problem. It is therefore very important to consider wastewater management as a critical component in achieving future water security through integrated water resources management [8]. This is particularly true as wastewater is the only source of additional water that actually increases in quantity as population and water consumption grow.

## **3. Wastewater crisis**

With increasing urbanization and changing lifestyles, increasing amounts of wastewater is being generated and where these are not sufficiently treated, freshwater bodies are continuously threatened [9]. Achieving the Sustainable Development Goal 6, which targets improved waste‐ water management, thus puts immense economic pressure especially on poor countries [10]. Inadequate infrastructure and sustainable management systems for the increasing volume of produced wastewater are at the heart of wastewater crisis in developing countries. As a result of inadequate infrastructure of wastewater treatment in most of the big cities where half of the world's population lives, the majority of wastewaters are discharged into the environment with‐ out any form of treatment, harming both the ecosystem and humans [9, 11]. Over half of the world's hospital beds are occupied by people suffering from diseases caused by contaminated water, and more people die as a result of polluted water than are killed by all forms of war. In many developing countries, an estimated 1.8 million children under 5 years old die every year due to water‐related disease [12].

Wastewater damages the ecosystem in many ways: For example, wastewater may contain high levels of nutrients such as nitrates and phosphates. When water bodies receive excess amounts of these nutrients, it may stimulate excessive plant growth, which may release toxins into the water bodies, leading to oxygen depletion and causing what is known as de‐oxygen‐ ated dead zones. This phenomenon decreases biodiversity and changes species composition and dominance, as well as decreasing water quality for reuse [11, 13]. Another example is the impact of wastewater on the climate around the globe: Wastewater treatment‐related emis‐ sions of methane and nitrous oxide (powerful global warming gases) could rise by 50% and 25%, respectively, between 1990 and 2020 [1].

According to the fourth World Water Development Report by UNESCO [14], only 20% of glob‐ ally produced wastewater receives proper treatment. Treatment capacity typically depends on the income level of the country; thus, in high‐income countries, the treatment capacity reaches up to 70% of the generated wastewater compared to 8% in low‐income countries [15]. Meeting the wastewater treatment challenge is thus not a luxury but a prudent, practical and transformative act, able to maintain public health and secure ecosystem health.

While so far, wastewater has mostly been seen as a treatment challenge, a paradigm shift toward its recognition as a resource for sustainable development is emerging. In this sense, wastewater can be reframed from being a problem to be disposed of to being a resource with social and economic value [5]. This shift offers wastewater to become part of an integrated, full life cycle, ecosystem‐based management system that operates across the three main dimen‐ sions of sustainable development, that is, its social, economic and environmental pillars [1, 16].

#### **4. Wastewater as a resource**

focus of most wastewater research has been on the technical aspects and improvements in terms of water quality and on minimizing the environmental and health impacts, without paying sufficient attention to their basic social and sustainability dimensions. Recent research has shown that ignoring broader social issues that impact the adoption of sustainable solu‐ tions prolongs global environmental problems as well as unjust public health and social con‐ ditions [5]. Thus, more attention is needed to the social aspects of wastewater management

The development of human societies is heavily dependent on the availability of water of suitable quality in adequate quantities. However, the demand for water is ever increasing due to population growth, technological advancement, industrial expansion, pollution and urbanization, which put great stress on the natural water cycle [1]. These demands were met by constructing ever‐larger dams, which in turn affect both water quality and quantity [6]. Moreover, the available freshwater supplies are not evenly distributed, and there is a growing competition for water from different sectors, including industry, agriculture, power genera‐ tion, domestic use, etc. As a result, one‐third of the world's population is currently experienc‐ ing water scarcity. In water‐scarce regions and countries, inequity in access to water resources is increasing because of competition for limited resources, and this particularly affects poor people [7]. However, the focus on freshwater without enough attention to its end products (wastewater) will exacerbate the water quality problem. It is therefore very important to consider wastewater management as a critical component in achieving future water security through integrated water resources management [8]. This is particularly true as wastewater is the only source of additional water that actually increases in quantity as population and

With increasing urbanization and changing lifestyles, increasing amounts of wastewater is being generated and where these are not sufficiently treated, freshwater bodies are continuously threatened [9]. Achieving the Sustainable Development Goal 6, which targets improved waste‐ water management, thus puts immense economic pressure especially on poor countries [10]. Inadequate infrastructure and sustainable management systems for the increasing volume of produced wastewater are at the heart of wastewater crisis in developing countries. As a result of inadequate infrastructure of wastewater treatment in most of the big cities where half of the world's population lives, the majority of wastewaters are discharged into the environment with‐ out any form of treatment, harming both the ecosystem and humans [9, 11]. Over half of the world's hospital beds are occupied by people suffering from diseases caused by contaminated water, and more people die as a result of polluted water than are killed by all forms of war. In

strategies.

**2. The Global Demand for Water**

254 Physico-Chemical Wastewater Treatment and Resource Recovery

water consumption grow.

**3. Wastewater crisis**

Depending on the treatment or lack of it, as well as the degree of dilution, wastewater can be rich in resources such as nutrients, inorganic and organic compounds as well as energy, mak‐ ing it worthwhile for recovery and reuse. On the other hand, it can also be rich in chemical and microbial contaminants, and the improper use of untreated wastewater can have adverse effect on both human health and environment [13]. Wise wastewater management can there‐ fore be a positive addition to the environment with significant returns in terms of enhancing food security, creating livelihood opportunities, climate change adaptation and sustainable ecosystem [1].

Successful examples of this paradigm shift can be found around the globe. There have been dramatic successes in using treated water for drinking purposes; for example, in Namibia, 35% of all drinking water is treated wastewater, and in Singapore, 30% of all water used is reclaimed water (and this percentage is increasing) [17]. The United States of America has also been seen several successes in treating wastewater for drinking purposes.

Wastewater can also be treated to provide energy. Various forms of energy can be recovered from wastewater and its biosolids, with biogas being the most prominent. It can be combusted on‐site for heat or electricity generation, cleaned and sold to local natural gas providers or as fuel for vehicles [18–20]. Wastewater treatment plants are increasingly generating their own energy, which is an important achievement because energy consumption is a major cost in treatment plants. Another example from low‐income countries is the transformation of fecal sludge (and other organic waste) into dry fuel like briquettes [21, 22]. The most common materials, however, that are recovered from wastewater are the water itself, which can be used for irrigation and its crop nutrients and biosolids as fertilizer. The use of fecal sludge as fertilizer is a well‐known practice, especially from septic treatment plants given the low con‐ tamination within household‐based on‐site sanitation systems, compared to biosolids recov‐ ered from wastewater treatment plants. Some treatment processes recover nutrients, such as the N‐ and P‐rich struvite, from wastewater during treatment rather than from the final products of the treatment [23–25].

## **5. The overlooked social dimension**

The focus of most wastewater‐related research has been on the technical aspects of the problem and improvements in terms of water quality and in minimizing environmental and health impacts, with very limited attention to its basic social and cultural sustainability dimensions [5, 9]. While, with increasing urbanization, wastewater treatment has moved fur‐ ther away from the household and its social roots, three types of campaigns (i) against open defecation, (ii) for the promotion of water‐saving dry toilets and (iii) for using reclaimed water for drinking made it clear that sanitation depends strongly on social habits and accep‐ tance. Where treatment is not keeping pace with population growth, and environmental pollution is threatening public health, the social dimension of wastewater management becomes obvious. Recognizing the role of the social base for wastewater management from risk reduction to reuse can have major implications, for example, on the choice and effective‐ ness of the technologies employed. Yet, usually, only limited information is available on the social perspective [5].

Wastewater management strategies have been traditionally driven by considerations of efficiency, safety, and cost‐effectiveness. Even technology choices are often made by finance institutions outside the country, especially in low‐income countries, often favoring "Northern" technology options. The emphasis on costs and benefits in this context would be acceptable if, in addition, other relevant factors could be included in the decision‐making process by adopting a holistic methodology that includes the voices of all stakeholders and an analysis of sociological factors. Unless a holistic methodology is adopted, even cutting‐ edge technology might impede progress toward sustainable development, as the example of Toowoomba shows (Box 1). Likewise, the Singaporean success story would have had a very different outcome if public buy‐in for wastewater reuse for drinking purposes had not been secured [5].

Queensland's Toowoomba in Australia is an often cited case illustrating the strength of public opinion regarding wastewater use. A plan to turn wastewater into drinking water failed in Toowoomba at a referendum in 2006, although water scarcity in the community was severe, to the point that water use for gardening was completely prohibited in the "Garden City." With no major river nearby, the community water supply had to be pumped uphill. During several years of drought, the 140,000 residents of Toowoomba and surrounding areas endured tough water restrictions. Local officials considered that the city had no choice but to treat and use parts of its wastewater for drinking water, and given the water crisis, they expected the program would be acceptable. However, the proposal met with fierce oppo‐ sition from the community. In 2006, the residents of Toowoomba voted strongly against treating and using 25% of the city's wastewater. They relied instead on water piped from Brisbane's Wivenhoe Dam, at a cost to ratepayers of nearly \$100 million more than the reuse program would have cost.

The Toowoomba proposal was an indirect wastewater use program, in which highly treated wastewater would be passed through an environmental buffer before being treated again, as part of the drinking water system. The public poll was accompanied by two dynamic cam‐ paigns building on the "yuck" and "fear" factors on one side, and social and financial argu‐ ments on the other. In the end, 62% of those polled opposed the project.

Sources: Ref. [5].

Wastewater can also be treated to provide energy. Various forms of energy can be recovered from wastewater and its biosolids, with biogas being the most prominent. It can be combusted on‐site for heat or electricity generation, cleaned and sold to local natural gas providers or as fuel for vehicles [18–20]. Wastewater treatment plants are increasingly generating their own energy, which is an important achievement because energy consumption is a major cost in treatment plants. Another example from low‐income countries is the transformation of fecal sludge (and other organic waste) into dry fuel like briquettes [21, 22]. The most common materials, however, that are recovered from wastewater are the water itself, which can be used for irrigation and its crop nutrients and biosolids as fertilizer. The use of fecal sludge as fertilizer is a well‐known practice, especially from septic treatment plants given the low con‐ tamination within household‐based on‐site sanitation systems, compared to biosolids recov‐ ered from wastewater treatment plants. Some treatment processes recover nutrients, such as the N‐ and P‐rich struvite, from wastewater during treatment rather than from the final

The focus of most wastewater‐related research has been on the technical aspects of the problem and improvements in terms of water quality and in minimizing environmental and health impacts, with very limited attention to its basic social and cultural sustainability dimensions [5, 9]. While, with increasing urbanization, wastewater treatment has moved fur‐ ther away from the household and its social roots, three types of campaigns (i) against open defecation, (ii) for the promotion of water‐saving dry toilets and (iii) for using reclaimed water for drinking made it clear that sanitation depends strongly on social habits and accep‐ tance. Where treatment is not keeping pace with population growth, and environmental pollution is threatening public health, the social dimension of wastewater management becomes obvious. Recognizing the role of the social base for wastewater management from risk reduction to reuse can have major implications, for example, on the choice and effective‐ ness of the technologies employed. Yet, usually, only limited information is available on the

Wastewater management strategies have been traditionally driven by considerations of efficiency, safety, and cost‐effectiveness. Even technology choices are often made by finance institutions outside the country, especially in low‐income countries, often favoring "Northern" technology options. The emphasis on costs and benefits in this context would be acceptable if, in addition, other relevant factors could be included in the decision‐making process by adopting a holistic methodology that includes the voices of all stakeholders and an analysis of sociological factors. Unless a holistic methodology is adopted, even cutting‐ edge technology might impede progress toward sustainable development, as the example of Toowoomba shows (Box 1). Likewise, the Singaporean success story would have had a very different outcome if public buy‐in for wastewater reuse for drinking purposes had not

products of the treatment [23–25].

social perspective [5].

been secured [5].

**5. The overlooked social dimension**

256 Physico-Chemical Wastewater Treatment and Resource Recovery

**Box 1:** Community resistance to wastewater reuse

A primary shortcoming in wastewater reuse is the lack of a combined sociotechnological planning and design methodology to identify and deploy the most sustainable solution in a given geographic and cultural context. The best practice, once a treatment or reuse technology has been developed, is to get early stakeholder buy‐in and identify the best way to implement the technology in a participatory manner that is socially acceptable from the local perspec‐ tive [26–28]. Stakeholders can be included in the decision‐making process in different ways, including facilitating positive social learning processes, minimizing and resolving conflicts and, most importantly, using local knowledge and community participation [4].

#### **5.1. Community participation and public involvement**

Successful employment of appropriate technologies requires deep understanding of the social dynamics of the community in which they are applied [29–31]. This is only achieved through effective public involvement and community participation. Public involvement is best achieved through participation and involvement of users in all parts of the project cycle, from planning and design to implementation and decision‐making, which produces more efficient and sustainable projects/outcomes [32]. In a sense, when communities have influence and control over decisions that affect them, they have a greater stake in the outcomes and are more committed to ensuring success.

Public involvement is of particular relevance when it comes to wastewater reuse, which is associated with major social concerns, including impacts on public health and safety, impacts on environmental quality as well as the benefits and risks of reuse. Thus, having an effective public involvement strategy from the planning phase to full implementation leads to greater acceptance and facilitates the implementation process of the wastewater reuse scheme. In other words, community participation can assure the social viability of the wastewater reuse practices [33–35]. Effective public involvement begins with early contact with potential users through the actual inclusion of all stakeholders and can involve educational and public awareness programs, the formation of advisory committees, and holding public workshops to discuss the benefits and risks of reuse [5, 36]. According to Ashley et al. [30], publicity, including advertisement in the media, education and inclusion of all stakeholders (politicians, experts and general public) in the decision‐making process are the key elements for successful design and implementation of wastewater schemes. Gibson and Apostolidis [37] argue that the best way to involve the general public and to gain its support and acceptance is through successful demonstration projects.

For community participation to be as inclusive and effective as possible, the diversity of people within the same community should be acknowledged and dealt with. Communities are made up of individuals of different genders and groups of people who command different levels of power and wealth. Within each community, there are always competing interest groups. For example, there are rich and the poor, the farmers who have fields and livestock to water and the landless farmhands with children to care for, marginalized groups and members of socioeco‐ nomic minorities, housewives who need water for drinking and household and businessmen who own industries that require water. Thus, perception studies are a key component of any social analysis [26, 38].

#### **5.2. Social perception and public acceptance**

Even when wastewater is treated using advanced technologies and health risks are carefully addressed and controlled, irrespective of all scientific evidence, social perception remains the driver of the success or failure of wastewater reuse schemes. Depending on public percep‐ tions, impressions and attitudes, the development of a wastewater scheme can be supported or constrained. Negative public perception can prevent well‐planned projects from moving forward. On the other hand, positive public perception, which leads to greater acceptance, is the key element for successful implementation of wastewater recycling [5, 39]. Experience shows that the local communities have rejected a number of wastewater recycling projects by the governments and water boards around the world as a result of inadequate community consultation which led to negative public perception [40].

The degree of acceptance of wastewater reuse varies widely depending on the reuse purposes and is influenced by many factors, such as the degree of contact; expressions of disgust; educa‐ tion; risk awareness; the degree of water scarcity or availability of alternative water sources; calculated costs and benefits; trust and knowledge; issues of choice; attitudes toward the envi‐ ronment; economic considerations; involvement in decision‐making; the source of water to be recycled; and experience with treated wastewater. Other factors that depend on the region and case include cultural, religious, educational and/or socioeconomic factors [5, 27, 35, 41–43].

Education and the level of physical contact (potable/no potable reuse) are the most influential factors that have been frequently associated with levels of acceptance of treated wastewater. In Kuwait and Greece, for example, the willingness to accept recycled water increased with edu‐ cational levels [44, 45]. However, as much as education and knowledge support public accep‐ tance, nevertheless, direct exposure to the recycled water strongly influences its acceptance [46, 47]. For example, potable use is usually rejected due to health concerns. Wastewater use in agriculture generally is preferred to potable use, but more distant uses, such as landscape irrigation, are the most preferred [48, 49].

Several authors have investigated the association of sociodemographic descriptors with the acceptance of treated wastewater. The D'Angelo report [50] indicated that the acceptability of using recycled water in agriculture is higher for nonedible crops than for edible crops. For edible crops, the preference is for crops that must be peeled prior to human consumption, such as oranges and sweet corn. A relevant study [51] reported that the public's acceptance of reuse increases as the degree of human contact with the recycled water increases, with 97% and 96% of the public supporting wastewater reuse for irrigation and for toilet flush‐ ing, respectively, whereas only 20%–30% support potable reuse. Another study conducted by Friedler and Lahav [39] to determine the attitudes of the Israeli urban public toward possible urban reuse revealed that the majority of participants supported options perceived as low contact, such as irrigation of public parks (96%), sidewalk landscaping (95%) and use in the construction industry (94%), while higher contact reuse options, such as commercial launder‐ ettes (60%), found less support. According to Bruvold [52], the degree of human contact has a greater effect when people were asked about general use options, whereas when the specific use scheme was used, other factors such as health, environment, treatment, distribution and conservation had greater impact on people's perceptions. Therefore, he argues that it is essen‐ tial to weigh the different objectives of the recycling options in coordination with people's/ users' acceptability and preference and select the recycling projects which are most likely to be accepted by the community and therefore make the project implementation successful.

#### **5.3. Gender roles and implications**

on environmental quality as well as the benefits and risks of reuse. Thus, having an effective public involvement strategy from the planning phase to full implementation leads to greater acceptance and facilitates the implementation process of the wastewater reuse scheme. In other words, community participation can assure the social viability of the wastewater reuse practices [33–35]. Effective public involvement begins with early contact with potential users through the actual inclusion of all stakeholders and can involve educational and public awareness programs, the formation of advisory committees, and holding public workshops to discuss the benefits and risks of reuse [5, 36]. According to Ashley et al. [30], publicity, including advertisement in the media, education and inclusion of all stakeholders (politicians, experts and general public) in the decision‐making process are the key elements for successful design and implementation of wastewater schemes. Gibson and Apostolidis [37] argue that the best way to involve the general public and to gain its support and acceptance is through

For community participation to be as inclusive and effective as possible, the diversity of people within the same community should be acknowledged and dealt with. Communities are made up of individuals of different genders and groups of people who command different levels of power and wealth. Within each community, there are always competing interest groups. For example, there are rich and the poor, the farmers who have fields and livestock to water and the landless farmhands with children to care for, marginalized groups and members of socioeco‐ nomic minorities, housewives who need water for drinking and household and businessmen who own industries that require water. Thus, perception studies are a key component of any

Even when wastewater is treated using advanced technologies and health risks are carefully addressed and controlled, irrespective of all scientific evidence, social perception remains the driver of the success or failure of wastewater reuse schemes. Depending on public percep‐ tions, impressions and attitudes, the development of a wastewater scheme can be supported or constrained. Negative public perception can prevent well‐planned projects from moving forward. On the other hand, positive public perception, which leads to greater acceptance, is the key element for successful implementation of wastewater recycling [5, 39]. Experience shows that the local communities have rejected a number of wastewater recycling projects by the governments and water boards around the world as a result of inadequate community

The degree of acceptance of wastewater reuse varies widely depending on the reuse purposes and is influenced by many factors, such as the degree of contact; expressions of disgust; educa‐ tion; risk awareness; the degree of water scarcity or availability of alternative water sources; calculated costs and benefits; trust and knowledge; issues of choice; attitudes toward the envi‐ ronment; economic considerations; involvement in decision‐making; the source of water to be recycled; and experience with treated wastewater. Other factors that depend on the region and case include cultural, religious, educational and/or socioeconomic factors [5, 27, 35, 41–43].

Education and the level of physical contact (potable/no potable reuse) are the most influential factors that have been frequently associated with levels of acceptance of treated wastewater. In

successful demonstration projects.

258 Physico-Chemical Wastewater Treatment and Resource Recovery

**5.2. Social perception and public acceptance**

consultation which led to negative public perception [40].

social analysis [26, 38].

As mentioned above, successful community participation is better achieved by acknowledging the diversity of people within the community. This includes gender, age, education level, power, wealth and so on [53]. In this context, it is very important to acknowledge the differences of inter‐ ests and roles between men and women as different stakeholders. There are a number of gender aspects which influence how both genders are involved in and benefit from improvements to the water. In many developing countries, women have limited access to education and other resources and services, have heavier workloads, are more constrained by poor health, have a lower social status, and are poorly represented in decision‐making at both household and community levels [54]. Thus, balanced attention is needed in the form of distinctions between what women and men know, do and decide and what the effects are for them, their families and communities (Box 2).

In general, women are most vulnerable to water‐related disasters, including water scarcity and bad water quality. Many infectious diseases are associated with poor water quality, and these are reported as being among the fifth biggest killer of women worldwide, causing more deaths than AIDS, diabetes or breast cancer [55]. Dirty water and poor sanitation are also at the root of problems such as maternal and child mortality and sexual violence. Many women in developing countries give birth at home without access to clean water, exposing themselves and their babies to infections. More than 50 million primary‐school‐aged girls in develop‐ ing countries are not in school because they are required to fetch water and firewood [56]. Thus, it becomes a necessity to bring women frequently on the scene for consultation and allow their full participation in wastewater management. Implementing a gender‐sensitive approach produces more effective, efficient and affordable outcomes. Including women in water and wastewater management planning often makes for fewer oversights in technical planning and improves resource and financial management, as well as allowing for greater transparency [54].

Acknowledging gender roles and differences not only contributes to the success of a project, but offers planning options to optimize the overall social and economic development and reduces competition and conflicts over water resources. In most societies, the provision of water for the fulfillment of fundamental human needs has always been women's responsibility, yet their par‐ ticipation in decision‐making is very limited if there at all. To bring about constructive change, more efforts are needed to better understand the gender implications in water sector [57, 58].

Thoughtful safety interventions must be gender sensitive. In many cultures, women carry the main responsibility for hygiene and health, also vis‐à‐vis greywater or wastewater use as reported, for example, from Jordan, Vietnam and Tunisia. The strong connection between women and water use at household level offers significant potential for innovative training approaches to improve the social acceptance of safe water reuse, as recently demonstrated in Jordan. Also the use of protective clothing should be gender‐specific. In Vietnam, women were observed wearing protective gloves and boots more consistently than men. The differ‐ ences were attributed to the gendered work division on the farm, with men walking around the farms much more than women, and where protective clothing constrained men's move‐ ments. Sources [5, 59–61].

**Box 2:** Gender roles

## **6. Key improvement areas**

There is little known about public perceptions of wastewater reuse in the literature, and it is mostly documented in a limited number of locations, that is, the United States, Australia and Western Europe. Still a lot of more studies are required at national and subnational context in order to avoid outcomes being transferred from one country to another, which is always inap‐ propriate due to the range of factors that influence public acceptance from country to another, including culture, religion, economy, climate and water availability [40].

In general, public acceptance of reuse is not straightforward, but it is always easier when water scarcity is already affecting the public, so that they perceive wastewater reuse as a solution rather than a problem [5, 62]. However, for greater acceptance, public and private concerns and benefits must be aligned. Public concerns about risks are to be weighed against the benefits of using treated water. The dialogue should be built on mutual trust to provide the right climate for negotiation and conflict resolution [5].

Certain social factors have always been associated with poor acceptance of wastewater reuse, including the lack of coordination between the authorities involved in planning; inadequate community consultation; lack of trust in technology; social pressure and fear of social back‐ lash; and fear of losing markets in case of wastewater reuse in irrigation [41, 63, 64]. Another factor is overlooking the gender dimension [65].

and their babies to infections. More than 50 million primary‐school‐aged girls in develop‐ ing countries are not in school because they are required to fetch water and firewood [56]. Thus, it becomes a necessity to bring women frequently on the scene for consultation and allow their full participation in wastewater management. Implementing a gender‐sensitive approach produces more effective, efficient and affordable outcomes. Including women in water and wastewater management planning often makes for fewer oversights in technical planning and improves resource and financial management, as well as allowing for greater

Acknowledging gender roles and differences not only contributes to the success of a project, but offers planning options to optimize the overall social and economic development and reduces competition and conflicts over water resources. In most societies, the provision of water for the fulfillment of fundamental human needs has always been women's responsibility, yet their par‐ ticipation in decision‐making is very limited if there at all. To bring about constructive change, more efforts are needed to better understand the gender implications in water sector [57, 58].

Thoughtful safety interventions must be gender sensitive. In many cultures, women carry the main responsibility for hygiene and health, also vis‐à‐vis greywater or wastewater use as reported, for example, from Jordan, Vietnam and Tunisia. The strong connection between women and water use at household level offers significant potential for innovative training approaches to improve the social acceptance of safe water reuse, as recently demonstrated in Jordan. Also the use of protective clothing should be gender‐specific. In Vietnam, women were observed wearing protective gloves and boots more consistently than men. The differ‐ ences were attributed to the gendered work division on the farm, with men walking around the farms much more than women, and where protective clothing constrained men's move‐

There is little known about public perceptions of wastewater reuse in the literature, and it is mostly documented in a limited number of locations, that is, the United States, Australia and Western Europe. Still a lot of more studies are required at national and subnational context in order to avoid outcomes being transferred from one country to another, which is always inap‐ propriate due to the range of factors that influence public acceptance from country to another,

In general, public acceptance of reuse is not straightforward, but it is always easier when water scarcity is already affecting the public, so that they perceive wastewater reuse as a solution rather than a problem [5, 62]. However, for greater acceptance, public and private concerns and benefits must be aligned. Public concerns about risks are to be weighed against the benefits of using treated water. The dialogue should be built on mutual trust to provide

including culture, religion, economy, climate and water availability [40].

the right climate for negotiation and conflict resolution [5].

transparency [54].

260 Physico-Chemical Wastewater Treatment and Resource Recovery

ments. Sources [5, 59–61].

**6. Key improvement areas**

**Box 2:** Gender roles

In order to fill the gap in knowledge regarding the social dimension of wastewater reuse, extensive social research into public perceptions of wastewater reuse is needed. Some of the priority areas to focus on are as follows (1) to understand judgment strategies that shape pub‐ lic decisions to support or reject wastewater reuse; (2) to identify factors influencing people's risk perceptions; (3) to investigate the role of trust in the authorities and the limits in scientific knowledge in people's decision‐making processes to either accept or reject the reuse; (4) to examine how factors such as health, environment, treatment, distribution and conservation can affect people's willingness to use recycled water; (5) to examine people's sensitivity with regard to the disgust emotion or "yuck" factor; (6) to understand the impact of the source of waste‐ water on people's decisions; (7) to understand how the economic advantages in using recycled water can facilitate public acceptance; and (8) to identify possible environmental justice issues that may affect public acceptance [66, 67].

With regard to gender implications on community participation and public acceptance, greater women's participation is needed through effective gender mainstreaming strategies. Obstacles to women's participation generally include lack of confidence, family commitment including child care, heavy workload and time constraints, traditional values and stereotypes, fears of men and husbands who prevent women from participating (many women said that their husbands do not support their participation in public life) [68, 69].

Apart from the social and cultural issues, another reality with regard to gender mainstream‐ ing is the lack of general awareness of the significance of gender factors in water and waste‐ water management, which applies to both leaders and decision‐makers who work in water management programs. Another shortcoming is the lack of gender‐disaggregated data, which is the only way to move forward from principles to practice in gender mainstreaming [70, 71].

In order to fill the gender gap in wastewater reuse, investing time and effort in awareness rais‐ ing on the different needs and impacts for women and men at all levels is part of the necessary training for all professionals in the wastewater sector. Nevertheless, gender mainstreaming is a continuous process and a holistic approach, which cannot be achieved by a single training session.

Some of the key issues to focus on are as follows: (1) to acknowledge both men's and women's roles and responsibilities, energy, experience and knowledge in contributing to the effective‐ ness of wastewater reuse programs as well as identifying their different needs and priori‐ ties; (2) to mainstream gender throughout all projects' cycle from planning and design to implementation as well as related policy; (3) to ensure women's participation in consultation committees and educational workshops in terms of timing and allocation of these meetings by taking into consideration their family responsibilities (e.g., domestic work), otherwise, women may choose not to participate to avoid conflict with their responsibilities; and (4) "at institutional level" to train the technical staff working in research and development to inte‐ grate gender dimensions into the socioeconomic aspects of research work, in order to address the differential impacts of structural interventions and the appropriation of new technologies [53, 54, 72].

## **7. Conclusions**

With increasing pressures on water resources, wastewater recycling and reuse have rapidly become an imperative for integrated water management strategies. However, along with the technology advancement in wastewater treatment, societal factors such as public perception, public acceptance and the dimension of gender have great implications on the success of wastewater reuse.

Adopting a sociotechnological approach by means of considering all social factors together with technology in wastewater recycling results in great improvements in terms of effective‐ ness and efficiency as the infrastructure will be more widely used and optimally sustained by all user groups including women and men. It will also contribute to the overall development of the society by increasing consumption, production, income, environmental security, health and overall family welfare, along with securing water resources when addressing the societal issues of the service delivered. Another gain of the sociotechnological approach in water sec‐ tor is the sustainability of the service, in the sense that equal participation of all stakeholders in research and project implementation can increase the potential, flexibility and creative innova‐ tion in responding to water insecurity.

## **Acknowledgements**

The authors would like to acknowledge the World Academy of Science (TWAS) for making this work possible through TWAS‐UNESCO Scheme, as well as the support of the CGIAR Research Program on Water, Land and Ecosystems.

## **Author details**

Dalia Saad1,2\*, Deirdre Byrne1 and Pay Drechsel<sup>2</sup>


## **References**

women may choose not to participate to avoid conflict with their responsibilities; and (4) "at institutional level" to train the technical staff working in research and development to inte‐ grate gender dimensions into the socioeconomic aspects of research work, in order to address the differential impacts of structural interventions and the appropriation of new technologies

With increasing pressures on water resources, wastewater recycling and reuse have rapidly become an imperative for integrated water management strategies. However, along with the technology advancement in wastewater treatment, societal factors such as public perception, public acceptance and the dimension of gender have great implications on the success of

Adopting a sociotechnological approach by means of considering all social factors together with technology in wastewater recycling results in great improvements in terms of effective‐ ness and efficiency as the infrastructure will be more widely used and optimally sustained by all user groups including women and men. It will also contribute to the overall development of the society by increasing consumption, production, income, environmental security, health and overall family welfare, along with securing water resources when addressing the societal issues of the service delivered. Another gain of the sociotechnological approach in water sec‐ tor is the sustainability of the service, in the sense that equal participation of all stakeholders in research and project implementation can increase the potential, flexibility and creative innova‐

The authors would like to acknowledge the World Academy of Science (TWAS) for making this work possible through TWAS‐UNESCO Scheme, as well as the support of the CGIAR

and Pay Drechsel<sup>2</sup>

2 Institute for Gender Studies, University of South Africa, Pretoria, South Africa

[53, 54, 72].

**7. Conclusions**

262 Physico-Chemical Wastewater Treatment and Resource Recovery

wastewater reuse.

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Research Program on Water, Land and Ecosystems.

\*Address all correspondence to: dalianono@gmail.com

1 International Water Management Institute, Colombo, Sri Lanka

**Acknowledgements**

**Author details**

Dalia Saad1,2\*, Deirdre Byrne1


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Water Project.


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