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## **Meet the editors**

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 COM-SATS Institute of Information Technology, Lahore, Pakistan. She discovered innovative and low-cost processes for the treatment of wastewater. She is the author of sci-

entific manuscripts, books, book chapters and 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

**Preface XI**





## Preface

Chapter 6 **Comparative Assessment of Pharmaceutical Removal from**

Chapter 7 **Pulp Mill Wastewater: Characteristics and Treatment 119**

Chapter 8 **Molecular Biomonitoring of Microbial Communities in Tannery**

Chapter 9 **Application of Mixed Microbial Culture Biofilms for Manganese**

Tonči Rezić, Iva Rezić, Michaela Zeiner and Božidar Šantek

**vulgaris and Scenedesmus obliquus 99**

María Noel Cabrera

**Chemicals 141**

**Wastewater 179**

**Section 3 Biofilm for Wastewater Treatment 157**

**Rotating Tubular Bioreactor 159**

Terelle Ramcharan and Ajay Bissessur

Chapter 11 **Biohydrogen Production from Wastewaters 197**

**Section 4 Bioenergy for Resource Recovery 195**

**Treatment (POME) 229**

Chapter 10 **Electrocoagulative and Biological Treatment of Laundry**

Stomeo

**VI** Contents

**Wastewater by the Microalgae Chlorella sorokiniana, Chlorella**

Carla Escapa Santos, Ricardo Nuno de Coimbra, Sergio Paniagua Bermejo, Ana Isabel García Pérez and Marta Otero Cabero

**Wastewater Treatment Plant for the Removal of Retanning**

Adey Feleke Desta, Joyce Nzioki, Solomon Maina and Francesca

**(II), Cobalt (II), and Chromium (VI) Biosorption by Horizontal**

Periyasamy Sivagurunathan, Gopalakrishnan Kumar, Arivalagan Pugazhendhi, Guangyin Zhen, Takuro Kobayashi and Kaiqin Xu

Eduardo Lucena Cavalcante de Amorim, Leandro Takano Sader,

**Membrane Anaerobic System (UMAS) for Palm Oil Mill Effluent**

Chapter 12 **Valorization of Glucose-Based Wastewater Through Production of Hydrogen, Volatile Fatty Acids and Alcohols 211**

Chapter 13 **Production of Biogas and Performance Evaluation of Ultrasonic**

Abdurahman Hamid Nour and Azhari Hamid Nour

Lucas Rodrigues Ramos and Edson Luiz Silva

Industrialization and rapid increase in global population resulted in the increase of hazard‐ ous chemicals including radioactive materials, pesticides, fertilizers, pharmaceuticals, petro‐ chemicals, dyes, paints, heavy metals, surfactants and detergents in the environment. These organic and inorganic substances find their way via air or soil to water and are the constant source of toxicity to flora and fauna resulting in the extinction of species. One of the major challenges facing mankind today is the provision of clean water and food to a vast majority of the population around the world. Nature has an amazing ability to cope with small amounts of water pollution, but the treatment of billions of gallons of wastewater and sew‐ age produced every day is required to be treated before releasing it back to the environment. Wastewater contains precious chemicals and substances, which can be recovered using dif‐ ferent chemical and biochemical reactions. The main focus of the book is to describe effec‐ tive methods for wastewater management and its treatment, reuse and recycling along with the recovery of valuable substances and energy.

Volume I focuses on the bioremediation of wastewater and is divided into four sections. Sec‐ tion 1 'Wastewater, Management and Monitoring' emphasizes on the micropollutants enter‐ ing into the environment after conventional wastewater treatment facilities of industrial, agricultural and domestic wastewaters. The occurrence of these persistent pollutants poses deleterious effects on human and environmental health. The fate and persistence of these chemicals after conventional treatment processes are discussed. Simple solution for the treatment of wastewater and recovery of water as resource using microbiological method is a viable option. This increases biomass and reduces water, air and soil pollution. Mitigating environmental risks of wastewater reuse for agriculture is also discussed with the support of experimental studies. An interesting book chapter on 'Micro-Based Strategy for Plant Nu‐ trient Management' is included in this section, which highlights the importance of slow leaching of nutrients for plant uptake. It provides modest solution to readers and farmers for the use of mixture of microbial consortia in soil, which not only helps to reduce leaching of nutrients in groundwater but also reduces ground water pollution along with the cost of fertilizers.

Section 2 'Hazards and Treatment of Organic Compounds in Wastewater' describes studies about the hazards and treatments of wastewater containing antibiotics, pharmaceuticals, cy‐ clic aromatic compounds and bleaching agents.

Persistent antibiotics in wastewater develop resistant microorganisms. This poses great fi‐ nancial and research burden to find alternate antibiotics. Their fate and treatment provides insight about the process. Similarly, the presence of pharmaceutical compounds and the emerging contaminants in wastewater cause physiological responses in nontarget organ‐ isms. This section covers the identification of efficient microorganisms, which is the key fac‐ tor for biological treatment of wastewater containing such contaminants. Identification of relationship between *Bacteroidetes* and *Proteobacteria* for treatment of xenobiotic compounds in the wastewater of tanning industry conducted on field-scale reactor is an important part of this section. A comprehensive review of pulping technologies and biological treatment of wastewater is included. Here, we also focus on known and emerging pollutants.

Section 3 'Biofilms for Wastewater Treatment' is designated for the application of microbial culture for bioabsorption of metals and treatment of surfactants. Parameters effecting bio‐ sorption of metals on biofilm are optimized using pilot-scale horizontal tubular bioreactor. The study about electrocoagulation and biological treatment of laundry wastewater investi‐ gates the possibility of recycling and reusing wastewater from laundry run-offs.

Section 4 'Bioenergy as Resource Recovery' draws the attention of readers and researchers towards the recovery of biogas, hydrogen, volatile fatty acids and alcohols during anaerobic treatment of wastewaters from carbohydrates present in wastewaters. This covers the dis‐ cussions about the synthesis of biofuel using anaerobic fluidized bed reactor for hydrogen synthesis from glucose-based wastewater. Another interesting study is the production of bi‐ ogas and its performance evaluation using ultrasonic membrane anaerobic system (UMAS). Integrated technology of UMAS is an attractive solution for treatment of palm oil wastewa‐ ter along with resource generation. This section also describes two important pathways for the production of hydrogen, which are light and dark fermentation reactions. Recent ad‐ vancements on biohydrogen production technologies from wastewater with respect to inoc‐ ulum development, process optimization, scale-up and challenges are discussed in detail. Bioremediation of wastewater is the low-cost solution. However, its efficiency is effected by environmental conditions. Therefore, physico-chemical treatment of wastewater is another efficient option, which will be covered in Volume II.

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 editior 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**

isms. This section covers the identification of efficient microorganisms, which is the key fac‐ tor for biological treatment of wastewater containing such contaminants. Identification of relationship between *Bacteroidetes* and *Proteobacteria* for treatment of xenobiotic compounds in the wastewater of tanning industry conducted on field-scale reactor is an important part of this section. A comprehensive review of pulping technologies and biological treatment of

Section 3 'Biofilms for Wastewater Treatment' is designated for the application of microbial culture for bioabsorption of metals and treatment of surfactants. Parameters effecting bio‐ sorption of metals on biofilm are optimized using pilot-scale horizontal tubular bioreactor. The study about electrocoagulation and biological treatment of laundry wastewater investi‐

Section 4 'Bioenergy as Resource Recovery' draws the attention of readers and researchers towards the recovery of biogas, hydrogen, volatile fatty acids and alcohols during anaerobic treatment of wastewaters from carbohydrates present in wastewaters. This covers the dis‐ cussions about the synthesis of biofuel using anaerobic fluidized bed reactor for hydrogen synthesis from glucose-based wastewater. Another interesting study is the production of bi‐ ogas and its performance evaluation using ultrasonic membrane anaerobic system (UMAS). Integrated technology of UMAS is an attractive solution for treatment of palm oil wastewa‐ ter along with resource generation. This section also describes two important pathways for the production of hydrogen, which are light and dark fermentation reactions. Recent ad‐ vancements on biohydrogen production technologies from wastewater with respect to inoc‐ ulum development, process optimization, scale-up and challenges are discussed in detail. Bioremediation of wastewater is the low-cost solution. However, its efficiency is effected by environmental conditions. Therefore, physico-chemical treatment of wastewater is another

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 editior 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

**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

wastewater is included. Here, we also focus on known and emerging pollutants.

gates the possibility of recycling and reusing wastewater from laundry run-offs.

efficient option, which will be covered in Volume II.

VIII Preface

who always supported and encouraged me throughout my life.

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

Prof. Ahmad was the 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 publica‐ tions *Aluminium Alloys - New Trends in Fabrication and Applications*, *New Trends in Alloy Devel‐ opment, 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 Wastewater Treat‐ ment and Resource Recovery*.

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

InTech Book Department Team

**Wastewater Management and Monitoring**

#### **Treatment of Organic Recalcitrant Contaminants in Wastewater Treatment of Organic Recalcitrant Contaminants in Wastewater**

Asmita Gupta and Indu Shekhar Thakur Asmita Gupta and Indu Shekhar Thakur

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

http://dx.doi.org/10.5772/66346

#### **Abstract**

Research has shown that a myriad of contaminants enter the environment through industrial and domestic sources on a daily basis. The biodegradable compounds often get degraded or mineralized by various physical, chemical or biological processes, whereas the recalcitrant organic contaminants either are transformed or get dispersed and persist in the receiving environments, and to an extent much greater than was earlier estimated. Many chemical compounds that were not previously included as pollutants can now be detected at much higher concentrations globally. The effect of most of these emerging contaminants on human and environment health is still unknown. Therefore, there is an urgent need to study the fate of these persistent compounds so as to better understand and manage their ecological and health effects.

**Keywords:** Wastewater, organic contaminants, recalcitrant, biodegradation, sorption

### **1. Introduction**

Water adversely affected in quality by anthropogenic activities is, typically called wastewater. Wastewater is generally collected and treated by various processes at centralized facilities, referred to as wastewater treatment plants (WWTPs). There can be several sources contributing towards wastewater generation, including domestic, industrial and agricultural. As there are various sources of wastewater generation, so are the compounds present in them. Wastewater, thus, is a cocktail of chemicals—the class, structure, biodegradability, toxicity and human and environmental impact of most of which are still unknown.

© 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.

Some of the wastewater contaminants, including aromatics, pharmaceuticals, pesticides,33 chlorinated congeners and plasticizers, pose deleterious effects on human and environmental health, even at trace levels [1]. Some of their harmful effects include impairment and/or abnormality in physiological processes, including reproductive impairment, increased risk of cancer in aquatic and terrestrial species, development of antibiotic‐resistant bacterial strains and increase in effluent toxicity post‐treatment plausibly owing to the synergistic or antago‐ nistic toxic effects of such recalcitrant chemical mixtures. Still unknown are the environmental effects of many emerging contaminants.

While most of the easily degradable wastewater contaminants are removed by conventional treatment methods, compounds that remain even in the treated effluent are recalcitrant and hence persist in the receiving environments, causing environmental and health problems. Low concentrations of such recalcitrants in large volumes of wastewater make their efficient treatment and removal very difficult by the conventional treatment processes including activated carbon, chemical precipitation, ionic exchange resins and membrane filtration [2]. Such processes have other disadvantages such as high plant operation and maintenance cost, accumulation and disposal issues of concentrated sludge, use of excessive chemicals, low sensitivity towards target compounds and accumulation of concentrated sludge and their disposal problems [3]. Removal of some of the organic recalcitrants is not effected even by the traditional biological processes, including activated sludge and trickling filters, employing microorganisms as these biorecalcitrants may result in death of the microbial population, thus reducing the efficiency of or halting the treatment process. Advanced treatment methods such as a pre‐separation step or post‐treatment of recalcitrants using potent and specialized microbial strains need to be employed for the efficient removal of such persistent organic pollutants from effluent [2].

Hence, there is a need for better understanding of the occurrence, behaviour and fate of organic contaminants during sewage treatment processes. The present paper reviews liter‐ ature about the fate of some of the recalcitrant organic contaminants during the various treatment processes.

#### **2. Status of wastewater generation**

Better management of wastewater at regional and global level requires up‐to‐date informa‐ tion on the status of sewage generation and treatment. Globally, a complete sewage genera‐ tion and treatment data are available for only 55 countries, 37% of it being recent (2008– 2012) [4]. There is a generation of about 15, 644 millions litre per day (MLD) of sewage from 35 metropolitan cities in India, out of which only 8040 MLD (51.4%) is the existing treatment capacity. While 3800 MLD is the municipal sewage generation in the national capital region of Delhi, the city has a treatment capacity of only about 2300 MLD. Rest 31% sewage is dis‐ charged into the environment untreated [5].

## **3. Wastewater treatment processes**

Some of the wastewater contaminants, including aromatics, pharmaceuticals, pesticides,33 chlorinated congeners and plasticizers, pose deleterious effects on human and environmental health, even at trace levels [1]. Some of their harmful effects include impairment and/or abnormality in physiological processes, including reproductive impairment, increased risk of cancer in aquatic and terrestrial species, development of antibiotic‐resistant bacterial strains and increase in effluent toxicity post‐treatment plausibly owing to the synergistic or antago‐ nistic toxic effects of such recalcitrant chemical mixtures. Still unknown are the environmental

While most of the easily degradable wastewater contaminants are removed by conventional treatment methods, compounds that remain even in the treated effluent are recalcitrant and hence persist in the receiving environments, causing environmental and health problems. Low concentrations of such recalcitrants in large volumes of wastewater make their efficient treatment and removal very difficult by the conventional treatment processes including activated carbon, chemical precipitation, ionic exchange resins and membrane filtration [2]. Such processes have other disadvantages such as high plant operation and maintenance cost, accumulation and disposal issues of concentrated sludge, use of excessive chemicals, low sensitivity towards target compounds and accumulation of concentrated sludge and their disposal problems [3]. Removal of some of the organic recalcitrants is not effected even by the traditional biological processes, including activated sludge and trickling filters, employing microorganisms as these biorecalcitrants may result in death of the microbial population, thus reducing the efficiency of or halting the treatment process. Advanced treatment methods such as a pre‐separation step or post‐treatment of recalcitrants using potent and specialized microbial strains need to be employed for the efficient removal of such persistent organic

Hence, there is a need for better understanding of the occurrence, behaviour and fate of organic contaminants during sewage treatment processes. The present paper reviews liter‐ ature about the fate of some of the recalcitrant organic contaminants during the various

Better management of wastewater at regional and global level requires up‐to‐date informa‐ tion on the status of sewage generation and treatment. Globally, a complete sewage genera‐ tion and treatment data are available for only 55 countries, 37% of it being recent (2008– 2012) [4]. There is a generation of about 15, 644 millions litre per day (MLD) of sewage from 35 metropolitan cities in India, out of which only 8040 MLD (51.4%) is the existing treatment capacity. While 3800 MLD is the municipal sewage generation in the national capital region of Delhi, the city has a treatment capacity of only about 2300 MLD. Rest 31% sewage is dis‐

effects of many emerging contaminants.

4 Biological Wastewater Treatment and Resource Recovery

pollutants from effluent [2].

**2. Status of wastewater generation**

charged into the environment untreated [5].

treatment processes.

Various processes are employed for the removal of wastewater contaminants depending on their type and level in the influent. Municipal wastewater is mostly treated in sewage treat‐ ment plants (STPs) which use various treatment processes including physical, chemical and biological. Wastewater treatment and discharge are done according to regional and nation‐ al regulations and standards. Wastewater treatment is done with the purpose of producing a pollutant‐ and toxicity‐free effluent which can safely be discharged into the environment [6]. Three main stages are involved in wastewater treatment, viz., primary or physic‐chemi‐ cal, secondary or biological and tertiary or advanced treatment.


The most important aerobic treatment system is the activated‐sludge process, based on the maintenance and recirculation of a complex biomass composed by micro‐organisms able to absorb and adsorb the organic matter carried in the wastewater. Other biological treat‐ ment processes such as expanded granular sludge bed (EGSB) reactor and upflow anaero‐ bic sludge blanket (UASB) are also employed for wastewater treatment. Synthetic membranes and micro‐filtration are now commonly being used as tertiary treatment tech‐ nologies.

## **4. Fate of organic recalcitrant contaminants in wastewater treatment**

#### **4.1. Pathways of contaminant removal**

There has been a radical increase in the occurrence and concentration of organic contaminants in wastewater and sludge as a result of an increase in the demand and industrial production of synthetic organic chemicals. Point discharge sources including discharges from industrial users or manufacturers and diffuse discharge sources such as commercial and domestic premises or run‐off after aerial deposition are some of the major contributors to the loading of organic contaminant in sewage. The following are some of the pathways (**Figure 1**) through which organic contaminants may be transformed or degraded during sewage treatment:


**Figure 1.** Some of the pathways involved in transformation of organic contaminant in wastewater treatment.

While some compounds may get completely degraded or mineralized in the process of treatment, some others are partially degraded and form breakdown products and a few other recalcitrant compounds may remain unaffected and persist in the effluent even after treatment. The occurrence of these synthetic organic contaminants in wastewater may be either in solution or sorbed onto solids. The hydrophobic or lipophilic nature of many organic contaminants result into their getting adsorbed on solid particles during wastewater treatment, eventually resulting in their accumulation in the sludge solids, sometimes at concentrations much higher than in the untreated wastewater [10, 11].

Structural composition of the organic residues may also provide information about their biodegradation pathways. For instance, biodegradation of unbranched and long‐chained hydrocarbons is easier as compared to the short‐chained or highly branched molecules. Biodegradation of unsaturated aliphatic compounds is generally more favoured than their saturated analogues. Molecules having highly polar groups and linkages tend to react by nucleophilic displacement (such as hydrolysis) [12]. Petrasek et al. [13] reported the association of recalcitrant and toxic chloro‐organic pentachlorophenol (PCP) with the sludge solids, and considerable degradation of phenolic compounds having polar groups.

#### **4.2. Processes involved in contaminant removal**

**4. Fate of organic recalcitrant contaminants in wastewater treatment**

There has been a radical increase in the occurrence and concentration of organic contaminants in wastewater and sludge as a result of an increase in the demand and industrial production of synthetic organic chemicals. Point discharge sources including discharges from industrial users or manufacturers and diffuse discharge sources such as commercial and domestic premises or run‐off after aerial deposition are some of the major contributors to the loading of organic contaminant in sewage. The following are some of the pathways (**Figure 1**) through which organic contaminants may be transformed or degraded during sewage treatment:

**Organic contaminants**

**Adsorp�on onto primary sludge solids**

**Ion exchange adsorp�on**

> **Output in Aqueous Effluent**

**Figure 1.** Some of the pathways involved in transformation of organic contaminant in wastewater treatment.

Soluble

Recalcitrant

**Detoxifica�on Destruc�ve Methods**

> Recalcitrant

**Evapora�on/Air Stripping**

**Anaerobic or aerobic biodegrada�on**

**4.1. Pathways of contaminant removal**

6 Biological Wastewater Treatment and Resource Recovery

**•** Air stripping

**•** Sorption

**•** Volatilization

**•** Biodegradation

**•** Chemical degradation

Several researches have been made to study the removal efficiency of various contaminants by different wastewater treatment processes. Partitioning of hydrophobic contaminants of influent onto settled primary sludge solids may take place during the primary sedimentation process in the primary clarifiers. Bulk organic components of wastewater such as cellulose, proteins and carbohydrates get biodegraded during the secondary treatment involving aerobic processes such as trickling filters, activated‐sludge process, oxidation ponds or anaerobic processes resulting in sludge digestion. Transformation or loss of some of the synthetic recalcitrant organic contaminants may also take place during the secondary treatment processes. Polysaccharides, proteins and fats occur in two phases during the anaerobic digestion process. First phase (acid phase) involves hydrolysis of polysaccharides to form mono‐ and disaccharides, of proteins to form amino acids, and of fats resulting in the formation of long‐chain fatty acids, and volatile acids such as formic, acetic and butyric acid. Second phase (methanogenic phase) results in the reduction of the volatile acids to methane and carbon dioxide [12, 14]. In one study involving a generalized model for the presentation of fate of organic compounds in an activated‐sludge process, it was demonstrated that the phase distribution of xenobiotic chemicals depended quantitatively upon their physico‐chemical properties and the operating conditions of wastewater treatment. The study also showed the removal of hydrophobic chemicals of wastewater, mostly by the process of sorption onto sludge particles followed by their transfer to the sludge‐processing units. Meanwhile, advec‐ tive transport into the final effluent and biodegradation was shown to be the common mechanism for the removal of hydrophilic compounds of wastewater. The model also pre‐ dicted an increase in the effluent concentration of complex organics such as substituted phthalates, high molecular weight (HMW) polycyclic aromatic hydrocarbons (PAHs) and dioxins with increasing solids retention time (SRT) during the operation of wastewater treatment plant [15].

#### **4.3. Common classes of contaminants found in wastewaters**

Although wastewaters contain a multitude of contaminants, yet they can be broadly grouped under different classes on the basis of their chemical structure. A total of 129 specific pollutants including heavy metals and specific organic chemicals have been defined by the US Clean Water Act as "Priority Pollutants". Municipal Environmental Research Laboratory (MERL), US EPA, conducted a comprehensive research programmes on the occurrence and fate of priority pollutants present in wastewater and sludge. The study assessed the fate and behav‐ iour of 22 harmful organics including phenols, pesticides, poly aromatic hydrocarbons and phthalates in the conventional water treatment systems and demonstrated up to 95–98% removal of organic compounds from the liquid phase. Many such organic compounds were found to have been partitioned onto the solid phases of primary and return activated sludges. Similar results were reported in other studies as well [16, 17]. In one study, the highest degree of enrichment of PAHs was observed in the primary sludge and phthalates such as bis‐ (ethylhexyl) and di‐n‐octyl phthalate were found to be among the most recalcitrant compounds present in wastewaters [13]. Wild and Jones [18] reported the occurrence of volatile chemicals, such as benzene, in sewage sludge, possibly as a result of their sorption over organic substances present in the sludge. Based on the reported literature, the following description discusses the fate of some common classes of organic compounds occurring in wastewaters (**Figure 2**).

**Figure 2.** Classes of organic contaminants commonly found in wastewater.

#### *4.3.1. Phthalic acid esters*

dioxins with increasing solids retention time (SRT) during the operation of wastewater

Although wastewaters contain a multitude of contaminants, yet they can be broadly grouped under different classes on the basis of their chemical structure. A total of 129 specific pollutants including heavy metals and specific organic chemicals have been defined by the US Clean Water Act as "Priority Pollutants". Municipal Environmental Research Laboratory (MERL), US EPA, conducted a comprehensive research programmes on the occurrence and fate of priority pollutants present in wastewater and sludge. The study assessed the fate and behav‐ iour of 22 harmful organics including phenols, pesticides, poly aromatic hydrocarbons and phthalates in the conventional water treatment systems and demonstrated up to 95–98% removal of organic compounds from the liquid phase. Many such organic compounds were found to have been partitioned onto the solid phases of primary and return activated sludges. Similar results were reported in other studies as well [16, 17]. In one study, the highest degree of enrichment of PAHs was observed in the primary sludge and phthalates such as bis‐ (ethylhexyl) and di‐n‐octyl phthalate were found to be among the most recalcitrant compounds present in wastewaters [13]. Wild and Jones [18] reported the occurrence of volatile chemicals, such as benzene, in sewage sludge, possibly as a result of their sorption over organic substances present in the sludge. Based on the reported literature, the following description discusses the fate of some common classes of organic compounds occurring in wastewaters (**Figure 2**).

Contaminant classes

Phthalic acid esters

**Figure 2.** Classes of organic contaminants commonly found in wastewater.

Personal care products

Chlorinate d congeners

Pharmace u�cals

PAHs

**4.3. Common classes of contaminants found in wastewaters**

treatment plant [15].

8 Biological Wastewater Treatment and Resource Recovery

Phthalates have a high environmental significance owing to their high production volumes as well as their eco‐toxicological effects especially on aquatic fauna including molluscs, crusta‐ ceans and amphibians. They have been reported to cause biological effects even at very low levels of exposure, varying in the range of ng L‐1 to μg L‐1 [19, 20]. Microbial degradation of phthalates under aerobic and anaerobic conditions has been previously reported [21]. The difference in the biodegradability of various phthalates could possibly be due to the steric effect of their side ester chains that hinders the binding of hydrolytic enzymes to the phthalates thus inhibiting their hydrolysis [22]. In a previous study on the occurrence of phthalates in raw and treated wastewater of WWTPs, it was found that most of the studied phthalates were present in post‐treated water samples, bis(2‐ethylbenzyl) phthalate (DEHP) being the most abundant. Also, biotransformation and adsorption onto sludge solids (that directly depend on the molecular weight and lipophilic nature of the compound) were shown to be the possible pathways of phthalate removal from liquid phase during wastewater treatment [23]. Roslev et al. [24] studied the degradation of four different phthalic acid esters in an activated‐sludge process, and showed an almost 96% association of DEHP (showing the least biodegradation among the four phthalates) with the wastewater suspended solids. The study also revealed a 7–9% recovery of the influent phthalate esters in the effluent. Also, aerobic and anoxic‐ denitrifying conditions were found to be less favourable for biodegradation of phthalate esters as compared to the alternating aerobic‐anoxic conditions.

#### *4.3.2. Polycyclic aromatic hydrocarbons*

PAHs are among the most mutagenic, carcinogenic and toxic class of organic contaminants some of which have also been included in the US‐EPA and EU list of priority pollutants [25]. The presence of PAHs in the environment is commonly attributed to various anthropogenic activities such as petroleum refining, power and heat generation from coal production, and chemical manufacturing [26]. A study on the fate of PAHs and other volatile organic com‐ pounds (VOCs) during wastewater treatment by the conventional activated‐sludge process (CASP) and the membrane bioreactors (MBRs) concluded that aromatic VOCs were removed mainly by volatilization and with comparable removal efficiencies for both treatment proc‐ esses, that is, CASP and MBRs. On the other hand, removal efficiency for PAHs was found to be enhanced in case of MBRs [27]. In another study conducted by Zhang et al. [28], the occurrence, behaviour and fate of 18 PAHs in a coking wastewater treatment plant was investigated and it was found that mostly high molecular weight PAHs were present in the raw coking wastewater, while 3–6 ring PAHs were the predominant PAHs detected in the effluent. There was detection of PAHs such as pyrene, phenanthrene and fluoranthene in the gas samples and pyrene, fluoranthene, chrysene and benzo[k]fluoranthene in sludge. While there was almost 97% removal for all the PAHs during treatment, the percent removal of PAHs from the liquid phase varied in a range of 47–92% in the biological stage. It was also observed that low molecular weight (LMW) PAHs were mostly removed in the aerobic tanks and following the mechanism of transformation, whereas their HMW counterparts were mainly removed in anaerobic tank. While transformation was observed to be the most common mechanism of removal of LMW PAHs from wastewaters, adsorption onto sludge solids was mainly responsible for the removal of HMW PAHs from the liquid phase.

#### *4.3.3. Chlorinated congeners*

Chlorinated congeners including polychlorinated biphenyls and polychlorinated pesticides are very toxic to human and environment health and are mostly added into the environment by industrial and domestic sources. Their presence has commonly been reported in wastewa‐ ter, surface water bodies as well as in sediments. Biologically mediated reductive dehaloge‐ nation process is one of the common pathways of degradation of these chlorinated contaminants during wastewater treatment. The less investigated reductive dechlorination process has also been identified as one of the possible pathways for the transformation of specific contaminants during anaerobic digestion of sludge. Previous studies have reported the formation of intermediates such as 1, 2 ,4‐trichlorobenzene and pentachlorobenzene, 1, 2, 4, 5‐tetrachlorobenzene and final products such as dichlorobenzene isomers and 1, 3 ,5‐ trichlorobenzene during the reductive dechlorination of hexachlorobenzene. The formation of 2, 4‐dichlorophenol and 4‐chlorophenol as intermediates and phenol as the end product during reductive dechlorination of 2, 4‐dichlorophenoxy acetate has similarly been reported [29].

While some of the chlorinated congeners such as polychlorinated biphenyls, have been known for long [30], some others have recently been documented as toxic contaminants including pharmaceuticals such as diclofenac and pesticides 4‐hydroxychlorothalonil and clomazone [31, 32]. The detection of such chlorinated contaminants, some of which are also endocrine‐ disrupting and toxic to biota, in effluent and receiving water bodies is a matter of concern [33]. The concentrations of chlorinated congeners in effluent have been reported to be much lower than in the influent, indicating their efficient removal by various physical, chemical or biological processes operational during the treatment of wastewater [34]. Nevertheless, there have been reports indicating the presence of chlorinated contaminants such as triclosan and triclocarban in effluent of STPs, and eventually in the downstream water bodies and sedi‐ ments [35, 36], thus pointing towards a need for upgradation of treatment mechanisms for their efficient removal. In a study conducted on the efficiency of aerobic and anaerobic processes in organic contaminant removal during treatment processes, it was concluded that a sequential system using a combination of both oxidative and reductive processes was probably the most efficient for the removal of recalcitrant organics. Highly chlorinated and volatile organohalogen compounds were found to degrade appreciably only under anaerobic conditions, while being resistant to oxidative degradation under aerobic conditions [37].

#### *4.3.4. Pharmaceutical compounds*

Pharmaceutical compounds are another class of emerging contaminants that have gained growing concerns in the past two decades mostly because of their less known health and environmental effects and ever‐increasing usage and unchecked release into the environment. Metabolic excretion post consumption and improper disposal techniques are the main sources of these compounds in the environment. In a study conducted to investigate the presence of some common pharmaceutical compounds and fluoroquinolones (one of the "priority pollutants" having potential hazardous effects on the aquatic life) in two wastewater treatment plants in Spain, frequent detection of pharmaceuticals such as analgesics, anti‐inflammatories and lipid regulators in effluent and incomplete elimination of most of the fluoroquinolones posttreatment was observed. The results also demonstrated higher efficiency of membrane bioreactor technique in removing pharmaceutical compounds as compared to the activated‐ sludge process [38]. Similar findings have been reported by other workers as well [39, 40].

#### *4.3.5. Personal care products*

mechanism of removal of LMW PAHs from wastewaters, adsorption onto sludge solids was

Chlorinated congeners including polychlorinated biphenyls and polychlorinated pesticides are very toxic to human and environment health and are mostly added into the environment by industrial and domestic sources. Their presence has commonly been reported in wastewa‐ ter, surface water bodies as well as in sediments. Biologically mediated reductive dehaloge‐ nation process is one of the common pathways of degradation of these chlorinated contaminants during wastewater treatment. The less investigated reductive dechlorination process has also been identified as one of the possible pathways for the transformation of specific contaminants during anaerobic digestion of sludge. Previous studies have reported the formation of intermediates such as 1, 2 ,4‐trichlorobenzene and pentachlorobenzene, 1, 2, 4, 5‐tetrachlorobenzene and final products such as dichlorobenzene isomers and 1, 3 ,5‐ trichlorobenzene during the reductive dechlorination of hexachlorobenzene. The formation of 2, 4‐dichlorophenol and 4‐chlorophenol as intermediates and phenol as the end product during reductive dechlorination of 2, 4‐dichlorophenoxy acetate has similarly been reported [29].

While some of the chlorinated congeners such as polychlorinated biphenyls, have been known for long [30], some others have recently been documented as toxic contaminants including pharmaceuticals such as diclofenac and pesticides 4‐hydroxychlorothalonil and clomazone [31, 32]. The detection of such chlorinated contaminants, some of which are also endocrine‐ disrupting and toxic to biota, in effluent and receiving water bodies is a matter of concern [33]. The concentrations of chlorinated congeners in effluent have been reported to be much lower than in the influent, indicating their efficient removal by various physical, chemical or biological processes operational during the treatment of wastewater [34]. Nevertheless, there have been reports indicating the presence of chlorinated contaminants such as triclosan and triclocarban in effluent of STPs, and eventually in the downstream water bodies and sedi‐ ments [35, 36], thus pointing towards a need for upgradation of treatment mechanisms for their efficient removal. In a study conducted on the efficiency of aerobic and anaerobic processes in organic contaminant removal during treatment processes, it was concluded that a sequential system using a combination of both oxidative and reductive processes was probably the most efficient for the removal of recalcitrant organics. Highly chlorinated and volatile organohalogen compounds were found to degrade appreciably only under anaerobic conditions, while being resistant to oxidative degradation under aerobic conditions [37].

Pharmaceutical compounds are another class of emerging contaminants that have gained growing concerns in the past two decades mostly because of their less known health and environmental effects and ever‐increasing usage and unchecked release into the environment. Metabolic excretion post consumption and improper disposal techniques are the main sources of these compounds in the environment. In a study conducted to investigate the presence of some common pharmaceutical compounds and fluoroquinolones (one of the "priority

mainly responsible for the removal of HMW PAHs from the liquid phase.

*4.3.3. Chlorinated congeners*

10 Biological Wastewater Treatment and Resource Recovery

*4.3.4. Pharmaceutical compounds*

There has been a recent concern over the toxic and ecological impact of personal care products (PCPs). Although there have been several reports on the assessment of concentrations of these chemicals in the environment [41–43], less work has been done to know their fate in the environment. In one assessment of the efficiency of various treatment processes for the removal of pharmaceuticals and personal care products, it was concluded that membrane bioreactor and activated‐sludge process with nitrogen treatment were the most efficient processes for the treatment of such compounds [44].

## **5. Conclusion**

Wastewater treatment facilities such as wastewater treatment plants, or domestic septic systems, which have been operating on the conventional technologies, are often inefficient in treating such a cocktail of compounds ranging from simple to complex and recalcitrant organic compounds. Thus, these centralized facilities, discharging treated effluent, which may still be contaminated with household chemicals, pharmaceuticals and biogenic hormones, into the environment end up being a source of pollutants for the receiving water bodies. Also, the sewage sludge generated at the STPs, often having a high accumulation of recalcitrant and hydrophobic contaminants, acts as a sink of such contaminants in the treatment facilities but a major source of organic recalcitrants when directly used as manure.

Such unchecked disposal and use of sewage and sludge into the environment or their direct application for domestic or agriculture purposes could lead to exposure of toxic contaminants to biological systems, possibly resulting in adverse metabolic responses. Advanced treatment technologies such as membrane bioreactors and sequential system using a combination of both oxidative and reductive processes were found to be more effective in the removal of various organic recalcitrant compounds. Therefore, implementation of such treatment technologies and addition of tertiary treatment techniques to the conventional methods, for the removal of such persistent contaminants,have become quintessential.

Thus, the occurrence of persistence organic contaminants in the effluent and sludge posttreat‐ ment and ambiguity about their fate pose a serious environmental challenge. Therefore, much research is still needed to identify the source, behaviour and sink as well as their ecological and health effects.

## **Acknowledgements**

This paper was supported by the research grants of the Jawaharlal Nehru University, New Delhi, India. The author (Gupta A) also thanks the University Grants Commission (UGC), New Delhi, Government of India, for providing Research Fellowship.

### **Author details**

Asmita Gupta\* and Indu Shekhar Thakur

\*Address all correspondence to: asmitagupta5@gmail.com

School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India

#### **References**


[9] Randall CW, Sen D. Full‐scale evaluation of an integrated fixed‐film activated sludge (IFAS) process for enhanced nitrogen removal. Water Sci. Technol. Elsevier; 1996; 33(12): 155–162. doi: 10.1016/0273‐1223(96)00469‐6

**Acknowledgements**

12 Biological Wastewater Treatment and Resource Recovery

**Author details**

Asmita Gupta\*

**References**

This paper was supported by the research grants of the Jawaharlal Nehru University, New Delhi, India. The author (Gupta A) also thanks the University Grants Commission (UGC),

New Delhi, Government of India, for providing Research Fellowship.

School of Environmental Sciences, Jawaharlal Nehru University, New Delhi, India

taining simulated wastewater. Process Biochem. 2005; 40(2):885–890.

dioxide loaded Amberlite XAD‐7 resin. Anal. Sci. 2002; 18:1345–1349.

[1] Lai HT, Hou JH, Su CI, Chen CL. Effects of chloramphenicol, florfenicol, and thiam‐ phenicol on growth of algae *Chlorella pyrenoidosa, Isochrysis galbana*, and *Tetraselmis*

[2] Padmavathy S, Sandhya S, Swaminathan K, Subrahmanyam YV, Chakrabarti T, Kaul SN. Microaerophilic‐aerobic sequential batch reactor for treatment of azo dyes con‐

[3] Balaji T, Matsunaga H. Adsorption characteristics of As(III) and As(V) with titanium

[4] Sato T, Qadir M, Yamamoto S, Endo T, Zahoor A. Global, regional, and country level need for data on wastewater generation, treatment, and use. Agricul. Water Manage.

[5] CPCB. Inventorization of sewage treatment plants, Control of urban pollution Series: CUPS//2015. Central Pollution Control Board, India. 2015. Available from: http:// cpcb.nic.in/upload/NewItems/NewItem\_210\_Inventorization\_of\_Sewage‐Treat‐

[6] Khopkar SM. Environmental Pollution Monitoring and Control. New Delhi: New Age

[7] EPA. Package Plants. Wastewater Technology Fact Sheet. Washington, DC. Document

[8] EPA. Primer for Municipal Waste Water Treatment Systems. Washington, DC. Docu‐

and Indu Shekhar Thakur

\*Address all correspondence to: asmitagupta5@gmail.com

*chui*. Ecotoxicol. Environ. Safe. 2009; 72:329–334.

International; 2004. p. 299. ISBN81‐224‐1507‐5.

2013; 130:1–13. ISSN 0378‐3774

no. EPA. 2000; 832‐F‐00‐016.

ment no. EPA. 2004; 832‐R‐04‐001.

ment\_Plant.pdf


[36] Gautam P, Carsella JS, Kinney CA. Presence and transport of the antimicrobials ticlocarban and triclosan in a wastewater‐dominated stream and freshwater environ‐ ment. Water Res. 2014; 48(1):247–256.

[24] Roslev P, Vorkamp K, Aarup J, et al. Degradation of phthalate esters in an activated

[25] Arfsten DP, Schaeffer DJ, Mulveny DC. The effects of near ultraviolet radiation on the toxic effects of polycyclic aromatic hydrocarbons in animals and plants: A review.

[26] Suess MJ. The environmental load and cycle of polycyclic aromatic hydrocarbons. Sci.

[27] Fatone F, Di Fabio S, Bolzonella D, Cecchi F. Fate of aromatic hydrocarbons in Italian municipal wastewater systems: an overview of wastewater treatment using conven‐ tional activated‐sludge processes (CASP) and membrane bioreactors (MBRs). Water

[28] Zhang W, Wei C, Chai X et al. The behaviours and fate of polycyclic aromatic hydro‐ carbons (PAHs) in a coking wastewater treatment plant. Chemosphere. 2012; 88:174–

[29] Schwartzenbach RP. Groundwater contamination by volatile halogenated alkanes: abiotic formation of volatile sulphur compounds under anaerobic conditions. Environ.

[30] Gutierrez AG, McIntyre AE, Perry R, Lester JN. Behaviour of persistent organochlorine micropollutants during primary sedimentation of waste water. Sci. Total Environ. 1984;

[31] Hug C, Ulrich N, Schulze T, Brack W, Krauss M. Identification of novel micropollutants in wastewater by a combination of suspect and non‐target screening. Environ. Pollut.

[32] Nelson ED, Do H, Lewis RS, Carr SA. Diurnal variability of pharmaceutical, personal care product, estrogen and alkylphenol concentrations in effluent from a tertiary

[33] Krzmarzick MJ, Novak PJ. Removal of chlorinated organic compounds during waste‐ water treatment: achievements and limits. Appl. Microbiol. Biotechnol. 2014; 98(14):

[34] Langford KH, Lester JN. Fate and behavior of endocrine disrupters in wastewater treatment processes. In: Lester JN, Birkett JW (eds) Endocrine Disrupters in Wastewater

[35] Zhang Q‐Q, Zhao J‐L, Liu Y‐S, Li B‐G, Ying G‐G. Multimedia modeling of the fate of triclosan and triclocarban in the Donjian River Basin, South China and comparison with

wastewater treatment facility. Environ. Sci. Technol. 2011; 45:1228–1234.

and Sludge Treatment Processes. London: CRC Press; 2003. 103–143.

field data. Environ Sci: Process Impacts. 2013; 15(11):2142–2152.

sludge wastewater treatment plant. Water Res. 2007; 41(5):969–976.

Ecotox. Environ. Safe. 1996; 33:1–24.

Total Environ. 1976; 6:239–250.

14 Biological Wastewater Treatment and Resource Recovery

Sci. Technol. 1985; 19:322–327.

Res. 2011; 45(1):93–104.

182.

39:27–47.

2014; 184:25–32.

6233–6242.


#### **Application of Macrobiological Methods in the Settlement Wastewater Treatment Application of Macrobiological Methods in the Settlement Wastewater Treatment**

Dragan Milićević, Slaviša Trajković and Milan Gocić Dragan Milićević, Slaviša Trajković and Milan Gocić

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

http://dx.doi.org/10.5772/65369

#### **Abstract**

The approach to solving the problem of water protection is changing in the world, and the opinion that wastewater is a resource instead of waste is now prevalent with research being directed in the direction of simpler, energetically more rational and more economically acceptable technological solutions for wastewater treatment, primarily in the field of biotechnology, especially there where favorable climate conditions and the use of large land areas are available. The mechanism of wastewater treatment by macrobiological methods is simple and is reduced to extraction of certain substances from wastewater directly with plants or through the food chain with animals and their concentration into macrobiological living stations. Macrobiological living stations are extracted from the water in the form of biomass by simple mechanical methods, and in that way the final removal of nutrients and other substances from the water is completed. The produced biomass can be used as food or feed, with mandatory sanitary inspection, or as an emergent in biomass production. This paper presents the principles of application of macro biologic methods in wastewater treatment and the experience gained through the research at the Faculty of Civil Engineering of Niš and at the waste water treatment facilities in Sokobanja.

**Keywords:** wastewater treatment, macrobiological methods, resource and energy po‐ tential

## **1. Introduction**

The sudden technological and industrial development, tumulous demographic growth and rapid urbanization especially in the last two decades pose humanity with four big problems: water, food, energy and environment. The problem of water is especially pronounced because

© 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.

it is implicitly present in the remaining three problems, the production of food and energy are decisively dependent on water and the key problems of environment protection are protection of water quality and protection from the harmful effects of water.

Even today, the water crisis is well underway and according to predictions, by the middle of this century it will develop into a crisis of global proportions. The gap between available reserves of water and the increasing need for it on one and the pollution of water resources on the other hand are more and more pronounced with each day so rational use of water resources and their protection from further pollution, today and especially in the near future are developing as the main global problem.

The world today has many high technologies for wastewater treatment available, primarily of physicochemical nature, which allow for wastewater to be treated to a very high level and can satisfy all strict and more rigorous demands set regarding quality of treated effluents released into water recipients. However, the time of intensive development of these technologies was also the time of cheap energy, but today, when the evidence crisis is evident these technologies are too expensive even for the most developed countries of the world. That is why, in the world, the approach to solving the wastewater treatment problem is increasingly changing and intensive research is conducted in the direction of cheaper technologies for treatment of wastewater and protection of water from pollution.

In the last three decades, a special interest in the world is aroused by the potential of using the macrobiological methods in the waste water treatment, whose application as of natural and not artificial waste water treatment processes provide effluents of demanded quality in an economically acceptable way in technically simple objects.

## **2. Macrobiological methods in the wastewater treatment**

Until the energy crisis which emerged in the 1970s the leading approach to secondary (bio‐ logical) wastewater treatment was the philosophy of destruction of organic matter, and tertiary treatment, removal of nutrients from wastewaters and treated waters was mainly connected with complex and expensive chemical technological processes.

The hint of an energy crisis demanded directing towards cheaper technologies for treatment of wastewater and protection of water resources from eutrophication. The opinion that wastewater is a resource instead of waste became increasingly prevalent and research was intensively directed towards simpler and economically more acceptable technological solutions, primarily in the field of biotechnology especially in conditions where the use of large land areas is possible. The tendencies of questioning the philosophy of destruction of organic matter, or their mineralisation, have become more widespread, present in the previous technological practice of wastewater treatment and accepting the philosophy of synthesis of organic matter and nutrients into higher forms, which brings up the numerous matter in wastewaters [nitrogen (N), phosphorus (P), etc.] which can be interesting and useful as a raw material.

In accordance to the philosophy of the synthesis of organic matter and nutrients into higher forms, the possibility to use the macrobiological methods for the wastewater treatment attracted interest around the world.

Macrobiological methods of wastewater treatment present aerobic processes of synthesis in the direction of more complex organic matter which is easily removed from water in the form of biomass with noticeable reduction of energy needed for functioning of the system and encompass a whole series of macrobiological unit operations, the list of which, with constant research conducted around the world, continues to grow. These methods are applied as natural instead of artificial processes of wastewater treatment and they provide effluents of demanded quality in an economically acceptable manner in technically simple objects.

Intensive research in this area began in the 1970s in the world [1], and almost at the same time in Serbia [2, 3]. The starting results were obtained under the direction of Prof. Dr Lazar Ignjatović, in the period between 1975 and 1979, from the Faculty of Civil Engineering, University of Niš, Serbia and the wastewater treatment plant (WWTP) in Sokobanja, which was used as training ground for the staff of the faculty, through the project "The influence of accumulation on the change of ecosystem", with the research being continued in the period of 1978–1988, through the multidisciplinary project under the same name within the Fulbright program. In order to control the nutrient into effluent, numerous macrobiological unit operations were tested as laboratory models and then brought to the level of macro model, namely the part of an already existing wastewater treatment plant in Sokobanja, in cooperation with reputable experts from the USA for certain areas.

## **3. Treatment mechanism**

it is implicitly present in the remaining three problems, the production of food and energy are decisively dependent on water and the key problems of environment protection are protection

Even today, the water crisis is well underway and according to predictions, by the middle of this century it will develop into a crisis of global proportions. The gap between available reserves of water and the increasing need for it on one and the pollution of water resources on the other hand are more and more pronounced with each day so rational use of water resources and their protection from further pollution, today and especially in the near future are

The world today has many high technologies for wastewater treatment available, primarily of physicochemical nature, which allow for wastewater to be treated to a very high level and can satisfy all strict and more rigorous demands set regarding quality of treated effluents released into water recipients. However, the time of intensive development of these technologies was also the time of cheap energy, but today, when the evidence crisis is evident these technologies are too expensive even for the most developed countries of the world. That is why, in the world, the approach to solving the wastewater treatment problem is increasingly changing and intensive research is conducted in the direction of cheaper technologies for treatment of

In the last three decades, a special interest in the world is aroused by the potential of using the macrobiological methods in the waste water treatment, whose application as of natural and not artificial waste water treatment processes provide effluents of demanded quality in an

Until the energy crisis which emerged in the 1970s the leading approach to secondary (bio‐ logical) wastewater treatment was the philosophy of destruction of organic matter, and tertiary treatment, removal of nutrients from wastewaters and treated waters was mainly connected

The hint of an energy crisis demanded directing towards cheaper technologies for treatment of wastewater and protection of water resources from eutrophication. The opinion that wastewater is a resource instead of waste became increasingly prevalent and research was intensively directed towards simpler and economically more acceptable technological solutions, primarily in the field of biotechnology especially in conditions where the use of large land areas is possible. The tendencies of questioning the philosophy of destruction of organic matter, or their mineralisation, have become more widespread, present in the previous technological practice of wastewater treatment and accepting the philosophy of synthesis of organic matter and nutrients into higher forms, which brings up the numerous matter in wastewaters [nitrogen (N), phosphorus (P), etc.] which can be interesting and useful as a raw

of water quality and protection from the harmful effects of water.

developing as the main global problem.

18 Biological Wastewater Treatment and Resource Recovery

wastewater and protection of water from pollution.

economically acceptable way in technically simple objects.

**2. Macrobiological methods in the wastewater treatment**

with complex and expensive chemical technological processes.

material.

In the wastewater treatment performed by the macrobiological methods the macrobiological living stations are used. The macrobiological living stations is the term for all the higher plants or animals with all the characteristics of living organisms, including with sexual and sexless procreation.

Living stations, either plant or animal, are mainly made of water starting from 73% in carp, 80% in terrestrial macrophytes, even up to 95% in hydrophytes. The exact percentage of water depends on the type of living station, the weight and the age of the station in the moment of sample processing.

The largest part of the living stations dry mass (usually over 90% in plants) is made of three basic elements: carbon (C), oxygen (O) and hydrogen (H), taken directly from the water or air, while a smaller part (around 10% in plants) is made of all the other elements.

The plant or animal cannot complete its life cycle in the absence of any of the necessary elements, which must be available directly or through suitable enzyme activity, provided that there are no antagonistic or toxic effects of the other elements. From 92 natural mineral known elements around 60 are found in living stations. From the 60, around 16 are considered essential for growth of plants, and approximately the same number is necessary for growth of aquatic animals of interest. It should be considered that although they do not need other elements, plants and animals accumulate some elements not significant for growth and development.

Wastewater treatment is completed through bioconcentration, or accumulation of substance from the environment and concentration in a biological station, directly in plants or indirectly, through the food chain, in animals. Because living matter is formed by a few biogenic elements either plastic: hydrogen (H), oxygen (O), carbon (C) and nitrogen (N) or oligo elements: zinc (Zn), copper (Cu), iron (Fe) and magnesium (Mg) in the presence of phosphorus (P), natrium (Na) and potassium (K), chloride (Cl) and manganese (Mn) the factor of bioconcentration of the substances from the environment in the organism is important along with the dynamic of the process, or the rate of bioconcentration of the substance of interest through time.

Numerous factors affect the growth, development and reproduction of the living stations. Some of them are crucial while others are less significant, depending on the phase of growth of the living station. The earliest developmental phases are the most sensitive in all living stations. Light and temperature, along with activational energies have crucial effect on the dynamic of the process. Plant species of areas with temperate continental climate are active when the temperature of the water reaches above 15–17°C, while tropical species are mainly active from 20 to 24°C.

If the natural conditions of the environment are favorable, application of macrobiological stations is possible in natural conditions, perfectly cheap because free energy of the Sun is used as an energy source, and the water itself serves as a collector with the least loss.

If the natural conditions of the environment are unfavorable, the influential factors can be put under control and then we have artificial environmental conditions. Application in green‐ houses in periods of unfavorable conditions of the environment, with the addition of light and thermal energy, is possible. In those cases great economic effects of use in natural conditions decrease. There where geothermal energy is available, its use may be rational for extended work over the whole year.

It is clear that the wastewater treatment mechanism is simple. It consists of taking certain substances from the water (whereby the water is rid of this substance) and its bioconcentration into microbiological living stations. By removing larger macrobiological living stations in the form of biomass from the water with simple mechanical methods the nutrients and other substances are finally removed from the water.

The final disposition of biomass, depending on its nature, is performed by the standard transport means. If the biomass is used as a nutrient or food (with necessary sanitary control) it has market value which considerably exceeds the transportation costs.

It is a highly clean technology in wastewater treatment using clean energy (solar and/or geothermal) with the final product being usable biomass. The civil engineering objects are usually made of soil, relatively simple and followed by a minimal equipment fond, which significantly influences low investment costs. In investment costs the land makes as significant item but if there is unsuitable agricultural or commercial land close to the settlement available, ideal conditions for the application of these technologies are acquired.

## **4. Macrobiological living stations**

animals of interest. It should be considered that although they do not need other elements, plants and animals accumulate some elements not significant for growth and development.

Wastewater treatment is completed through bioconcentration, or accumulation of substance from the environment and concentration in a biological station, directly in plants or indirectly, through the food chain, in animals. Because living matter is formed by a few biogenic elements either plastic: hydrogen (H), oxygen (O), carbon (C) and nitrogen (N) or oligo elements: zinc (Zn), copper (Cu), iron (Fe) and magnesium (Mg) in the presence of phosphorus (P), natrium (Na) and potassium (K), chloride (Cl) and manganese (Mn) the factor of bioconcentration of the substances from the environment in the organism is important along with the dynamic of

Numerous factors affect the growth, development and reproduction of the living stations. Some of them are crucial while others are less significant, depending on the phase of growth of the living station. The earliest developmental phases are the most sensitive in all living stations. Light and temperature, along with activational energies have crucial effect on the dynamic of the process. Plant species of areas with temperate continental climate are active when the temperature of the water reaches above 15–17°C, while tropical species are mainly

If the natural conditions of the environment are favorable, application of macrobiological stations is possible in natural conditions, perfectly cheap because free energy of the Sun is used

If the natural conditions of the environment are unfavorable, the influential factors can be put under control and then we have artificial environmental conditions. Application in green‐ houses in periods of unfavorable conditions of the environment, with the addition of light and thermal energy, is possible. In those cases great economic effects of use in natural conditions decrease. There where geothermal energy is available, its use may be rational for extended

It is clear that the wastewater treatment mechanism is simple. It consists of taking certain substances from the water (whereby the water is rid of this substance) and its bioconcentration into microbiological living stations. By removing larger macrobiological living stations in the form of biomass from the water with simple mechanical methods the nutrients and other

The final disposition of biomass, depending on its nature, is performed by the standard transport means. If the biomass is used as a nutrient or food (with necessary sanitary control)

It is a highly clean technology in wastewater treatment using clean energy (solar and/or geothermal) with the final product being usable biomass. The civil engineering objects are usually made of soil, relatively simple and followed by a minimal equipment fond, which significantly influences low investment costs. In investment costs the land makes as significant item but if there is unsuitable agricultural or commercial land close to the settlement available,

it has market value which considerably exceeds the transportation costs.

ideal conditions for the application of these technologies are acquired.

as an energy source, and the water itself serves as a collector with the least loss.

the process, or the rate of bioconcentration of the substance of interest through time.

active from 20 to 24°C.

20 Biological Wastewater Treatment and Resource Recovery

work over the whole year.

substances are finally removed from the water.

For a living being to qualify as a macrobiological living station, respectively, a technological element in wastewater or sludge treatment it has to satisfy special criteria:


Great attention should be given to the question if the macrobiological living station can, for a longer period of time, survive and reproduce in natural habitat on its own and if the species is invasive. This is of great importance, from the aspect of possible ecological effects if a macrobiological living station finds its way into the natural environment, out of the object where it is used under control.

On the basis of the mentioned criteria for the macrobiological unit operations, only a small number of plant and animal species can be qualified. They predominantly originate in the tropical zones.

**Figure 1.** Floating macrophytes: *Eichornia crassipes*, *Pistia stratiotes* and *Salvinia* [3].

Macrobiological living stations, which qualify as a technological element in wastewater treatment technology, are classified by groups that are floating macrophytes, fish, mussels, earthworms, etc. The list of macrobiological living stations is very wide, but it does not encompass all possibilities because in this area intensive research is present.

The list of possible floating macrophytes should be made of hydrophytes without woody tissues, especially ones which float on the surface of the water. These plants cannot adapt to the change of the water level so for their normal growth and development the water level must be kept approximately constant.

The representatives of this plant group use food directly in the shape of dissolved nutrients in the water. Some do that only through the leaves, e.g., floating crystalwort (*Riccia fluitans L.*) to some extent lemna (*Lemna minor L.* and *Lemna trisulca L.*), while other especially larger species do that through their roots which hang in the water, e.g., water hyacinth (*Eichhornia crassipes Martius*), water lettuce (*Pistia stratiotes L*) and salvinia (*Salvinia natans L.* and *Salvinia auriculata Aublet*) (**Figure 1**).

In terms of fish, fast‐growing species capable of consuming large quantities of food are of interest. Also of interest are food pyramids because of the choice of fish species, especially because of interrelationships in polyculture composition. Phytophagous species have a special role because they lean directly on the primary production in the aquatorium in the food chain.

**Figure 2** shows a food chain and fish species of predominant interest: silver carp (*Hypophthal‐ michthys molitrix*), grass carp (*Ctenopharyngodon idella*), bighead carp (*Hypophthalmichthys nobilis*) and common carp (*Cyprinus carpio*). The list can be expanded with some tropical fish species, e.g., tilapia (*Tilapia aurea*) and thai catfish (*Clarias batrachus Linnaeus*). The use of tropical fish species applies the same demands as the use of tropical plants.

**Figure 2.** Fish‐food pyramid and the primary interest species (based on Ref. [3]).

Fish, as poikilothermic animals, because of poor adjustment to sudden temperature changes, must not be rapidly transferred from one water environment to the other if the water temper‐ ature difference is greater than 2°C, because this leads to temperature shock and death in most species. The fish must be transferred carefully, because being thrown in the water during transfer leads to bursting of the swim bladder.

The role of mussels in removal of suspended and colloid material from wastewater deserves great attention from the researchers. It was experimented only with one species of mussel from the temperate climate belt in the Faculty of Civil Engineering of Niš. It was experimented with the zebra mussel (*Dreissena polymorpha*) (**Figure 3**). It should be noted that this species of mussel may pose not only an ecological threat, but also great danger to the hydrotechnical systems and objects, so it should be used in strictly closed systems.

**Figure 3.** Mussels – *Dreissena polymorpha* [4], *Unio pictorum* [5] and *Anodonta cygnea* [6].

The representatives of this plant group use food directly in the shape of dissolved nutrients in the water. Some do that only through the leaves, e.g., floating crystalwort (*Riccia fluitans L.*) to some extent lemna (*Lemna minor L.* and *Lemna trisulca L.*), while other especially larger species do that through their roots which hang in the water, e.g., water hyacinth (*Eichhornia crassipes Martius*), water lettuce (*Pistia stratiotes L*) and salvinia (*Salvinia natans L.* and *Salvinia auriculata*

In terms of fish, fast‐growing species capable of consuming large quantities of food are of interest. Also of interest are food pyramids because of the choice of fish species, especially because of interrelationships in polyculture composition. Phytophagous species have a special role because they lean directly on the primary production in the aquatorium in the food chain.

**Figure 2** shows a food chain and fish species of predominant interest: silver carp (*Hypophthal‐ michthys molitrix*), grass carp (*Ctenopharyngodon idella*), bighead carp (*Hypophthalmichthys nobilis*) and common carp (*Cyprinus carpio*). The list can be expanded with some tropical fish species, e.g., tilapia (*Tilapia aurea*) and thai catfish (*Clarias batrachus Linnaeus*). The use of tropical

fish species applies the same demands as the use of tropical plants.

**Figure 2.** Fish‐food pyramid and the primary interest species (based on Ref. [3]).

Fish, as poikilothermic animals, because of poor adjustment to sudden temperature changes, must not be rapidly transferred from one water environment to the other if the water temper‐ ature difference is greater than 2°C, because this leads to temperature shock and death in most

*Aublet*) (**Figure 1**).

22 Biological Wastewater Treatment and Resource Recovery

Species of interest would also probably be the painter's mussel (*Unio pictorum*), swan mussel (*Anodonta cygnea*) (**Figure 3**), Eastern Asiatic freshwater clam (*Anodonta (Sinanodonta) woodi‐ ana*) and large far eastern mussel species (order *Cristaria*) whose shell can grow over 30 cm. This area is open for research with warning if nonendemic species are in question.

Two species of earthworms are of interest as macrobiological living stations in the technology of sludge treatment: red Californian earthworm (*Eisenia fetida*) and red earthworm (*Lumbricus rubellus*) (**Figure 4**). For further research the red tiger earthworm (*Eisenia andrei*) (**Figure 4**) is interesting because it can treat rich organic waste in massive amounts. Probably the European nightcrawler (*Eisenia hortensis* or *Dendrobaena veneta*) would also be of interest but it is consid‐ ered an invasive species which should be used in strictly controlled conditions without being allowed into the natural environment [7].

**Figure 4.** Earthworms – *Eisenia fetida* [8], *Lumbricus rubellus* [9] and *Eisenia andrei* [10].

For the needs of hydroponics unit terrestrial plants are used, such as tomatoes, e.g., the American flowerpot tomatoes (*Licopersicum esculentum*), leafy vegetables, e.g., chard (*Beta vulgaris*), corn for silage and similar species. The list is very long because numerous fruits or vegetables can be used.

Carefully composed polycultures (bigger number of different species of macrobiological living stations in a unique aquatorium) have a bigger effect on the quality of effluents than a monoculture (a single species of a microbiological living station in a unique aquatorium). The reason being that a monoculture drains a narrow circle of substances and because of that it has limited effect in removing nutrients and wastewater treatment. The advantage is given to monocultures only in the case of final biomass derivation if it is used for human or animal consumption or partial wastewater treatment.

If ambient conditions favor some of the members of the polyculture, it spontaneously comes to suppression of the other members and the formation of a monoculture, namely population of a macrobiological living station which ambient and other factors provide the most suitable conditions. In those cases instead of insisting on polycultures the transition on a series of monoculture basins is expedient.

It should be mentioned that successive application of monoculture basins enlarges the investment costs. But continual additional introduction of macrobiological living stations from external sources, for polyculture maintenance, is usually more expensive than amortization of bigger investment costs in more basins.

In all cases parent clusters under optimal conditions must be ensured. This is optimally in the shape of a macrobiological living station bank on a regional level, for example botanical gardens or zoos, organized on a wider administrative area.

## **5. Objects and system design**

Working on the choice of unit operations, their synthesis into the technological process and objects and system design is complex engineering work which demands professional experi‐ ence along with team work of participants of the system design. Designing objects and the system for wastewater treatment starts from the available information on the wastewater quality, defining the type and concentration of the unwanted substances and the needed removable level. Based on the analysis, the technological scheme of the wastewater treatment system is defined.

Based on information on the amount of wastewater and its variation, hydraulic and process loads are defined. If the variations are big, the problem of synchronized hydraulic and process loads must be solved by choosing adequate modular object units. In this phase, decisions are made on the choice of macrobiological unit operations and the choice is made between mono and polycultures.

It should be kept in mind that for synthesis of macrobiological unit operations into the treatment processes, aside from macrobiological, standard (classic) unit operations are often incorporated with the purpose of bringing characteristics of wastewaters on the effluent of wanted quality. Although any wanted level of wastewater quality may be achieved through a planned combination of unit operations, the choice is made under clear economical conditions. As with any modern biological system of wastewater treatment, primary treatment must always precede a system based on macrobiological methods. For primary treatment of wastewater the use of a highly efficient (tubular) settlement tank from which the primary sludge is processed by anaerobic decomposition in digesters and on vermiculture (VF) fields next to smaller wastewater treatment plants is recommended.

Carefully composed polycultures (bigger number of different species of macrobiological living stations in a unique aquatorium) have a bigger effect on the quality of effluents than a monoculture (a single species of a microbiological living station in a unique aquatorium). The reason being that a monoculture drains a narrow circle of substances and because of that it has limited effect in removing nutrients and wastewater treatment. The advantage is given to monocultures only in the case of final biomass derivation if it is used for human or animal

If ambient conditions favor some of the members of the polyculture, it spontaneously comes to suppression of the other members and the formation of a monoculture, namely population of a macrobiological living station which ambient and other factors provide the most suitable conditions. In those cases instead of insisting on polycultures the transition on a series of

It should be mentioned that successive application of monoculture basins enlarges the investment costs. But continual additional introduction of macrobiological living stations from external sources, for polyculture maintenance, is usually more expensive than amortization of

In all cases parent clusters under optimal conditions must be ensured. This is optimally in the shape of a macrobiological living station bank on a regional level, for example botanical

Working on the choice of unit operations, their synthesis into the technological process and objects and system design is complex engineering work which demands professional experi‐ ence along with team work of participants of the system design. Designing objects and the system for wastewater treatment starts from the available information on the wastewater quality, defining the type and concentration of the unwanted substances and the needed removable level. Based on the analysis, the technological scheme of the wastewater treatment

Based on information on the amount of wastewater and its variation, hydraulic and process loads are defined. If the variations are big, the problem of synchronized hydraulic and process loads must be solved by choosing adequate modular object units. In this phase, decisions are made on the choice of macrobiological unit operations and the choice is made between mono

It should be kept in mind that for synthesis of macrobiological unit operations into the treatment processes, aside from macrobiological, standard (classic) unit operations are often incorporated with the purpose of bringing characteristics of wastewaters on the effluent of wanted quality. Although any wanted level of wastewater quality may be achieved through a planned combination of unit operations, the choice is made under clear economical conditions.

consumption or partial wastewater treatment.

24 Biological Wastewater Treatment and Resource Recovery

monoculture basins is expedient.

bigger investment costs in more basins.

**5. Objects and system design**

system is defined.

and polycultures.

gardens or zoos, organized on a wider administrative area.

When secondary treatment is in question, unlike classic technologies with microbiological population with which secondary treatment is made of a microbiological unit and a secondary settlement tank, with macrobiological methods there is no need for a secondary settlement tank. The reason being the lack of secondary (biological) sludge because the transformations of materials from wastewaters, through the food chain, are done into the biomass of the macrobiological living station.

**Figure 5.** Scheme of the human settlement waste water treatment facility [3, 11]. IS – inlet structure; ET – efficient settle‐ ment tank; BH – basin for sanitary hydrophytocultures; BA – basin for sanitary aquacultures; SD – sludge digester; VF – vermiculture fields.

Based on the technological scheme, after choosing unit operations and defining modular units, the technological scheme of the system with the basic hydraulic and technological calculations is designed. This results in a horizontal plan of the objects and their height scheme from the entrance to the exit of the treatment plant.

After the place and the role of some objects, their sizes and height positions are defined, the design of the objects for application of macrobiological unit operations is reduced to civil engineering design of objects. For object design, the knowledge of unit of macrobiological living station, design information and characteristic technical details of the object is required. Knowledge of civil engineering design, stability and dimension of constructions and civil engineering regulations is also required.

For application of macrobiological unit operations, two tendencies are present:


The example for the first approach for smaller settlements is given in the follow‐up. The scheme of a wastewater treatment plant is given in **Figure 5** with object marks. Each of the objects is described in more detail with needed design information.

The shown scheme is applicable for settlements without industrial and toxic wastewater. The scheme incorporates wastewater treatment and sludge stabilization so that they can be disponated into the natural environment without negative ecological effects behind the treatment plant. This is ecologically clean technology.

Depending on the ability and concert of the operator the removal of suspended solids is from 80% to above 95%, and this applies for putrescible matter too. The reduction of bacteria is above 99% so the water can be used for irrigation in semiarid areas without danger.

This technological scheme is more favorable in the level of efficiency and the produced biomass if the climate conditions are warmer and insolation is more intensive. In areas of temperate continental climate the starting hypothesis is the disposal of solar energy during the whole year or geothermal energy, if continued work of the system with low cost investment and maintenance is desired during the whole year.

The technological scheme and all unit operations are checked on the wastewater treatment plant in Sokobanja, Serbia which served as a pilot treatment plant with the process scale of 1:1, under realistic conditions. In the follow‐up, description and instructions for some objects are given.

#### **5.1. IS – inlet structure**

Inlet structure (**Figure 6**) serves for removal of large suspended matter and measurement of flow of wastewater. It is made of a channel with a grid which continues to the Parshall flume. The space between iron flat bars is 2–5 cm. The slope is 1:2 to 1:3 for easier cleaning. For smaller treatment plants the cleaning is done manually with loading of handcarts and daily transport to burial of the material from the grid to a suitable location in the treatment plant area. For bigger treatment plants the cleaning of the grid is automatic and the transport of material is off the grid to the landfill.

**Figure 6.** Inlet structure and efficient settlement tank on the WWTP Sokobanja.

Based on the known, defined hydraulic load, standard hydraulic calculation of width of the channel for defined level and loss is completed.

### **5.2. ET – efficient settlement tank**

Based on the technological scheme, after choosing unit operations and defining modular units, the technological scheme of the system with the basic hydraulic and technological calculations is designed. This results in a horizontal plan of the objects and their height scheme from the

After the place and the role of some objects, their sizes and height positions are defined, the design of the objects for application of macrobiological unit operations is reduced to civil engineering design of objects. For object design, the knowledge of unit of macrobiological living station, design information and characteristic technical details of the object is required. Knowledge of civil engineering design, stability and dimension of constructions and civil

**•** For smaller agglomerations, especially with seasonal problems, macrobiological unit operations are synthesized into complex, with cheaper investment and maintenance objects. **•** For bigger agglomerations behind classical treatment plants these methods are used for

The example for the first approach for smaller settlements is given in the follow‐up. The scheme of a wastewater treatment plant is given in **Figure 5** with object marks. Each of the objects is

The shown scheme is applicable for settlements without industrial and toxic wastewater. The scheme incorporates wastewater treatment and sludge stabilization so that they can be disponated into the natural environment without negative ecological effects behind the

Depending on the ability and concert of the operator the removal of suspended solids is from 80% to above 95%, and this applies for putrescible matter too. The reduction of bacteria is above

This technological scheme is more favorable in the level of efficiency and the produced biomass if the climate conditions are warmer and insolation is more intensive. In areas of temperate continental climate the starting hypothesis is the disposal of solar energy during the whole year or geothermal energy, if continued work of the system with low cost investment and

The technological scheme and all unit operations are checked on the wastewater treatment plant in Sokobanja, Serbia which served as a pilot treatment plant with the process scale of 1:1, under realistic conditions. In the follow‐up, description and instructions for some objects are

Inlet structure (**Figure 6**) serves for removal of large suspended matter and measurement of flow of wastewater. It is made of a channel with a grid which continues to the Parshall flume. The space between iron flat bars is 2–5 cm. The slope is 1:2 to 1:3 for easier cleaning. For smaller

For application of macrobiological unit operations, two tendencies are present:

polishing of effluent quality with nutrient removal (tertial treatment).

99% so the water can be used for irrigation in semiarid areas without danger.

described in more detail with needed design information.

treatment plant. This is ecologically clean technology.

maintenance is desired during the whole year.

given.

**5.1. IS – inlet structure**

entrance to the exit of the treatment plant.

26 Biological Wastewater Treatment and Resource Recovery

engineering regulations is also required.

The primary settlement tank is based on the system of a highly efficient tubular settlement tank (**Figure 6**) which includes a separator of oil and grease into compact construction.

Domestic wastewater treated on this type of settlement tank with a process load not greater than 0.6 l/s po m2 (horizontal area of settlement tank) is of such quality that without further treatment it can go on macrobiological units. Water is kept in the settlement tank shortly, 15– 25 minutes and there is no danger of transit into septic state which is of great importance for the effluent quality. The sludge from the settlement tank is pumped into the sludge digester (SD) for further treatment.

Based on the known, defined hydraulic load, the calculation is completed by standard procedure for primary, mechanical wastewater treatment.

In the case of different primary treatment or no treatment, the quality of wastewater should be brought to an acceptable one for macrobiological living stations which will be used in further treatment.

### **5.3. BH – basin for sanitary hydrophytocultures**

In this basin (**Figure 7**), the dissolved and colloid matter from the wastewater is transformed into biomass of floating macrophytes under the influence of solar energy. Basin depth of 0.4– 0.6 m with a protective bank or edge of 0.2 m above the water level is recommended. The insertion of young macrophytes can be done with monocultures or polycultures depending on whether there is previous experience with the wastewater being treated. In highly polluted wastewater, *Eichhornia crassipes* is the most active species; in medium‐polluted wastewaters, *Pistia stratiotes* should be given advantage and in the least polluted wastewaters, *Salvinia* is most appropriate. The quality of the wastewater, the choice of macrobiological living station and process loads define the dimensions of the basin. For domestic wastewater and daily specific consumption of 250 l/person from 3 to 5 m2 /PE of area under the hydrophytoculture *Eichhornia crassipes* is needed, while *Pistia stratiotes* demands double of that value.

**Figure 7.** Basin for sanitary hydrophytocultures and basin for sanitary aquacultures on the WWTP Sokobanja.

The growth dynamic of green biomass of floating macrophytes and area coverage of the basin in green biomass is in function of plant quality, insolation and temperature. For temperate climate conditions, based on the research done in the wastewater treatment plant in Sokobanja, the basin area coverage for the *Pistia stratiotes* biomass is 2–25 kg/m2 and for *Eichhornia crassipes* biomass it is 5–35 kg/m2 [3, 12]. The growth dynamic of green biomass of larger floating macrophytes is well presented by the exponential equation B = Bo ekt in which B is the probable green biomass after a certain time t in kg, Bo is initial green biomass in kg, k is the rate of growth in 1 day and t is time in days [3, 12]. The rates of growth for *Eichhornia crassipes* are 0.130 for 30°C, 0.052 for 20°C and 0.015 for 15°C and for *Pistia stratiotes* 0.061 for 30°C, 0.026 for 20°C and 0.010 for 15°C [3, 12].

For basin coverage with green biomass of 20 kg/m2 and average temperature of 20°C, daily wet biomass growth about 5 t/ha for *Pistia stratiotes* and about 10 t/ha for *Eichhornia crassipes*. Biomass extraction in smaller treatment plants is manual. The biomass dries on a bank if fresh, green biomass is not used. Mechanized extraction of biomass is possible and economically justified in bigger treatment plants.

The work of this part of the treatment plant is connected to temperature and insolation conditions and under natural climate conditions in areas of temperate continental climate is possible in the period between May and October. In the case of continued process during the year greenhouses and introduction of additional thermal and solar energy are needed.

#### **5.4. BA – basin for sanitary aquacultures**

0.6 m with a protective bank or edge of 0.2 m above the water level is recommended. The insertion of young macrophytes can be done with monocultures or polycultures depending on whether there is previous experience with the wastewater being treated. In highly polluted wastewater, *Eichhornia crassipes* is the most active species; in medium‐polluted wastewaters, *Pistia stratiotes* should be given advantage and in the least polluted wastewaters, *Salvinia* is most appropriate. The quality of the wastewater, the choice of macrobiological living station and process loads define the dimensions of the basin. For domestic wastewater and daily

*Eichhornia crassipes* is needed, while *Pistia stratiotes* demands double of that value.

**Figure 7.** Basin for sanitary hydrophytocultures and basin for sanitary aquacultures on the WWTP Sokobanja.

The growth dynamic of green biomass of floating macrophytes and area coverage of the basin in green biomass is in function of plant quality, insolation and temperature. For temperate climate conditions, based on the research done in the wastewater treatment plant in Sokobanja, the basin area coverage for the *Pistia stratiotes* biomass is 2–25 kg/m2 and for *Eichhornia crassipes* biomass it is 5–35 kg/m2 [3, 12]. The growth dynamic of green biomass of larger floating macrophytes is well presented by the exponential equation B = Bo ekt in which B is the probable green biomass after a certain time t in kg, Bo is initial green biomass in kg, k is the rate of growth in 1 day and t is time in days [3, 12]. The rates of growth for *Eichhornia crassipes* are 0.130 for 30°C, 0.052 for 20°C and 0.015 for 15°C and for *Pistia stratiotes* 0.061 for 30°C, 0.026

wet biomass growth about 5 t/ha for *Pistia stratiotes* and about 10 t/ha for *Eichhornia crassipes*. Biomass extraction in smaller treatment plants is manual. The biomass dries on a bank if fresh, green biomass is not used. Mechanized extraction of biomass is possible and economically

The work of this part of the treatment plant is connected to temperature and insolation conditions and under natural climate conditions in areas of temperate continental climate is

/PE of area under the hydrophytoculture

and average temperature of 20°C, daily

specific consumption of 250 l/person from 3 to 5 m2

28 Biological Wastewater Treatment and Resource Recovery

for 20°C and 0.010 for 15°C [3, 12].

justified in bigger treatment plants.

For basin coverage with green biomass of 20 kg/m2

Under the effect of solar energy, through primary production the process of nutrient removal and transformation into the biomass of fish is completed. Phytoplankton and zooplankton which have used the nutrients are transformed through the food chain of herbivore and carnivore fish into a high value protein.

Average basin depth of 0.3–0.7 m is recommended (**Figure 7**). Young fish are inserted into the basin in spring, most often with polycultures of herbivore fish with the addition of carp. Depending on the input water quality, dimensioning of the basin and aquapolyculture composition are completed.

If primary water treatment is done through tubular settlement tanks by the usual specific consumption, hydraulic load of Q ≤50 m3 /ha for a day is recommended. Stocking is done with 200–400 kg/ha by polyculture (silver carp 65–50%, bighead carp 22–30%, grass carp 8–10% and river and ponds common carp 5–10%).

Production of biomass for a season depends on the success of plant management, and it ranges from 1.200 to 2.000 kg/ha for a season of 200 days.

#### **5.5. SD – sludge digester**

Sludge digestion (**Figure 8**) is desirable for hygienic and esthetic reasons, although it is not necessary if macrobiological treatment of the sludge is completed.

**Figure 8.** Sludge digester and vermiculture field on the WWTP Sokobanja.

The amount of sludge and digester dimensions is calculated by the process for classical problem solutions. The calculation of sludge pumps and pipes, with notice that the pipes should not be under Ø 200 mm because of sludge flow resistance, is done by standard procedure. The same applies for the use of biogas.

If hydrophytocultures are used as an emergent for biogas production, the volume of the digester is to be increased by 20%. Construction of a lateral opening with a nonreturn flap for insertion of the biomass into the digester is mandatory, along with a spiral access ramp or a lift.

#### **5.6. VFs – vermiculture fields**

On the vermiculture field (**Figure 8**) sludge treatment into highly valued hummus is complet‐ ed, using *Eisenia fetida* or *Lumbricus rubellus* living stations.

If fresh sludge was put on vermiculture fields, no matter the addition of wood chips or cut paper for moisture reduction, odor and insects may appear.

The excess of earthworms is returned to the basin for sanitary aquacultures and can be used as food for the fish with a goal of increasing their growth.

Turning plant biomass into hummus if by far more rational than gasification and it gives a valuable commercial product, it especially increases the quality of the total produced material.

Macrobiological unit operations are especially important for a greater number of smaller agglomerations, in which rational expansion of wastewater treatment systems is possible so they give an effluent of high quality. The use of macrobiological unit operations, along with other classic operations in the technological scheme, allows not only a cheap but a technically simple and safe solution to nutrient (nitrogen, phosphorus) removal and BOD5 reduction from wastewater of settlements without industrial wastewater.

## **6. Usage of the resource and energy potential of waste waters**

The macrobiological living stations use from the waste water the nutrients and other elements which are a part of the biomass for their growth and development. The produced biomass has a practical value, so the nutrients and other matter in the waste water are not only harmful matter to be removed from the waste water, but are also the resource for production of the biomass. Regarding that the macrobiological living stations should belong to the fast‐growing species with the short reproductive cycle, the quantity of biomass produced in the waste water treatment process, are huge, as a rule.

Based on previous research some possibilities for the use of biomass were noticed, but this is an area which is yet to be thoroughly researched.

Floating macrophytes can be widely used as biomass [3], especially when water hyacinth (*Eichornia crassipes*) and water lettuce (*Pistia stratiotes*) are in question [13–15].

Water hyacinth and water lettuce mixed with sludge are great material for hummus produc‐ tion, especially for winter cover and thermal protection of the vermiculture.

Water hyacinth combined and composted with manure gives better quality material for vermiculture nutrition with acceleration of the population dynamic because the root system is the ideal habitat for laying cocoons and reproduction.

If hydrophytocultures are used as an emergent for biogas production, the volume of the digester is to be increased by 20%. Construction of a lateral opening with a nonreturn flap for insertion of the biomass into the digester is mandatory, along with a spiral access ramp or a

On the vermiculture field (**Figure 8**) sludge treatment into highly valued hummus is complet‐

If fresh sludge was put on vermiculture fields, no matter the addition of wood chips or cut

The excess of earthworms is returned to the basin for sanitary aquacultures and can be used

Turning plant biomass into hummus if by far more rational than gasification and it gives a valuable commercial product, it especially increases the quality of the total produced material.

Macrobiological unit operations are especially important for a greater number of smaller agglomerations, in which rational expansion of wastewater treatment systems is possible so they give an effluent of high quality. The use of macrobiological unit operations, along with other classic operations in the technological scheme, allows not only a cheap but a technically simple and safe solution to nutrient (nitrogen, phosphorus) removal and BOD5 reduction from

The macrobiological living stations use from the waste water the nutrients and other elements which are a part of the biomass for their growth and development. The produced biomass has a practical value, so the nutrients and other matter in the waste water are not only harmful matter to be removed from the waste water, but are also the resource for production of the biomass. Regarding that the macrobiological living stations should belong to the fast‐growing species with the short reproductive cycle, the quantity of biomass produced in the waste water

Based on previous research some possibilities for the use of biomass were noticed, but this is

Floating macrophytes can be widely used as biomass [3], especially when water hyacinth

Water hyacinth and water lettuce mixed with sludge are great material for hummus produc‐

(*Eichornia crassipes*) and water lettuce (*Pistia stratiotes*) are in question [13–15].

tion, especially for winter cover and thermal protection of the vermiculture.

lift.

**5.6. VFs – vermiculture fields**

30 Biological Wastewater Treatment and Resource Recovery

ed, using *Eisenia fetida* or *Lumbricus rubellus* living stations.

paper for moisture reduction, odor and insects may appear.

as food for the fish with a goal of increasing their growth.

wastewater of settlements without industrial wastewater.

treatment process, are huge, as a rule.

an area which is yet to be thoroughly researched.

**6. Usage of the resource and energy potential of waste waters**

Water hyacinth is ideal for nutrition of nutria (coypu) and they would rather be fed the hyacinth than beet leaves. Detailed information about the possibilities of application of floating macrophytes in livestock keeping, including its possible application in silage, is not available.

Fish as biomass can be used as food in human nutrition, especially if higher quality species are in question: grass carp (*Ctenopharyngodon idella*), silver carp (*Hypophthalmichthys molitrix*), bighead carp (*Aristichthys nobilis*), walking catfish (*Clarias batrachus*) and common carp (*Cyprinus carpio*). The possible danger from bioconcentration of heavy metals and pesticide, as well as quarantine in connection with epidemiological control of food quality should be mentioned.

Mussel meat, which is easily removed with hot water, is rich in proteins and is eaten by poultry and pigs in dried form. The shell of some mussels, e.g., painter's mussel, can be used as material for nacre products.

Worms, both species: red Californian earthworms (*Eisenia fetida*) and red earthworms (*Lumbricus rubellus*) which are interesting as macrobiological living stations in the sludge and manure treatment technology can be widely used as a biomass.

Earthworms can directly be used as food for poultry, pigs and pets (birds, tortoises, iguanas, snakes, fish in aquariums). They can also be used for production of protein flour, a high quality protein component in dry condition which is added to fish flour or feed. Commercially, the use of earthworms for the nutrition of aquacultures of fish is very favorable, especially for cultivation and fattening of fastgrowing Clarias catfish and cultivation of trout.

Even more profitable is placement of earthworms as bait (in fishing or for attracting wild birds) through specialized stores for hunter and anglers.

Earthworms feed on detritus, decomposing organic matter, and as secretion finely crushed material appears, relatively stable vermicompost, namely humus. Humus has great value as a natural fertilizer because it improves the structure of the ground and reduces/eliminates the need for chemical fertilizers. It is used in flower and vegetable cultivation, for nursery gardens, orchards and lawns, for topsoiling of surfaces or as a component in soil devastated because of the use of chemical fertilizers. If there was no pesticide, heavy metals or toxic substances in the starting material, the humus can be used in production of healthy food because, aside from having a positive effect on the ground, it has a positive effect on various plants and crop plants.

Regarding that, when larger floating macrophytes are in question, the amounts of biomass produced daily in the process of wastewater treatment are quite large, the possibility of using said biomass as raw material for production of biogas is of great importance. The produced biogas can further be used for combined production of electrical and heat energy.

Combined production of electrical and heat energy [combined heat and power (CHP)], also called cogeneration, is the production of electrical power out of the natural gas, biogas and waste matter disposal site gas, with the simultaneous usage of waste heat which is otherwise lost in the industrial process. Modern cogeneration systems today reach efficiency above 90%, that is why cogeneration presents the most efficient and economically most justified way of reducing high energy costs in industrial plants and municipal objects.

In wastewater treatment plants with an anaerobic reactor (digester) for sludge stabilization biogas occurs as a mixture of combustible and noncombustible gases with the average composition of (in cubic %): methane 55–75%, carbon‐dioxide 25–45%, other gases like hydrogen, oxygen, carbon‐monoxide, nitrogen, hydrogen‐sulfide, ammonia and water vapor [16]. The efficiency of biogas production is provided by maintaining temperature, pH value, by mixing and removal of oxygen and toxic matter.

Production of biogas can be assessed based on the following practical and experimental information [16]:


The heat power of biogas depends on the methane content and for the average content of 65% methane it is equal to 6.4 kWh/m3 [16]. That is how it is possible to produce 2.5 kWh of electrical and 3.3 kWh of heat energy from 1 m3 of biogas with the reduction of CO2 emission above 50% in a practical operation on a cogeneration plant with gas motors (**Figure 9**) [16].

**Figure 9.** Usage of biogas at the facilities for waste water treatment (based on Ref. [17]). 1 – biomass; 2 – sludge thicken‐ ing; 3 – anaerobic digester; 4 – gas torch; 5 – biogas; 6 – gas tank; 7 – gas engine; 8 – heat exchanger; 9 – exhaust; 10 – heat energy; 11 – electrical energy; 12 – agricultural fertilizers.

In wastewater treatment plants with a basin for sanitary phytocultures, the produced biomass of floating macrophytes, either processed through a digester for biogas production increase, used in cogeneration plants with a gas motor, or directly burned in cogeneration plants with an indirect gas turbine process, the amounts of produced electrical or heat energy can be multiply increased relative to plants with classic technologies, which of course directly depends from the available basin area for sanitary hydrophytocultures and the daily growth of the biomass of floating macrophytes.

waste matter disposal site gas, with the simultaneous usage of waste heat which is otherwise lost in the industrial process. Modern cogeneration systems today reach efficiency above 90%, that is why cogeneration presents the most efficient and economically most justified way of

In wastewater treatment plants with an anaerobic reactor (digester) for sludge stabilization biogas occurs as a mixture of combustible and noncombustible gases with the average composition of (in cubic %): methane 55–75%, carbon‐dioxide 25–45%, other gases like hydrogen, oxygen, carbon‐monoxide, nitrogen, hydrogen‐sulfide, ammonia and water vapor [16]. The efficiency of biogas production is provided by maintaining temperature, pH value,

Production of biogas can be assessed based on the following practical and experimental

**•** on municipal wastewater treatment plants the average production of biogas 25 l/PE per day;

**•** with industrial wastewater (sugar refineries, molasses processing, potato processing, fruit juice production, dairy farms, breweries, paper and cellulose) the average methane pro‐

The heat power of biogas depends on the methane content and for the average content of 65% methane it is equal to 6.4 kWh/m3 [16]. That is how it is possible to produce 2.5 kWh of electrical

**Figure 9.** Usage of biogas at the facilities for waste water treatment (based on Ref. [17]). 1 – biomass; 2 – sludge thicken‐ ing; 3 – anaerobic digester; 4 – gas torch; 5 – biogas; 6 – gas tank; 7 – gas engine; 8 – heat exchanger; 9 – exhaust; 10 –

in a practical operation on a cogeneration plant with gas motors (**Figure 9**) [16].

/kg CSB with the methane fraction in biogas being 60–80%.

of biogas with the reduction of CO2 emission above 50%

reducing high energy costs in industrial plants and municipal objects.

by mixing and removal of oxygen and toxic matter.

32 Biological Wastewater Treatment and Resource Recovery

information [16]:

duction is 0.20–0.30 m3

and 3.3 kWh of heat energy from 1 m3

heat energy; 11 – electrical energy; 12 – agricultural fertilizers.

The dry mass of *Pistia stratiotes* is 4.9% and for *Eichhornia crassipes* 4.6% from the green mass for the leaf part of the plant (the variations of the root mass are great). Based on the literature information each kilogram of *Eichhornia crassipes* dry mass gives 370 l of biogas, whose heating value is around 6.1 kWh/m3 [13]. For *Pistia stratiotes,* keep in mind that the structure of biomass is similar to the previously mentioned plant.

In **Table 1**, the values of biomass growth, biogas amount and electrical and heat energy, which can be produced from the biogas are shown for *Pistia stratiotes* and *Eichhornia crassipes,* calculated based on the previously stated experimental data. All values are given in ha of basin area under sanitary hydrophytocultures by day.


**Table 1.** Biomass growth, biogas amount and energies which can be produced for larger floating macrophytes [18].

As it may be concluded on the basis of the displayed values, at the facilities for waste water treatment with the basin for sanitary hydrophytocultures, significant quantities of electric and thermal energy can be obtained through the cogeneration.

Part of the produced electrical energy can be used for the plants' own needs, and extras can be forwarded into the ED network, while the produced heat energy can be used for maintaining the temperature in the digester, to ensure the efficiency of biogas production. The heat energy can also be used for providing favorable conditions (air temperature from min. 20°C and area coverage from min. 20 kg/m2 ) in greenhouses for application of these technologies in our climate conditions during the whole year, which in turn provides constant growth of floating macrophyte biomass and annuls the seasonal character of macrobiological methods.

Considering the global climate changes, the Kyoto protocol predicts the possibility that developed countries invest in modernization of industrial and energy power plants and reduction of carbon dioxide emission and other gases which cause the greenhouse effect on the territories of undeveloped and developing countries. As introduction of cogeneration in wastewater treatment plants by macrobiological methods and the usage of surpluses of electrical and heat energy in the energy system is in accordance with the Kyoto protocol, it would allow receiving of exceptionally favorable credits and investments for energy and ecology sector, and is as such of great importance, especially for developing countries, which are yet to solve the problem of settlement and industrial wastewater and the building of plants for their treatment.

Considering that the yields, which amount to a few dozen tons by ha daily with larger floating macrophytes in favorable insolation and temperature conditions, are of fantastic size , research of their value of use in animal husbandry, and even more in energetics is extremely significant.

## **7. Conclusion**

Based on previously achieved results, it is evident that macrobiological unit operations will in the future find their place in the technology of wastewater treatment for multiple reasons:


**•** the final product is biomass which has value of use as food or feed, or as an emergent in biogas production, which in turn affects the reduction, even complete annulment of wastewater treatment costs.

The energy crisis, which is deepening day by day, high prices of energy, materials and the workforce, the demand for low investment and operation costs, and the sharper requests set in regards of discharge treated wastewater into recipients make the application of macrobio‐ logical methods in wastewater treatment around the world even today, and especially in the near future come to the fore and intensify the research in this extremely important area.

## **Author details**

As it may be concluded on the basis of the displayed values, at the facilities for waste water treatment with the basin for sanitary hydrophytocultures, significant quantities of electric and

Part of the produced electrical energy can be used for the plants' own needs, and extras can be forwarded into the ED network, while the produced heat energy can be used for maintaining the temperature in the digester, to ensure the efficiency of biogas production. The heat energy can also be used for providing favorable conditions (air temperature from min. 20°C and area

climate conditions during the whole year, which in turn provides constant growth of floating

Considering the global climate changes, the Kyoto protocol predicts the possibility that developed countries invest in modernization of industrial and energy power plants and reduction of carbon dioxide emission and other gases which cause the greenhouse effect on the territories of undeveloped and developing countries. As introduction of cogeneration in wastewater treatment plants by macrobiological methods and the usage of surpluses of electrical and heat energy in the energy system is in accordance with the Kyoto protocol, it would allow receiving of exceptionally favorable credits and investments for energy and ecology sector, and is as such of great importance, especially for developing countries, which are yet to solve the problem of settlement and industrial wastewater and the building of plants

Considering that the yields, which amount to a few dozen tons by ha daily with larger floating macrophytes in favorable insolation and temperature conditions, are of fantastic size , research of their value of use in animal husbandry, and even more in energetics is extremely significant.

Based on previously achieved results, it is evident that macrobiological unit operations will in the future find their place in the technology of wastewater treatment for multiple reasons:

**•** instead of destruction of material, namely the stopping of natural processes we are going towards the philosophy of synthesis of organic matter into higher levels of biomass;

**•** the processes of synthesis use natural energy sources (sun, heat) and on that basis they

**•** the objects are relatively simple civil engineering objects made of land followed by a minimal

**•** there is no biological sludge and no parts of the object which represent the secondary settlement tank and sludge line, which significantly affects the relief of wastewater treat‐

present ecologically highly "clean" technologies;

equipment fond, which greatly impacts low investment costs;

macrophyte biomass and annuls the seasonal character of macrobiological methods.

) in greenhouses for application of these technologies in our

thermal energy can be obtained through the cogeneration.

coverage from min. 20 kg/m2

34 Biological Wastewater Treatment and Resource Recovery

for their treatment.

**7. Conclusion**

ment costs;

Dragan Milićević\* , Slaviša Trajković and Milan Gocić

\*Address all correspondence to: drgara@gaf.ni.ac.rs

The Faculty of Civil Engineering and Architecture of the University of Niš, Aleksandra Medvedeva, Niš, Serbia

## **References**


#### **Microbe-Based Strategy for Plant Nutrient Management Microbe-Based Strategy for Plant Nutrient Management**

Shaon Ray Chaudhuri, Madhusmita Mishra, Sonakshi De, Biswajit Samal, Amrita Saha, Srimoyee Banerjee, Abhinandan Chakraborty, Antara Chakraborty, Sonali Pardhiya, Deepak Gola, Joyeeta Chakraborty, Sourav Ghosh, Kamlesh Jangid, Indranil Mukherjee, Mathummal Sudarshan, Rajib Nath and Ashoke Ranjan Thakur Shaon Ray Chaudhuri, Madhusmita Mishra, Sonakshi De, Biswajit Samal, Amrita Saha, Srimoyee Banerjee, Abhinandan Chakraborty, Antara Chakraborty, Sonali Pardhiya, Deepak Gola, Joyeeta Chakraborty, Sourav Ghosh, Kamlesh Jangid, Indranil Mukherjee, Mathummal Sudarshan, Rajib Nath and Ashoke Ranjan Thakur Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

http://dx.doi.org/10.5772/67307

#### **Abstract**

[8] File: Eisenia foetida R.H. (2).JPG [Internet]. Available from: https://commons.wikime‐

[9] Discover Life. Lumbricus rubellus Hoffmeister, 1843. [Internet]. Available from: http://

[10] The Composter Newsletter. Wriggle your way into my heart! [Internet]. Available from: http://archive.constantcontact.com/fs057/1101915097096/archive/1102821409563.html

[11] Milićević D., Milenković S., Nikolić V. Application of macrobiological methods in settlements wastewater treatment, Water and Sanitation Equipment; 2006; 1:31–37 (in

[12] Milićević D. Modeling system to use nutrients from the water using macrophytes [Unpublished master thesis]. Civil Engineering faculty of Nis, Serbia; 1994. (in Serbian)

[13] Center for New Crops & Plant Products, at Purdue University. Eichornia crassipes (Mart.) Solms [Internet]. Available from: https://hort.purdue.edu/newcrop/duke\_ener‐

[14] Lareo L., Bressani R. Possible utilization of the water hyacinth in nutrition and industry [Internet]. Available from: http://archive.unu.edu/unupress/food/8F044e/8F044E0c.

[15] Sanmuga Priya E., Senthamil Selvan P. Water hyacinth (Eichhornia crassipes) – An efficient and economic adsorbent for textile effluent treatment – A review. Arabian Journal of Chemistry, Available online 13 March 2014 [Internet]. Available from: http://

[16] MPD TRADE, Belgrade. Production and characteristics of biogas [Internet]. Available

[17] MPD TRADE, Belgrade. The emergence and use of biogas [Internet]. Available from:

[18] Milenkovic S., Milicevic D., Nikolic V. Application of macro biologic methods in wastewater treatment and usage of its energy and resource potential. International Conference on Water Observation And Information System For Decision Support

www.sciencedirect.com/science/article/pii/S1878535214000562

BALWOIS, Ohrid, Republic of Macedonia, 23–26 May 20066

from: http://www.kogeneracija.rs/biogas.html

http://www.biogas.rs/nastanak.html

dia.org/wiki/File:Eisenia\_foetida\_R.H.\_(2).JPG

36 Biological Wastewater Treatment and Resource Recovery

Serbian)

htm

gy/ Eichornia\_crassipes.html

www.discoverlife.org/mp/20q?search=Lumbricus+rubellus

The rapid industrialization and urbanization of developing countries such as India have encroached on cultivable lands to meet the demands of an ever-increasing population. The altered land use patterns with increased fertilizer use has increased crop yields with leaching of major portion of the applied nutrients from the soil. Nitrates and phosphates are the agricultural pollutants that are discharged into aquifers due to anthropogenic reasons causing severe environmental and health problems. Production of these nutrients requires energy and finite resources (rock phosphate, which has gradually depleting reserves). An alternative management strategy would be to sequester excess nutrients within a biomass that is reused for agriculture. Two discrete enriched microbial consortia with the potential of simultaneous nitrate and phosphate sequestration upon application as biofertilizer restricted them within the plant root zone, ensuring prevention of eutrophication through leaching while making it available for uptake by plants. The nutrient accumulated biomass enhanced the crop yield by 21.88% during mung bean cultivation with maintained elemental content and other nutritional qualities. The major drawback of conventional biofertilizer application (slow release and action) could be overcome using this formulation leading to environmental protection, crop yield enhancement and soil fertility maintenance post-cultivation.

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 © 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.

Attribution License (http://creativecommons.org/licenses/by/3.0), which permits unrestricted use, distribution,

**Keywords:** nitrate accumulation, plant growth promotion, phosphate accumulation, phosphatase activity, microbial consortium

#### **1. Introduction**

In developing countries like India, rapid industrialization and urbanization have led to encroachment of cultivable lands. The agricultural practices are being gradually modified to increase the food production so as to meet the need of the ever-increasing population. The significant increase in the use of inorganic and organic fertilizers as well as alterations in the land use pattern has led to high yield of crops. But the major disadvantage that emerged out of such practices is the gradual leaching of nutrients and harmful chemicals in the soil and water. Nitrate is one such common agricultural pollutant discharged into the aquifers. Other potential sources of nitrate are the geological processes like eruptions, flood and land silting, irregular rainfall and stream flow patterns, natural process of plant decay and organic residues, anthropogenic sources of land practices, traditional agricultural practices like dry farming, marginal irrigation, large scale flood plain farming and application of fertilizers, leaching from paddy and tea cultivation, sewage infiltration, reuse of agricultural land for human settlement, industrial chemical spills and landfill leachates [1–10]. Nitrate pollution has thus emerged as a global problem and happens to be the second most dangerous pollutant after the pesticides [11, 12]. In marine environment, it induces plankton bloom destroying the native flora and fauna of the region [13]. In humans, it causes condition known as methemoglobinemia (blue baby syndrome) in infants and disorders of central nervous system, cardiovascular system as well as gastrointestinal system while posing to be carcinogenic [14].

The permissible nitrate level in ground water (10 mg/l for NO3 –N and 45 mg/l for NO3 ) has been demarcated by "United States Environmental Protection Agency (EPA)." Some of the conventional methods for nitrate removal from water include distillation, reverse osmosis and ion exchange. These processes are quite complex as well as expensive which limits their application during scale up of processes. Bioremediation appears as a desired alternative [15–17], but the major limitation for its application is the longer retention time as compared to the physicochemical processes. Lately the membrane technology of denitrification has been blended with biological immobilization techniques to achieve efficient operation. This combination helps minimize the associated problem while making the process economically viable [18]. Electro bioremediation where effect of electric field is observed on pollutant reduction has also been studied [19–21]. Nitrate reduction by biological means has been reported to be carried out in fluidized expanded bed bioreactors [22], submerged membrane bioreactor [23], continuous flow bioreactors [24] as well as packed bed reactor [25] with PVS tubes [26], alginate [27], K- Carrageenan [28] and microbial cellulose [29] as immobilization matrices. It could either be through assimilatory or dissimilatory pathway. An alternative pathway of nitrate removal is through nitrate accumulation as evident in Isolates of **genus** Beggiatoa, Thiomargarita and Thioploca, as well as one species of Bacillus [30].

Phosphate is another essential plant growth nutrient which is lost in wastewater from domestic, industrial (dairy as well as detergent) and agricultural sectors [31]. It also causes eutrophication upon seepage into the surface and ground water bodies. Phosphate is derived from rock phosphate whose reserves are limited [32]. Thus, it is desirable to sequester the phosphate from the wastewater for reuse instead of indiscriminate use of rock phosphate [32]. Phosphate accumulation is already reported in bacteria, but nitrate accumulation in bacteria is relatively rare. It is in the genus *Beggiatoa*, *Thioploca* and *Thiomargarita* that nitrate accumulation is observed in intracellular vacuoles [33–35]. Only recently nitrate accumulation from wastewater has been reported in the genus *Bacillus* [36]. Since nitrate and phosphate are both essentials for agriculture, but only a small fraction (12–30%) [7] of the applied nutrients is utilized by the plant, thus it becomes essential to trap these nutrients for reuse as well as environmental protection.

In order to address this upcoming environmental challenge, an alternative plant nutrient management strategy was developed with the following approach: (i) isolation and characterization of microbial consortium with ability to simultaneously accumulate nitrate and phosphate; (ii) utilize these microbes to prevent nutrient leaching from soil; and (iii) utilize these microbes with intracellular accumulated nutrients as biofertilizer.

## **2. Consortia development and characterization**

**Keywords:** nitrate accumulation, plant growth promotion, phosphate accumulation,

In developing countries like India, rapid industrialization and urbanization have led to encroachment of cultivable lands. The agricultural practices are being gradually modified to increase the food production so as to meet the need of the ever-increasing population. The significant increase in the use of inorganic and organic fertilizers as well as alterations in the land use pattern has led to high yield of crops. But the major disadvantage that emerged out of such practices is the gradual leaching of nutrients and harmful chemicals in the soil and water. Nitrate is one such common agricultural pollutant discharged into the aquifers. Other potential sources of nitrate are the geological processes like eruptions, flood and land silting, irregular rainfall and stream flow patterns, natural process of plant decay and organic residues, anthropogenic sources of land practices, traditional agricultural practices like dry farming, marginal irrigation, large scale flood plain farming and application of fertilizers, leaching from paddy and tea cultivation, sewage infiltration, reuse of agricultural land for human settlement, industrial chemical spills and landfill leachates [1–10]. Nitrate pollution has thus emerged as a global problem and happens to be the second most dangerous pollutant after the pesticides [11, 12]. In marine environment, it induces plankton bloom destroying the native flora and fauna of the region [13]. In humans, it causes condition known as methemoglobinemia (blue baby syndrome) in infants and disorders of central nervous system, cardiovascular system as well as gastrointestinal system while posing to

been demarcated by "United States Environmental Protection Agency (EPA)." Some of the conventional methods for nitrate removal from water include distillation, reverse osmosis and ion exchange. These processes are quite complex as well as expensive which limits their application during scale up of processes. Bioremediation appears as a desired alternative [15–17], but the major limitation for its application is the longer retention time as compared to the physicochemical processes. Lately the membrane technology of denitrification has been blended with biological immobilization techniques to achieve efficient operation. This combination helps minimize the associated problem while making the process economically viable [18]. Electro bioremediation where effect of electric field is observed on pollutant reduction has also been studied [19–21]. Nitrate reduction by biological means has been reported to be carried out in fluidized expanded bed bioreactors [22], submerged membrane bioreactor [23], continuous flow bioreactors [24] as well as packed bed reactor [25] with PVS tubes [26], alginate [27], K- Carrageenan [28] and microbial cellulose [29] as

–N and 45 mg/l for NO3

) has

phosphatase activity, microbial consortium

38 Biological Wastewater Treatment and Resource Recovery

The permissible nitrate level in ground water (10 mg/l for NO3

**1. Introduction**

be carcinogenic [14].

Nitrate broth (Himedia M439) was used as the medium of choice for isolation of nitrate reducing microbial consortium. Two types of inoculum were used under both aerobic and anaerobic condition (in an atmosphere of carbon dioxide and nitrogen) at 37°C. The first type was the soil from East Calcutta Wetland (ECW) (22°27′ N, 88°27′E) which is known as the world's largest waste dumping ground and natural waste recycling center [37]. The reason for selecting soil from East Calcutta Wetland as the inoculum was that it was expected to harbor microbes with rich diversity as well as bioremedial ability. Since cultivation is the ongoing practice in this area, efficient strains with potential for promoting plant growth are expected to inhabit this area. The other inoculum was the biomass from a low-level radioactive waste treating microbial biofilm bioreactor removing mainly nitrate [38, 39]. This was expected to contain efficient nitrate reducers/accumulators due to its constant exposure to nitrate. Nitrate removal from the medium by the bacteria was set as the primary criteria for the selection of consortium. After 48 h of incubation, the nitrate concentration [40, 41] in the cell-free medium was checked. Of the four different combinations tested, two consortia were found to be efficient: anaerobic consortium from ECW (NB1) and aerobic consortium from bioreactor biomass (BN7). They demonstrated 96 and 97.44% nitrate removal in 12 and 4 h by NB1 and BN7, respectively [39]. Another interesting feature of BN7 was its simultaneous accumulation of nitrate and phosphate from medium.

Both the cultures were also tested for phosphate removing ability as per standard procedure [30, 32] and demonstrated 23.88 and 48.2% removal with 565 and 1.14mg per gram wet weight of polyphosphate in NB1 and BN7, respectively. NB1 reduced 75–90% nitrate within a pH range of 5–12 with the maximum at pH 10 while that of BN7 was a range of 6–11 [39]. The optimum temperature range for NB1 was 30–40°C and that for BN7 was 25–37°C [39].

The effect of metals [viz., zinc (ZnSO4 ), cobalt (CoCl·6H2 O), lead {Pb(NO3 )} and copper (CuSO4 ·5H2 O)] on the nitrate reduction efficiency of NB1 and BN7 consortia was checked at two different concentrations, that is, 0.1 and 0.5 mM. It was compared to the reduction in the absence of metal salts (control) in both cases. The experiments were repeated thrice. The aerobic culture exhibiting growth along with nitrate reduction in the presence of different metal salts was checked for metal accumulation within the biomass using energy-dispersive X-ray fluorescence (EDXRF) analysis [39, 40]. While chromium (Cr), strontium (Sr) and cadmium (Cd) salts were inhibitory for the growth of the anaerobic consortium NB1 even at a concentration of 0.1 mM, the consortium showed growth in up to 0.5 mM concentration of copper (Cu), lead (Pb), cobalt (Co) and zinc (Zn). Being an anaerobic consortium, it was better preserved as glycerol stock while retaining its nitrate removal activity up to 24 days rather than stab or lyophilized culture as compared to BN7 [39].

16S rDNA based molecular characterization of both the consortia were done as per prior report [42]. The sequences obtained were subjected to NCBI nucleotide BLAST analysis, and novel sequences were submitted to GenBank. These sequences were then subjected to phylogenetic analysis using neighbor joining method. The rarefaction curves were drawn, and the richness (Shannon diversity index) and evenness (equitability index) of the population were determined as per standard procedure [37, 43, 44]. Mothur analysis was conducted using the data.

At the molecular level, NB1 was composed of novel organisms (GenBank JN626182-JN626198 and JN665074-JN665081) with closest identity in the ratio of 44:37:19 with *Pseudomonas* sp., *E. coli* and uncultured bacterium (**Figure 1a**–**c**) with poor diversity (Shannon diversity index 0.417) of evenly distributed population (equitability index 0.873). *Pseudomonas* sp. might be involved in nitrate removal as well as phosphate accumulation. BN7 on the other hand was composed of *Pseudomonas* sp.:*Azoarcus* sp.:uncultured bacterium: *Bacillus* sp. in the ratio of 20:31:46:3% in terms of 16S rDNA sequence similarity of its clones (GenBank GU644465 to GU644489). Like any enriched consortium in selective medium, BN7 reflected poor diversity (Shannon diversity index 0.39) of evenly distributed microbes (equitability index 0.83). Genus Pseudomonas and Bacillus were involved in phosphate accumulation and nitrate reduction [39].

(BN7). They demonstrated 96 and 97.44% nitrate removal in 12 and 4 h by NB1 and BN7, respectively [39]. Another interesting feature of BN7 was its simultaneous accumulation of

Both the cultures were also tested for phosphate removing ability as per standard procedure [30, 32] and demonstrated 23.88 and 48.2% removal with 565 and 1.14mg per gram wet weight of polyphosphate in NB1 and BN7, respectively. NB1 reduced 75–90% nitrate within a pH range of 5–12 with the maximum at pH 10 while that of BN7 was a range of 6–11 [39]. The optimum temperature range for NB1 was 30–40°C and that for BN7 was

), cobalt (CoCl·6H2

at two different concentrations, that is, 0.1 and 0.5 mM. It was compared to the reduction in the absence of metal salts (control) in both cases. The experiments were repeated thrice. The aerobic culture exhibiting growth along with nitrate reduction in the presence of different metal salts was checked for metal accumulation within the biomass using energy-dispersive X-ray fluorescence (EDXRF) analysis [39, 40]. While chromium (Cr), strontium (Sr) and cadmium (Cd) salts were inhibitory for the growth of the anaerobic consortium NB1 even at a concentration of 0.1 mM, the consortium showed growth in up to 0.5 mM concentration of copper (Cu), lead (Pb), cobalt (Co) and zinc (Zn). Being an anaerobic consortium, it was better preserved as glycerol stock while retaining its nitrate removal activity up to 24 days rather than stab or lyophilized culture as compared to

16S rDNA based molecular characterization of both the consortia were done as per prior report [42]. The sequences obtained were subjected to NCBI nucleotide BLAST analysis, and novel sequences were submitted to GenBank. These sequences were then subjected to phylogenetic analysis using neighbor joining method. The rarefaction curves were drawn, and the richness (Shannon diversity index) and evenness (equitability index) of the population were determined as per standard procedure [37, 43, 44]. Mothur analysis was conducted

At the molecular level, NB1 was composed of novel organisms (GenBank JN626182-JN626198 and JN665074-JN665081) with closest identity in the ratio of 44:37:19 with *Pseudomonas* sp., *E. coli* and uncultured bacterium (**Figure 1a**–**c**) with poor diversity (Shannon diversity index 0.417) of evenly distributed population (equitability index 0.873). *Pseudomonas* sp. might be involved in nitrate removal as well as phosphate accumulation. BN7 on the other hand was composed of *Pseudomonas* sp.:*Azoarcus* sp.:uncultured bacterium: *Bacillus* sp. in the ratio of 20:31:46:3% in terms of 16S rDNA sequence similarity of its clones (GenBank GU644465 to GU644489). Like any enriched consortium in selective medium, BN7 reflected poor diversity (Shannon diversity index 0.39) of evenly distributed microbes (equitability index 0.83). Genus Pseudomonas and Bacillus were involved in phosphate accumulation and nitrate

O)] on the nitrate reduction efficiency of NB1 and BN7 consortia was checked

O), lead {Pb(NO3

)} and copper

nitrate and phosphate from medium.

40 Biological Wastewater Treatment and Resource Recovery

The effect of metals [viz., zinc (ZnSO4

25–37°C [39].

·5H2

(CuSO4

BN7 [39].

using the data.

reduction [39].

**Figure 1.** Phylogenetic trees constructed using neighbor joining method for the clones from the consortium NB1 showing maximum similarity with uncultured bacterium (a), Pseudomonas (b) and *E. coli* (c).

Mothur analysis revealed saturation of screening of the consortia which were different from each other (**Figure 2**; **Tables 1** and **2**).

**Figure 2.** Rarefaction curve drawn for the consortium BN7 and NB1 reflecting saturation of screening for both the consortiums.


**Table 1.** Libshuff comparison showing that both libraries have a very different community structure.



**Table 2.** Diversity indices calculated for both the consortia.

## **3. Soil leaching**

**Figure 2.** Rarefaction curve drawn for the consortium BN7 and NB1 reflecting saturation of screening for both the

**Comparison dCXYScore Significance** BN7-NB1 0.0206 <0.0001 NB1-BN7 0.0121 <0.0001

**Diversity index @ 0.01 BN7 NB1** N 25 25 S 13 7 Simpson (1/D) 18.75 3.03 95% LCI 12.90 1.96 95% HCI 34.32 6.69 Shannon (H) 2.47 1.41 95% LCI 2.22 0.99 95% HCI 2.72 1.82 Hmax 2.84 1.67 Chao 15.00 8.00 95% LCI 13.29 7.09 95% HCI 26.96 17.68 Ace 16.25 10.08 95% LCI 14.49 7.45 95% HCI 20.07 28.24

**Table 1.** Libshuff comparison showing that both libraries have a very different community structure.

consortiums.

42 Biological Wastewater Treatment and Resource Recovery

An experimental tub of dimension 18 cm × 12 cm × 17 cm (l × b × h respectively) (**Figure 3**), with surface area of 216 cm2 and volume 3672 cm3 filled up with 8.095 kg of soil, was set up for studying nitrate leaching in soil. In order to study the leaching process, outlets were made along the breadth of the tub at different heights of 3, 7, 11, 15 and 17 cm from the surface of the soil which facilitated in sample collection which in turn were assessed for the nitrate concentration [37, 38].

**Figure 3.** Schematic representation of the apparatus (soil filled tub) used for soil leaching experiment.

The experiment was carried out in four sets. For the first set (control), leaching of nitrate from soil in the presence of the native soil microbial population was tested. For this, water was poured into the soil filled tub. As the water seeped down, samples were collected from each outlet and analyzed for nitrate concentration [37, 38]. For the second and third set, the soil was inoculated with 100 ml of seed culture of BN7 and NB1, respectively. The system was left for 48 h for the consortium to colonize in the soil. Finally after 48 h, the leaching experiment was repeated as reported above to assess the nitrate released from the soil into the seepage water collected at different heights as a result of the interaction of soil native microbial population with the applied microbial consortia separately. For the fourth set, the combination of BN7 and NB1 in 1:1 ratio was applied and the experiment was repeated as in case of set two and three. The leaching of nitrate with and without external microbial consortium application was analyzed from the above experiments. This study was repeated thrice. In case of control, the soil interaction with the native microbial population as reflected through nitrate leaching was analyzed. In case of BN7 and NB1, these consortia were applied separately and the mixed impact of these consortia with the existing native soil microbial population was studied on the extent of nitrate leaching in water with traversed soil depth. In case of NB1 + BN7, the joint interaction of all the three consortium on nitrate leaching in soil was analyzed. From the results, it was observed that the application of the mixed formulation prevented leaching of nitrate from the soil resulting in decrease in the incidences of eutrophication due to soil nitrate leaching as documented in **Table 3**. It results in substantial reduction in nitrate leaching.


**Table 3.** Tabular representation of the nitrate leaching from soil in the presence of different microbial consortia.

The correlation coefficient values indicate strong correlation between the depth of soil traversed by the applied water and the extent of nitrate leached in the presence of all the four treatments. Moreover, the prevention of leaching was complete at 11 cm of soil depth, indicating immobilization of nitrate in that zone. If this nitrate is made available to plants then this being the root zone for most of the plant, the productivity is expected to rise and the soil fertility is expected to be maintained. Also the phosphate accumulated inside as polyphosphate upon being released could be solubilized by the phosphatase released by the bacteria and made available to the plants. Both these phenomena are expected to strengthen the ability of this consortium (NB1 + BN7) to function as a biofertilizer. The nitrate and phosphate concentration in agricultural runoff could also be reduced by these microbes.

## **4. Plant growth promoting activity**

each outlet and analyzed for nitrate concentration [37, 38]. For the second and third set, the soil was inoculated with 100 ml of seed culture of BN7 and NB1, respectively. The system was left for 48 h for the consortium to colonize in the soil. Finally after 48 h, the leaching experiment was repeated as reported above to assess the nitrate released from the soil into the seepage water collected at different heights as a result of the interaction of soil native microbial population with the applied microbial consortia separately. For the fourth set, the combination of BN7 and NB1 in 1:1 ratio was applied and the experiment was repeated as in case of set two and three. The leaching of nitrate with and without external microbial consortium application was analyzed from the above experiments. This study was repeated thrice. In case of control, the soil interaction with the native microbial population as reflected through nitrate leaching was analyzed. In case of BN7 and NB1, these consortia were applied separately and the mixed impact of these consortia with the existing native soil microbial population was studied on the extent of nitrate leaching in water with traversed soil depth. In case of NB1 + BN7, the joint interaction of all the three consortium on nitrate leaching in soil was analyzed. From the results, it was observed that the application of the mixed formulation prevented leaching of nitrate from the soil resulting in decrease in the incidences of eutrophication due to soil nitrate leaching as documented in **Table 3**. It results in substantial

The correlation coefficient values indicate strong correlation between the depth of soil traversed by the applied water and the extent of nitrate leached in the presence of all the four treatments. Moreover, the prevention of leaching was complete at 11 cm of soil depth, indicating immobilization of nitrate in that zone. If this nitrate is made available to plants then this being the root zone for most of the plant, the productivity is expected to rise and the soil fertility is expected to be maintained. Also the phosphate accumulated inside as polyphosphate upon being released could be solubilized by the phosphatase released by the bacteria and made available to the plants. Both these phenomena are expected to strengthen the ability of this consortium (NB1 + BN7) to function as a biofertilizer. The nitrate and phosphate concen-

**Table 3.** Tabular representation of the nitrate leaching from soil in the presence of different microbial consortia.

– 0.94 0.82 – 0.88 – 0.79 –

tration in agricultural runoff could also be reduced by these microbes.

**Level Concentration of nitrate in seepage water at different levels in ppm**

**concentration (fold change)**

A 3 0 92.34 – 0 – 0 – B 7 4.8 5.4 12.5 0 −100 0 −100 C 11 28.25 255.53 804.53 123.68 337.8 0 −100 D 15 75.1 425.7 466.84 154.82 106.15 4.36 −94.2 E 17 110.65 1160.27 948.59 120.6 8.99 12.83 −88.41

**NB1 Difference in concentration (fold change)**

**BN7 + NB1 Difference in** 

**concentration (fold change)**

**Control BN7 Difference in** 

reduction in nitrate leaching.

44 Biological Wastewater Treatment and Resource Recovery

**Distance from soil surface (cm)**

Correlation coefficient

Production of phytostimulator like ammonia, hydrogen cyanide (as plant protector), indole acetic acid, gibberellic acid (as plant hormones), phosphatase (to solubilize inorganic phosphate) and siderophore was tested for both the consortiums as per standard procedure [45]. NB1 produced 5.2 mg/100 ml and BN7 produced 1.64 mg/100 ml of ammonia with no hydrogen cyanide and siderophore production by either of them. Indole acetic acid (550 μg/ml) was produced by NB1 only. Both NB1 and BN7 produced enzyme phosphates, which were quantified to be 9.12 and 8.7 U/ml, respectively, with a final pH change to 4.11 and 6.3.

Since the consortium (NB1 + BN7) possessed plant growth promoting characters and also prevented leaching from soil, thereby making soil nutrients available to plants, both (NB1 and BN7) were tested for its effect on germination following soil application at the time of sowing, and the data were analyzed as per the standard protocol [45]. The data represent the combined effect of the native soil microbial population with the applied consortium. In order to assess the effect of only the combined consortia (NB1 + BN7) on germination in mung bean, the germination trial was repeated in germination tray using sterile soilrite mix kel006 (soil-free medium by Keltech Energies Limited, Bangaluru, India) and compared with that of control (uninoculated sterile soilrite). Application of either consortium improved the germination percentage, germination index and vigor index relative to the untreated control (**Table 4**).


**Table 4.** Represents data for germination trial with and without consortium application.

Even without any supporting microbes in the soil-free medium (Soilrite mix), this combination (NB1 + BN7) enhanced *Vigna radiata* (mung bean) germination (98%) as compared to the control (78%).

The consortia (NB1, BN7, NB1 + BN7) were further tested during pot trial (at Maulana Abul Kalam Azad University of Technology, India) and field trial for *Vigna radiata* var Samrat (developed by Indian Institute of Pulse Research, Kanpur, India) from Feb 2013 to May 2013 (spring/summer cultivation). The culture was applied only once at the time of sowing. For field trial, randomized block design with four replicates was carried out at Bidhan Chandra Krishi Viswavidyalaya Seed farm, Kalyani, Nadia, West Bengal, India as well as at State Department of Science and Technology facility, Salt Lake, Kolkata, West Bengal, India. The sowing was done in the north south orientation in February 2013. The seeds post-germination were subjected to thinning on the 8th day post-sowing such that each 1 m<sup>2</sup> area contains a total of 40 plants (4 rows of 10 plants each). The inoculum applied on the day of sowing for field trial was 3.68 × 109 cells per plot (1 m × 1 m). The following parameters were monitored: plant height, number of branches, 50% flowering, 100% flowering, number of flowers, pod initiation, number of pods/plant, pod length, weight/pod, seeds/ pod and weight of 100 seeds. In order to compare the data of the above-mentioned agronomic parameters as well as yield with that of conventional agriculture, simultaneously four (1 m × 1 m) plots were treated with chemical fertilizer. The chemical fertilizer (12.59 g) was applied in the ratio of N:P:K equals 20:40:40 (urea:single super phosphate:murated potash) for each 1 m × 1 m area. The total yield per hectare for each of the applications was monitored with respect to control (unfertilized). When applied together (NB1 + BN7) in field trials, the consortium significantly improved plant growth as compared to separate application (**Table 5**).


**Table 5.** Agronomic parameters for mung bean cultivation following chemical and biofertilizer application as compared to control (unfertilized) condition.

For every parameter, the combined application of NB1 + BN7 exhibited a better effect. Notably, the calculated yield per hectare was highest for NB1 + BN7 (2582.5 kg/ha) followed by chemical fertilizer (2017.5 kg/ha), BN7 (1802.5 kg/ha), NB1 (799.6 kg/ha) and the control (710.05 kg/ ha). Thus, it offers potential advantage in meeting the increased food requirement in today's limited availability of land for agriculture. In addition, the consortia NB1 + BN7 also maintained soil fertility as revealed during the pot trial (**Table 6**).

In addition, each consortium (NB1, BN7, NB1 + BN7) could remove hydrocarbons such as metacil, pesticide and servo (lubricant) from the soil, suggesting that it has potential use in oil spill bioremediation.


India. The sowing was done in the north south orientation in February 2013. The seeds post-germination were subjected to thinning on the 8th day post-sowing such that each 1

eters were monitored: plant height, number of branches, 50% flowering, 100% flowering, number of flowers, pod initiation, number of pods/plant, pod length, weight/pod, seeds/ pod and weight of 100 seeds. In order to compare the data of the above-mentioned agronomic parameters as well as yield with that of conventional agriculture, simultaneously four (1 m × 1 m) plots were treated with chemical fertilizer. The chemical fertilizer (12.59 g) was applied in the ratio of N:P:K equals 20:40:40 (urea:single super phosphate:murated potash) for each 1 m × 1 m area. The total yield per hectare for each of the applications was monitored with respect to control (unfertilized). When applied together (NB1 + BN7) in field trials, the consortium significantly improved plant growth as compared to separate

For every parameter, the combined application of NB1 + BN7 exhibited a better effect. Notably, the calculated yield per hectare was highest for NB1 + BN7 (2582.5 kg/ha) followed by chemical fertilizer (2017.5 kg/ha), BN7 (1802.5 kg/ha), NB1 (799.6 kg/ha) and the control (710.05 kg/ ha). Thus, it offers potential advantage in meeting the increased food requirement in today's limited availability of land for agriculture. In addition, the consortia NB1 + BN7 also main-

**Table 5.** Agronomic parameters for mung bean cultivation following chemical and biofertilizer application as compared

In addition, each consortium (NB1, BN7, NB1 + BN7) could remove hydrocarbons such as metacil, pesticide and servo (lubricant) from the soil, suggesting that it has potential use in oil

tained soil fertility as revealed during the pot trial (**Table 6**).

area contains a total of 40 plants (4 rows of 10 plants each). The inoculum applied on the

**Control NB1 BN7 NB1 + BN7 Chemical**

37.86 ± 4.79 38.87 ± 10.27 40.25 ± 9 38.99 ± 6.79 31.34 ± 8.57

7.8 ± 0.63 7.9 ± 0.8 8.2 ± 1.3 8.9 ± 0.99 8 ± 1.41

4.12 ± 3.09 10.25 ± 3.87 12.89 ± 4.98 11.85 ± 6.23 3.87 ± 2.69

0.41 ± 0.12 0.58 ± 0.23 0.53 ± 0.18 0.77 ± 0.22 0.53 ± 0.11

3 ± 0.005 3.7 ± 0.45 3.59 ± 0.86 4.34 ± 0.46 4.27 ± 0.01

Pod length (cm) 6.33 ± 0.86 7.65 ± 0.67 7.71 ± 1.31 8.07 ± 1.12 7.83 ± 1.05

Seeds per pod 4 ± 1.58 4 ± 0.83 5 ± 1.15 7 ± 1.3 10 ± 0.83

cells per plot (1 m × 1 m). The following param-

m<sup>2</sup>

application (**Table 5**).

**Parameters Treatments**

Height of plants

Number of pods per plant

Weight per pod

Weight of 100 seeds (g)

Number of branches

(cm)

(g)

spill bioremediation.

to control (unfertilized) condition.

day of sowing for field trial was 3.68 × 109

46 Biological Wastewater Treatment and Resource Recovery

**Table 6.** Soil nutritional quality analysis pre- and post-cultivation of mung bean during pot trial using standard methods..

## **5. Seed quality analysis**

The seeds were lyophilized for 24 h and manually ground in the mortar and pestle; 0.2 g ground material was pelleted using Pelletizer (Technolab, Kbr Press) at 110 kg/cm2 . The mineral content of the pellets was assessed using energy-dispersive X-ray fluorescence (Jordan Valley EX–3600) analysis as per reported protocol [46, 47] at University Grant Commission-Department of Atomic Energy facility, Kolkata Center, India (**Table 7**).


The commercially available fertilizer (Urea: Single Super Phosphate: Murated Potash) was applied in ratio of N:P:K equals 20:40:40 whereas in case of microbial biomass (N:P—2.52:1.51), 3.68 × 109 cells were added per plot (1 m × 1 m). The lyophilized seeds were manually grounded, and 0.25 g of the powder was converted into pellet and was analyzed by EDXRF for mineral content.

**Table 7.** Represents the elemental content of the seeds grown during control (unfertilized), chemical fertilizer as well as biofertilizer treatment.

The nutritional quality analysis like moisture [IS:4333(Part-II):2002], total protein (AOAC 920.87), available carbohydrate (AOAC 986.25), fat (AOAC 963.15), energy (Analytical Chemistry of Food by CS James:1995), ash content (AOAC 941.12), sugar (AOAC 923.09) and fiber (AOAC 985.29) content was carried out at SGS India Private Limited, Kolkata, India as per standard protocol (**Table 8**).

The statistical validation for the variation in elemental content of the seeds grown using varying treatments was carried out using single-factor ANOVA in Microsoft excel 2007. Here, the two hypotheses were as follows: null hypothesis H0 : no difference in elemental content with difference in treatment; alternative hypothesis H1 : significant difference in elemental content with difference in treatment. The level of significance was fixed at 5%. Based on a singlefactor ANOVA, a significant variation was observed in the elemental content of the seeds produced after the treatments, especially in the Zn, Mn and Cu content between the control and NB1 + BN7 seeds. This clearly suggests that the consortium produces more elementally


stable seeds. However, the overall nutritional quality of the seeds was maintained regardless of the treatment. The consortium exhibited similar trends for *Cicer arietinum* (chick pea) and *Abelmoschus esculentus* (ladies finger) cultivations.

**Table 8.** The nutritional quality of the seeds following cultivation under control (unfertilized), chemical fertilizer as well as consortium (NB1, BN7, NB1 + BN7) treatment.

## **6. Conclusion**

**5. Seed quality analysis**

48 Biological Wastewater Treatment and Resource Recovery

**Elements mg/kg (ppm)**

P 4242.09 ± 475.2

K 13,538.33 ± 491.76

S 2165.53 ± 288.35

Ca 2034.13 ± 149.41

by EDXRF for mineral content.

biofertilizer treatment.

per standard protocol (**Table 8**).

two hypotheses were as follows: null hypothesis H0

difference in treatment; alternative hypothesis H1

The seeds were lyophilized for 24 h and manually ground in the mortar and pestle; 0.2 g

eral content of the pellets was assessed using energy-dispersive X-ray fluorescence (Jordan Valley EX–3600) analysis as per reported protocol [46, 47] at University Grant Commission-

> 3741.01 ± 481.4947

11,807.17 ± 773.6117

2037.44 ± 118.75 1575.90 ±

1714.23 ± 79.81 1777.90 ±

1416.79 ± 574.18

10,943.22 ± 1349.72

118.02

396.11

**Control NB1 BN7 NB1 + BN7 Chemical p-Value Recommended** 

The nutritional quality analysis like moisture [IS:4333(Part-II):2002], total protein (AOAC 920.87), available carbohydrate (AOAC 986.25), fat (AOAC 963.15), energy (Analytical Chemistry of Food by CS James:1995), ash content (AOAC 941.12), sugar (AOAC 923.09) and fiber (AOAC 985.29) content was carried out at SGS India Private Limited, Kolkata, India as

The commercially available fertilizer (Urea: Single Super Phosphate: Murated Potash) was applied in ratio of N:P:K

The lyophilized seeds were manually grounded, and 0.25 g of the powder was converted into pellet and was analyzed

**Table 7.** Represents the elemental content of the seeds grown during control (unfertilized), chemical fertilizer as well as

The statistical validation for the variation in elemental content of the seeds grown using varying treatments was carried out using single-factor ANOVA in Microsoft excel 2007. Here, the

with difference in treatment. The level of significance was fixed at 5%. Based on a singlefactor ANOVA, a significant variation was observed in the elemental content of the seeds produced after the treatments, especially in the Zn, Mn and Cu content between the control and NB1 + BN7 seeds. This clearly suggests that the consortium produces more elementally

: no difference in elemental content with

: significant difference in elemental content

. The min-

**by USDA**

0.003 3670.00

0.18 12,460.00

0.05 NA

0.04 1320.00

cells were added per plot (1 m × 1 m).

ground material was pelleted using Pelletizer (Technolab, Kbr Press) at 110 kg/cm2

Zn 37.21 ± 2 44.57 ± 2.05 27 ± 3.02 29.06 ± 2.43 34.23 ± 2.58 0.04 26.8 Fe 68.34 ± 2.25 71.92 ± 1.66 68.45 ± 6.89 70.71 ± 0.57 67.21 ± 4.41 0.04 67.4 Mn 12.42 ± 0.44 12.74 ± 1.56 13.65 ± 1.43 15.46 ± 1.50 13.30 ± 0.64 0.02 10.35 Cu 13.30 ± 0.45 15.19 ± 0.56 15.66 ± 1.02 14.62 ± 1.39 14.49 ± 1.30 0.21 9.41

Department of Atomic Energy facility, Kolkata Center, India (**Table 7**).

4604.71 ± 50.2 2429.97 ±

13,830.88 ± 415.3

2341.02 ± 63.25

2071.45 ± 214.95

619.20

9651.83 ± 1546.293

1692.56 ± 199.5616

1650.99 ± 410.549

equals 20:40:40 whereas in case of microbial biomass (N:P—2.52:1.51), 3.68 × 109

The aim of this study was to develop an alternative strategy for plant nutrient management through microbial intervention. The objective of prevention of leaching of nitrate from soil was achieved through application of a 1:1 mixture of NB1 and BN7. It also ensured retention of nitrate within the root zone of soil. Being accumulators of nitrate and phosphate as well as producers of phytohormones with phosphatase activity, they could enhance germination while making the phosphate available for plant uptake. Thus, a single combination has the desired properties of a biofertilizer like phytohormone production, supplying of nutrients (nitrate and phosphate) resulting in higher yield of nutritionally enriched seeds. The unique selling points of this bioformulation are as follows: (i) its 21.88 times greater productivity (in case of mung bean) as compared to chemical fertilizer application and (ii) maintenance of soil fertility post-cultivation. Hereby, the remaining objections of multinutrient sequestration and reuse were effectively achieved. The wide range of pH and metal tolerance makes these consortia suitable for environmental application under varied conditions. These unique features of BN7 as well as NB1 + BN7 have been filed as Indian Patents 518/KOL/2011 dated April 11, 2011 and 203/KOL/2013 dated Feb 21, 2013. By this method, the nitrate concentration from agricultural runoff could be reduced substantially by using these microbes. All these properties point towards the future application of this innovation for bioremediation through nutrient sequestration from agricultural runoff as well as effluents and its reuse as biofertilizer with potential for environmental protection and agricultural sustenance.

## **Acknowledgements**

Authors would like to acknowledge Mr. Vivek Singh for pot trial experiments using soilfree medium; Mr. Arpan Pal as well as Late Mr. Sourav Chakraborty for assistance during field trial experiments. The authors acknowledge Maulana Abul Kalam Azad University of Technology for the laboratory and computational facility. The authors would like to thank the following granting agencies for financial assistance: Ministry of Human Resource Development, Government of India (GOI) under the FAST scheme for providing the publication fee and fund for outsourcing services; Department of Biotechnology, GOI for M.Tech student fellowship; Indian Council for Agricultural Research, GOI under the NFBSFARA scheme for field trial related expenditure and seed quality analysis; Department of Atomic Energy under the BRNS scheme for consortium development and characterization as well as the World Bank under the TEQIP II scheme for student fellow.

## **Author details**

Shaon Ray Chaudhuri1,2\*, Madhusmita Mishra3 , Sonakshi De3 , Biswajit Samal3 , Amrita Saha2 , Srimoyee Banerjee1,3, Abhinandan Chakraborty3 , Antara Chakraborty3 , Sonali Pardhiya3 , Deepak Gola3 , Joyeeta Chakraborty3 , Sourav Ghosh1,3, Kamlesh Jangid4 , Indranil Mukherjee2 , Mathummal Sudarshan5 , Rajib Nath6 and Ashoke Ranjan Thakur1

\*Address all correspondence to: shaonraychaudhuri@tripurauniv.in

1 Department of Microbiology, Tripura University, Suryamaninagar, On lien from Maulana Abul Kalam Azad University of Technology, West Bengal (Formerly known as West Bengal University of Technology), India

2 Centre of Excellence in Environmental Technology and Management, Maulana Abul Kalam Azad University of Technology, West Bengal (Formerly known as West Bengal University of Technology), India

3 Department of Biotechnology, Maulana Abul Kalam Azad University of Technology, West Bengal (Formerly known as West Bengal University of Technology), India

4 Microbial Culture Collection, National Centre for Cell Science, Pune, India

5 Laboratory of Trace Elements, Inter University Consortium, Kolkata, Kolkata Centre, West Bengal, India

6 Faculty of Agronomy, Bidhan Chandra Krishi Viswavidyalaya, Mohanpur, Nadia, West Bengal, India

## **References**

agricultural runoff could be reduced substantially by using these microbes. All these properties point towards the future application of this innovation for bioremediation through nutrient sequestration from agricultural runoff as well as effluents and its reuse as biofertilizer

Authors would like to acknowledge Mr. Vivek Singh for pot trial experiments using soilfree medium; Mr. Arpan Pal as well as Late Mr. Sourav Chakraborty for assistance during field trial experiments. The authors acknowledge Maulana Abul Kalam Azad University of Technology for the laboratory and computational facility. The authors would like to thank the following granting agencies for financial assistance: Ministry of Human Resource Development, Government of India (GOI) under the FAST scheme for providing the publication fee and fund for outsourcing services; Department of Biotechnology, GOI for M.Tech student fellowship; Indian Council for Agricultural Research, GOI under the NFBSFARA scheme for field trial related expenditure and seed quality analysis; Department of Atomic Energy under the BRNS scheme for consortium development and characterization as well

, Sonakshi De3

, Sourav Ghosh1,3, Kamlesh Jangid4

and Ashoke Ranjan Thakur1

1 Department of Microbiology, Tripura University, Suryamaninagar, On lien from Maulana Abul Kalam Azad University of Technology, West Bengal (Formerly known as West Bengal

2 Centre of Excellence in Environmental Technology and Management, Maulana Abul Kalam Azad University of Technology, West Bengal (Formerly known as West Bengal University of

3 Department of Biotechnology, Maulana Abul Kalam Azad University of Technology, West

5 Laboratory of Trace Elements, Inter University Consortium, Kolkata, Kolkata Centre, West

6 Faculty of Agronomy, Bidhan Chandra Krishi Viswavidyalaya, Mohanpur, Nadia, West

, Antara Chakraborty3

, Biswajit Samal3

, Amrita Saha2

, Sonali Pardhiya3

, Indranil Mukherjee2

,

,

,

with potential for environmental protection and agricultural sustenance.

as the World Bank under the TEQIP II scheme for student fellow.

Shaon Ray Chaudhuri1,2\*, Madhusmita Mishra3

Srimoyee Banerjee1,3, Abhinandan Chakraborty3

, Joyeeta Chakraborty3

, Rajib Nath6

\*Address all correspondence to: shaonraychaudhuri@tripurauniv.in

Bengal (Formerly known as West Bengal University of Technology), India

4 Microbial Culture Collection, National Centre for Cell Science, Pune, India

**Acknowledgements**

50 Biological Wastewater Treatment and Resource Recovery

**Author details**

Deepak Gola3

Mathummal Sudarshan5

Technology), India

Bengal, India

Bengal, India

University of Technology), India


[27] Konovalova V, Nigmatullin R, Dmytrenko G, Pobigay G. Spatial sequencing of microbial reduction of chromate and nitrate in membrane bioreactor. Bioprocess and Biosystems Engineering. 2008 Oct 1;31(6):647–53. doi:10.1007/s00449-008-0215-7

[15] Pinar G, Duque E, Haidour A, Oliva J, Sanchez-Barbero L, Calvo V, Ramos JL. Removal of high concentrations of nitrate from industrial wastewaters by bacteria. Applied and Environmental Microbiology. 1997 May 1;63(5):2071–3 (Accessed 10th September 2016).

[16] Awadallah RM, Soltan ME, Shabeb MS, Moalla SM. Bacterial removal of nitrate, nitrite and sulphate in wastewater. Water Research. 1998 Oct 31;32(10):3080–4.doi:10.1016/

[17] Eckford RE, Fedorak PM. Planktonic nitrate-reducing bacteria and sulfate-reducing bacteria in some western Canadian oil field waters. Journal of Industrial Microbiology and

[18] McAdam EJ, Judd SJ. Immersed membrane bioreactors for nitrate removal from drinking water: cost and feasibility. Desalination. 2008 Oct 31;231(1):52–60. doi:10.1016/j.

[19] Choi JH, Maruthamuthu S, Lee HG, Ha TH, Bae JH. Nitrate removal by electro-bioremediation technology in Korean soil. Journal of Hazardous Materials. 2009 Sep

[20] Parvanova-Mancheva T, Beschkov V. Microbial denitrification by immobilized bacteria *Pseudomonas denitrificans* stimulated by constant electric field. Biochemical Engineering

[21] Pawels R, Haridas A, Jose BT. Biological sulphate reduction with hydrogen in a jet loop biofilm reactor. International Journal of Scientific and Research Publications. 315. http://www.

[22] Hollo J, Czako L. Nitrate removal from drinking water in a fluidized-bed biological denitrification bioreactor. Acta Biotechnologica. 1987 Jan 1;7(5):417–23. doi:10.1002/

[23] Li H, Yang M, Zhang Y, Yu T, Kamagata Y. Nitrification performance and microbial community dynamics in a submerged membrane bioreactor with complete sludge retention. Journal of Biotechnology. 2006 May 3;123(1):60–70. doi:10.1016/j.jbiotec.2005.10.001

[24] Kesserű P, Kiss I, Bihari Z, Polyák B. Biological denitrification in a continuous-flow pilot bioreactor containing immobilized *Pseudomonas butanovora* cells. Bioresource

[25] Van Rijn J, Tal Y, Schreier HJ. Denitrification in recirculating systems: theory and applications. Aquacultural Engineering. 2006 May 31;34(3):364–76.doi:10.1016/j.

[26] Cao GM, Zhao QX, Sun XB, Zhang T. Integrated nitrogen removal in a shell-and-tube co-immobilized cell bioreactor. Process Biochemistry. 2004 Jun 30;39(10):1269–73.

Technology. 2003 Mar 31;87(1):75–80. doi:10.1016/S0960-8524(02)00209-2

ijsrp.org/research-paper-0613/ijsrp-p1858.pdf (Accessed 10th September 2016).

Biotechnology. 2002 Aug 1;29(2):83–92. doi:10.1038/sj.jim.7000274

Journal. 2009 May 15;44(2):208–13. doi:10.1016/j.bej.2008.12.005

15;168(2):1208–16. doi:10.1016/j.jhazmat.2009.02.162

S0043-1354(98) 00069-4

52 Biological Wastewater Treatment and Resource Recovery

desal.2007.11.038

abio.370070509

aquaeng.2005.04.004

doi:10.1016/S0032-9592(03)00256-5


[49] Trivedi RK, Goel PK. Chemical and Biological methods for water pollution studies. Karad, India: Environmental Publication; 1984.

[38] Mishra M, Jain S, Thakur AR, RayChaudhuri S. Microbial community in packed bed bioreactor involved in nitrate remediation from low level radioactive waste. Journal of

[39] Ray Chaudhuri S, Sharmin J, Banerjee S, Jayakrishnan U, Saha A, Mishra M, Ghosh M, Mukherjee I, Banerjee A, Jangid K, Sudarshan M, Chakraborty A, Ghosh S, Nath R, Banerjee M, Singh SS, Saha AK, Thakur AR. Novel microbial system developed from low level radioactive waste treatment plant for environmental sustenance. In: Management of radioactive and hazardous waste. InTech; 2016. ISBN 978-953-51-4764-0

[40] Cataldo DA, Maroon M, Schrader LE, Youngs VL. Rapid colorimetric determination of nitrate in plant tissue by nitration of salicylic acid 1. Communications in Soil Science &

[41] Adarsh VK, Mishra M, Chowdhury S, Sudarshan M, Thakur AR, Chaudhuri SR. Studies on metal microbe interaction of three bacterial isolates from east Calcutta wetland.

[42] RayChaudhuri S, Thakur AR. Microbial DNA extraction from sample of varied origin. Current Science. 2006 12: 1697–1700. http://www.researchgate.net/publication/237225066\_Microbial\_DNA\_extraction\_from\_samples\_of\_varied\_origin/

[43] Nasipuri P, Pandit G, Thakur AR, Chaudhuri SR. Microbial consortia from taptapani hot water springs for mining effluent treatment. American Journal of Microbiology.

[44] Nasipuri P, Pandit GG, Thakur AR, Chaudhuri SR. Comparative study of soluble sulfate reduction by bacterial consortia from varied regions of India. American Journal of

[45] Chakraborty U, Chakraborty BN, Chakraborty AP, Sunar K, Dey PL. Plant growth promoting rhizobacteria mediated improvement of health status of tea plants. Indian Journal of Biotechnology. 2013 Jan 1;12(1):20–31. doi:10.2135/cropsci1973.0011183X001300060 013x. http://nopr.niscair.res.in/bitstream/123456789/16536/1/IJBT%2012%281%29%20

[46] Chowdhury S, Mishra M, Adarsh VK, Mukherjee A, Thakur AR, Chaudhuri SR. Novel metal accumulator and protease secretor microbes from East Calcutta Wetland. American Journal of Biochemistry Biotechnology. 2008;4(3):255–64.doi:10.3844/ajbbsp.2008.255.264

[47] Raychaudhuri S, Salodkar S, Sudarshan M, Thakur AR. Integrated resource recovery at East Calcutta Wetland: how safe is these? American Journal of Agricultural and

[48] Sacheti AK. Agricultural meteorology: instructional cum practical manual. New Delhi:

Environmental Sciences. 2010;6(2):152–8. doi:10.3844/ajessp.2010.152.158

Plant Analysis. 1975 Jan 1;6(1):71–80. doi:10.1080/00103627509366547

file/3deec525bd625c8a61.pdf (Accessed 8th June 2014).

2011;1(3):23–9.doi:10.3844/ajmsp.2010.23.29

20-31.pdf (Accessed 10th September 2016).

Biological Science. 2007. doi:10.3844/ajabssp.2007.75.80

National Council of Educational Research and Training; 1985, p. 53.

Basic Microbiology. 2014 Mar 1;54(3):198–203. doi:10.1002/jobm.201200676.

DOI:10.5772/63323.

doi:10.3844/ojbsci.2007.80.88

54 Biological Wastewater Treatment and Resource Recovery


#### **Mitigating Environmental Risks of Wastewater Reuse for Agriculture Mitigating Environmental Risks of Wastewater Reuse for Agriculture**

Ahmed Al‐Busaidi and Mushtaque Ahmed Ahmed Al‐Busaidi and Mushtaque Ahmed

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

http://dx.doi.org/10.5772/65871

#### **Abstract**

The study was aimed to maximize and optimize treated wastewater reuse in conjunction with surface and ground waters resources. Moreover, environmental, agronomic and economic components were also considered. The project was funded by USAID and implemented in three countries (Oman, Tunisia and Jordan). In Oman, the study was done at Sultan Qaboos University experimental station field. Four types of waters (A: 50% of treated wastewater with 50% of groundwater, B: 100% of groundwater, C: 25% of groundwater with 75% of treated wastewater, and D: 100% of treated wastewater) were used to grow three different crops (okra, maize and sweet corn). Results showed no significant differences in soil physical and chemical properties with treatments irrigated with treated wastewater as compared to groundwater. On other hand, some chemical properties significantly increased (p<0.05) when treated wastewater was applied such as soil total carbon and some major elements (N, K, Mg). Crop physical analysis showed significant increases in plant productivity when plants were irrigated with treated wastewater and values of chemical properties were within the international standards. Crop biological analysis showed no effect on crop quality and all tested crops were free from any microbial contamination.

**Keywords:** treated wastewater, soil, fruits, yield, heavy metals

## **1. Introduction**

Drought and overexploitation of conventional water resources present a critical problem in many regions of the world, especially the Middle East [1]. Therefore, water resources including nonconventional water should be well managed. Usage of treated wastewater (TWW) on agriculture can save fresh water resources and minimize the applications of chemical fertilizers. In many parts of the world, treated wastewater has been successfully used for irrigation, and

© 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.

many researchers have recognized its benefits [2, 3]. The continuous use of treated wastewater in irrigation increases the total soluble salts in the soil. The cation exchange capacity values are increased by increasing the period of using treated wastewater for irrigation, especially in the surface layer(0–30 cm).Moreover,Fe,Zn,Cu,Mn,Pb, andCowere increasedbyirrigationusing treated wastewater as compared to virgin soil [4]. The use of treated wastewater for irrigation increasedthemitoticindexofdividedcells,chromosomeabnormality,andcontentsofindividual amino acids. However, no differences in the profile of protein bands were observed between controlandtreatedwastewaterirrigationplants [5].Whereas theaccumulationsofheavymetals in the ediblepart of someplants weredetected which adversely affect human andanimal health through the food chain [6].

Many countries have included wastewater reuse as an important component of water resources planning. Some countries like Oman have a national policy to reuse all treated wastewater effluents and have already made considerable progress toward this end [7]. Sultanate of Oman is one of the Middle East countries that is considered as the driest or semidriest region in the world with rapidly developing economy and a high population growth [8]. Soil and ground‐ water (GW) resources of good quality irrigation water have become limited. Rainfall is scanty to support crop production with annual mean rainfall of 100 mm. Therefore, its agriculture is almost fully dependent on groundwater [9]. Water resources augmentation together with conservation has been adopted by the government to combat the water shortage problem. The rapid development of Oman's urbanization, increase in population, and increase in agricul‐ tural production has led to high demand for water and urgent need to use treated wastewater as an alternative source of freshwater in agriculture. However, treated wastewater may contain high concentrations of salts, heavy metals, pathogens, and emerging pollutants with unknown effects on the ecological system [10]. High concentrations of heavy metals in plant fruits could affect human health and cause many environmental problems. However, the conjunctive use of treated wastewater and groundwater resources could be employed, helping to safeguard farmer's income and sustain food production. Despite this promising option, more research and education efforts are needed to ensure proper use of treated wastewater for agricultural production. Therefore, the study aimed to optimize treated wastewater reuse in conjunction with groundwater by taking into consideration their quantity and quality, in addition to the agronomic, environmental, and economic components.

### **2. Materials and methods**

The field work was done in plots at the Agricultural Experiment Station, Sultan Qaboos University. Twelve plots (2.5 × 3.5 m each) were designed and sweet corn, okra, and maize crops were grown during the study. The plots were irrigated with four types of waters (A: 50% groundwater and 50% treated wastewater; B: 100% groundwater; C: 75% treated wastewater and 25% groundwater; and D: 100% treated wastewater) as shown in **Figure 1**. Plants were daily irrigated based on evapotranspiration (ETc). Soil samples were taken before and at the end of the study at a depth of 0–30 cm. Whereas plant samples were taken when the crop was mature and ready for analysis.

**Figure 1.** Field experimental design with different treatments.

Plants growth and yield of each crop irrigated by different waters were monitored. Fruits quality and quantity were assessed. Samples from soil and plants were taken for different physical, chemical, and biological analyses. All physicochemical analysis for soil and plants were done in soil and water labs (SQU) following standard methods [11] and using inductively coupled plasma (ICP) instrument for metals analysis. Soil and plant nitrogen (N) were analyzed in Rumais Research Laboratory (Ministry of Agriculture and Fisheries). Whereas biological analyses for crop samples were done in Muscat Municipality Laboratory.

The data were analyzed statistically using the analysis of variance (ANOVA) and the means were compared at the probability level of 5% using the least significant difference [12].

## **3. Result and discussion**

many researchers have recognized its benefits [2, 3]. The continuous use of treated wastewater in irrigation increases the total soluble salts in the soil. The cation exchange capacity values are increased by increasing the period of using treated wastewater for irrigation, especially in the surface layer(0–30 cm).Moreover,Fe,Zn,Cu,Mn,Pb, andCowere increasedbyirrigationusing treated wastewater as compared to virgin soil [4]. The use of treated wastewater for irrigation increasedthemitoticindexofdividedcells,chromosomeabnormality,andcontentsofindividual amino acids. However, no differences in the profile of protein bands were observed between controlandtreatedwastewaterirrigationplants [5].Whereas theaccumulationsofheavymetals in the ediblepart of someplants weredetected which adversely affect human andanimal health

Many countries have included wastewater reuse as an important component of water resources planning. Some countries like Oman have a national policy to reuse all treated wastewater effluents and have already made considerable progress toward this end [7]. Sultanate of Oman is one of the Middle East countries that is considered as the driest or semidriest region in the world with rapidly developing economy and a high population growth [8]. Soil and ground‐ water (GW) resources of good quality irrigation water have become limited. Rainfall is scanty to support crop production with annual mean rainfall of 100 mm. Therefore, its agriculture is almost fully dependent on groundwater [9]. Water resources augmentation together with conservation has been adopted by the government to combat the water shortage problem. The rapid development of Oman's urbanization, increase in population, and increase in agricul‐ tural production has led to high demand for water and urgent need to use treated wastewater as an alternative source of freshwater in agriculture. However, treated wastewater may contain high concentrations of salts, heavy metals, pathogens, and emerging pollutants with unknown effects on the ecological system [10]. High concentrations of heavy metals in plant fruits could affect human health and cause many environmental problems. However, the conjunctive use of treated wastewater and groundwater resources could be employed, helping to safeguard farmer's income and sustain food production. Despite this promising option, more research and education efforts are needed to ensure proper use of treated wastewater for agricultural production. Therefore, the study aimed to optimize treated wastewater reuse in conjunction with groundwater by taking into consideration their quantity and quality, in addition to the

The field work was done in plots at the Agricultural Experiment Station, Sultan Qaboos University. Twelve plots (2.5 × 3.5 m each) were designed and sweet corn, okra, and maize crops were grown during the study. The plots were irrigated with four types of waters (A: 50% groundwater and 50% treated wastewater; B: 100% groundwater; C: 75% treated wastewater and 25% groundwater; and D: 100% treated wastewater) as shown in **Figure 1**. Plants were daily irrigated based on evapotranspiration (ETc). Soil samples were taken before and at the end of the study at a depth of 0–30 cm. Whereas plant samples were taken when the crop was

through the food chain [6].

58 Biological Wastewater Treatment and Resource Recovery

agronomic, environmental, and economic components.

**2. Materials and methods**

mature and ready for analysis.

#### **3.1. Heavy metals in irrigation water**

Growing conditions and the irrigation water are the most important parameters controlling plant life. **Table 1** demonstrates heavy metal concentrations in the irrigation waters that were used in the study. Comparing the used waters with national and international standards, it can be seen that elements concentrations mentioned in **Table 1** had lower values than applied standards. However, long‐term application of some waters may accumulate some harmful elements in soil and plant tissues if mismanagement occurs. In some studies, it was found that wastewaters could carry appreciable amounts of trace toxic metals [13, 14] and concentrations of trace metals in sewage effluents vary from one city to another [15]. Although the concen‐ tration of heavy metals in sewage effluents are low, long‐term use of these wastewaters on agricultural lands often results in the build‐up of elevated levels of these metals in soils [15]. The results of Rattan et al. [15] reported high amount of Cr, Cu, Pb, Co, Ni, Mn, Cd, Fe, Zn, and As in sewage effluents compared to groundwater. Whereas soil organic matter was also increased in soil samples irrigated with sewage effluents compared to groundwater.


\* Summary of U.S. EPA guidelines for water reuse for irrigation [16].

**Table 1.** Comparing heavy metals concentration (mg/l) in the irrigation waters with national and international standards\* .

#### **3.2. Soil physicochemical properties**

Quality of irrigation water could affect soil physicochemical properties. It could improve the soil quality by adding more nutrients or degrading the soil by adding toxic salts. Soil organic matter and total carbon are usually interconnected parameters and they are good indicators for soil fertility. In our study, some of them were found to be high in treated wastewater (TWW) compared to groundwater (GW) treatments. It is an expected result since treated wastewater is usually rich in nitrogen and other nutrients, which enrich soil and enhance plant growth (**Figure 2a** and **b**).

**Figure 2.** (a) Soil organic matter and (b) soil total carbon.

The presence of more nutrients (salts) in treated wastewater helps in keeping more water in plant root zone compared to groundwater (**Figure 3**). Nutrients as salts increased water viscosity and reduced evaporation process and as a result more water can be kept in the root zone [17].

**Figure 3.** Soil moisture of all treatments.

**Water Mn Fe Zn Cu Cr Cd Pb Ni B** Groundwater 0.002 0.013 0.013 0.008 <0.002 <0.001 <0.001 <0.001 0.295 Treated wastewater 0.002 0.016 0.064 0.024 <0.002 <0.001 0.066 <0.001 0.508 EPA Standard 0.200 5.000 5.000 0.500 0.100 0.010 0.100 0.100 0.750 FAO Standard 0.200 5.000 2.000 0.200 0.100 0.010 0.500 0.200 0.750 Omani Standard 0.500 5.000 5.000 1.000 0.050 0.010 0.200 0.100 0.750

**Table 1.** Comparing heavy metals concentration (mg/l) in the irrigation waters with national and international

Quality of irrigation water could affect soil physicochemical properties. It could improve the soil quality by adding more nutrients or degrading the soil by adding toxic salts. Soil organic matter and total carbon are usually interconnected parameters and they are good indicators for soil fertility. In our study, some of them were found to be high in treated wastewater (TWW) compared to groundwater (GW) treatments. It is an expected result since treated wastewater is usually rich in nitrogen and other nutrients, which enrich soil and enhance plant growth

The presence of more nutrients (salts) in treated wastewater helps in keeping more water in plant root zone compared to groundwater (**Figure 3**). Nutrients as salts increased water viscosity and reduced evaporation process and as a result more water can be kept in the root

Summary of U.S. EPA guidelines for water reuse for irrigation [16].

**3.2. Soil physicochemical properties**

60 Biological Wastewater Treatment and Resource Recovery

**Figure 2.** (a) Soil organic matter and (b) soil total carbon.

\*

standards\* .

(**Figure 2a** and **b**).

zone [17].

Treated wastewater has a good amount of nutrients (salts). Therefore, it will add more salts to the irrigated soil and increase soil salinity compared to soil irrigated with freshwater (**Figure 4**). Salts are usually managed and reduced by leaching process.

**Figure 4.** Soil electrical conductivity in all treatments.

In addition to organic matter, treated wastewaters have higher values for several nutrients compared to groundwater (**Figure 5**). These nutrients can improve soil fertility and later support plant growth. The variations in some elements' concentrations between treatments could be due to original nutrients concentrations in the soil and absorbance of those metals during plant growth. Mohammed and Mazahareh [10] found that treated wastewater irriga‐ tion increased soil salinity, soil phosphorous, potassium, iron, and manganese levels. They noticed that soil organic matter increased only in the topsoil.

**Figure 5.** Mean concentration of some elements in soil samples.


\* Means in the column with same letter indicate no difference at Duncan's Multiple Range Test at *p* < 0.05.

**Table 2.** Mean concentration of heavy metals (mg/l) in soil samples\* .

Checking soil for microelements (heavy metals) concentrations, it can be seen in **Table 2** that all values of heavy metals for both treatments (treated wastewater and groundwater) were very close to each other. However, some significant differences were found between some treatments which could be an indicator for long‐term changes in soil chemical properties which is also found in Bansal et al. [18] and Palaniswami and Sree Ramulu [19] studies when they applied wastewater for long period. Rattan et al. [20] observed a build‐up of Zn, Pb, Ni, Mn, Fe, Cu, Cr, Co, and As in the sewage‐irrigated soils, over the well water‐irrigated ones. Significant effect of irrigation through sewage water was observed in case of studied metals. There has been an enormous build‐up in the available Fe content in the sewage‐irrigated soils. Soils irrigated with groundwater and sewage water showed higher level of Cu and Zn. However, some sewage‐irrigated soils accumulated more than 70 mg kg‐1 total Zn, which could cause a phytotoxicity problem [20]. Whereas Berry et al. [21] found that soil zinc and copper were not significantly affected by wastewater irrigation.

#### **3.3. Crop physicochemical analysis**

From **Table 3**, it can be seen that treated wastewater gave the best yield for all three crops compared to groundwater. The good supply of different nutrients from treated wastewater enhanced plant growth and improved plant productivity. Abohassan et al. [22] and Stewart et al. [23] have identified the beneficial effects of treated sewage water on some trees grown in Saudi Arabia and Australia. Shafiq et al. [24] found an increase of 24, 45, and 68% in maize total fresh biomass, dry yield, and grain yield irrigated by treated wastewater compared to groundwater. Same finding was also reported by Harati [25] in maize plants.


**Table 3.** Average weight (kg) of some crops grown in the study.

**Figure 5.** Mean concentration of some elements in soil samples.

62 Biological Wastewater Treatment and Resource Recovery

**Table 2.** Mean concentration of heavy metals (mg/l) in soil samples\*

\*

**Treatment Mn Cd Fe Zn B Ba Cr Co Pb Ni** 50%TWW 0.018a 0.001a 0.330a 0.026a 0.166c 0.118a 0.043a 0.058a 0.196b 0.005a 100%GW 0.018a 0.001a 0.331a 0.001b 0.171b 0.123a 0.039a 0.060a 0.219a 0.011a 75%TWW 0.016a 0.001a 0.334a 0.001b 0.088d 0.087b 0.041a 0.061a 0.220a 0.005a 100%TWW 0.018a 0.001a 0.345a 0.003b 0.309a 0.110a 0.039a 0.062a 0.234a 0.001a

Means in the column with same letter indicate no difference at Duncan's Multiple Range Test at *p* < 0.05.

.

Checking soil for microelements (heavy metals) concentrations, it can be seen in **Table 2** that all values of heavy metals for both treatments (treated wastewater and groundwater) were very close to each other. However, some significant differences were found between some treatments which could be an indicator for long‐term changes in soil chemical properties which is also found in Bansal et al. [18] and Palaniswami and Sree Ramulu [19] studies when they applied wastewater for long period. Rattan et al. [20] observed a build‐up of Zn, Pb, Ni, Mn, Fe, Cu, Cr, Co, and As in the sewage‐irrigated soils, over the well water‐irrigated ones. Significant effect of irrigation through sewage water was observed in case of studied metals. There has been an enormous build‐up in the available Fe content in the sewage‐irrigated soils. Soils irrigated with groundwater and sewage water showed higher level of Cu and Zn. However, some sewage‐irrigated soils accumulated more than 70 mg kg‐1 total Zn, which could

**Figure 6.** Percentage of total carbon in maize plant leaves.

Maize leaves were the best indicator for carbon content. Therefore, it can be seen from **Figure 6** that treated wastewater got the highest values compared to other treatments. It could be a reflection for what was found in water and soil samples. In a similar study done by Abd‐ Elfattah et al. [6], they found significant differences in metal content of plant leaves grown in soils irrigated with treated wastewater and plant leaves grown in soils irrigated with Nile water of both seasons.


\* Means in the column with same letter indicate no difference at Duncan's Multiple Range Test at *p* < 0.05.

**Table 4.** Heavy metals concentration (mg/l) in tested crops\* .


**Table 5.** Guideline for safe limits of heavy metals in plants (mg/kg).

For soil, usually there is a direct relationship between salts found in the irrigation water and irrigated land. Whereas, for plants, root selectivity and present of salts in different forms could play a role in elements movement and translocation from soil to plant. From **Table 4**, it can be seen that concentrations of many elements were significantly (*p* < 0.05) different from one treatment to other. However, microelements in the edible parts of all crops grown in the field were not that high and they were within the international standards (**Table 5**). Same results were reported by Abdelrahman et al. [26] when they observed no significant difference between fresh and treated wastewater with regards to heavy metals accumulation in grown crops. Moreover, this finding was supported by Pescod [27] study, when he concluded that the concentrations of heavy metals in seeds were within normal level when treated wastewater effluent was used. Such results make it clear that heavy metal in soil are not readily bio‐ available for crop uptake and do not represent a threat to quality of crop consumption.

In general, treated wastewater contains variable amounts of nutrient elements and heavy metals. Availability and translocation of these elements to and within the plant tissues is highly dependent on the environmental conditions as well as their concentration and ratios in the plant organs [28]. Same results were also found by Mahdi et al. [29] when they reported that concentration of nutrient elements of different crops indicated that the crop nutrient uptake is affected by tree age and species. Longer exposure to treated wastewater did not indicate major effects on fruit minerals, including heavy metals. Sampling over longer period of time is needed to confirm the changes in nutrient composition over time. Therefore, in the present study it can be seen that treated wastewater treatment sometime got the highest values for heavy metals which could be an indication for heavy metal accumulation with long‐term application if treated wastewater is used without proper management. This prediction could be similar to Abd‐Elfattah et al. [6] findings when they found a significant difference in fruit contents of heavy metals and trace elements (Pb, Cd, Ni, Cu, Mn, Fe, Zn) between fruits produced by treated wastewater compared with Nile water in both seasons. The accumulation of heavy metals in the edible part of plant was detected which adversely affects human and animal health through the food chain [6].

Finally, the findings of this study are supported by many researches. As such, Omran et al. [31] found no significant problems with orange trees when they were irrigated with treated sewage water. Furthermore, in Hamad et al. [32] study, toxicity problems for some metals (Cd, Hg, Cr, Pb) in tested crops due to irrigation with treated wastewater was not observed. In the Sultanate of Oman it was found that treated sewage water did not cause any phyto‐toxicity symptoms in date palm leaves and fruits [33]. Therefore, it can be concluded that proper management of wastewater irrigation and periodic monitoring of soil and plant quality parameters are required to ensure successful safe long‐term wastewater irrigation [34].

#### **3.4. Crop biological analysis**

**Treatment Element conc. (mg/l)**

64 Biological Wastewater Treatment and Resource Recovery

\*

*Source:* CPCB [30].

**Okra Mn Cd Fe Zn B Cr Co Pb Ni** 50%TWW 0.173d 0.001a 1.224c 0.357b 0.521c 0.060a 0.069a 0.252b 0.006b 100%GW 0.190c 0.001a 1.365b 0.364b 0.336d 0.071a 0.075a 0.229c 0.013b 75%TWW 0.242b 0.001a 2.372a 0.482a 0.745b 0.083a 0.087a 0.255b 0.127a 100%TWW 0.263a 0.001a 1.177d 0.495a 0.862a 0.057a 0.073a 0.300a 0.014b Sweet corn Mn Cd Fe Zn B Cr Co Pb Ni 50%TWW 0.177b 0.001a 1.295b 0.329c 0.073c 0.122b 0.091b 0.222c 0.068b 100%GW 0.204a 0.001a 1.582a 0.613a 0.492a 0.215a 0.100 a 0.191d 0.104a 75%TWW 0.152c 0.001a 1.584a 0.301c 0.122b 0.061c 0.070c 0.240b 0.011c 100%TWW 0.127d 0.001a 0.889c 0.400b 0.062c 0.064c 0.072c 0.444a 0.003d Maize Mn Cd Fe Zn B Cr Co Pb Ni 50%TWW 0.457b 0.001a 2.365b 0.256a 0.903a 0.136d 0.064a 0.210c 0.037a 100%GW 0.463a 0.001a 2.362b 0.219c 0.717c 0.146c 0.074a 0.213c 0.047a 75%TWW 0.366d 0.001a 2.279d 0.189d 0.454d 0.151b 0.075a 0.280a 0.040a 100%TWW 0.393c 0.001a 2.483a 0.226b 0.832b 0.181a 0.073a 0.241b 0.052a

Means in the column with same letter indicate no difference at Duncan's Multiple Range Test at *p* < 0.05.

.

**Standards/elements Cd Cu Pb Zn As Ni Cr** WHO/FAO (2007) 0.2 40 5 60 – – – European Union (EU 2006) 0.2 – 0.3 – 0.4 – 2.3 Indian Standard (Awashthi, 2000) 1.5 30 2.5 50 – 1.5 20

For soil, usually there is a direct relationship between salts found in the irrigation water and irrigated land. Whereas, for plants, root selectivity and present of salts in different forms could play a role in elements movement and translocation from soil to plant. From **Table 4**, it can be seen that concentrations of many elements were significantly (*p* < 0.05) different from one treatment to other. However, microelements in the edible parts of all crops grown in the field were not that high and they were within the international standards (**Table 5**). Same results were reported by Abdelrahman et al. [26] when they observed no significant difference between fresh and treated wastewater with regards to heavy metals accumulation in grown crops. Moreover, this finding was supported by Pescod [27] study, when he concluded that the concentrations of heavy metals in seeds were within normal level when treated wastewater effluent was used. Such results make it clear that heavy metal in soil are not readily bio‐ available for crop uptake and do not represent a threat to quality of crop consumption.

**Table 4.** Heavy metals concentration (mg/l) in tested crops\*

**Table 5.** Guideline for safe limits of heavy metals in plants (mg/kg).

Usually microbial analyses are the direct indicators for microbial contaminations in different crops. In this study, the edible part of grown crops was checked by Muscat Municipality laboratory and different microbes were analyzed such as coliform bacteria, *Escherichia coli*, and *Salmonella* spp. All tested samples were free from any microbial contamination. This finding was supported by Mexican and Tunisian studies where sewage effluent at different levels of treatment has been employed to irrigate various crops. It has been used with no serious effect on man and plants [35].

## **4. Conclusion**

The use of treated wastewater for irrigation is increasingly being considered as a technical solution to save fresh groundwater, minimize soil degradation, and improve soil fertility. In this study, usage of treated wastewater irrigation as compared to groundwater did not affect significantly some soil physical and chemical properties. Whereas some chemical properties such as major elements (N, K, Mg) and total carbon were significantly increased when treated wastewater was applied. Concentrations of heavy metals were increased in soils irrigated with treated wastewater compared to groundwater. The differences in heavy metals concentrations of all treatments were small and data of all treatments was close to each other.

Treated wastewater is a rich source of nutrients and provides most nutrients that are necessary for crop growth. Therefore, treated wastewater improved significantly plant productivity compared to groundwater treatments. Whereas small increase was noticed with some chemical properties of plants irrigated with treated wastewater compared to groundwater. However, all measured values were within the international standards. Biologically, all tested crops were free from any microbial contaminations. In general, most crops gave higher yield with wastewater irrigation and reduced the need for chemical fertilizers, resulting in net cost savings to farmers. Therefore, it can be concluded that treated wastewater is an important source of water for agricultural production and to avoid any health or environmental prob‐ lems, quality of treated wastewater should be monitored with time.

## **Acknowledgements**

The authors like to thank the staff from Sultan Qaboos University and Ministry of Agriculture and Fisheries for their support in collecting and analyzing soil, water, and plant samples. Special thanks to USAID (FABRI) for their financial support.

## **Author details**

Ahmed Al‐Busaidi\* and Mushtaque Ahmed

\*Address all correspondence to: ahmed99@squ.edu.om

Department of Soils, Water and Agricultural Engineering, College of Agricultural & Marine Sciences, Sultan Qaboos University, Muscat, Oman

## **References**


[4] Selem MM, Amir S, Abdel‐Aziz SM, Kandil MF, Mansour SF: Effect of irrigation with treated waste water on some chemical characteristics of soils and plants. Egypt. J. Soil Sci. 2000; 40 (1/2): 49–59.

wastewater was applied. Concentrations of heavy metals were increased in soils irrigated with treated wastewater compared to groundwater. The differences in heavy metals concentrations

Treated wastewater is a rich source of nutrients and provides most nutrients that are necessary for crop growth. Therefore, treated wastewater improved significantly plant productivity compared to groundwater treatments. Whereas small increase was noticed with some chemical properties of plants irrigated with treated wastewater compared to groundwater. However, all measured values were within the international standards. Biologically, all tested crops were free from any microbial contaminations. In general, most crops gave higher yield with wastewater irrigation and reduced the need for chemical fertilizers, resulting in net cost savings to farmers. Therefore, it can be concluded that treated wastewater is an important source of water for agricultural production and to avoid any health or environmental prob‐

The authors like to thank the staff from Sultan Qaboos University and Ministry of Agriculture and Fisheries for their support in collecting and analyzing soil, water, and plant samples.

Department of Soils, Water and Agricultural Engineering, College of Agricultural & Marine

[1] World Bank: Renewable Energy Desalination: An Emerging Solution to Close the Water Gap in the Middle East and North Africa. Washington, DC: World Bank. 2012.

[2] Mujeriego R: Salal Golf course irrigation with reclaimed waste water. Sci. Technol. 1991;

[3] Levine A, Sanot A: Recovering sustainable water from waste water. Environ sci.

of all treatments were small and data of all treatments was close to each other.

lems, quality of treated wastewater should be monitored with time.

Special thanks to USAID (FABRI) for their financial support.

and Mushtaque Ahmed

\*Address all correspondence to: ahmed99@squ.edu.om

Sciences, Sultan Qaboos University, Muscat, Oman

Technol. 2004; 38 (11): 201A–208A.

**Acknowledgements**

66 Biological Wastewater Treatment and Resource Recovery

**Author details**

Ahmed Al‐Busaidi\*

**References**

24 (9): 161–172.


[32] Hamad I, Nizam AA, Suleiman MS: Biological detection of contamination in vegetables irrigated with brackish and underground water. In: Proceedings of the Arab Regional Workshop on the Use of Saline, Brackish, and Treated Sewage Effluent Water in Irrigated Agriculture, Muscat, Sultanate of Oman. 1995. pp. 481–531.

[18] Bansal RL, Nayyar VK, Takkar PN: Accumulation and bioavailability of Zn, Cu, Mn and Fe in soils polluted with industrial waste water. J. Indian Soc. Soil Sci. 1992; 39: 795–

[19] Palaniswami C, Sree Ramulu US: Effects of continuous irrigation with paper factory

[20] Rattan RK, Datta SP, Chhankar PK, Suribabu K, Singh AK: Long term impact of irrigation with sewage effluents on heavy metal contents in soils, crops and ground‐

[21] Berry WL, Wallace A, Lunt OR: Ultilization of municipal wastewater for culture of

[22] Abohassan AA, Kherallah IE, Kandeel SA: Effect of sewage effluent irrigation regimes on wood quality of *Presopi juliflera* grown in Riyadh Region. Arab Gulf J. Sci. Res. 1988;

[23] Stewart HTL, Hopmans P, Flinn DW: Nutrient accumulation in trees and soil following irrigation with municipal effluent. Australian Environ. Pollut. 1990; 63(2): 155–177.

[24] Shafiq M, Hassan I, Hassan Z: Influence of irrigation methods on the productivity of summer maize under saline/sodic environment. Asian Journal of Plant Sciences. 2002;

[25] Harati M: Study on heavy metal accumulation in different parts of corn irrigated by

[26] Abdelrahman HA, Alkhamisi SA, Ahmed M, Ali H: Effects of Treated Wastewater Irrigation on Element Concentrations in Soil and Maize Plants. Conference Proceeding.

[27] Pescod M: Wastewater treatment and use in agriculture. Bull. FAO #47 (125) (Rome).

[28] Izzo K, Scagnozzi A, Belligno A, Navari‐Izzo F: Influence of NaCl treatment or Ca, K, Na interrelations in maize shoots. In: M.A.C. Frageso and M.L. Beusichem (Eds.), Optimization of Plant Nutrition. Kluwer Academic Publishers, Dordrecht, The

[29] Mahdi OE, Salama SB, Consolacion EC, Al‐Solomi M: Effect of treated sewage water on the concentration of certain nutrient elements in date palm leaves and fruits.

[30] CPCB (Central Pollution Control Board): Parivesh, Newsletter from CPCB. 2002.

[31] Omran MS, Waly TM, Abo‐Eleinain EM, Elnashar BMB: Effect of sewage irrigation on yield, tree components, and heavy metals accumulation in navel orange trees. Biol.

Available at /http://www.cpcb.delhi.nic.in/legislation/ch15dec02a.htms

sewage in south of Tehran. MSc Thesis. Tehran University. 2003.

effluent on soil properties. J. Indian. Soc. Soil Sci. 1994; 42: 139–140.

water—a case study. Agric. Ecosyst. Environ. 2004; 109: 310–322.

horticultural crops. Hort. Sci. 1980; 15: 169–171.

Sultan Qaboos University, Oman. 2011.

Netherlands. 1993, pp. 577–582.

Wastes BIWAED. 1988; 23: 17–24.

Commun. Soil Sci. Plant Anal. 1998; 29: 5–6.

799.

68 Biological Wastewater Treatment and Resource Recovery

136 (l): 45–53.

1 (6): 678–680.

1992.


**Hazards and Treatment of Organic Compounds in Wastewater**

## **Spreading of Antibiotic Resistance with Wastewater Spreading of Antibiotic Resistance with Wastewater**

Sadik Dincer and Esra Sunduz Yigittekin Sadik Dincer and Esra Sunduz Yigittekin

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

http://dx.doi.org/10.5772/66188

#### **Abstract**

The recent statistics show that the world's population is rapidly increasing. This increase negatively affects the water resources and it increases the water demand progressively. Along with the increase in the world's population, the insensible use of water resources, pollution, and drought lead to the increasing reduction of water resources. Due to these factors, all countries, primarily developed countries, have started looking for new water resources. This search has been extended to extraterrestrial water. However, the exist‐ ing technology and opportunities direct countries toward the purification of wastewater rather than searching for new water resources. For the reasons outlined above, purifica‐ tion and recycling of wastewater become important. In addition to the natural resistance of microorganisms against antibiotics, a resistance also arises because of the unconscious and overuse of antibiotics. This resistance spreads through wastewater progressively. Antibiotic resistance shows an increase according to the scientific data. In order to prevent the resistance, it is of capital importance to treat the wastewater in which the domestic pollution burden is high. In this study, the role of domestic wastewater in the occurrence and spread of antibiotic resistance will be revealed.

**Keywords:** antibiotics, spreading antibiotic resistance, water, wastewater, domestic wastewater

### **1. Introduction**

Water contains millions of microscopic living beings within itself. The plenty amount of water is accessible on our planet for living beings to maintain their vital activities.

Along with 14 billion m<sup>3</sup> of water, 97.5% of it is salty water, 2.6% of it is freshwater, and 0.8% of the total amount of water is present as freshwater in the state of constant vaporization, pre‐ cipitation, and flow. Water scarcity is indicated as one of the main problems of the twenty‐first century in the whole world, and for this reason, lives of many people depend on the right usage

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

of water. People need water primarily for civic, industrial, and agricultural areas. However, water is regarded as a limited source. For the fact that water resources become insufficient and decrease in quality creates serious concerns. The population increase, urbanization, agri‐ cultural practices, and industrialization increase the water demand. Wastewater treatment is built for the purpose of reducing the pollution by removing pathogens, nutrients, and biode‐ gradable substances, and protecting public health and the environment. Furthermore, with the increase in water demand, the recycling of wastewater has been brought into question [1].

Wastewaters are divided into two groups such as domestic and industrial. Domestic waste‐ water can originate from house, workplace, and hospital because of its content. The complete treatment of this type of wastewater is impossible even if it goes through many stages. This situation causes many problems. The emergence and spread of antibiotic‐resistant bacteria are the leading reasons for this problem and they make humans and animals sick.

According to American Centers for Disease Control and Prevention (CDC), antibiotic resistance constitutes one of the most important health problems of a country. In America, it is estimated that 2 million people become sick and 700,000 people die worldwide because of resistant bacte‐ ria. In a report of 2013, CDC indicated that the usage of antibiotics in the production of food ani‐ mals causes the emergence of resistant *Campylobacter* that is contagious to humans. Resistance genes can be transferred between zoonotic bacteria types, among the bacteria species, through food chain and contact with feces of ill animals and contaminated environment [2].

Antimicrobial resistance causes many problems in humans and animals in the case of the spread through wastewater, spread wastewater treatment output's being low, and the usage of these waters in agricultural practices and irrigation fields, the emergence of many antibi‐ otic‐resistant microorganisms.

## **2. Significance of water**

Water is an essential substance which is necessary for vital activities such as nourishment, circulation, respiration, excretion, and reproduction to occur in every period of human life. At the same time, water itself is a habitat as one of the basic elements in nature while forming a habitat. The presence of water in a habitat and its quality are extremely important for life [3].

As the most important one of all natural sources for all living beings, water is a habitat and it contains millions of microscopic living beings. It constitutes approximately three‐quarters of the Earth [4].

Water is crucial for the life of living beings as the most common natural resource on the Earth. Seventy‐five percent of the Earth's surface, seventy percent of the human body, and seventy‐ eight percent of the blood consist of water [5].

Ninety‐seven percent of the water body on the Earth consists of oceans and seas, two per‐ cent of lakes, rivers, and underground waters, and one percent of glaciers and snows. Water has been used during the development of civilizations for many purposes such as personal hygiene, agricultural irrigation, industrial production, and electric power production [6].

Water submerges more than 70% of the Earth as fresh and salty water and these environments are defined as aquatic environments [7]. Salty water constitutes more than 96% of the water on the Earth. More than 68% of the present freshwater is found in ice and glaciers. In this way, it is considered that water stays in stock. Thirty percent of freshwater consists of underground water. Two‐thirds of underground waters are located deeper than 800 m. Surface freshwater sources such as rivers and lakes constitute 93,100 km<sup>3</sup> (22,300 cubic miles), which is 1/700 of 1% of all water on the Earth [8].

The minerals, salts, and sulfates contained in the water are very important along with its other characteristics. The presence of these substances in a certain amount in water is essential for life while their presence in small or greater quantities continually affects life in a negative way. At the same time, water is a habitat. The pollution of this environment creates danger for life [9, 10].

An increase in industrialization, urbanization, and population that started at the beginning of the twentieth century, and an increase in the use of natural resources have caused the emer‐ gence of the problems called the environmental pollution jeopardizing human life. An increase in the variety and amount of solid and liquid wastes that are disposed into the environment causes air and water pollution [6].

Water is a component which is significant for the life cycle of all living beings on the Earth. While three‐quarters of the Earth is covered by water, two‐thirds of the human body is cov‐ ered by water. This rate significantly affects all living beings while being important for both the Earth and human beings. Water has many important roles from systems in the living organism to cellular functions. Even a small decrease in the amount of water can endanger life.

Since the existence of the Earth, all civilizations have settled in the places around or close to water. This shows us that water is a functional substance which is completely life‐oriented. The decrease in the amount of water arises from both the environmental pollution and unconscious consump‐ tion. Because of this decrease, countries are in search of new sources and wars break out.

## **3. Water consumption**

of water. People need water primarily for civic, industrial, and agricultural areas. However, water is regarded as a limited source. For the fact that water resources become insufficient and decrease in quality creates serious concerns. The population increase, urbanization, agri‐ cultural practices, and industrialization increase the water demand. Wastewater treatment is built for the purpose of reducing the pollution by removing pathogens, nutrients, and biode‐ gradable substances, and protecting public health and the environment. Furthermore, with the increase in water demand, the recycling of wastewater has been brought into question [1]. Wastewaters are divided into two groups such as domestic and industrial. Domestic waste‐ water can originate from house, workplace, and hospital because of its content. The complete treatment of this type of wastewater is impossible even if it goes through many stages. This situation causes many problems. The emergence and spread of antibiotic‐resistant bacteria

are the leading reasons for this problem and they make humans and animals sick.

food chain and contact with feces of ill animals and contaminated environment [2].

otic‐resistant microorganisms.

74 Biological Wastewater Treatment and Resource Recovery

**2. Significance of water**

eight percent of the blood consist of water [5].

the Earth [4].

According to American Centers for Disease Control and Prevention (CDC), antibiotic resistance constitutes one of the most important health problems of a country. In America, it is estimated that 2 million people become sick and 700,000 people die worldwide because of resistant bacte‐ ria. In a report of 2013, CDC indicated that the usage of antibiotics in the production of food ani‐ mals causes the emergence of resistant *Campylobacter* that is contagious to humans. Resistance genes can be transferred between zoonotic bacteria types, among the bacteria species, through

Antimicrobial resistance causes many problems in humans and animals in the case of the spread through wastewater, spread wastewater treatment output's being low, and the usage of these waters in agricultural practices and irrigation fields, the emergence of many antibi‐

Water is an essential substance which is necessary for vital activities such as nourishment, circulation, respiration, excretion, and reproduction to occur in every period of human life. At the same time, water itself is a habitat as one of the basic elements in nature while forming a habitat. The presence of water in a habitat and its quality are extremely important for life [3]. As the most important one of all natural sources for all living beings, water is a habitat and it contains millions of microscopic living beings. It constitutes approximately three‐quarters of

Water is crucial for the life of living beings as the most common natural resource on the Earth. Seventy‐five percent of the Earth's surface, seventy percent of the human body, and seventy‐

Ninety‐seven percent of the water body on the Earth consists of oceans and seas, two per‐ cent of lakes, rivers, and underground waters, and one percent of glaciers and snows. Water has been used during the development of civilizations for many purposes such as personal hygiene, agricultural irrigation, industrial production, and electric power production [6].

Population increase on the Earth leads to the decrease of water and pollution of clean and potable water. This will cause water scarcity in the future. Rivers and lakes constitute most of the water that people use daily. The pollution of these water sources will create water short‐ age. The amount of water that meets our needs is 0.25% of all water sources on the Earth [11].

About 97.39% of 1384.109 km<sup>3</sup> water on the Earth is found in the oceans and seas. The remain‐ ing 2.01% consist of glaciers and 0.60% consists of underground water, lakes, and rivers. This situation shows that the available freshwater supply constitutes quite a small amount of all water sources on the Earth [12].

The world's population that is approximately 6 billion is able to use 54% of the renewable surface and underground water supplies. It is considered that this rate will increase to 70% with the population's increase as the conditions of use remain the same. At the same time, it is estimated that 90% of the present freshwater sources will be used with the increase of life standards and the increase of water usage per person. For other living beings, there will be 10% of the available water supply. It is indicated that there will not be enough water for environmental and ecological functions because of the population increase and unconscious use of water resources [13].

Water resources are also used in a sectorial aspect besides meeting daily needs. The usage of water is classified as agricultural, industrial, and domestic sectors [14].

Sixty‐nine percent of the freshwater resources on the Earth are used for agriculture, twenty‐ three percent are used for industry, and eight percent are used for domestic purposes. These rates differ from continent to continent. For instance, while the rates of agricultural, indus‐ trial, and domestic usage of water in Africa are 88, 5, and 7%, respectively, these rates in Europe are 33, 54 and 13%, respectively [15].

Water consumption in the world has increased 10‐fold since 1900. In the studies conducted, it is determined that water consumption will increase 17% in agriculture, 20% in industry, and 70% in domestic consumption in 2015. Moreover, it is told that 20% of 6 billion world popula‐ tion is deprived of clean water resources. The water amount per person decreased to 7300 m<sup>3</sup> in 2000 while it was 16,800 m<sup>3</sup> in 1950 [16–18]

While it is estimated that world population will be 8 billion, it is considered that water consump‐ tion per person will decrease to 4800 m<sup>3</sup> in 2025. This decrease in consumption will arise from water resources shortage. Furthermore, the present available water resources will be polluted in 2025 and then water will not be provided [19]. It is estimated that the curve of the increase in water demand and the curve of the decrease of clean water resources will intersect in 2023.

Recent studies show that population growth will increase the consumption of water. However, this situation is inversely proportional to the number of water resources. Due to the decrease of clean and available water resources, the quantity of water per person has decreased. The reason for these situations is water scarcity which arises from the unconscious use and pollution.

## **4. Water pollution**

Water pollution has a negative effect on public health and ecology because of the degradation of water quality and natural balance. Water pollutants contain surplus metal, some radio‐ active isotopes, nitrogen, phosphorus, sodium, and other beneficial and necessary elements along with especially some faecal originated pathogenic bacteria, parasites, and viruses which can be human or animal originated.

Mixing of any organic, inorganic, radioactive, or biologic substance that inhibits or disturbs the usage of water resources by impairing their quality into water is called water pollution [20, 21].

The reasons of water pollution are particularly domestic, industrial, agricultural, physical, chemical, radioactive, and microbial pollution.

Domestic wastewater contains organic and inorganic substances that are suspended, colloi‐ dal, and dissolved. Domestic wastewater consists of organic foods such as too much carbon, nitrogen, and phosphorus and highly concentrated microorganisms [22]. With the increase in urbanization, the flow of domestic waste into water through sewerage system also increased. In particular, detergents which are used in washing machines, oils poured out into lavabo, and the dispersion of wastes that should be accumulated in dustbins and recycled into the environment cause water pollution [23]. The characteristics of industrial wastewater differ from industry to industry [22].

it is estimated that 90% of the present freshwater sources will be used with the increase of life standards and the increase of water usage per person. For other living beings, there will be 10% of the available water supply. It is indicated that there will not be enough water for environmental and ecological functions because of the population increase and unconscious

Water resources are also used in a sectorial aspect besides meeting daily needs. The usage of

Sixty‐nine percent of the freshwater resources on the Earth are used for agriculture, twenty‐ three percent are used for industry, and eight percent are used for domestic purposes. These rates differ from continent to continent. For instance, while the rates of agricultural, indus‐ trial, and domestic usage of water in Africa are 88, 5, and 7%, respectively, these rates in

Water consumption in the world has increased 10‐fold since 1900. In the studies conducted, it is determined that water consumption will increase 17% in agriculture, 20% in industry, and 70% in domestic consumption in 2015. Moreover, it is told that 20% of 6 billion world popula‐ tion is deprived of clean water resources. The water amount per person decreased to 7300 m<sup>3</sup>

While it is estimated that world population will be 8 billion, it is considered that water consump‐

water resources shortage. Furthermore, the present available water resources will be polluted in 2025 and then water will not be provided [19]. It is estimated that the curve of the increase in water demand and the curve of the decrease of clean water resources will intersect in 2023. Recent studies show that population growth will increase the consumption of water. However, this situation is inversely proportional to the number of water resources. Due to the decrease of clean and available water resources, the quantity of water per person has decreased. The reason for these situations is water scarcity which arises from the unconscious

Water pollution has a negative effect on public health and ecology because of the degradation of water quality and natural balance. Water pollutants contain surplus metal, some radio‐ active isotopes, nitrogen, phosphorus, sodium, and other beneficial and necessary elements along with especially some faecal originated pathogenic bacteria, parasites, and viruses which

Mixing of any organic, inorganic, radioactive, or biologic substance that inhibits or disturbs the usage of water resources by impairing their quality into water is called water pollution [20, 21]. The reasons of water pollution are particularly domestic, industrial, agricultural, physical,

in 2025. This decrease in consumption will arise from

water is classified as agricultural, industrial, and domestic sectors [14].

in 1950 [16–18]

use of water resources [13].

76 Biological Wastewater Treatment and Resource Recovery

in 2000 while it was 16,800 m<sup>3</sup>

use and pollution.

**4. Water pollution**

can be human or animal originated.

chemical, radioactive, and microbial pollution.

tion per person will decrease to 4800 m<sup>3</sup>

Europe are 33, 54 and 13%, respectively [15].

Apart from domestic and industrial wastewaters that are discharged into water sources with‐ out being treated, the unconscious fertilization and unconscious usage of agricultural pesti‐ cides are also the reasons for pollution. These pollutants become crucial with negative effects on the water resources regarded as inadequate according to the world average, environmen‐ tal, and public health and in terms of economy [6]. In the fields close to water, the incorrect ploughing mixes into water through the wind and causes pollution in water [3]. An increase in the usage of synthetic manure and pesticides in agriculture and industry and chemical substances that are used in the industry create a risk of water pollution [23].

Industrial establishments cause physical pollution. Power plants, steel, paper, car, plastic, and packing factories, which are big industrial establishments, throw environmentally hazardous solid and liquid substances. These substances are mostly toxic like arsenic, phenol, cyanide, chromium, and cadmium [24].

The chemical pollution of water began to cause critical health problems. It is estimated that in the future one of the most important water pollution problems will be the pollution caused by chemicals. Main metals causing chemical pollution are copper, zinc, mercury, nitrate, and phosphate [23].

Radioactive pollution in water can result from research agencies, hospitals, and some indus‐ trial fields. Radioactivity increases because of testing nuclear weapons. Therefore, rain water is getting dirty and, as a consequence, surface water is exposed to radioactive pollution [25].

Water might also be polluted by some pathogenic bacteria, parasites, and viruses that can originate from humans and animals [6]. Microorganisms that cause water pollution in terms of hygiene generally originate from diseases or human and animal excrements and urine being a porter [26]. An increase in bacteria population leads to bacterial pollution as a result of the decomposition of organic substances that are accumulated in the sea and inland waters or mixing of various pollutants apart from sewerage [25]. In stored waters, there is a lot of bacterial communities like the members of *Pseudomanas* sp.*, Micrococcus* sp.*, Achromobacter* sp.*, Streptomyces* sp., and especially *Enterobacteriaceae.* The members of *Enterobacteriaceae* do not reproduce in clean water and their natural habitat is not potable water. Coliform bacteria are important in terms of human health and as an indicator of water pollution [24]. Their pres‐ ence indicates that stool is mixed up with water through sewage directly or indirectly in one or more phases starting with the raw material to the transfer of water [27].

When it comes to water pollution, microbial pollution comes to mind first, even though there are many reasons of pollution. The most important of them is the pollution that arises from domestic and industrial wastes. Since domestic wastewater contains sewage waters and detergents, it also causes the indirect microbial and chemical pollution.

## **5. Wastewaters and classification**

Wastewaters are formed as a result of the pollution of water used in households and indus‐ trial establishments [28].

For the waters that are disposed by being used in households or industry, "wastewater" defi‐ nition is used. Wastewaters demonstrate biological, chemical, and physical pollution. While biological pollution consists of bacteria, fungi, parasites, and virus particles, and chemical pollution consists of toxic substances, decomposed organic substances, and phosphor, physi‐ cal pollution consists of color, scent, foaming, temperature increase, and suspended matters. Heavy metals contain colorants that belong to the group of chemical pollutants and include industrial wastes and some pesticides [29].

#### **5.1. Classification of wastewaters**

Wastewaters are classified into two groups as domestic and industrial.

#### *5.1.1. Domestic wastewaters*

Wastewaters that originate from the dirty water from households and workplaces and do not include the industrial content of factories are called domestic wastewaters. Although their pollution rate is low they contain a high level of oily compounds, proteins, particles, chemical oxygen demand (COD), and detergents. For this reason, domestic wastewaters have a com‐ plex structure (**Table 1**) [30].

Domestic wastewaters are the waters that contain dirty looking and colorful soluble and insol‐ uble matters from food wastes, kitchen lavabos, bathrooms, washing, and dishing machines and the matters that have organic and inorganic content and 99% of water [31].


**Table 1.** Physical, chemical, and biological components of domestic wastewater [32].

Domestic wastewaters contain suspended, colloidal, and dissolved organic and inorganic sub‐ stances. As well as this pollution arises from sewerages and detergents, it can also originate primarily from households and business enterprises. Moreover, domestic wastewaters con‐ tain pathogenic microorganisms such as bacteria, helminth, protozoa, and viruses. This situa‐ tion increases the pollution rate of waters and indicates that water treatment is absolute. The indicator of the treatment's necessity is that some bacteria include R‐plasmid. Since R‐plas‐ mid ensures antibiotic resistance to bacteria, untreated domestic wastewaters cause antibi‐ otic‐resistant bacteria to infect people and animals and create disease.

One of the most significant wastewaters that belong to domestic wastewaters is hospital‐ acquired wastewaters.

#### *5.1.1.1. Hospital‐acquired wastewaters*

domestic and industrial wastes. Since domestic wastewater contains sewage waters and

Wastewaters are formed as a result of the pollution of water used in households and indus‐

For the waters that are disposed by being used in households or industry, "wastewater" defi‐ nition is used. Wastewaters demonstrate biological, chemical, and physical pollution. While biological pollution consists of bacteria, fungi, parasites, and virus particles, and chemical pollution consists of toxic substances, decomposed organic substances, and phosphor, physi‐ cal pollution consists of color, scent, foaming, temperature increase, and suspended matters. Heavy metals contain colorants that belong to the group of chemical pollutants and include

Wastewaters that originate from the dirty water from households and workplaces and do not include the industrial content of factories are called domestic wastewaters. Although their pollution rate is low they contain a high level of oily compounds, proteins, particles, chemical oxygen demand (COD), and detergents. For this reason, domestic wastewaters have a com‐

Domestic wastewaters are the waters that contain dirty looking and colorful soluble and insol‐ uble matters from food wastes, kitchen lavabos, bathrooms, washing, and dishing machines

**Physical properties Chemical components Biological components**

and the matters that have organic and inorganic content and 99% of water [31].

**Organics Inorganics Gases** Solid matters Carbohydrates pH Methane Living cells Heat Oil and grease Nitrogen Oxygen Plants Color Pesticides Phosphorus Hydrogen Single cells Smell Phenols Alkalinity Sulfur Viruses

Sulfur

Toxic components

Proteins Chlorides Surface active agent Heavy metals

**Table 1.** Physical, chemical, and biological components of domestic wastewater [32].

detergents, it also causes the indirect microbial and chemical pollution.

Wastewaters are classified into two groups as domestic and industrial.

**5. Wastewaters and classification**

78 Biological Wastewater Treatment and Resource Recovery

industrial wastes and some pesticides [29].

**5.1. Classification of wastewaters**

*5.1.1. Domestic wastewaters*

plex structure (**Table 1**) [30].

trial establishments [28].

Hospital wastewaters contain micro and macro pollutants that come from various sources such as operating rooms, laboratories, investigation units, polyclinics, and drug use. The most important macro pollutants are bacteria and viruses while the most important micro pol‐ lutants are antibiotics, heavy metals (Hg, Pt, Gd, etc.), hormones and detergents/antiseptics. While microbiologic quality is determined for the usage of water, faecal pollution is identi‐ fied by biologic and chemical indicators. In the content of biological indicators, there are total coliforms, faecal coliforms, faecal streptococcus, and *Clostridium perfringens*. Total coliforms are in the form of aerobic and facultative anaerobic, asporogenic and Gram‐negative bacte‐ ria. Faecal coliforms, the marker of the pollution of water in which faecal coliforms and total coliform bacteria are found and which indicates the presence of pathogenic bacteria with human or animal excrements represent the presence of pathogenic bacteria and limited virus contamination. In hospital wastewaters, antibiotics such as ciprofloxacin, erythromycin, and sulfamethoxazole are found in high numbers in accredited adsorbable organic halide (AOX) and paracetamol. In municipal sewage, antibiotics such as ofloxacin and erythromycin are found in high numbers in AOX, paracetamol, and ibuprofen which is an analgesic [33].

In order for drugs to be stored longer and be easier to take, they must be quite durable and of high mobility quality in the liquid‐phase while being produced. For this reason, active substances in drugs and biotransformation products lead to various factors by accumulating in the ecosystem. A lot of drugs such as antibacterial drugs, antibiotics, antifebrile, anodyne, synthetic steroids, cholesterol medicines, beta blockers and cytostatic drugs are the drugs detected in the ecosystem by studies performed [34].

Various drugs are used for various purposes during the treatment, protection, and development of human and animal diseases. These drugs cannot completely metabolize and they are removed from the body as they are or as a by‐product in the form of ordure, urine, sweat, etc. [35].

In order for living beings to be treated, protected against microorganisms and infections and become resistant, many drugs should be taken. After the functions of these drugs in the body are over, they are removed through liver and kidneys. Medicine taken reaches the maximum level in the blood and when it starts to decrease the excretion also begins. While the excretion periods of drugs such as painkillers and antibiotics out of body are different, antibiotics are not removed for a long time. Drugs are removed from the body as urine, ordure, or metabolized product. In this way, they mix into wastewaters through the sewer system.

Drugs mix into wastewaters not only through excretion. With the disposal of unused drugs in households and hospitals, they also mix into wastewaters through the sewage.

The medicines found in wastewaters cannot be completely refined through the refinement. One of the biggest reasons for this situation is that hospital wastewaters directly mix with domestic waters without pretreatment. This affects primarily potable waters, underground waters, lakes, and rivers in a negative way.

Medicine remnants that mix into the potable water as a result of the inadequate refinement of domestic wastewaters negatively affect living beings in many ways. This effect arises espe‐ cially from antibiotics. Antibiotics that enter the body through water cause pathogenic micro‐ organisms to become resistant.

Due to the negative outcomes on living beings, hospital wastewaters should be refined before they are transferred to domestic wastewaters.

#### *5.1.2. Industrial wastewaters*

Pollution in the environment that originates from unavailable or economically unvalued wastes in the industrial system is called industrial pollution. The accumulation of permanent and toxic organic substances in industrial wastewaters creates serious problems. The facts that these wastewaters are not discharged into the receiving environment, pollutants are not biodegraded, and they have a toxic influence upon living beings create many troubles [36].

Industrial wastewaters comprise of various resources such as refrigerant waters, process wastewaters, and domestic qualified wastewaters. Because of this content, the refinement of industrial wastewaters becomes crucial [11].

Since industrial wastewaters contain heavy metal content, they create the most crucial envi‐ ronmental problem of the present day. Wastewaters containing heavy metals are the waters that are generally acidic and have a low biochemical oxygen demand (BOD) value. Aquatic life is affected by mixing of wastewaters into the receiving environment. Because of this situ‐ ation, expensive refinement systems are needed in order to use water resources as potable water sources. Heavy metals contained in wastewaters make the mud impossible to use for agricultural purposes by affecting the refinement efficiency of domestic wastewaters. For this reason, the discharge of industrial wastewaters with heavy metal content into the sewer sys‐ tem has an important role [37].

#### **5.2. Wastewater treatment**

The treatment of domestic wastewaters takes place in three stages as mechanic, biologic, and chemical.

Physical treatment covers the refinement of solid matters in wastewaters. This treatment stage comprises of four units as grid/sieve/grinder, sand catcher/oil slinger, preliminary settling, and flotation [38].

The biologic treatment contains the stage in which organic matters contained in waste‐ waters are refined. This treatment happens along with the decrease in the organic matter amount by using and decomposing of organic substances as a nutrition substance by micro‐ organisms. Domestic wastewater generally decreases nutrition and organic substances such as nitrogen and phosphorus contained in it. Biologic treatment helps microorgan‐ isms such as fungi, algae, protozoa, and metazoans and organisms belonging to bacteria and archaea. The most used processes in biologic treatment are activated sludge processes, air‐conditioned lagoons, trickling filters, revolving biodiscs, and stabilization pools. Basic operations of this treatment are nitrification, denitrification, dephosphorization, waste sta‐ bilization and eliminating organics which are measured especially as BOD<sup>5</sup> and COD in wastewater [39, 40, 38].

different, antibiotics are not removed for a long time. Drugs are removed from the body as urine, ordure, or metabolized product. In this way, they mix into wastewaters through

Drugs mix into wastewaters not only through excretion. With the disposal of unused drugs in

The medicines found in wastewaters cannot be completely refined through the refinement. One of the biggest reasons for this situation is that hospital wastewaters directly mix with domestic waters without pretreatment. This affects primarily potable waters, underground

Medicine remnants that mix into the potable water as a result of the inadequate refinement of domestic wastewaters negatively affect living beings in many ways. This effect arises espe‐ cially from antibiotics. Antibiotics that enter the body through water cause pathogenic micro‐

Due to the negative outcomes on living beings, hospital wastewaters should be refined before

Pollution in the environment that originates from unavailable or economically unvalued wastes in the industrial system is called industrial pollution. The accumulation of permanent and toxic organic substances in industrial wastewaters creates serious problems. The facts that these wastewaters are not discharged into the receiving environment, pollutants are not biodegraded, and they have a toxic influence upon living beings create many troubles [36]. Industrial wastewaters comprise of various resources such as refrigerant waters, process wastewaters, and domestic qualified wastewaters. Because of this content, the refinement of

Since industrial wastewaters contain heavy metal content, they create the most crucial envi‐ ronmental problem of the present day. Wastewaters containing heavy metals are the waters that are generally acidic and have a low biochemical oxygen demand (BOD) value. Aquatic life is affected by mixing of wastewaters into the receiving environment. Because of this situ‐ ation, expensive refinement systems are needed in order to use water resources as potable water sources. Heavy metals contained in wastewaters make the mud impossible to use for agricultural purposes by affecting the refinement efficiency of domestic wastewaters. For this reason, the discharge of industrial wastewaters with heavy metal content into the sewer sys‐

The treatment of domestic wastewaters takes place in three stages as mechanic, biologic, and

Physical treatment covers the refinement of solid matters in wastewaters. This treatment stage comprises of four units as grid/sieve/grinder, sand catcher/oil slinger, preliminary settling,

households and hospitals, they also mix into wastewaters through the sewage.

the sewer system.

waters, lakes, and rivers in a negative way.

80 Biological Wastewater Treatment and Resource Recovery

they are transferred to domestic wastewaters.

industrial wastewaters becomes crucial [11].

tem has an important role [37].

**5.2. Wastewater treatment**

chemical.

and flotation [38].

organisms to become resistant.

*5.1.2. Industrial wastewaters*

Micropollutants that are common in water resources cannot be effectively eliminated with the present treatment systems and environmental impacts in the receiving environment. In par‐ ticular, antibiotics and pharmaceutics are released into the environment after their production and consumption so they create a threat in the receiving environment. Conventional treat‐ ment processes, primarily biologic treatment systems, remain insufficient in the elimination of antibiotics. In order to remove antibiotics that are resistant to biodegradation, advanced oxidation processes with a high oxidation potential should be used [41].

In wastewater treatment establishments, antibiotics are generally removed from the envi‐ ronment by biodegradation and sorption with activated sludge. Antibiotic‐resistant bacteria spread in nature thanks to the removing methods of activated sludge containing antibiotic‐ resistant organisms such as agricultural practices or burying into the pit [42].

Antibiotics are used to help the growth of animals along with the treatment of human and animal diseases. Antibiotics that enter the body are removed without being metabolized at the rates reaching 90%. For this reason, the main source of antibiotic pollution in nature is the antibiotics in human and animal faeces. In recent studies, it is determined that antibiotics are found in animal faeces and domestic wastewater sewage sludge besides various com‐ partments. By regarding physical and chemical properties, antibiotics can reach sediments, soil, and underground water. It is determined that conventional treatment methods remain insufficient in the removal of low concentrated antibiotics in water. The high concentration of antibiotics in the environment causes the degradation of ecological balance by creating a toxic effect on microorganisms, and their low concentration causes pathogenic and nonpathogenic bacteria to gain antibiotic resistance. For this reason, in order to remove antibiotic pollution, alternative treatment methods are necessary [43].

Gao et al. [44] determined 14 antibiotics in total in the wastewater, and 18 antibiotics in the activated sludge in their study. In the activated sludge, floroquinons, and ofloxacin were determined at the highest rate. Wastewater treatment establishments cannot remove antibiotics completely and the removal rate ranges from 34 to 72%. The amount of anti‐ biotics in water is determined to be higher in winter months in comparison with spring and fall months. At the same time, antibiotic remnants have an adverse effect on very dif‐ ferent organisms in nature (they encourage reproduction). Because of the low treatment effect, wastewater treatment establishments are the major source of antibiotics in aquatic environments.

According to Li et al. [45], the removal activity of target antibiotics from water changed between 32 and 78% through conventional treatment. With the advanced treatment methods, the removal rate of target antibiotics became 85–100% and pollution probability of antibiotics decreased. In addition to this, in the risk assessment, the effects of ofloxacin and erythromycin on microorganisms in water are investigated by refining it. The majority of antibiotics cannot be absorbed or metabolized in the body. Moreover, the large part pass into the sewage system through urine and faeces and it comprises a significant part of the antibiotic source in nature.

In the study conducted by Zhang et al. [46], the elimination mechanism of three β‐lactam, two fluoroquinolones, and two macrolide antibiotics was investigated in the wastewater treatment establishment, which has four different treatment methods among six wastewater treatment establishments in China, Dalian. In this study, fluoroquinolones and macrolide antibiotics were determined as dominant antibiotics at the exit of wastewater treatment establishment and in coastal waters. It is revealed that β‐lactams are removed through biodegradation, for fluoroquinolones pretreatment is more effective than biologic treatment, and macrolide con‐ centration increases dramatically after biological treatment. The reason for this is that macro‐ lides that are covered by faeces particles are revealed [46].

Xu et al. [47] examined antibiotics and their resistant genes in a water treatment establishment in Beijing, China and the situation of the river into which water was discharged in their study. A total of 13 antibiotic resistance genes (ARGs) were examined. SuI‐arg was found at the highest rate among all antibiotic resistance genes (ARGs). ARG quantity in the wastewater treatment establishment is higher than in the river. According to the correlation analysis, there is a posi‐ tive relationship between tetracyclines and tetargs in water. This correlation could not be per‐ formed between suI‐args and sulfonamides. A negative relationship was observed between the concentration of quinolone genes and enrofloxacin. When ARG abundance of the waters that are treated in the treatment establishment is examined, treatment establishment causes resistant genes to increase. Results show that treatment establishments have a function of a warehouse for resistance genes. As a result, treated water needs advanced treatment before it is sent to the natural aquatic environment. In the study, three antibiotic groups were studied as tetracycline, sulfonamides, and quinones that are known for their permanence in the aquatic environment. Tetracyclines are removed at the rate of 87.9% in sludge elimination establishments. In the elimi‐ nation of tetracycline, biodegradation, and adsorption have an important role. In sludge elimi‐ nation establishments, teta, tetm, tetw, and teto genes are the ones that are mostly found [47].

### **6. Antibiotics**

Antibiotics are bioactive substances that kill or stunt the growth of the microorganism and have a high effect on synthetic or biological origin [48].

Antibiotics that are naturally obtained from plants and their extracts and are used for medical purposes have been brought into use as a result of Paul Ehrlich's studies in 1908. Paul Ehrlich revealed some chemical substances that are harmful to some bacteria and are less harmful to the host cells by investigating them [49].

Antibiotics are produced in nature by bacteria or fungi. The production of antibiotics by these living beings and their release into the environment result from their food competition with other species. Therefore, they produce antibiotics in their environment which extinguish other microorganisms or inhibit their growth. Antibiotics do not affect fungi, viruses, and protozoa since they are active only in bacterial infections. At the present time, antibiotics are produced synthetically. The microorganisms the production of which has provided the invention of antibiotics are fungi [50].

As antibiotics can be broad‐spectrum affecting numerous bacteria, they can also be narrow‐ spectrum affecting limited bacteria. Furthermore, antibiotics with bactericide effect have an effect on bacteria by killing bacteria and antibiotics with bacteriostatic effect have an effect on bacteria by stopping their reproduction [51].

Although it has not been a long time since antibiotics have come into use, a rapid increase has been observed in their development. However, many problems have occurred during and after the consumption of these drugs. One of the main problems among them is bacterial resistance that develops against antibiotics.

There are a lot of reasons that bacteria develop resistance to antibiotics. The most important among them is that antibiotics are used without need and unconsciously.

Drugs that are used most frequently and in an excessive amount in the world are antibiotics. This usage also covers the unnecessary and unconscious use besides the proper use for treat‐ ment. The use for wrong purposes, misuse, and unnecessary use of antibiotics lead to bacte‐ rial resistance. For this reason, information about the usage of antibiotics should be given and the excessive and unnecessary usage should be prevented.

#### **6.1. Classification of antibiotics**

According to Li et al. [45], the removal activity of target antibiotics from water changed between 32 and 78% through conventional treatment. With the advanced treatment methods, the removal rate of target antibiotics became 85–100% and pollution probability of antibiotics decreased. In addition to this, in the risk assessment, the effects of ofloxacin and erythromycin on microorganisms in water are investigated by refining it. The majority of antibiotics cannot be absorbed or metabolized in the body. Moreover, the large part pass into the sewage system through urine and faeces and it comprises a significant part of the antibiotic source in nature.

In the study conducted by Zhang et al. [46], the elimination mechanism of three β‐lactam, two fluoroquinolones, and two macrolide antibiotics was investigated in the wastewater treatment establishment, which has four different treatment methods among six wastewater treatment establishments in China, Dalian. In this study, fluoroquinolones and macrolide antibiotics were determined as dominant antibiotics at the exit of wastewater treatment establishment and in coastal waters. It is revealed that β‐lactams are removed through biodegradation, for fluoroquinolones pretreatment is more effective than biologic treatment, and macrolide con‐ centration increases dramatically after biological treatment. The reason for this is that macro‐

Xu et al. [47] examined antibiotics and their resistant genes in a water treatment establishment in Beijing, China and the situation of the river into which water was discharged in their study. A total of 13 antibiotic resistance genes (ARGs) were examined. SuI‐arg was found at the highest rate among all antibiotic resistance genes (ARGs). ARG quantity in the wastewater treatment establishment is higher than in the river. According to the correlation analysis, there is a posi‐ tive relationship between tetracyclines and tetargs in water. This correlation could not be per‐ formed between suI‐args and sulfonamides. A negative relationship was observed between the concentration of quinolone genes and enrofloxacin. When ARG abundance of the waters that are treated in the treatment establishment is examined, treatment establishment causes resistant genes to increase. Results show that treatment establishments have a function of a warehouse for resistance genes. As a result, treated water needs advanced treatment before it is sent to the natural aquatic environment. In the study, three antibiotic groups were studied as tetracycline, sulfonamides, and quinones that are known for their permanence in the aquatic environment. Tetracyclines are removed at the rate of 87.9% in sludge elimination establishments. In the elimi‐ nation of tetracycline, biodegradation, and adsorption have an important role. In sludge elimi‐ nation establishments, teta, tetm, tetw, and teto genes are the ones that are mostly found [47].

Antibiotics are bioactive substances that kill or stunt the growth of the microorganism and

Antibiotics that are naturally obtained from plants and their extracts and are used for medical purposes have been brought into use as a result of Paul Ehrlich's studies in 1908. Paul Ehrlich revealed some chemical substances that are harmful to some bacteria and are less harmful to

lides that are covered by faeces particles are revealed [46].

82 Biological Wastewater Treatment and Resource Recovery

have a high effect on synthetic or biological origin [48].

the host cells by investigating them [49].

**6. Antibiotics**

Antibiotics are separated into two groups according to their effect on microorganisms:


### **6.2. Mechanisms of action of antibiotics**

#### *6.2.1. The ones that inhibit cell wall synthesis*

Bacteria are prokaryote microorganisms. They do not have real nucleus but they have cell walls. Cell walls protect bacteria from the external environment and antimicrobials. Cell wall contains pores 1–2 nm in diameter that is convenient to the transition of substances found in the external environment and nonselective. In short, they are not semipermeable. The transition of antimicrobials depends on the structure of the cell wall and molecular size of the drug.

Human cells have no cell wall. Thus, antibiotics (Penicillins and Beta‐lactams) in this group cannot spoil the adhesion of human cells. These antibiotics affect either by adhering to Penicillin‐Binding Proteins (PBP) or by spoiling the synthesis of cell wall without adhering to PBP (**Table 2**) [51].



**Table 2.** Classification of antibiotics [53].

**The ones that inhibit the cell wall synthesis Beta‐Lactams:**

PBP (**Table 2**) [51].

84 Biological Wastewater Treatment and Resource Recovery

Human cells have no cell wall. Thus, antibiotics (Penicillins and Beta‐lactams) in this group cannot spoil the adhesion of human cells. These antibiotics affect either by adhering to Penicillin‐Binding Proteins (PBP) or by spoiling the synthesis of cell wall without adhering to

The ones that inhibit cytoplasm membrane permeability Polymyxins

The ones that inhibit ribosome's protein synthesis Tetracyclines

The ones that effect bacteria's genetic material break DNA and RNA Fluoroquinolones

Penicillines Cephalosporins

Cycloserine Ristocetin Bacitracin Teicoplanin Vancomycin

Gramicidin Nystatin

Amphotericin B Candicein

Hexachlorophene Cationic detergents

Aminoglycosides Macrolides Amphenicols Lincosamides Fucidicasid

Rifamycins Nalidixicasid Metronidazole Actinomycins

Monobactams (Aztreonam)

Carbapenems (imipenem, Meropenem)

Ketoconazole and other antifungal imidazols Fluconazole and other antifungal trizols

#### *6.2.2. The ones that inhibit the protein synthesis*

These chemotherapeutic drugs are generally broad‐spectrum and have a bacteriostatic effect. Tetracyclines which belong to this antibiotic group prevent the adhesion of t‐RNA to ribo‐ somes. As human ribosomes (60S + 40S) and bacterial ribosomes (50S + 30S) are structurally different, these antibiotics that show an effect by adhering to ribosomes do not affect human ribosomes and protein synthesis (**Table 2**) [51].

#### *6.2.3. The ones that inhibit nucleic acid synthesis*

The most important antibiotics which belong to this group are rifampicin and quinones. Rifampicin inhibits the transcription (RNA inhibition dependent on DNA). Quinones inhibit the formation of supercolid (DNA gyrase inhibitors).

Topoisomerases which are used in human DNA and RNA synthesis and enzymes which are used in the nucleic acid synthesis of microorganisms are different. For this reason, these antibiotics do not have a toxic effect on human cells (**Table 2**) [51].

#### *6.2.4. The ones that increase cytoplasmic membrane permeability*

These antimicrobials create an effect by splitting the membrane substances in bacteria, inhib‐ iting sterol synthesis in fungi or spoiling the permeability by binding sterols.

The cytoplasmic membrane of human cell bears a resemblance with cytoplasmic membranes of bacteria and fungi. Therefore, these antibiotics can have a toxic effect on human cells when they are used in a systemic way (**Table 2**) [51].

#### *6.2.5. The ones with antimetabolic activity*

Antibiotics in this group are generally bacteriostatic. The ones that are broadly known are the drugs such as sulfonamides, sulfons, para‐amino salicylic acid (PAS), ethambutols, and isoniazid. Sulfonamides and Sulfons stop the function of PAS and para‐amino benzoic acid (**Table 2**) [51].

#### **6.3. Basic antibiotic groups**

#### *6.3.1. Beta‐lactams*

Antibiotics containing beta‐lactam circle which is found in the nucleus and is responsible for the antibacterial effect of molecules are called beta‐lactam antibiotics. The beta‐lactam circle is a saturated circle, which comprises one nitrogen and three carbons. Antibiotics in this group have bactericide effect by influencing the cell wall which consists of the murine of bacteria. Penicillin and Ampicillin are the most known antibiotics in the beta‐lactam group [52].

#### *6.3.2. Vancomycin*

They have an effect on multiresistant bacteria [54]. These antibiotics inhibit cell wall synthe‐ sis by stopping RNA synthesis in bacteria, break the continuity of the peptidoglycan chain and spoil the cytoplasmic membrane structure. Vancomycin which has a narrow antibacterial spectrum affects Gram (+) cokes and *Clostridiums* [51].

#### *6.3.3. Tetracycline*

These antibiotics inhibit protein synthesis by adhering to 30S subunit of the microorganism ribo‐ some. Tetracycline which affects both Gram (+) and Gram (‐) bacteria is broad‐spectrum and has bacteriostatic effect [55, 56]. Tetracycline affects numerous and various bacteria types. It is also effective against *Rickettsia* sp., *Chlamydia* sp., *Spirochaete* sp., *Mycoplasma* sp., *Leptospira* sp., and some protozoa [51].

#### *6.3.4. Aminoglycosides*

Aminoglycosides inhibit protein synthesis in ribosomes by adhering to 30S subunit of bacte‐ rial ribosomes. Moreover, they cause the misreading of genetic code that m‐RNA has. These antibiotics are narrow‐spectrum and have bactericide effect. They are effective only in aerobe bacteria as they are dependent on oxygen in the membrane cell [51].

#### *6.3.5. Macrolides*

These antibiotics inhibit protein synthesis that is dependent on RNA in bacteria. They provide this effect by preventing the continuity of the peptide chain and adhesion of t‐RNA by bind‐ ing 70S ribosome to 50S subunit. Bacteriostatic macrolides have an intense effect against Gram (+) cokes and bacillus [57].

#### *6.3.6. Chloramphenicol*

Chloramphenicol is the first broad‐spectrum antibiotic. These antibiotics inhibit peptidyl transferase enzyme by binding bacterial ribosomes to 50S subunit and thus they inhibit pro‐

tein synthesis in a reversible way. They are sensitive to coke, aerobe, anaerobe Gram (+) bacilli, and most of the Gram (‐) bacteria. Furthermore, these antibiotics inhibit protein synthesis of bacteria in the tissue by transferring into the tissue [55, 58].

#### *6.3.7. Quinolones*

*6.2.5. The ones with antimetabolic activity*

86 Biological Wastewater Treatment and Resource Recovery

spectrum affects Gram (+) cokes and *Clostridiums* [51].

**6.3. Basic antibiotic groups**

*6.3.1. Beta‐lactams*

*6.3.2. Vancomycin*

*6.3.3. Tetracycline*

some protozoa [51].

*6.3.4. Aminoglycosides*

*6.3.5. Macrolides*

(+) cokes and bacillus [57].

*6.3.6. Chloramphenicol*

Antibiotics in this group are generally bacteriostatic. The ones that are broadly known are the drugs such as sulfonamides, sulfons, para‐amino salicylic acid (PAS), ethambutols, and isoniazid. Sulfonamides and Sulfons stop the function of PAS and para‐amino benzoic acid (**Table 2**) [51].

Antibiotics containing beta‐lactam circle which is found in the nucleus and is responsible for the antibacterial effect of molecules are called beta‐lactam antibiotics. The beta‐lactam circle is a saturated circle, which comprises one nitrogen and three carbons. Antibiotics in this group have bactericide effect by influencing the cell wall which consists of the murine of bacteria. Penicillin and Ampicillin are the most known antibiotics in the beta‐lactam group [52].

They have an effect on multiresistant bacteria [54]. These antibiotics inhibit cell wall synthe‐ sis by stopping RNA synthesis in bacteria, break the continuity of the peptidoglycan chain and spoil the cytoplasmic membrane structure. Vancomycin which has a narrow antibacterial

These antibiotics inhibit protein synthesis by adhering to 30S subunit of the microorganism ribo‐ some. Tetracycline which affects both Gram (+) and Gram (‐) bacteria is broad‐spectrum and has bacteriostatic effect [55, 56]. Tetracycline affects numerous and various bacteria types. It is also effective against *Rickettsia* sp., *Chlamydia* sp., *Spirochaete* sp., *Mycoplasma* sp., *Leptospira* sp., and

Aminoglycosides inhibit protein synthesis in ribosomes by adhering to 30S subunit of bacte‐ rial ribosomes. Moreover, they cause the misreading of genetic code that m‐RNA has. These antibiotics are narrow‐spectrum and have bactericide effect. They are effective only in aerobe

These antibiotics inhibit protein synthesis that is dependent on RNA in bacteria. They provide this effect by preventing the continuity of the peptide chain and adhesion of t‐RNA by bind‐ ing 70S ribosome to 50S subunit. Bacteriostatic macrolides have an intense effect against Gram

Chloramphenicol is the first broad‐spectrum antibiotic. These antibiotics inhibit peptidyl transferase enzyme by binding bacterial ribosomes to 50S subunit and thus they inhibit pro‐

bacteria as they are dependent on oxygen in the membrane cell [51].

Quinolones affect bacteria by inhibiting DNA gyrase. This effect prevents DNA replication and creates bactericide impact. Moreover, the bacteria that are exposed to this antibiotic do not divide and die from stretching abnormally. They are effective in most Gram (‐) bacteria and Gram (‐) bacteria [59].

## *6.3.8. Trimethoprim*‐*sulfamethoxazole*

It is also known as cotrimoxazole. When Sulfamethoxazole (STX) is a sulfonamide, Trime‐ thoprim (TMP) is a diaminopiriminid which inhibits bacterial dihydrofolate reductase com‐ petitively. They affect many Gram (+) and Gram (‐) bacteria by causing unnoticeable synergistic bactericide effect when both drugs are used separately.

As a rule, the maximum synergistic activity of both antibacterial drugs, Trimethoprim (TMP), and Sulfamethoxazole (STX), occurs in bacteria types which are sensitive to both drugs. In the determination of the activity, sensitivity to TMP is more important [60].

#### **6.4. Antibiotic resilience**

Antibiotic resilience is simply the ability to resist against any antibiotic which spoils the reproduction function of a microorganism or causes its death. Resistance concerns the microorganism, patient, antibiotic, and environment or all of them. Resistance has no con‐ nection with virulence [61].

Antibiotic resistance spreads in three ways in bacteria:


The resistance that microorganisms show against antibiotics is classified in two groups as natural (phenotypic) and acquired (genotypic).

#### *6.4.1. Natural resistance*

Natural resistance is the situation that occurs when the microorganism cannot carry the struc‐ ture affected by the drug as its quality or it cannot reach the target due to the structure of the drug. This resistance is not hereditary besides it is the key feature of bacteria and it is not related to the use of drugs.

For instance, microorganisms such as L‐forms of bacteria and *Mycoplasma* that have no mem‐ brane have a natural resistance to antibiotics such as penicillin which inhibit the cell wall synthesis. Another example is that vancomycin cannot affect Gram (‐) bacteria due to the fact that it cannot pass from adventitia [62].

#### *6.4.2. Acquired resistance*

Depending on the change in bacteria's genetic characteristic, it is the resistance which occurs as a result of taking DNA series that have resistance gene from another bacterium through transformation, transduction, or conjugation as it can be through mutations in a plasmid, chromosome, or transposon DNA. Furthermore, these bacteria can gain resistance against antibiotics to which they have been sensitive before [63]. Genetic originated resistance is examined in two groups as chromosomal and extrachromosomal.

#### *6.4.2.1. Chromosomal resistance*

It occurs as a result of mutations which happen spontaneously in the bacterial chromosome. Spontaneous mutations arise from some physical or chemical factors. Consequently, struc‐ tural changes occur in the bacterial cell. In this situation, changes can happen in the drug's target in the cell or permeability of the cell to the drug can decrease [62].

#### *6.4.2.2. Extrachromosomal resistance*

Bacteria have extrachromosomal resistance plasmids that are called extrachromosomal ele‐ ments, transposons that are active elements found on the chromosomes and bring chromo‐ somes new antibiotic resilience, integrons, and antibiotics.

#### *6.4.2.2.1.+ Plasmids*

The structures that can be inside bacteria or outside the chromosomes in the DNA structure, bring some qualities to these bacteria and keep these qualities under control genetically are called plasmids.

Plasmids can have virulence factors besides resistance genes against antimicrobics and heavy metals. Plasmids which have resistance genes are called R‐plasmids. R‐plasmids transfer the resistant gene package by passing into other bacteria through transformation, transduction, and conjugation. Thus, they provide the spread of resistance [64].

#### *6.4.2.2.2. Transposons*

Transposons are the structures which can settle in different places in the bacterial chromosome or can be transferred from chromosome to plasmid, from plasmid to plasmid, from plasmid to DNA or bacteriophage. These structures are DNA series found over the replicon like a chromosome, plasmid or bacteriophage as they cannot replicate by themselves. They have an active role in the spread of the multiple drug resistant isolates of transposons by revealing them in a short time [62, 65].

#### *6.4.2.2.3. Integrons*

Integrons are active DNA elements which have the ability to capture genes, which codify antibi‐ otic‐resistant genes in enteric bacteria, with specific recombination. These genes that are captured by integrons are called gene tapes. Gene tapes are active genetic elements which comprise of only one gene and recombination zone which is free, little‐alkali and called the 59‐base element. As well as these gene tapes may not present in integrons at all, there can be 100 of them [66].

#### *6.4.3. Cross‐resistance*

*6.4.2. Acquired resistance*

88 Biological Wastewater Treatment and Resource Recovery

*6.4.2.1. Chromosomal resistance*

*6.4.2.2. Extrachromosomal resistance*

*6.4.2.2.1.+ Plasmids*

*6.4.2.2.2. Transposons*

them in a short time [62, 65].

*6.4.2.2.3. Integrons*

plasmids.

Depending on the change in bacteria's genetic characteristic, it is the resistance which occurs as a result of taking DNA series that have resistance gene from another bacterium through transformation, transduction, or conjugation as it can be through mutations in a plasmid, chromosome, or transposon DNA. Furthermore, these bacteria can gain resistance against antibiotics to which they have been sensitive before [63]. Genetic originated resistance is

It occurs as a result of mutations which happen spontaneously in the bacterial chromosome. Spontaneous mutations arise from some physical or chemical factors. Consequently, struc‐ tural changes occur in the bacterial cell. In this situation, changes can happen in the drug's

Bacteria have extrachromosomal resistance plasmids that are called extrachromosomal ele‐ ments, transposons that are active elements found on the chromosomes and bring chromo‐

The structures that can be inside bacteria or outside the chromosomes in the DNA structure, bring some qualities to these bacteria and keep these qualities under control genetically are called

Plasmids can have virulence factors besides resistance genes against antimicrobics and heavy metals. Plasmids which have resistance genes are called R‐plasmids. R‐plasmids transfer the resistant gene package by passing into other bacteria through transformation, transduction,

Transposons are the structures which can settle in different places in the bacterial chromosome or can be transferred from chromosome to plasmid, from plasmid to plasmid, from plasmid to DNA or bacteriophage. These structures are DNA series found over the replicon like a chromosome, plasmid or bacteriophage as they cannot replicate by themselves. They have an active role in the spread of the multiple drug resistant isolates of transposons by revealing

Integrons are active DNA elements which have the ability to capture genes, which codify antibi‐ otic‐resistant genes in enteric bacteria, with specific recombination. These genes that are captured

examined in two groups as chromosomal and extrachromosomal.

target in the cell or permeability of the cell to the drug can decrease [62].

somes new antibiotic resilience, integrons, and antibiotics.

and conjugation. Thus, they provide the spread of resistance [64].

It is the situation when some microorganisms are resistant both to some drugs and at the same time to other drugs that have a similar mechanism. This resilience can be seen between struc‐ turally similar drugs like erythromycin and kanamycin as it can be seen between completely different drugs like erythromycin and lincomycin [62].

## **7. Antibiotic resistance in aquaculture and agriculture**

Antibiotic concentrations below curative doses cause antibiotic resistance in many patient groups especially in critically ill patients [67].

The emergence of antibiotic‐resistant bacteria is seen as an important health problem. For, thousands of patients die because of resistant bacteria. All efforts are concentrated on the decrease of existing antibiotic‐resistant bacteria and antibiotic usage [2].

Rapidly developing antibiotic‐resistant bacteria force public health services and health centers. American Centers for Disease Control and Prevention and Food and Agricultural Organization stated that antibiotic resilience has seriousness over the world. According to the predictions, 700,000 people die because of antibiotic resistance in a year. With the changes in temperature and rain regime, climate‐sensitive bacteria and diseases will increase and spread to new regions, consequently, the situation will worsen [68].

Determination frequency and antibiotic concentrations are generally higher in January and May. The reason for this is that low‐flow and low‐temperature conditions cause antibiotics to be trapped by sediments. Antibiotic quantities vary per region. The highest quantities are found in estuaries and places where sewage is disposed. Antibiotic usage is more than 100,000–200,000 tonnes over the world and more than 25,000 tonnes in China. 80–90% of these antibiotics are released into nature through human urine and faeces. Pharmacologically active compounds in animal manure are used as a fertilizer in agriculture, and in conclusion, these compounds are accumulated in soil or mix into surface or underground waters [69].

Antibiotics are used as an environmental pollutant, in the treatment of diseases in a broad sense, in the protection and treatment of diseases in veterinary, and as growth promotive in aquaculture and agriculture [42].

Veterinary drugs are used for the protection and treatment of animal diseases and are one of the important components of environmental pollution as a result of intensive agricultural and aquaculture actions. Veterinary drugs are among the potentials of chemical pollutants and they have a biological effect in low concentrations like other drugs. While the annual usage of veterinary antibiotics in the United States reaches 11,000 tonnes, China follows it by 6000 tonnes. These quantities contain not only drugs with therapeutic purposes but also antibiotics which are used to promote production. In Europe, France leads these rates with 1064 tonnes, Holland follows it with 514 tonnes, and England with 403 tonnes. The most used antibiotics are tetracyclines, sulfonamides, β‐lactams, and macrolides. The presence of veterinary antibiotics in nature causes the emergence of antibiotic‐resistant bacteria and nontarget microorganisms are affected by drinking potable water that contains antibiotic remnants or by consumption of animal or herbal foods that contain antibiotics. Mixing of veterinary drugs into nature may cause the development of single, multiple, and cross‐resis‐ tance in pathogens, commensals, and nonpathogens. Most of the veterinary drugs are feebly absorbed in the animal intestine. The remaining large quantity is removed with faeces. A small combination of these drugs removed undergoes a change, conjugates with polar mol‐ ecules or remains the same. Consequently, these drugs can be detected in natural environ‐ ments such as animal manure, soil, surface, and underground water resources. The major source of veterinary drugs in nature is biological remnants and the usage of dirty animal faeces in fertilization [70].

The usage of wastewaters for agricultural and other purposes by treating them provides many advantages such as the formation of alternative water resources, prevention of the pollution of surface and underground waters, and reduction of fertilizer usage. However, along with its advantages, it also has negative effects on public health and the environment. In order to minimize these effects, risks that origin from pathogens and chemicals that emerge from the wastewater usage should be evaluated well [71].

Waters that are polluted in many ways are treated by many methods with the progress of tech‐ nology. The usage of wastewaters as irrigation waters by putting them through pretreatment or delivering into the land is one of these methods. Causing soil pollution by water pollution occurs in this method. Wastewaters from various resources pollute the soil and they have various effects on soil pollution [72].

Domestic wastewaters can be used in forests, pastures, lawns by being pretreated. The removal of wastewaters by using them in the irrigation of lands in this way creates serious health prob‐ lems. Moreover, bacteria and pollutants in wastewaters are harmful to human health by being absorbed by the soil and reaching underground waters when the buffering effect decreases [72].

In the sector of aquaculture, antibiotics are intensively used to treat fish and protect it from diseases. Antibiotics that are applied to fish cause fish pathogens and zoonotic fish bacteria to gain resistance to antibiotics. Zoonotic fish bacteria which develop antibacterial resistance create danger for people and cause infections that are hard to treat [73].

The misuse of antibiotics affects human health directly or indirectly and complicates the treatment of fish diseases. Its direct effect is that fish bacteria and zoonotic fish bacteria gain resistance. These strains which are resistant create refractory infections when they infect peo‐ ple. The indirect antibiotic resilience occurs with the transfer of resistance plasmids in bacteria to human pathogens. In this way, human pathogens that gain resistance create resistant infec‐ tions in people. Also in the studies conducted, it is revealed that multiple antibiotic resistance genes are transferred from fish pathogens to human pathogens [73].

In August 2011, 20 antibiotics that were taken from 20 different samples' regions taken from sediment and aquatic organisms in Dalian coastline were examined. Tetracyclines are domi‐ nant antibiotics in sea water. Sulfonamides are dominant antibiotics in sediment and aquatic organisms. Industrial aquaculture is the most significant reason for the pollution of coasts in developed and developing countries because of the intensive antibiotic usage. Antibiotic usage in China comprises the quarter of antibiotic usage over the world [74].

The state of 37 antibiotics was examined on 6 aquaculture farms around Hailing Island. Sulfamethoxazole, salinomycin, and trimethoprim were detected at the highest rate in water; ox tetracycline was detected at the highest rate in shrimp larva pools, enrofloxacin was detected at the highest rate in feed samples, and erythromycin was detected at the highest rate in sediment [75].

Wastewater usage in agriculture and land irrigations can be described as wastewater recy‐ cling. This usage brings many problems even if it is very economical. The usage of wastewater in agricultural activities by pretreating it is not enough to eliminate these results. Especially, domestic wastewaters constitute an enormous danger because of their content. Sewage and hospital wastewaters are the reasons for this danger. Antibiotics cause infections and anti‐ biotic resistance in people besides the fact that they decrease the productivity in agriculture with the bacterial and parasite microorganisms they contain.

In aquaculture, which is a method used in fish farming, it is possible that bacterial and para‐ site infections occur. Therefore, antibiotics are used for the treatment and protection from infections. Antibiotic usage complicates treatment as well as it creates antibacterial resistance in fish and people.

## **8. Suggestion**

antibiotics which are used to promote production. In Europe, France leads these rates with 1064 tonnes, Holland follows it with 514 tonnes, and England with 403 tonnes. The most used antibiotics are tetracyclines, sulfonamides, β‐lactams, and macrolides. The presence of veterinary antibiotics in nature causes the emergence of antibiotic‐resistant bacteria and nontarget microorganisms are affected by drinking potable water that contains antibiotic remnants or by consumption of animal or herbal foods that contain antibiotics. Mixing of veterinary drugs into nature may cause the development of single, multiple, and cross‐resis‐ tance in pathogens, commensals, and nonpathogens. Most of the veterinary drugs are feebly absorbed in the animal intestine. The remaining large quantity is removed with faeces. A small combination of these drugs removed undergoes a change, conjugates with polar mol‐ ecules or remains the same. Consequently, these drugs can be detected in natural environ‐ ments such as animal manure, soil, surface, and underground water resources. The major source of veterinary drugs in nature is biological remnants and the usage of dirty animal

The usage of wastewaters for agricultural and other purposes by treating them provides many advantages such as the formation of alternative water resources, prevention of the pollution of surface and underground waters, and reduction of fertilizer usage. However, along with its advantages, it also has negative effects on public health and the environment. In order to minimize these effects, risks that origin from pathogens and chemicals that emerge from the

Waters that are polluted in many ways are treated by many methods with the progress of tech‐ nology. The usage of wastewaters as irrigation waters by putting them through pretreatment or delivering into the land is one of these methods. Causing soil pollution by water pollution occurs in this method. Wastewaters from various resources pollute the soil and they have various effects

Domestic wastewaters can be used in forests, pastures, lawns by being pretreated. The removal of wastewaters by using them in the irrigation of lands in this way creates serious health prob‐ lems. Moreover, bacteria and pollutants in wastewaters are harmful to human health by being absorbed by the soil and reaching underground waters when the buffering effect decreases [72].

In the sector of aquaculture, antibiotics are intensively used to treat fish and protect it from diseases. Antibiotics that are applied to fish cause fish pathogens and zoonotic fish bacteria to gain resistance to antibiotics. Zoonotic fish bacteria which develop antibacterial resistance

The misuse of antibiotics affects human health directly or indirectly and complicates the treatment of fish diseases. Its direct effect is that fish bacteria and zoonotic fish bacteria gain resistance. These strains which are resistant create refractory infections when they infect peo‐ ple. The indirect antibiotic resilience occurs with the transfer of resistance plasmids in bacteria to human pathogens. In this way, human pathogens that gain resistance create resistant infec‐ tions in people. Also in the studies conducted, it is revealed that multiple antibiotic resistance

create danger for people and cause infections that are hard to treat [73].

genes are transferred from fish pathogens to human pathogens [73].

faeces in fertilization [70].

90 Biological Wastewater Treatment and Resource Recovery

on soil pollution [72].

wastewater usage should be evaluated well [71].

Urbanization, an increase in industry and population, increases the water demand with each passing day. The most important need is water and nutrition's existence depends on water. For this reason, the amount of water, as well as its presence, is important for living beings.

The increase of water usage, its unconscious use, the involvement in pollutant activities, and not taking precautions against pollution have a negative influence on water amount. In this situation, people's awareness should be increased, and pollution pretending precautions should be taken. Besides these situations, water reutilization can be provided by treatment.

The reutilization of wastewater by treatment increases water amount and creates some sources for the use of living beings. This brings positive situations as well as many negative situations along with it.

Wastewater causes pathogenic factors in living beings because of its content as well as it causes antibiotic‐resistant bacteria and the spread of pathogen microorganisms.

There are many ways for antibiotic resistance to occur and spread. The leading factors among them are the excessive and unconscious usage of antibiotics, the usage of broad‐spectrum antibiotics, its accumulation in sewers through taking it to the body and urinating it, espe‐ cially giving hospital sewer's accumulation to treatment facilities without its pretreatment.

The result of the insufficient treatment of wastewater treatment facilities is that the amount of antibiotics remains and an increase occurs as well. Due to the chemical structure of antibiot‐ ics, it may not come to light before entering the wastewater treatment. Antibiotics have the tendency to hold on to sediments due to their structure. In the stages of wastewater treatment, as a result of the decomposition of sediments, antibiotics come out. In this situation, the prob‐ lems of not treating antibiotics arise. The usage of these waters for agricultural purposes also causes antibiotic resistance to spread.

Antibiotics are used for the purposes of treating diseases in humans and also for the same purposes in animals. In this situation, this can cause the emergence and spread of antibiotic‐ resistant bacteria.

To prevent the emergence and spread of antibiotic‐resistant bacteria, first the awareness of people of antibiotic use should be raised. The usage of antibiotics in human and animal treat‐ ment should be reduced. Other waters that belong to the group of all sewage and wastewater, especially hospital sewage, should be pretreated before being discharged to wastewater treat‐ ments with biological treatment. Mechanic, chemical, and thermal treatment processes are included in pretreatment. Many pretreatment processes such as the process of oxidation, ther‐ mophilic pretreatment, sludge disintegration, ozonation, photocatalytic pretreatment, physi‐ cochemical pretreatment, and ultrasonic method should be used. Since wastewater treatment systems used fall short in some cases, new systems and pretreatment systems for antibiotic treatment should be developed.

## **Author details**

Sadik Dincer\* and Esra Sunduz Yigittekin\*

\*Address all correspondence to: sdincer@cu.edu.tr and esra‐gokyuzu@hotmail.com

Biology Department, Science and Letter Faculty, Cukurova University, Adana, Turkey

## **References**


antibiotics, its accumulation in sewers through taking it to the body and urinating it, espe‐ cially giving hospital sewer's accumulation to treatment facilities without its pretreatment.

The result of the insufficient treatment of wastewater treatment facilities is that the amount of antibiotics remains and an increase occurs as well. Due to the chemical structure of antibiot‐ ics, it may not come to light before entering the wastewater treatment. Antibiotics have the tendency to hold on to sediments due to their structure. In the stages of wastewater treatment, as a result of the decomposition of sediments, antibiotics come out. In this situation, the prob‐ lems of not treating antibiotics arise. The usage of these waters for agricultural purposes also

Antibiotics are used for the purposes of treating diseases in humans and also for the same purposes in animals. In this situation, this can cause the emergence and spread of antibiotic‐

To prevent the emergence and spread of antibiotic‐resistant bacteria, first the awareness of people of antibiotic use should be raised. The usage of antibiotics in human and animal treat‐ ment should be reduced. Other waters that belong to the group of all sewage and wastewater, especially hospital sewage, should be pretreated before being discharged to wastewater treat‐ ments with biological treatment. Mechanic, chemical, and thermal treatment processes are included in pretreatment. Many pretreatment processes such as the process of oxidation, ther‐ mophilic pretreatment, sludge disintegration, ozonation, photocatalytic pretreatment, physi‐ cochemical pretreatment, and ultrasonic method should be used. Since wastewater treatment systems used fall short in some cases, new systems and pretreatment systems for antibiotic

\*Address all correspondence to: sdincer@cu.edu.tr and esra‐gokyuzu@hotmail.com

Turkish Journal of Scientific Reviews. 2013; 6 (1): 58–62. ( in Turkish)

Biology Department, Science and Letter Faculty, Cukurova University, Adana, Turkey

[1] Polat A. Su Kaynaklarının Sürdürülebilirliği İçin Arıtılan Atıksuların Yeniden Kullanımı.

[2] Centner TJ. Recent government regulations in the United States seek to ensure the effec‐ tiveness of antibiotics by limiting their agricultural use. Environment International.

[3] Akin M, Akin G. Importance of water, water potential in turkey, water basins and water pollution. Ankara University Faculty of Language, History and Geography Journal.

causes antibiotic resistance to spread.

92 Biological Wastewater Treatment and Resource Recovery

treatment should be developed.

Sadik Dincer\* and Esra Sunduz Yigittekin\*

resistant bacteria.

**Author details**

**References**

2016; 94: 1–7.

2007; 47, 2:105–118.


[39] Chojnacka K. Using bisorption to enrich the biomoass of Choleralla vulgaris with micro‐ elements to be used as mineral feed supplement. World Journal of Microbiology and Biotechnology.2007; 23:1139–1147.

[23] Guler C, Cobanoglu Z. Su Kirliliği. 1st ed. Ankara: Çevre Sağlığı Temel Kaynak Dizisi;

[24] Ozarslan A. Determinetion of fecal coliform levels and antibiotics resistance frequency

[26] Demirekin H. Sensitivity analysis of environmental problems in Isparta city [thesis].

[27] Dincer S, Matyar F, Sonmez N. Seyhan nehrinin fekal kirlilik duzeyi ve fekal koliform‐ larin antibiyotik hassasiyetleri. 12. Biyoteknoloji Kongresi. 2001; Ayvalık, 252–255. ( in

[28] Samsunlu A. Wastewater Treatment. 3rd ed. Istanbul: Birsen Publisher; 2011. 647 p.

[29] San NO. Treatment of wastewater with heavy metal and reactive dye by *Rhodotorula sp.*

[30] Gunes E. Anaerobic treatment of domestic wastewaters in upflow anaerobic sludge bed (UASB) reactor at temperate conditions and chemical post‐treatment applications [the‐

[31] Dogan M, Saylak M. Su Kimyası. Kayseri: Erciyes University Publications; 2000. 120 p.

[32] Metcalf E. Wastewater Engineering, Treatment and Reuse. 4th ed. New York: McGraw‐

[33] Yasar A, Dogan EC, Arslan A. Macro and micro pollutants and treatment options in hospital wastewaters. Erciyes University, Journal of Institute of Science and Technology.

[34] Saygi S. Cevre ve insan sagligi yönünden ilaç atiklarinin onemi. Marmara Pharmaceutical

[35] Sonmez G, Isik M. Sulardaki Ilac Kalintilarinin Ileri Oksidasyon Yontemleri Ile Giderimi. Turkish Journal of Scientific Reviews. 2013; *6*(1):68–73. SSN: 1308‐0040, E‐ISSN:

[36] Turker C. Treatability of segregated textile effluents containing auxiliary chemicals [the‐

[37] Turkman A, Aslan S, Ege I. Lead removal from wastewaters by natural zeolites. Dokuz Eylul University Faculty of Engıneerıng Journal of Engıneerıng Scıence. 2001; 3:2:13–19.

[38] Sinan RK. Estimation of primary treatment and biological treatment effluent parameters by artificial neural networks in domestic wastewater treatment plants [thesis]. Selcuk

Journal. 2012; 16: 82–90. DOI: 10.12991/201216406. ( in Turkish)

sis]. Istanbul: Istanbul Technical University Institute of Science; 2013.

of Adana drinking water [thesis]. Adana: Cukurova University; 2009.

[25] Tanyolaç J. Limnology. Ankara: Hatipoğlu Publications; 1993. 235 p.

Isparta: Suleyman Demirel University; 2001.

[thesis]. Ankara: Ankara University; 2007.

sis]. Istanbul: Istanbul Technical University; 2008.

1997. 47 p. ( in Turkish)

94 Biological Wastewater Treatment and Resource Recovery

Turkish)

( in Turkish)

Hill Education; 2003. 1771 p.

2016; *29*(2): 144–158.

2146‐0132. ( in Turkish)

University; 2010.


[69] Shi H, Yang Y, Liu M, Yan C, Yue H, Zhou J. Occurrence and distribution of antibiotics in the surface sediments of the Yangtze Estuary and nearby coastal areas. Marine Pollution Bulletin. 2014; 83(1): 317–323.

[53] Akkan AG. Antibiyotiklerin sınıflandırılması. I.U. Cerrahpasa Faculty of Medicine Continuing Medical Education Activities Antibiotics Practice Symposium; 2–3 May

[54] Strohl WA, Rouse H, Bruce DF. Lippincott's Illustrated Reviews. Microbiology. Nobel

[55] Alcamo EI. Fundamentals of Microbiology. 6th ed. Sudbury Massachusetts: Jones and

[56] Lefever Kee J, Hayes ER, McCuistion LE. Pharmacology: A Nursing Process Approach. 7th ed. Printed in The United States America: Elsevier Saundres; 2012. 983 p.

[58] Braibant M, Gilot P, Content J. The ATP binding cassette (ABC) transport systems of

[59] Algun U, Arisoy A, Gunduz T, Ozbakkaloglu B. The resistance of Pseudomonas aeruginosa strains to fluoroquınolone group of antibiotics. Indian Journal of Medical

[60] Rang HP, Dale MM, Ritter JM, Moore PK. 2006. Pharmacology. 5th ed. Edinburgh:

[61] Durupinar B. Antibiyotiklere dirençte yeni eğilimler. Klimik Journal. 2001; 14(2): 47–56.

[62] Yuce A. Antimikrobik İlaçlara Direnç Kazanma Mekanizmaları. Klimik Journal. 2001;

[64] Gur D, Tutar I, Vardar Unlu G. (2001). İsepamisinin hastane izolatı Gram‐negatif bak‐ terilere karsı in vitro etkisi. Turkish Journal of Hospital Infections. 2001; *5*(1):19. ( in

[65] Aygun G, et al. The antibiotic susceptibility patterns of *acinetobacter baumannii* strains isolated from nosocomial infections in intensive care unit. Journal of ANKEM. 2002;

[66] Roy PH. Integrons: Novel mobile genetic elements mediating antibiotic resistance in

[67] Carlier M, Stove V, Wallis SC, De Waele JJ, Verstraete AG, Lipman J, Roberts JA. Assays for therapeutic drug monitoring of β‐lactam antibiotics: A structured review.

[68] Centner TJ. Efforts to slacken antibiotic resistance: Labeling meat products from animals raised without antibiotics in the United States. Science of the Total Environment. 2016;

[63] Tanir G, Gol N. Antibiyotik Direnci. Klimik Journal. 1999; 2:12. ( in Turkish)

*Enterobacteria* and *Pseudomonas.* APUA Newsletter. 1995; 13(3): 1–4.

International Journal of Antimicrobial Agents. 2015; 46(4): 367–375.

[57] Allen N. Effects of macrolide antibiotics on ribosome function. In Macrolide

Mycobacterium tuberculosis. FEMS Microbiology. 2000; 24: 449–467

Antibiotics. Birkhäuser Verlag, Basel. 2002; 261–280 pp.

1997; Istanbul. pp. 52–62.( in Turkish)

Medicine Bookstores. 2006; 516:44–47.

Bartlett Publishers; 2001. 832 p.

96 Biological Wastewater Treatment and Resource Recovery

Microbiology. 2004; 22(2): 112–114.

( in Turkish)

Turkish)

16(1): 85–88.

563: 1088–1094.

14(2): 41–46. ( in Turkish)

Elsevier Churchill Livingstone; 2006. 797 p.


**Comparative Assessment of Pharmaceutical Removal from Wastewater by the Microalgae** *Chlorella sorokiniana***,** *Chlorella vulgaris* **and** *Scenedesmus obliquus* **Comparative Assessment of Pharmaceutical Removal from Wastewater by the Microalgae Chlorella sorokiniana, Chlorella vulgaris and Scenedesmus obliquus**

Carla Escapa Santos, Ricardo Nuno de Coimbra, Sergio Paniagua Bermejo, Ana Isabel García Pérez and Marta Otero Cabero Coimbra, Sergio Paniagua Bermejo, Ana Isabel García Pérez and Marta Otero Cabero Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

Carla Escapa Santos, Ricardo Nuno de

http://dx.doi.org/10.5772/66772

#### **Abstract**

In view of risks associated with the discharge of pharmaceuticals in the aquatic envi‐ ronment, the objective of this work was to assess the removal of paracetamol, salicylic acid and diclofenac from water by a microalgae‐based treatment. For a comparison purpose, the growth and kinetic parameters for the removal of drugs were determined for three different microalgae strains, namely *Chlorella sorokiniana*, *Chlorella vulgaris* and *Scenedesmus obliquus*. It was found that the drugs removal efficiency by these strains was related to their growth. Comparing the three pharmaceuticals, the salicylic acid was the most efficiently removed, especially by *S. obliquus* (>93% batch culture, >99% semicontinuous culture) and *C. sorokiniana* (>73% batch culture, >93% semicontinuous culture). Contrarily, paracetamol was the most poorly removed, the maximum efficien‐ cies being those attained by *C. sorokiniana* (>67% batch culture, >41% semicontinuous culture). On the other hand, diclofenac was efficiently removed only by *S. obliquus* (>98% batch culture, >79% semicontinuous culture). For the three considered drugs, *C. vulgaris* was the strain showing the lowest removal capacity. The large differences here revealed between microalgae strains regarding their removal capacity of pharma‐ ceuticals, pointed to the strain selection as a key issue for a successful application in wastewater treatment.

**Keywords:** emerging contaminants, wastewater treatment, phytoremediation, paracetamol, salicylic acid, diclofenac

© 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, 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.

## **1. Introduction**

Emerging contaminants (ECs) include a wide range of compounds and may be defined as naturally occurring, manufactured or man‐made chemicals or materials that have been found or are suspected to be present in various environmental compartments and whose toxicity or persistence are likely to significantly alter the metabolism of a living being [1]. Among them, pharmaceuticals have received considerable attention with respect to their environ‐ mental fate and toxicological properties over the last decade [2]. Pharmaceuticals represent an especially worrying class since they were designed to cause a physiological response and their presence in the environment may affect non‐target individuals and species [3]. This con‐ cern on pharmaceuticals presence in the aquatic environment has led to the recent consider‐ ation by European regulations within the Water Framework Directive (2000/60/EC) (WFD). The Commission proposal of 31 January 2012 foresaw the inclusion of three pharmaceuti‐ cals, namely diclofenac, 17‐beta‐estradiol (E2) and 17‐alpha‐ethinylestradiol (EE2) in the list of priority substances. Instead, by the EU Decision 2015/495, these compounds together with another estrogen (E1) and three antibiotics (azithromycin, clarithromycin and erythromycin) were finally included in the first watch list of substances to be monitored in all member states to support future reviews of the priority substances list [4].

Pharmaceuticals in domestic sewage or from hospital or industrial discharges end in municipal sewage treatment plants (STPs), but conventional wastewater treatments have been reported to be ineffective in the removal of such pollutants, with efficiency values of <5 to 40% [5]. In fact, STPs were not originally designed for the removal of pharmaceuticals due to the non‐existence of limiting regulations on their discharge [6, 7]. Consequently, STPs are important sources of such pollutants in the aquatic environment [8, 9]. In this regard, Verlicchi et al. [10], who reviewed the occurrence of 118 pharmaceuticals in the influent and effluent of 244 STPs, found that the occur‐ rence of some of them in the effluent discharged into surface water bodies may pose a medium‐ high (acute) risk to aquatic life. Among the studied pharmaceuticals, diclofenac was shown to have the highest average mass load (240 mg/1000 inhabitant) in the effluents of municipal STPs [10]. The removal efficiencies of diclofenac in conventional STPs have been reported to be about 17% [11], which translates into relative high concentrations in the corresponding effluents [12].

In the recent years, phytoremediation of waters by using photoautotrophic aquatic organ‐ isms such as algae has gained attention for the removal of both organic and inorganic pollutants [13–15]. Microalgae are characterized by high photosynthetic efficiency, high growth rates, wide adaptability and high potential to remove inorganic nutrients from the wastewater. The principal mechanism of algal nutrient removal is their uptake into the cell biomass [16]. The main advantages of using microalgae for nutrients removal during the ter‐ tiary treatment of wastewaters are the possibility of recycling the assimilated nitrogen and phosphorus into algal biomass as a fertilizer, as a source of products (e.g. paraffin, olefin, glycerol, protein, anti‐oxidant, pigment, plastic, etc.), or as biofuel, and also the generation of an oxygenated high‐quality effluent [17]. However, although the capability of microal‐ gae wastewater treatments systems to remove organic matter and nutrients has been deeply studied, little is known about the removal of ECs, such as pharmaceuticals, by algae. In fact, it has already been claimed the necessity of further studies on the removal of this sort of pol‐ lutants by algal systems [18].

In this context, the aim of this study was to determine and compare the potential of green micro‐ algae *Chlorella sorokiniana*, *Chlorella vulgaris* and *Scenedesmus obliquus* to remove paracetamol, salicylic acid and diclofenac from water. The strains used in this work were selected since they are known to have fast growth rates and potential for wastewater treatment due to their tolerance to the severe environmental conditions found in municipal wastewater and some industrial wastewaters [19].

## **2. Materials and methods**

**1. Introduction**

100 Biological Wastewater Treatment and Resource Recovery

Emerging contaminants (ECs) include a wide range of compounds and may be defined as naturally occurring, manufactured or man‐made chemicals or materials that have been found or are suspected to be present in various environmental compartments and whose toxicity or persistence are likely to significantly alter the metabolism of a living being [1]. Among them, pharmaceuticals have received considerable attention with respect to their environ‐ mental fate and toxicological properties over the last decade [2]. Pharmaceuticals represent an especially worrying class since they were designed to cause a physiological response and their presence in the environment may affect non‐target individuals and species [3]. This con‐ cern on pharmaceuticals presence in the aquatic environment has led to the recent consider‐ ation by European regulations within the Water Framework Directive (2000/60/EC) (WFD). The Commission proposal of 31 January 2012 foresaw the inclusion of three pharmaceuti‐ cals, namely diclofenac, 17‐beta‐estradiol (E2) and 17‐alpha‐ethinylestradiol (EE2) in the list of priority substances. Instead, by the EU Decision 2015/495, these compounds together with another estrogen (E1) and three antibiotics (azithromycin, clarithromycin and erythromycin) were finally included in the first watch list of substances to be monitored in all member states

Pharmaceuticals in domestic sewage or from hospital or industrial discharges end in municipal sewage treatment plants (STPs), but conventional wastewater treatments have been reported to be ineffective in the removal of such pollutants, with efficiency values of <5 to 40% [5]. In fact, STPs were not originally designed for the removal of pharmaceuticals due to the non‐existence of limiting regulations on their discharge [6, 7]. Consequently, STPs are important sources of such pollutants in the aquatic environment [8, 9]. In this regard, Verlicchi et al. [10], who reviewed the occurrence of 118 pharmaceuticals in the influent and effluent of 244 STPs, found that the occur‐ rence of some of them in the effluent discharged into surface water bodies may pose a medium‐ high (acute) risk to aquatic life. Among the studied pharmaceuticals, diclofenac was shown to have the highest average mass load (240 mg/1000 inhabitant) in the effluents of municipal STPs [10]. The removal efficiencies of diclofenac in conventional STPs have been reported to be about 17% [11], which translates into relative high concentrations in the corresponding effluents [12]. In the recent years, phytoremediation of waters by using photoautotrophic aquatic organ‐ isms such as algae has gained attention for the removal of both organic and inorganic pollutants [13–15]. Microalgae are characterized by high photosynthetic efficiency, high growth rates, wide adaptability and high potential to remove inorganic nutrients from the wastewater. The principal mechanism of algal nutrient removal is their uptake into the cell biomass [16]. The main advantages of using microalgae for nutrients removal during the ter‐ tiary treatment of wastewaters are the possibility of recycling the assimilated nitrogen and phosphorus into algal biomass as a fertilizer, as a source of products (e.g. paraffin, olefin, glycerol, protein, anti‐oxidant, pigment, plastic, etc.), or as biofuel, and also the generation of an oxygenated high‐quality effluent [17]. However, although the capability of microal‐ gae wastewater treatments systems to remove organic matter and nutrients has been deeply studied, little is known about the removal of ECs, such as pharmaceuticals, by algae. In fact,

to support future reviews of the priority substances list [4].

#### **2.1. Microorganisms and culture conditions**

The microalgae strains used in this study were *C. sorokiniana* CCAP 211/8 K from UTEX Culture Collection of Algae, *C. vulgaris* SAG 221‐12 from SAG Culture Collection of Algae and *S. obliquus* SAG 276‐1 from SAG Culture Collection of Algae. These microalgae strains are among the most commonly used for wastewater treatment have high growth rates and are able to grow under a wide range of conditions [19], which motivated their choice for this study.

The inoculum of each strain was cultivated in 250‐ml Erlenmeyer flasks in the standard culture medium Mann and Myers [20], which is composed of (per litre of distilled water): 1.2 g MgSO4 .7H2 O, 1.0 g NaNO3 , 0.3 CaCl2 , 0.1 g K2 HPO4 , 3.0 x 10−2 g Na2 EDTA, 6.0 x 10−3 g H3 BO3 , 2.0 x 10−3 g FeSO4 .7H2 O, 1.4 x 10−3 g MnCl2 , 3.3 x 10−4 g ZnSO4 .7H2 O, 7.0 x 10−6 g Co(NO3 ) 2 .6H2 O, 2.0 x 10−6 g CuSO4 .5H2 O. The inoculum was kept inside a vegetal culture cham‐ ber, where growth occurred under controlled temperature (25 ± 1°C), irradiance in the range of photosynthetically active radiation (175 µE m−2 s−1), photoperiod (12:12) and shaking (250 rpm).

Bubbling column photobioreactors (PBRs) with spherical bases (40 mm diameter and 300 mm height with 300 ml capacity) were used for the experimental setup, keeping an operating volume of 250 ml. In each PBR, the Mann and Myers culture medium was inoculated with the required volume of the corresponding pre‐cultured microalgae in order to have an initial concentration of about 3 × 106 cells ml−1.

During the experimental phase, the culture was aerated with filtered air (0.22‐µm sterile air‐ venting filter, MillexFG50‐Millipore), at a rate of 0.3 v/v/min, enriched with CO<sup>2</sup> at 7% v/v, which was injected on demand to keep a constant pH (pH = 7.5 ± 0.5), as controlled by a pH sensor. The irradiance supplied during this phase was 370 µE m−2 s−1, which was provided by eight fluorescent lamps (58 W, 2150 lumen, Philips, France). The photoperiod was maintained in 12:12 h light/dark and the temperature in 25 ± 1°C.

### **2.2. Experimental setup**

PBRs were operated in batch mode until the end of the exponential growth phase and then under semicontinuous mode till the growth parameters remained constant at the steady state. During the batch culture, an aliquot of 5 ml was daily taken from each PBR for the ana‐ lytical determinations, this volume being replaced with distilled water to keep the operation volume. During the semicontinuous culture, 30% of the culture volume was daily harvested and used for analysis, this volume being replaced with fresh medium.

For each strain of microalgae used in this work (*C. sorokiniana*, *C. vulgaris* and *S. obliquus*), three treatments were conducted: (i) a treatment with inoculated culture medium and 25 mg l−1 paracetamol (with *C. sorokiniana* PCS, *C. vulgaris* PCV, *S. obliquus* PSO), (ii) a treatment with inoculated culture medium and 25 mg l‐1 salicylic acid (with *C. sorokiniana* SCS, *C. vulgaris* SCV, *S. obliquus* SSO) and (iii) a treatment with inoculated culture medium and 25 mg l‐1 diclofenac (with *C. sorokiniana* DCS, *C. vulgaris* DCV, *S. obliquus* DSO). Also, the corresponding positive controls with inoculated culture medium (with *C. sorokiniana* CCS+, *C. vulgaris* CCV+ and *S. obliquus* CSO+) were run. The negative controls consisted of 25 mg l−1 paracetamol (CP−), salicylic acid (CS−) or diclofenac (CD−) in culture medium with no microalgae. For each strain, experiments were run in triplicate and under identical conditions in all the PBRs. Paracetamol (C8 H9 NO2 , ≥99%) was supplied by Roic Pharma, salicylic acid (C7 H6 O3 , ≥99%) by Panreac and diclofenac (C14H10Cl2NNaO2 , ≥99%) by Sigma‐Aldrich.

Throughout the experiments, the growth of the culture was daily monitored by the determi‐ nation of biomass concentration and cell density. The removal of pharmaceuticals was daily determined by the analysis of the remaining concentration of this drug in the culture medium. All analyses were conducted in triplicate.

#### **2.3. Analytical methods**

Biomass concentration (Cb) was determined by optical density at 680 nm (OD680) by spectro‐ photometric (UV/visible spectrophotometer BECKMAN DU640) and verified by dry weight. Preliminary studies were conducted to determinate the relationship between dry weight and OD680 for each strain; as shown in Eq. (1) for *C. sorokiniana*, in Eq. (2) for *C. vulgaris* and in Eq. (3) for *S. obliquus*:

$$\text{OD}\_{\text{Cs.800}} = 5.1834 \times \text{ C}\_b + 0.0128, \text{ R}^2 = 0.9983 \tag{1}$$

$$\text{OD}\_{\text{$$

$$\text{OD}\_{\text{S}\,0.60} = 2.0098 \times \, \text{C}\_{b} + 0.0451, \, \text{R}^{2} = 0.9915 \,\tag{3}$$

Dry weight measurements were performed by filtering 10 ml of culture through a 0.45 µm Whatman filter, which was then washed with 20 ml HCl (0.5 M) to dissolve precipitated salts. Then, the filtrate was dried in an oven at 80°C for 24 h. Additionally, the growth of the culture was measured as cell density (Nc) by cell counting with a Neubauer chamber.

The initial and remaining pharmaceuticals concentration in the culture medium was quantified by a Waters HPLC 600 equipped with a 2487 Dual λ Absorbance Detector. A Phenomenex Gemini‐NX C18 column (5 µm, 250 mm × 4.6 mm) was used for the sepa‐ ration. The wavelengths of detection were 246 nm for paracetamol, 236 nm for salicylic acid and 276 nm for diclofenac. The mobile phase consisted of a mixture of acetonitrile to water (30:70, v/v) for the analysis of paracetamol and a mixture of acetonitrile to water to orthophosphoric acid (70:30:0.1, v/v/v) for salicylic acid and diclofenac. HPLC qual‐ ity acetonitrile (CH3 CN) and orthophosphoric acid (H3 PO4 ) from Prolabo Chemicals and ultrapure water obtained by a Millipore System were used for the preparation of the mobile phase. Before use, each mixture was passed through a Millipore 0.45‐µm pore‐size filter and degasified in an ultrasound bath for 30 min. Before analysis, all the samples were centrifuged twice at 7500 rpm for 10 min (SIGMA 2‐16P centrifuge). For the chro‐ matographic analysis, the mobile phase flow rate was 1 ml min‐1 and the injection volume was 100 µl.

#### **2.4. Data analysis**

During the batch culture, an aliquot of 5 ml was daily taken from each PBR for the ana‐ lytical determinations, this volume being replaced with distilled water to keep the operation volume. During the semicontinuous culture, 30% of the culture volume was daily harvested

For each strain of microalgae used in this work (*C. sorokiniana*, *C. vulgaris* and *S. obliquus*), three treatments were conducted: (i) a treatment with inoculated culture medium and 25 mg l−1 paracetamol (with *C. sorokiniana* PCS, *C. vulgaris* PCV, *S. obliquus* PSO), (ii) a treatment with inoculated culture medium and 25 mg l‐1 salicylic acid (with *C. sorokiniana* SCS, *C. vulgaris* SCV, *S. obliquus* SSO) and (iii) a treatment with inoculated culture medium and 25 mg l‐1 diclofenac (with *C. sorokiniana* DCS, *C. vulgaris* DCV, *S. obliquus* DSO). Also, the corresponding positive controls with inoculated culture medium (with *C. sorokiniana* CCS+, *C. vulgaris* CCV+ and *S. obliquus* CSO+) were run. The negative controls consisted of 25 mg l−1 paracetamol (CP−), salicylic acid (CS−) or diclofenac (CD−) in culture medium with no microalgae. For each strain, experiments were run in triplicate and under

Throughout the experiments, the growth of the culture was daily monitored by the determi‐ nation of biomass concentration and cell density. The removal of pharmaceuticals was daily determined by the analysis of the remaining concentration of this drug in the culture medium.

Biomass concentration (Cb) was determined by optical density at 680 nm (OD680) by spectro‐ photometric (UV/visible spectrophotometer BECKMAN DU640) and verified by dry weight. Preliminary studies were conducted to determinate the relationship between dry weight and OD680 for each strain; as shown in Eq. (1) for *C. sorokiniana*, in Eq. (2) for *C. vulgaris* and in

Dry weight measurements were performed by filtering 10 ml of culture through a 0.45 µm Whatman filter, which was then washed with 20 ml HCl (0.5 M) to dissolve precipitated salts. Then, the filtrate was dried in an oven at 80°C for 24 h. Additionally, the growth of the culture

was measured as cell density (Nc) by cell counting with a Neubauer chamber.

H9 NO2

, ≥99%) by Panreac and diclofenac (C14H10Cl2NNaO2

, ≥99%) was supplied by Roic

+ 0.0128, *R*<sup>2</sup> = 0.9983 (1)

+ 0.0317, *R*<sup>2</sup> = 0.9958 (2)

+ 0.0451, *R*<sup>2</sup> = 0.9915 (3)

, ≥99%)

and used for analysis, this volume being replaced with fresh medium.

identical conditions in all the PBRs. Paracetamol (C8

H6 O3

All analyses were conducted in triplicate.

102 Biological Wastewater Treatment and Resource Recovery

OD *<sup>C</sup>*.*S*<sup>680</sup> = 5.1834 × *Cb*

OD *<sup>C</sup>*.*V*<sup>680</sup> = 2.7933 × *Cb*

OD *<sup>S</sup>*.*O*<sup>680</sup> = 2.0098 × *Cb*

Pharma, salicylic acid (C7

**2.3. Analytical methods**

Eq. (3) for *S. obliquus*:

by Sigma‐Aldrich.

Growth kinetics were resolved in OriginPro 8 using the classic model originally described by Verhulst [21] called logistic model, which has been proved to fit the growth of microalgae [22]. The logistic model fits to a sigmoidal curve that describes the relationship between microor‐ ganisms' growth and density in limited environmental conditions (Eq. (4)).

$$N = \frac{K}{1 + \frac{e^{\nu \sigma}}{e^{\nu \sigma}}} \tag{4}$$

Where *N* (g l−1) is the algal density at time *t* (h), *K* (g l−1) is the carrying capacity (the maximum algal density reached in the culture), *a* is a constant in the logistic model that refers to the relative position from the origin and indicates the duration of the lag phase and *r* (d−1) is the specific growth rate.

Furthermore, the kinetic curves for the removal of pharmaceuticals were fitted to the logistic model. In each case, the parameter *K* (g l−1) is the maximum removal capacity by the microal‐ gae in the culture. The parameter *a* is a constant in the logistic model that indicates the delay in the beginning of the target compounds removal and the parameter *r* (d−1) is the specific removal rate.

Finally, differences among the strains with respect to the kinetic parameters of growth and removal of pharmaceuticals were compared by a non‐parametric test using IBM SPPS Statistics 21. The comparison of means was performed by means of the U Mann‐Whitney test. Significance was defined at *p* ≤ 0.05.

For the removal of pharmaceuticals, the volumetric efficiency for each target compound was calculated as the difference between its average concentration in the influent (*C*inf) and in the effluent (*C*efflu) at every sampling day, considering the daily dilution rate of the cor‐ responding operation stage (D) (Eq. (5)). During the batch culture these efficiencies were cumulatively expressed as milligram per litre and as milligram per litre per day during the steady state of the semicontinuous culture:

$$\text{Volumetric efficiency} = \left(\mathbb{C}\_{\text{in}} - \mathbb{C}\_{\text{off}u}\right) \times D \tag{5}$$

The specific efficiency of the removed pharmaceuticals was calculated as the ratio between the volumetric efficiency and the biomass concentration (Cb) (Eq. (6)). Likewise, during the batch culture these efficiencies were cumulatively expressed as milligram per gram per biomass and as milligram per gram day during the steady state of the semicontinuous culture:

$$\begin{array}{ccccc} \circ & \circ & \circ & \circ & \circ\\ & & & & \circ & \circ \\ & & & & \text{Specific efficiency} \end{array}$$

$$\text{Specific efficiency} = \frac{\left(\mathbb{C}\_{\text{inf}} - \mathbb{C}\_{\text{eith}}\right) \times D}{\mathbb{C}\_{\text{s}}} \tag{6}$$

## **3. Results**

#### **3.1. Growth of the culture**

The growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* during the batch culture, of either the treatments or the controls, showed a typical sigmoidal growth of 8–10 days until reaching the steady state. On the other hand, during the semicontinuous mode, daily dilu‐ tion rates produced instability and the growth rate declined throughout several days until the growth parameters remained constant during the steady state. This instability is a typical behaviour in the microalgae culture when the growth conditions change and it is related with an adaptation phase (**Figures 1**–**3**).

**Figure 1.** Growth curves of *C. sorokiniana* (CCS+ ●, PCS ◯), *C. vulgaris* (CCV+ ■, PCV □) and *S. obliquus* (CSO+ ▲, PSO △) for the paracetamol treatments. Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Experiments were performed in triplicate and bars show standard derivations. Note: experimental points obtained during semicontinuous culture are connected with dashed lines.

Comparative Assessment of Pharmaceutical Removal from Wastewater by the Microalgae *Chlorella sorokiniana*... http://dx.doi.org/10.5772/66772 105

**Figure 2.** Growth curves of *C. sorokiniana* (CCS+, ●; SCS, ◯), *C. vulgaris* (CCV+, ■; SCV, □) and *S. obliquus* (CSO+, ▲; SSO, △) for the salicylic acid treatments. Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Experiments were performed in triplicate and bars show standard derivations. Note: experimental points obtained during semicontinuous culture are connected with dashed lines.

**Figure 3.** Growth curves of *C. sorokiniana* (CCS+, ●; DCS, ◯), *C. vulgaris* (CCV+, ■; DCV, □) and *S. obliquus* (CSO+, ▲; DSO, △) for the diclofenac treatments. Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Experiments were performed in triplicate and bars show standard derivations. Note: experimental points obtained during semicontinuous culture are connected with dashed lines.

#### *3.1.1. Growth of the culture under paracetamol addition*

cumulatively expressed as milligram per litre and as milligram per litre per day during the

 Volumetric efficiency = (*C*inf − *C*efflu) × *D* (5) The specific efficiency of the removed pharmaceuticals was calculated as the ratio between the volumetric efficiency and the biomass concentration (Cb) (Eq. (6)). Likewise, during the batch culture these efficiencies were cumulatively expressed as milligram per gram per biomass

The growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* during the batch culture, of either the treatments or the controls, showed a typical sigmoidal growth of 8–10 days until reaching the steady state. On the other hand, during the semicontinuous mode, daily dilu‐ tion rates produced instability and the growth rate declined throughout several days until the growth parameters remained constant during the steady state. This instability is a typical behaviour in the microalgae culture when the growth conditions change and it is related with

**Figure 1.** Growth curves of *C. sorokiniana* (CCS+ ●, PCS ◯), *C. vulgaris* (CCV+ ■, PCV □) and *S. obliquus* (CSO+ ▲, PSO △) for the paracetamol treatments. Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Experiments were performed in triplicate and bars show standard derivations. Note: experimental points obtained during semicontinuous culture are connected with dashed lines.

 \_\_\_\_\_\_\_\_\_\_\_\_\_ *C*b

(6)

and as milligram per gram day during the steady state of the semicontinuous culture:

Specific efficiency <sup>=</sup> (*C*inf <sup>−</sup> *<sup>C</sup>*efflu) <sup>×</sup> *<sup>D</sup>*

steady state of the semicontinuous culture:

104 Biological Wastewater Treatment and Resource Recovery

**3. Results**

**3.1. Growth of the culture**

an adaptation phase (**Figures 1**–**3**).

The microalgae growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* under the presence of paracetamol, the corresponding positive controls and their respective fittings to the logistic kinetic model are represented as values of biomass concentration versus time in **Figure 1**. The


differences among the treatments were analysed according to growth kinetic parameters, as shown in **Table 1**.

*C*bo , initial biomass; *N*co , initial number of cells; *C*bm, maximum biomass; *N*cm, maximum number of cells; *K*, carrying capacity; *a*, constant of logistic kinetic model; *r*, microalgae growth rate; *R*<sup>2</sup> , correlation coefficient.

**Table 1.** Experimental data (*C*bo , *N*co , *C*bm, *N*cm) and logistic model kinetic parameters (*K*, *a*, *r*) determined for the growth of positive controls and treatments with paracetamol of *C. sorokiniana* (CCS+, PCS), *C. vulgaris* (CCV+, PCV) and *S. obliquus* (CSO+, PSO), *n* = 3.

The addition of paracetamol increased the lag phase of the strains of the genus *Chlorella* com‐ pared with the positive controls (*p* ≤ 0.05), as it can be seen for the values of the parameter *a* in **Table 1**. However, in the case of *S. obliquus*, the addition of the drug did not modify the beginning of the exponential growth phase compared with the positive control (*p* > 0.05). Furthermore, there were significant differences among the treatments with paracetamol, *C. vulgaris* showed a quite longer lag phase than *S. obliquus* and this one than *C. sorokiniana*.

At the end of the batch culture, the biomass concentration was increased above 49% by the presence of paracetamol in the *C. sorokiniana* culture (CCS+, 1.40 ± 0.29 g l−1; PCS, 2.09 ± 0.02 g l −1) and was increased above 31% in the *C. vulgaris* culture (CCV+, 2.60 ± 0.15 g l−1; PCV, 3.42 ± 0.15 g l−‐1) compared with their positive control, as shown in **Figure 1** and confirmed by *K* values in **Table 1**. However, *S. obliquus* culture was not significantly modified by the addition of the drug and the maximum algal density reached in the treatment (PSO, 3.27 ± 0.15 g l−1) was similar to the positive control (CSO+, 3.46 ± 0.08 g l−1). In spite of the different response of the strains to the presence of paracetamol, there were not significant differences between PCV and PSO, even though the value reached for the parameter *K* in the case of CSO+ was signifi‐ cantly larger than for CCV+. Still, the carrying capacity of the PCS treatment was significantly lower than for PCV and PSO.

Respect to microalgae growth rate (*r*), there was significant differences between the paracetamol treatment for *C. vulgaris* (PCV, 1.08 ± 0.04 d−1) and the corresponding positive control (CCV+, 0.84 ± 0.06 d−1). However, likewise the *K* parameter, the growth rate was neither modified in the case of *S. obliquus* treatment (PSO, 1.12 ± 0.03 d−1) compared with the corresponding positive control (CSO+, 1.16 ± 0.07 d−1). Also, no significant differences were detected in the case of *C. sorokiniana* (CCS+, 0.94 ± 0.06 d−1; PCS, 0.96 ± 0.07 d−1). In addition, when comparing the treatments of the three strains, there were not significant differences between the paracetamol treatments of *C. vulgaris* and *S. obliquus* strains (PCV, PSO) despite there were significant differences between their respective positive controls (CCV+, CSO+).

#### *3.1.2. Growth of the culture under salicylic acid addition*

differences among the treatments were analysed according to growth kinetic parameters, as

**CCS+ PCS CCV+ PCV CSO+ PSO**

2.17 × 108 ± 4.77 × 107 ± 4.62 × 107 ±

0.20 × 107

0.01 × 107

, correlation coefficient.

, *C*bm, *N*cm) and logistic model kinetic parameters (*K*, *a*, *r*) determined for the

0.20 × 108

, initial number of cells; *C*bm, maximum biomass; *N*cm, maximum number of cells; *K*, carrying

(g l−1) 0.04 0.04 0.11 0.11 0.08 0.08

 (cell ml−1) 3.20 × 106 3.20 × 106 1.21 × 106 1.21 × 106 8.35 × 105 8.35 × 105 *C*bm (g l−1) 1.41 ± 0.29 2.05 ± 0.03 2.48 ± 0.11 3.35 ± 0.08 3.33 ± 0.06 3.09 ± 0.20

> 1.18 × 108 ±

0.09 × 108

*a* 3.77 ± 0.01 4.33 ± 0.21 4.47 ± 0.25 5.58 ± 0.03 5.45 ± 0.43 4.97 ± 0.31 *K* (g l−1) 1.40 ± 0.29 2.09 ± 0.02 2.60 ± 0.15 3.42 ± 0.15 3.46 ± 0.08 3.27 ± 0.15 *r* (d−1) 0.94 ± 0.06 0.96 ± 0.07 0.84 ± 0.06 1.08 ± 0.04 1.16 ± 0.07 1.12 ± 0.03 *R*<sup>2</sup> 0.9935 0.9939 0.9971 0.9968 0.9874 0.9886

The addition of paracetamol increased the lag phase of the strains of the genus *Chlorella* com‐ pared with the positive controls (*p* ≤ 0.05), as it can be seen for the values of the parameter *a* in **Table 1**. However, in the case of *S. obliquus*, the addition of the drug did not modify the beginning of the exponential growth phase compared with the positive control (*p* > 0.05). Furthermore, there were significant differences among the treatments with paracetamol, *C. vulgaris* showed a quite longer lag phase than *S. obliquus* and this one than *C. sorokiniana*.

growth of positive controls and treatments with paracetamol of *C. sorokiniana* (CCS+, PCS), *C. vulgaris* (CCV+, PCV)

At the end of the batch culture, the biomass concentration was increased above 49% by the presence of paracetamol in the *C. sorokiniana* culture (CCS+, 1.40 ± 0.29 g l−1; PCS, 2.09 ± 0.02 g

−1) and was increased above 31% in the *C. vulgaris* culture (CCV+, 2.60 ± 0.15 g l−1; PCV, 3.42 ± 0.15 g l−‐1) compared with their positive control, as shown in **Figure 1** and confirmed by *K* values in **Table 1**. However, *S. obliquus* culture was not significantly modified by the addition of the drug and the maximum algal density reached in the treatment (PSO, 3.27 ± 0.15 g l−1) was similar to the positive control (CSO+, 3.46 ± 0.08 g l−1). In spite of the different response of the strains to the presence of paracetamol, there were not significant differences between PCV and PSO, even though the value reached for the parameter *K* in the case of CSO+ was signifi‐ cantly larger than for CCV+. Still, the carrying capacity of the PCS treatment was significantly

Respect to microalgae growth rate (*r*), there was significant differences between the paracetamol treatment for *C. vulgaris* (PCV, 1.08 ± 0.04 d−1) and the corresponding positive control (CCV+, 0.84 ± 0.06 d−1). However, likewise the *K* parameter, the growth rate was neither modified in the case of *S. obliquus* treatment (PSO, 1.12 ± 0.03 d−1) compared with the corresponding positive control (CSO+, 1.16 ± 0.07 d−1). Also, no significant differences were detected in the case of *C. sorokiniana* (CCS+, 0.94 ± 0.06 d−1; PCS, 0.96 ± 0.07 d−1). In

shown in **Table 1**.

*N*cm (cell ml−1) 2.12 × 108

, initial biomass; *N*co

**Table 1.** Experimental data (*C*bo

and *S. obliquus* (CSO+, PSO), *n* = 3.

±

106 Biological Wastewater Treatment and Resource Recovery

4.20 × 108 ±

0.22 × 108

capacity; *a*, constant of logistic kinetic model; *r*, microalgae growth rate; *R*<sup>2</sup>

, *N*co

0.49 × 108

*C*b0

*N*c0

*C*bo

l

lower than for PCV and PSO.

The microalgae growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* under the presence of salicylic acid, the corresponding positive controls and their fittings to the logistic kinetic model, are represented as values of biomass concentration versus time in **Figure 2**. The differ‐ ences among the treatments were analysed according to growth kinetic parameters, as shown in **Table 2**.


*C*bo , initial biomass; *N*co , initial number of cells; *C*bm, maximum biomass; *N*cm, maximum number of cells; *K*, carrying capacity; *a*, constant of logistic kinetic model; *r*, microalgae growth rate; *R*<sup>2</sup> , correlation coefficient.

**Table 2.** Experimental data (*C*bo , *N*co , *C*bm, *N*cm) and logistic model kinetic parameters (*K*, *a*, *r*) determined for the growth of positive controls and treatments with salicylic acid of *C. sorokiniana* (CCS+, SCS), *C. vulgaris* (CCV+, SCV) and *S. obliquus* (CSO+, SSO), *n*=3 .

Regarding the parameter *a*, the addition of salicylic acid increased significantly the lag phase of the strains *C. vulgaris* and *S. obliquus* compared with the positive controls. Also, *C. sorokiniana* treatment showed a higher *a* value than the positive control, in spite of the differ‐ ence being not significant (**Table 2**). Comparing the treatments with salicylic acid, *C. vulgaris* showed a quite longer lag phase than *C. sorokiniana* and *S. obliquus*.

As it can be seen in **Figure 2**, the maximum algal density reached at the end of the batch culture was significantly higher in the treatments with salicylic acid for all strains here con‐ sidered as compared with the positive controls. The *C. sorokiniana* treatment increased their biomass concentration above 52% (CCS+, 1.40 ± 0.29 g l−1; SCS, 2.14 ± 0.13 g l−1), *C. vulgaris* above 18% (CCV+, 2.60 ± 0.15 g l−1; SCS, 3.09 ± 0.23 g l−1) and *S. obliquus* above 36% (CSO+, 3.46 ± 0.08 g l−1; SCS, 4.71 ± 0.30 g l−1) over their respective positive controls at the end of the batch culture. However, under salicylic acid, the carrying capacity of *S. obliquus* was significantly larger than those of *C. sorokiniana* and *C. vulgaris*.

The *C. vulgaris* growth rate was significantly increased under the presence of salicylic acid in comparison with the positive control (CCV+, 0.84 ± 0.06 d<sup>−</sup> 1; SCV, 1.69 ± 0.13 d−1). However, it was significantly reduced in the case of *C. sorokiniana* (CCS+, 0.94 ± 0.06 d−1; SCS, 0.77 ± 0.12 d−1) and *S. obliquus* (CSO+ 1.16 ± 0.07 d−1, SSO, 0.72 ± 0.01 d−1). Moreover, the growth rate of SSO was significantly lower than that of SCS and SCV.

#### *3.1.3. Growth of the culture under diclofenac addition*

The microalgae growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* under the presence of diclofenac, the corresponding positive controls and their respective fittings to the logistic kinetic model are represented as values of biomass concentration versus time in **Figure 3**. The differences among the treatments were analysed according to growth kinetic parameters, as shown in **Table 3**.


*C*bo , initial biomass; *N*co , initial number of cells; *C*bm, maximum biomass; *N*cm, maximum number of cells; *K*, carrying capacity; *a*, constant of logistic kinetic model; *r*, microalgae growth rate; *R*<sup>2</sup> , correlation coefficient.

**Table 3.** Experimental data (*C*bo , *N*co , *C*bm, *N*cm) and logistic model kinetic parameters (*K, a, r*) determined for the growth of positive controls and treatments with diclofenac of *C. sorokiniana* (CCS+, DCS), *C. vulgaris* (CCV+, DCV) and *S. obliquus* (CSO+, DSO), *n* = 3 .

There were significant differences respect the parameter *a* (*p* ≤ 0.05) between the positive con‐ trol and the corresponding treatment of each strain of microalgae, reaching higher values in the case of the treatments with diclofenac. Therefore, the presence of diclofenac produced a delayed response in the beginning of the exponential growth phase compared with the positive control. Comparing the treatments with diclofenac, *C. sorokiniana* showed a longer lag phase than *C. vulgaris* and *S. obliquus*

As it can be seen in **Figure 2**, the treatments with diclofenac achieved significantly higher biomass concentration than their respective positive controls. At the end of the batch culture, the *C. sorokiniana* treatment showed an increase of biomass concentration above 45% (CCS+, 1.58 ± 0.11 g l−1; DCS, 2.30 ± 0.03 g l−1), *C. vulgaris* above 35% (CCV+, 1.96 ± 0.13 g l−1; SCV, 2.65 ± 0.10 g l−1) and *S. obliquus* above 11% (CSO+, 1.34 ± 0.03 g l−1; SCS, 1.49 ± 0.05 g l−1) over their respective positive controls. The *C. vulgaris* treatment reached the highest *K* value, which was significantly higher than those determined for the *C. sorokiniana* and the *S. obliquus* treatments.

With respect to microalgae growth rate, there were significant differences between the posi‐ tive control and the corresponding treatment for the two strains of the genus *Chlorella* here used. The *C. sorokiniana* growth rate was significantly increased under the presence of this drug (CCV+, 0.72 ± 0.04 d−1; DCS, 0.96 ± 0.01 d−1). This significant increase was also confirmed for *C. vulgaris* (CCV+, 0.56 ± 0.00 d−1, DCV, 0.74 ± 0.01 d−1). However, no significant differences were determined in the case of *S. obliquus* (CSO+, 0.79 ± 0.03 d−1, DSO, 0.81 ± 0.09 d−1).

#### **3.2. Removal of pharmaceuticals**

The *C. vulgaris* growth rate was significantly increased under the presence of salicylic acid in

it was significantly reduced in the case of *C. sorokiniana* (CCS+, 0.94 ± 0.06 d−1; SCS, 0.77 ± 0.12 d−1) and *S. obliquus* (CSO+ 1.16 ± 0.07 d−1, SSO, 0.72 ± 0.01 d−1). Moreover, the growth rate of

The microalgae growth curves of *C. sorokiniana*, *C. vulgaris* and *S. obliquus* under the presence of diclofenac, the corresponding positive controls and their respective fittings to the logistic kinetic model are represented as values of biomass concentration versus time in **Figure 3**. The differences among the treatments were analysed according to growth kinetic parameters, as

**CCS+ DCS CCV+ DCV CSO+ DSO**

1.73 × 108 ± 5.15 × 107 ± 6.33 × 107 ±

0.32 × 107

0.38 × 107

, correlation coefficient.

0.22 × 108

, initial number of cells; *C*bm, maximum biomass; *N*cm, maximum number of cells; *K*, carrying

, *C*bm, *N*cm) and logistic model kinetic parameters (*K, a, r*) determined for the growth

(g l−1) 0.04 0.04 0.23 0.23 0.14 0.14

 (cell ml−1) 3.39×106 3.39×106 3.53×106 3.53×106 3.40×106 3.40×106 Cbm (g l−1) 1.53 ± 0.11 2.28 ± 0.03 1.69 ± 0.06 2.51 ± 0.13 1.27 ± 0.04 1.40 ± 0.05

> 7.91 × 107 ±

0.19 × 107

*a* 3.31 ± 0.16 4.24 ± 0.00 2.60 ± 0.05 3.57 ± 0.12 3.30 ± 0.24 3.76 ± 0.37 *K* (g l−1) 1.58 ± 0.11 2.30 ± 0.03 1.96 ± 0.13 2.65 ± 0.10 1.34 ± 0.03 1.49 ± 0.05 *r* (d−1) 0.72 ± 0.04 0.96 ± 0.01 0.56 ± 0.00 0.74 ± 0.01 0.79 ± 0.03 0.81 ± 0.09 *R*<sup>2</sup> 0.9907 0.9988 0.9804 0.9915 0.9890 0.9860

There were significant differences respect the parameter *a* (*p* ≤ 0.05) between the positive con‐ trol and the corresponding treatment of each strain of microalgae, reaching higher values in the case of the treatments with diclofenac. Therefore, the presence of diclofenac produced a delayed response in the beginning of the exponential growth phase compared with the positive control. Comparing the treatments with diclofenac, *C. sorokiniana* showed a longer

of positive controls and treatments with diclofenac of *C. sorokiniana* (CCS+, DCS), *C. vulgaris* (CCV+, DCV) and *S. obliquus*

As it can be seen in **Figure 2**, the treatments with diclofenac achieved significantly higher biomass concentration than their respective positive controls. At the end of the batch culture, the *C. sorokiniana* treatment showed an increase of biomass concentration above 45% (CCS+, 1.58 ± 0.11 g l−1; DCS, 2.30 ± 0.03 g l−1), *C. vulgaris* above 35% (CCV+, 1.96 ± 0.13 g l−1; SCV, 2.65 ± 0.10 g l−1) and *S. obliquus* above 11% (CSO+, 1.34 ± 0.03 g l−1; SCS, 1.49 ± 0.05 g l−1) over their respective positive controls. The *C. vulgaris* treatment reached the highest

1; SCV, 1.69 ± 0.13 d−1). However,

comparison with the positive control (CCV+, 0.84 ± 0.06 d<sup>−</sup>

SSO was significantly lower than that of SCS and SCV.

*3.1.3. Growth of the culture under diclofenac addition*

108 Biological Wastewater Treatment and Resource Recovery

±

4.19 × 108 ±

0.04 × 108

capacity; *a*, constant of logistic kinetic model; *r*, microalgae growth rate; *R*<sup>2</sup>

, *N*co

0.22 × 108

lag phase than *C. vulgaris* and *S. obliquus*

shown in **Table 3**.

*N*cm (cell ml−1) 2.49 × 108

, initial biomass; *N*co

**Table 3.** Experimental data (*C*bo

(CSO+, DSO), *n* = 3 .

Cb0

Nc0

*C*bo

The pharmaceutical concentration in each reactor was daily monitored and compared with the concentration of each pharmaceutical in the corresponding negative control. The con‐ centration of the pharmaceuticals here studied decreased over the time in the treatments with microalgae, either with *C. sorokiniana* (PCS, SCS, DCS), *C. vulgaris* (PCV, SCV, DCS) or *S. obliquus* (PSO, SSO, DSO). Meanwhile, no concentration reduction was observed in the neg‐ ative controls (CP−, CS−, CD−). Therefore, it may be assumed that the pharmaceuticals con‐ centration decrease in the microalgae treatments was due to the removal by the microalgae.

#### *3.2.1. Removal of paracetamol*

The removal curves of paracetamol by each strain of microalgae and the corresponding fittings to the logistic kinetic model during the batch mode are displayed in **Figure 4(a)**. In addition, differences among the treatments were analysed according to removal kinetic parameters, as shown in **Table 4**.

**Figure 4.** Volumetric efficiency in the removal of paracetamol by *C. sorokiniana* (PCS, ●), *C. vulgaris* (PCV, ■) and *S. obliquus* (PSO, ▲) during batch culture (a). Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Volumetric efficiency in the removal of paracetamol (b) at the steady state of the semicontinuous culture. Experiments were performed in triplicate and bars show standard derivations.

Regarding the parameter *a*, there were no significant differences between *C. vulgaris* and *S. obliquus* in the lag phase for the removal of paracetamol. However, *C. sorokiniana* showed a sig‐ nificantly longer response at the beginning of the removal of this drug than the other two strains.


**Table 4.** Logistic model kinetic parameters (*K, a, r*) determined for the removal of paracetamol, salicylic acid and diclofenac in the batch culture of *C. sorokiniana* (PCS, SCS, DCS), *C. vulgaris* (PCV, SCV, DCV) and *S. obliquus* (PSO, SSO, DSO). Volumetric efficiency and specific efficiency attained in the steady state of the semicontinuous culture. n=3.

The parameter *K* values in **Table 4** revealed that *C. sorokiniana* (PCS, 17.62 ± 0.91 mg l−1) reached a carrying capacity 2.8 times higher than *C. vulgaris* (PCV, 6.23 ± 0.02 mg l−1) and 1.7 times higher than *S. obliquus* (PSO, 10.41 ± 1.58 mg l−1). In the same way, the removal rates revealed significant differences among the treatments, with *C. sorokiniana* showing the quick‐ est removal (PCS, 1.01 ± 0.06 d−1) and *C. vulgaris* the slowest one (PCV, 0.77 ± 0.01 d−1), which is in agreement with the determined *K* values.

As a consequence of the different responses obtained for the removal parameters between the strains, at the end of the batch culture, efficiencies in the removal of paracetamol above 67% for *C. sorokiniana*, 21% for *C. vulgaris* and 40% for *S. obliquus* were achieved. These results evidenced a larger removal capacity of paracetamol by *C. sorokiniana*, followed by *S. obliquus*, and *C. vulgaris*.

The average volumetric efficiencies on the paracetamol removal by each strain at the steady stage of the semicontinuous culture are depicted as percentages in **Figure 4(b)**. The paracetamol volumetric efficiency reached values above 41% for *C. sorokiniana*, 12% for *C. vulgaris* and 9% for *S. obliquus*. Moreover, the ratios between the volumetric effi‐ ciency and the microalgae biomass are shown in **Table 4** as specific efficiencies. These results revealed that *C. sorokiniana* cells removed above 7.2 times more paracetamol than *C. vulgaris* and 8.4 times more than *S. obliquus* per gram of biomass. On the other hand, the paracetamol removal per gram of biomass was similar between *S. obliquus* and *C. vulgaris*.

#### *3.2.2. Removal of salicylic acid*

The parameter *K* values in **Table 4** revealed that *C. sorokiniana* (PCS, 17.62 ± 0.91 mg l−1) reached a carrying capacity 2.8 times higher than *C. vulgaris* (PCV, 6.23 ± 0.02 mg l−1) and 1.7 times higher than *S. obliquus* (PSO, 10.41 ± 1.58 mg l−1). In the same way, the removal rates revealed significant differences among the treatments, with *C. sorokiniana* showing the quick‐ est removal (PCS, 1.01 ± 0.06 d−1) and *C. vulgaris* the slowest one (PCV, 0.77 ± 0.01 d−1), which

**Table 4.** Logistic model kinetic parameters (*K, a, r*) determined for the removal of paracetamol, salicylic acid and diclofenac in the batch culture of *C. sorokiniana* (PCS, SCS, DCS), *C. vulgaris* (PCV, SCV, DCV) and *S. obliquus* (PSO, SSO, DSO). Volumetric efficiency and specific efficiency attained in the steady state of the semicontinuous

**Paracetamol PCS PCV PSO** *a* 4.49 ± 0.24 3.84 ± 0.01 3.19 ± 0.58 *K* (mg l−1) 17.62 ± 0.91 6.23 ± 0.02 10.41 ± 1.58 *r* (d−1) 1.01 ± 0.06 0.77 ± 0.01 0.86 ± 0.21 *R*<sup>2</sup> 0.9941 0.9827 0.9766 Volumetric efficiency (mg l−1 d−1) 3.13 ± 0.22 0.95 ± 0.05 0.72 ± 0.07 Specific efficiency (mg g biomass−1 d−1) 2.68 ± 0.26 0.32 ± 0.02 0.37 ± 0.03 **Salicylic acid SCS SCV SSO** *a* 10.20 ± 3.16 4.09 ± 0.87 4.11 ± 0.16 *K* (mg l−1) 17.68 ± 0.96 6.44 ± 0.63 24.67 ± 0.32 *r* (d−1) 4.07 ± 1.21 0.84 ± 0.17 0.76 ± 0.03 *R*<sup>2</sup> 0.9919 0.9947 0.9973 Volumetric efficiency (mg l−1 d−1) 6.98 ± 0.31 1.72 ± 0.15 7.55 ± 0.01 Specific efficiency (mg g biomass−1 d−1) 8.34 ± 1.21 0.67 ± 0.06 1.85 ± 0.02 **Diclofenac DCS DCV DSO** *a* 3.88 ± 0.62 3.23 ± 0.02 3.01 ± 0.38 *K* (mg l−1) 14.55 ± 0.73 15.52 ± 0.26 22.43 ± 0.20 *r* (d−1) 2.03 ± 0.33 1.44 ± 0.05 1.25 ± 0.19 *R*<sup>2</sup> 0.9626 0.9755 0.9690 Volumetric efficiency (mg l−1 d−1) 2.18 ± 0.39 1.53 ± 0.32 5.66 ± 0.39 Specific efficiency (mg g biomass−1 d−1) 1.73 ± 0.38 0.97 ± 0.19 5.21 ± 0.18

110 Biological Wastewater Treatment and Resource Recovery

As a consequence of the different responses obtained for the removal parameters between the strains, at the end of the batch culture, efficiencies in the removal of paracetamol above 67% for *C. sorokiniana*, 21% for *C. vulgaris* and 40% for *S. obliquus* were achieved. These results evidenced a larger removal capacity of paracetamol by *C. sorokiniana*, followed by *S. obliquus*,

is in agreement with the determined *K* values.

and *C. vulgaris*.

culture. n=3.

The removal curves of salicylic acid by each strain of microalgae and the corresponding fittings to the logistic kinetic model during the batch mode are displayed in **Figure 5(a)**. In addition, differences among the treatments were analysed according to removal kinetic parameters, as shown in **Table 4**.

**Figure 5.** Volumetric efficiency in the removal of salicylic acid by *C. sorokiniana* (SCS, ●), *C. vulgaris* (SCV, ■) and *S. obliquus* (SSO, ▲) during batch culture (a). Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Volumetric efficiency in the removal of salicylic acid (b) at the steady state of the semicontinuous culture. Experiments were performed in triplicate and bars show standard derivations.

In the case of *C. sorokiniana* there were significant differences respect the parameter *a*, which indicated that the beginning of the removal of salicylic acid had a delayed response as com‐ pared with the lag phase of *C. vulgaris* and *S. obliquus*.

The results obtained for the maximum removal capacity (*K* parameter) revealed that *S. obliquus* (SSO, 24.67 ± 0.32 mg l−1) removed 1.4 times more salicylic acid than *C. sorokiniana* (SCS, 17.68 ± 0.96 mg l−1) and 3.8 time more than *C. vulgaris* (SCV, 6.44 ± 0.63 mg l−1). In spite of salicylic acid removal efficiencies at the end of the batch culture being above 73% by *C. sorokiniana*, 25% by *C. vulgaris*, 93% by *S. obliquus*, the removal rate of *S. obliquus* was significantly lower (SSO, 0.76 ± 0.03 d−1) than that of *C. sorokiniana* (SCS, 4.07 ± 1.21 d−1).

The average salicylic acid volumetric efficiencies by each strain at the steady stage of the semicontinuous culture are depicted as percentages in **Figure 5(b)**. The paracetamol volu‐

metric efficiency did not showed significant differences between the strains *C. sorokiniana* and *S. obliquus*, reaching values above 93% for SCS and 99% for SSO. However, the salicylic acid volumetric efficiency of *C. vulgaris* (above 22%) was more than four times lower than by the other strains. Moreover, the obtained specific efficiencies revealed that *C. sorokiniana* removed above 12.4 times more salicylic acid than *C. vulgaris* and 4.5 times more than *S. obliquus* per gram of biomass (**Table 4**).

#### *3.2.3. Removal of diclofenac*

The removal curves of diclofenac by each strain of microalgae and the corresponding fittings to the logistic kinetic model during the batch mode are displayed in **Figure 6(a)**. In addition, differences among the treatments were analysed according to removal kinetic parameters, as shown in **Table 4**.

**Figure 6.** Volumetric efficiency in the removal of diclofenac by *C. sorokiniana* (DCS, ●), *C. vulgaris* (DCV, ■) and *S. obliquus* (DSO, ▲) during batch culture (a). Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Volumetric efficiency in the removal of salicylic acid (b) at the steady state of the semicontinuous culture. Experiments were performed in triplicate and bars show standard derivations.

The *a* values were similar (*p* > 0.05) for all the treatments, which indicated that the three strains showed the same delayed response in the removal of diclofenac. However, regarding the maximum removal capacity, there were significant differences between the treatment with *S. obliquus* (DSO, 22.43 ± 0.20 mg l−1), which removed 1.5 times more diclofenac than by *C. sorokiniana* (DCS, 14.55 ± 0.73 mg l−1) and 1.4 times more than *C. vulgaris* (DSO, 15.52 ± 0.26 mg l−1).

Concerning the removal rate, the obtained results revealed significant differences among the treatments. The quickest removal rate was attained by *C. sorokiniana* (DCS, 2.03 ± 0.33 d−1), with removal values 1.6 times higher than *S. obliquus* (DSO, 1.25 ± 0.19 d−1) and 1.4 times higher than *C. vulgaris* (DCS, 1.44 ± 0.05 d−1). Despite the differences between strains regarding the removal parameters, at the end of the batch culture, efficiencies above 65% for *C. sorokiniana*, 69% for *C. vulgaris* and 98% for *S. obliquus* were achieved.

The average volumetric efficiencies for the diclofenac removal in the steady stage of the semi‐ continuous culture are showed in **Figure 6(b)**. The volumetric efficiency for *S. obliquus* (above 79%) was 2.6 times higher than for *C. sorokiniana* and 3.7 times higher than for *C. vulgaris.*

Moreover, the ratios between the volumetric efficiency and the microalgae biomass are shown in **Table 4** as specific efficiencies. The determined values revealed that *S. obliquus* removed above 3.0 times more diclofenac than *C. sorokiniana* and above 5.4 times more than *C. vulgaris* per gram of biomass.

## **4. Discussion**

metric efficiency did not showed significant differences between the strains *C. sorokiniana* and *S. obliquus*, reaching values above 93% for SCS and 99% for SSO. However, the salicylic acid volumetric efficiency of *C. vulgaris* (above 22%) was more than four times lower than by the other strains. Moreover, the obtained specific efficiencies revealed that *C. sorokiniana* removed above 12.4 times more salicylic acid than *C. vulgaris* and 4.5 times more than *S. obliquus* per

The removal curves of diclofenac by each strain of microalgae and the corresponding fittings to the logistic kinetic model during the batch mode are displayed in **Figure 6(a)**. In addition, differences among the treatments were analysed according to removal kinetic parameters, as

The *a* values were similar (*p* > 0.05) for all the treatments, which indicated that the three strains showed the same delayed response in the removal of diclofenac. However, regarding the maximum removal capacity, there were significant differences between the treatment with *S. obliquus* (DSO, 22.43 ± 0.20 mg l−1), which removed 1.5 times more diclofenac than by *C. sorokiniana* (DCS, 14.55 ± 0.73 mg l−1) and 1.4 times more than *C. vulgaris* (DSO, 15.52 ± 0.26 mg l−1). Concerning the removal rate, the obtained results revealed significant differences among the treatments. The quickest removal rate was attained by *C. sorokiniana* (DCS, 2.03 ± 0.33 d−1), with removal values 1.6 times higher than *S. obliquus* (DSO, 1.25 ± 0.19 d−1) and 1.4 times higher than *C. vulgaris* (DCS, 1.44 ± 0.05 d−1). Despite the differences between strains regarding the removal parameters, at the end of the batch culture, efficiencies above 65% for *C. sorokini-*

**Figure 6.** Volumetric efficiency in the removal of diclofenac by *C. sorokiniana* (DCS, ●), *C. vulgaris* (DCV, ■) and *S. obliquus* (DSO, ▲) during batch culture (a). Dots correspond to experimental data and continuous lines correspond to fittings by the logistic kinetic model during batch culture. Volumetric efficiency in the removal of salicylic acid (b) at the steady state of the semicontinuous culture. Experiments were performed in triplicate and bars show standard derivations.

The average volumetric efficiencies for the diclofenac removal in the steady stage of the semi‐ continuous culture are showed in **Figure 6(b)**. The volumetric efficiency for *S. obliquus* (above 79%) was 2.6 times higher than for *C. sorokiniana* and 3.7 times higher than for *C. vulgaris.*

*ana*, 69% for *C. vulgaris* and 98% for *S. obliquus* were achieved.

gram of biomass (**Table 4**).

112 Biological Wastewater Treatment and Resource Recovery

*3.2.3. Removal of diclofenac*

shown in **Table 4**.

In view of the obtained results, it may be inferred that the presence of paracetamol, sali‐ cylic acid and diclofenac modified the growth parameters of the strains here studied. In most of the treatments, the addition of the pharmaceutical increased the biomass concentration, which may be explained by the fact that these pharmaceuticals were an additional source of organic carbon. It is well known that the genus *Chlorella* and *Scenedesmus* can have a mixotro‐ phic growth. However, *S. obliquus* did not show a significant increase of microalgae biomass under the addition of paracetamol or diclofenac. These results suggest that the other removal mechanisms, apart from metabolism, may be involved.

The fact that removal curves displayed a similar trend than growth curves points to the asso‐ ciation between the microalgae growth and the removal efficiency of pharmaceuticals.

In view of the obtained results, it may be concluded that paracetamol was more efficiently removed by *C. sorokiniana*, either per litre or per gram of biomass (>67% batch culture, >41% semicontinuous culture), in spite of the biomass concentration reached in the culture being the lowest one among the three strains. Also, the removal rate by *C. sorokiniana* was the fastest one, in spite of showing the lowest growth rate among the paracetamol treatments. However, the addition of paracetamol in the *C. sorokiniana* culture produced the largest increase in the biomass concentration compared with the corresponding positive control (>49%).

On the other hand, *S. obliquus* showed the highest salicylic acid removal capacity at the end of the batch culture (>93%) and also at the steady state of the semicontinuous culture (>99%). However, the removal rate by *S. obliquus* was the lowest one among the salicylic acid treat‐ ments. The highest removal rate was reached by *C. sorokiniana*, which showed a removal per gram of biomass 4.5 times larger than *S. obliquus.* Furthermore, the increase of biomass under the addition of salicylic acid was above 52% in the *C. sorokiniana* treatment, while for *S. obliquus* was above 36%.

Regarding diclofenac, despite *C. sorokiniana* cells attained a higher removal rate and the higher growth rate, it may be stated that *S. obliquus* was the strain that reached the highest removal efficiency (>98% batch culture, >79% semicontinuous culture) with more diclofenac removed either per litre or per gram of biomass.

Comparing the three pharmaceuticals, the salicylic acid was more efficiently removed, with *C. sorokiniana* and *S. obliquus* showing the highest efficiencies. Contrarily, the paracetamol was the less efficiently removed. In all cases, *C. vulgaris* showed the lowest efficiencies for the three pharmaceuticals. These results may be related with the specific strain char‐ acteristics, the mechanisms involved in the removal and the particular properties of each pharmaceutical.

As in this work, published results on the removal of ECs by microalgae have revealed dif‐ ferent efficiencies depending on the pollutant and on the microalgae strain. For example, Gattullo et al. [23] demonstrated that *Monoraphidium braunii* was able to remove up to 48% of bisphenol A with an initial concentration of 4 mg l−1. de Wilt et al. [14] reported removal efficiencies by *C. sorokiniana*, grown in wastewater streams, up to 60–100% for diclofenac, ibuprofen, paracetamol and metoprolol. However, under identical conditions, the removal of carbamazepine and trimethoprim was incomplete and did not exceed 30% and 60%, respec‐ tively [14]. Wang et al. [24] studied the removal of phenol by *Chorella* sp. culture, obtaining removal efficiencies up to 100% from an initial concentration of 500 mg l−1 in 7 days. Peng et al. [25] reported removals above 95% of progesterone by *S. obliquus* and *Chlorella pyrenoidosa*, nearly complete removal of norgestrel by *S. obliquus* and almost 40% of norgestrel by *C. pyrenoidosa*. Likewise, Hom‐Díaz et al. [15] studied the elimination of the hormones E2 and EE2 from anaerobic digestate centrate by the microalgae *Selenastrum capricornutum* and *Chlamydomonas reinhardtii*. After 7 days of culture, these authors [15] determined removals above 88% for E2 and above 60% for EE2. Furthermore, Matamoros et al. [26] studied the capability of microalgae‐based wastewater treatment systems to remove diclofenac, among other 25 emerging organic contaminants. These authors [26] determined diclofenac removal efficiencies above 82% under HRT of 4 days and above 92% under HRT of 8 days during the warm season (11–26°C, on a daily average). These efficiencies are higher than the here obtained under an HRT of 80 h and temperature of 25±1°C. Differences must be related, at least to some extent, to the fact that microalgae monocultures were used in this work while Matamoros et al. [26] worked with mixed microalgae strains present in the wastewater, mostly identified as *Stigeoclonium* sp., diatoms, *Chlorella* sp. and *Monoraphidium* sp.

## **5. Conclusions**

Among the here considered strains, *S. obliquus* displayed the highest removal efficiency for salicylic acid and diclofenac, while *C. sorokiniana* did it for paracetamol. On the other hand, *C. vulgaris* showed the lowest efficiencies for the three pharmaceuticals. Comparing the three pharmaceuticals, the salicylic acid was more efficiently removed while paracetamol removal was the less efficient. These differences may be related with the specific strain characteristics, the mechanisms involved in the removal and the particular properties of each pharmaceuti‐ cal. The obtained results pointed to the feasibility of using the microalgae here considered in bioremediation systems and revealed that this sort of studies are key for the selection of the strain, which depends on the application. Still, further research is needed to assess the mechanisms involved in the removal of pharmaceuticals by these strains.

## **Acknowledgements**

Authors thank University of León for funding given to MICROTRAT (project UXXI2016/00128). Carla Escapa and Sergio Paniagua acknowledge the Spanish Ministry of Educations, Culture and Sports for their PhD fellowships (FPU12/03073 and FPU14/05846, respectively). Marta Otero acknowledges University of León for the extension of her RYC‐2010‐05634 contract.

## **Author details**

As in this work, published results on the removal of ECs by microalgae have revealed dif‐ ferent efficiencies depending on the pollutant and on the microalgae strain. For example, Gattullo et al. [23] demonstrated that *Monoraphidium braunii* was able to remove up to 48% of bisphenol A with an initial concentration of 4 mg l−1. de Wilt et al. [14] reported removal efficiencies by *C. sorokiniana*, grown in wastewater streams, up to 60–100% for diclofenac, ibuprofen, paracetamol and metoprolol. However, under identical conditions, the removal of carbamazepine and trimethoprim was incomplete and did not exceed 30% and 60%, respec‐ tively [14]. Wang et al. [24] studied the removal of phenol by *Chorella* sp. culture, obtaining removal efficiencies up to 100% from an initial concentration of 500 mg l−1 in 7 days. Peng et al. [25] reported removals above 95% of progesterone by *S. obliquus* and *Chlorella pyrenoidosa*, nearly complete removal of norgestrel by *S. obliquus* and almost 40% of norgestrel by *C. pyrenoidosa*. Likewise, Hom‐Díaz et al. [15] studied the elimination of the hormones E2 and EE2 from anaerobic digestate centrate by the microalgae *Selenastrum capricornutum* and *Chlamydomonas reinhardtii*. After 7 days of culture, these authors [15] determined removals above 88% for E2 and above 60% for EE2. Furthermore, Matamoros et al. [26] studied the capability of microalgae‐based wastewater treatment systems to remove diclofenac, among other 25 emerging organic contaminants. These authors [26] determined diclofenac removal efficiencies above 82% under HRT of 4 days and above 92% under HRT of 8 days during the warm season (11–26°C, on a daily average). These efficiencies are higher than the here obtained under an HRT of 80 h and temperature of 25±1°C. Differences must be related, at least to some extent, to the fact that microalgae monocultures were used in this work while Matamoros et al. [26] worked with mixed microalgae strains present in the wastewater, mostly

identified as *Stigeoclonium* sp., diatoms, *Chlorella* sp. and *Monoraphidium* sp.

mechanisms involved in the removal of pharmaceuticals by these strains.

Among the here considered strains, *S. obliquus* displayed the highest removal efficiency for salicylic acid and diclofenac, while *C. sorokiniana* did it for paracetamol. On the other hand, *C. vulgaris* showed the lowest efficiencies for the three pharmaceuticals. Comparing the three pharmaceuticals, the salicylic acid was more efficiently removed while paracetamol removal was the less efficient. These differences may be related with the specific strain characteristics, the mechanisms involved in the removal and the particular properties of each pharmaceuti‐ cal. The obtained results pointed to the feasibility of using the microalgae here considered in bioremediation systems and revealed that this sort of studies are key for the selection of the strain, which depends on the application. Still, further research is needed to assess the

Authors thank University of León for funding given to MICROTRAT (project UXXI2016/00128). Carla Escapa and Sergio Paniagua acknowledge the Spanish Ministry of Educations, Culture and Sports for their PhD fellowships (FPU12/03073 and FPU14/05846,

**5. Conclusions**

114 Biological Wastewater Treatment and Resource Recovery

**Acknowledgements**

Carla Escapa Santos, Ricardo Nuno de Coimbra, Sergio Paniagua Bermejo, Ana Isabel García Pérez\* and Marta Otero Cabero\*

\*Address all correspondence to: ana.garcia@unileon.es and marta.otero@unileon.es

Department of Applied Chemistry and Physics, Institute of Environment, Natural Resources and Biodiversity (IMARENABIO), University of León, León, Spain

## **References**


[20] Mann J, Myers J: On Pigments Growth and Photosynthesis of *Phaeodactylum tricornutum*. Journal of Phycology. 1968;**4**:349–355. DOI: 10.1111/j.1529‐8817.1968.tb04707.x

[9] Pal A, He Y, Jekel M, Reinhard M, Gin K: Emerging contaminants of public health sig‐ nificance as water quality indicator compounds in the urban water cycle. Environment

[10] Verlicchi P, Aukidy M, Zambello E: Occurrence of pharmaceutical compounds in urban wastewater: Removal, mass load and environmental risk after a secondary treat‐ ment‐A review. Science of the Total Environment. 2012;**429**:123–155. DOI: 10.1016/j.

[11] Heberer T, Reddersen K, Mechlinski A: From municipal sewage to drinking water: Fate and removal of pharmaceutical residues in the aquatic environment in urban areas.

[12] Ashton D, Hilton M, Thomas K: Investigating the environmental transport of human pharmaceuticals to streams in the United Kingdom. Science of the Total Environment.

[13] Combarros R, Rosas I, Lavin A, Rendueles M, Diaz M: Influence of biofilm on activated carbon on the adsorption and biodegradation of salicylic acid in wastewater. Water, Air,

[14] de Wilt A, Butkovskyi A, Tuantet K, Leal L, Fernandes T, Langenhoff A, Zeeman G: Micropollutant removal in an algal treatment system fed with source separated wastewater streams. Journal of Hazardous Materials. 2016;**304**:84–92. DOI: 10.1016/j.

[15] Hom‐Díaz A, LLorca M, Rodríguez‐Mozaz S, Vicent T, Barceló D, Blanquez P: Microalgae cultivation on wastewater digestate: Beta‐estradiol and 17 alpha‐ethynylestradiol degradation and transformation products identification. Journal of Environmental

[16] Ritchmon A: Handbook of Microalgal Mass Culture. Boca Raton, Florida: CRC Press;

[17] Matamoros V, Uggetti E, García J, Bayona J: Assessment of the mechanisms involved in the removal of emerging contaminants by microalgae from wastewater: A labora‐ tory scale study. Journal of Hazardous Materials. 2016;**301**:197–205. DOI: 10.1016/j.

[18] Petrie B, Barden R, Kasprzyk‐Hordern B: A review on emerging contaminants in wastewaters and the environment: Current knowledge, understudied areas and rec‐ ommendations for future monitoring. Water Research. 2015;**72**:3–27. DOI: 10.1016/j.

[19] Beuckels A, Smolders E, Muylaert K: Nitrogen availability influences phosphorus removal in microalgae‐based wastewater treatment. Water Research. 2015;**77**:98–106.

International. 2014;**71**:46–62. DOI: 10.1016/j.envint.2014.05.025

Water Science and Technology. 2002;**46**:81–88.

2004;**333**:167–184. DOI: 10.1016/j.scitotenv.2004.04.062

and Soil Pollution. 2014;**225**:1858. DOI: 10.1007/s11270‐013‐1858‐9

Management. 2015;**155**:106–113. DOI: 10.1016/j.jenvman.2015.03.003

scitotenv.2012.04.028

116 Biological Wastewater Treatment and Resource Recovery

jhazmat.2015.10.033

1986. 528 p.

jhazmat.2015.08.050

watres.2014.08.053

DOI: 10.1016/j.watres.2015.03.018


## **Pulp Mill Wastewater: Characteristics and Treatment Pulp Mill Wastewater: Characteristics and Treatment**

María Noel Cabrera María Noel Cabrera

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

http://dx.doi.org/10.5772/67537

#### **Abstract**

The production of chemical pulp in recent times is 180 million tons per year; while the production of eucalyptus pulp has increased intensively, especially in the southern hemisphere. The pulp and paper industry has long been considered a large consumer of natural resources (wood and water) and one of the largest sources of pollution to the environment (air, water courses and soil). Important efforts are being made to reduce the pollutant levels and water consumption of the industry. The wastewater composition, and therefore, the efficiency of effluent treatments and characteristics of the discharges to water are strongly dependent on the applied technology and raw materials. Despite a large body of literature on softwood-based wastewater, few studies have examined the characteristics of kraft eucalyptus bleaching effluents and their behaviour in the different biological treatments. The largest secondary treatment systems today use the activated sludge process. Sixty to seventy-five per cent of all the biological effluent treatment plants within the pulp and paper industry use this kind of treatment system. This chapter reviews the current pulping technologies at mills and compares the chemical composition and biological treatment of wastewater between softwood and hardwood bleached pulps.

**Keywords:** pulp mills, hardwood, softwood, kraft pulping, ECF-TCF bleaching

## **1. Introduction**

The pulp and paper mill industry is an intensive consumer of water and natural resources (wood), discharging a variety of liquid, gaseous and solid wastes to the environment. Since the 1970s, a growing awareness of the effects of pulp and paper wastes in the ambience had prompted water and energy consumption levels and the loads of toxic compounds discharge to reduce. One of the most important implemented changes in this regard was made within the mill, wherein chlorine was completely substituted by, that is, chlorine dioxide as the bleaching

© 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, 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.

chemical agent. Another major issue was the implementation of secondary biological treatments. The wastewater composition and hence the effluent treatment efficiencies and characteristics of the discharges are strongly dependent on the technology applied and the raw materials. In the last 25 years, however, the global distribution of pulp producers has significantly changed and so have the species of wood used. Eucalyptus pulp production, for example, is becoming a leader in the hardwood pulp market; Brazil went from being a pulp consumer to a world leader in hardwood pulp production, and since 2008, it has been the fourth largest pulp producer in the world.

## **2. Wood pulp market**

Cellulose pulp is the main raw material in the production of different types of paper and paperboard. It is also used as the absorbent material in diapers and other sanitary products.

The global pulp market has changed intensely in recent years. A few decades ago, this industry was characterized as national character as a supply industry inputs for domestic production of paper and paperboard. Globalization has led to increased competitiveness in the international market, as new players have emerged both at the level of producers and consumers. Within the latter, the appearance of China and India have strongly modified cellulose demand worldwide [1].

**Figure 1** graphically shows the evolution of world's production of wood pulp between 1979 and 2013 according to the data published by FAO [2–6].

**Figure 1.** World pulp production 1979–2013. Data obtained from Ref. [2**].**

It is clearly shown that the world's wood pulp production increased to about 50% in this period, from 120 million tons in 1979 to nearly 180 million tons in 2013. Part of this growth can be explained by the explosive increase in production in non-traditional wood pulp producing regions such as Asia and South America. The main producing regions are still North America with 38% and Europe with 28%, even though in 2013, Asia produced 17% of the wood pulp and South America about 13% [6].

chemical agent. Another major issue was the implementation of secondary biological treatments. The wastewater composition and hence the effluent treatment efficiencies and characteristics of the discharges are strongly dependent on the technology applied and the raw materials. In the last 25 years, however, the global distribution of pulp producers has significantly changed and so have the species of wood used. Eucalyptus pulp production, for example, is becoming a leader in the hardwood pulp market; Brazil went from being a pulp consumer to a world leader in hardwood pulp production, and since 2008, it has been the fourth largest pulp producer in the world.

Cellulose pulp is the main raw material in the production of different types of paper and paperboard. It is also used as the absorbent material in diapers and other sanitary products. The global pulp market has changed intensely in recent years. A few decades ago, this industry was characterized as national character as a supply industry inputs for domestic production of paper and paperboard. Globalization has led to increased competitiveness in the international market, as new players have emerged both at the level of producers and consumers. Within the latter, the appearance of China and India have strongly modified cellulose

**Figure 1** graphically shows the evolution of world's production of wood pulp between 1979

**2. Wood pulp market**

120 Biological Wastewater Treatment and Resource Recovery

demand worldwide [1].

and 2013 according to the data published by FAO [2–6].

**Figure 1.** World pulp production 1979–2013. Data obtained from Ref. [2**].**

Wood pulp grades are categorized according to the pulping process, which can be classified as mechanical, semi-chemical and chemical pulps. In a mechanical process, logs or wood chips are mechanically grinded by abrasive action. In a chemical cooking process, a significant part of the wood components (mainly lignin) is chemically dissolved to obtain a solid compound with high cellulose fibre content. There are two main methods of chemical pulping: (1) sulphite pulping and (2) sulphate (kraft) pulping. The first process—sulphite cooking process—uses aqueous sulphur dioxide (SO<sup>2</sup> ) and a base of calcium, sodium, magnesium or ammonium. The kraft process uses a treatment comprising a mixture of sodium hydroxide and sodium sulphide, known as white liquor, at a high pressure and temperature. The semi-chemical pulping process combines chemical and mechanical methods, where wood chips are first softened or partially cooked with chemicals and then mechanically pulped [7].

**Figures 2** and **3** illustrate the different kinds of pulp produced in 1979 and 2013.

The rise in wood pulp production is due to an increase in chemical pulp production, as the production of mechanical pulp has declined in the same period. Mechanical pulping has the advantage of converting up to 95% of dry weight wood into pulp, although considerable

**Figure 2.** (a) World pulp production by type of pulp in 1979; (b) different kinds of chemical pulps produced in 1979 (Data from FAO [2])**.**

**Figure 3.** (a) World pulp production by type of pulp in 2013; (b) different kind of chemical pulps produced in 2013 (Data from FAO [6]).

amounts of energy are required to do so. The pulp obtained produces a highly opaque paper with good printability, but the physical properties are inferior than chemical pulps and yellowing when exposed to light. Moreover, mechanical pulps are mainly produced from softwood [7].

There are significant changes in the production of chemical pulp. The use of sulphite cooking process in pulp production compared to kraft pulping technology decreased steadily, from 60% in 1925 to 20% in 1967 and 9.2% in 1979 to only 2.4% in 2013 [6, 8]. The superiority of kraft pulping process is explained by the following facts: (1) all wooden materials including low-quality wood can be used as raw material; (2) superior fibre strength of pulp compared to other chemical pulping methods; (3) more simple chemical and energy recovery process; (4) scale of economy of kraft methods prevents competition and (5) low environmental risks in modern mills [9].

A second classification considers the type of wood used by distinguishing softwood or long fibre (produced mainly from pine and spruce) from hardwood or short fibre (produced from eucalyptus, birch, poplar, etc.) [10]. A gradual move from softwood to hardwood can be observed. In 2013, 56% of bleached kraft pulp was produced with long-fibre wood (softwood), while the remaining 44% was produced with short-fibre wood (hardwood) (according to data from Ref. [6]). In 1980, the production capacity of bleached kraft pulp corresponded to 63% of softwood pulp. The entry into the market of non-traditional producing countries such as Brazil, Indonesia, Spain and Portugal, significantly increased the production of hardwood pulp. Eucalyptus bleached pulp production is rapidly increasing (from 8 million tons in 2003 to nearly 15 million in 2015), and eucalyptus wood is thus considered to be the most important raw material of hardwood bleached market pulp in the world [11].

As kraft pulping is by far the most common process used these days, this chapter will focus in the wastewaters generated in this process.

## **3. Main processes description**

#### **3.1. Mechanical pulping**

The oldest method of mechanical pulping is the groundwood process. In this process, round logs are forced against a rotating pulp stone (revolving at peripheral speeds of 1000–1200 m/ min), under specified conditions of pressure and temperature. Atmospheric grinding, pressure grinding and thermo-grinding could be done according to the applied temperature and pressure. In all of them, the temperature levels obtained from the heat applied or from rubbing the logs on the stone soften and break down the fibres structure; and cracks the fibres from the wood matrix [7, 8].

Another common method is the refiner mechanical pulping (RMP). The wood chips are pulled between two rotating disks. Among them, thermomechanical pulping operates like RMP, but under higher temperature and pressure. The high temperature and pressure levels soften the lignin even more than frictional heat, making fibres separation easier. Thermomechanical pulp is stronger than refined mechanical pulp, and still retains the high-yield and cost-effectiveness of mechanical pulps [7].

### **3.2. Chemical pulping**

amounts of energy are required to do so. The pulp obtained produces a highly opaque paper with good printability, but the physical properties are inferior than chemical pulps and yellowing when exposed to light. Moreover, mechanical pulps are mainly produced from soft-

**Figure 3.** (a) World pulp production by type of pulp in 2013; (b) different kind of chemical pulps produced in 2013 (Data

There are significant changes in the production of chemical pulp. The use of sulphite cooking process in pulp production compared to kraft pulping technology decreased steadily, from 60% in 1925 to 20% in 1967 and 9.2% in 1979 to only 2.4% in 2013 [6, 8]. The superiority of kraft pulping process is explained by the following facts: (1) all wooden materials including low-quality wood can be used as raw material; (2) superior fibre strength of pulp compared to other chemical pulping methods; (3) more simple chemical and energy recovery process; (4) scale of economy of kraft methods prevents competition and (5) low environmental risks

A second classification considers the type of wood used by distinguishing softwood or long fibre (produced mainly from pine and spruce) from hardwood or short fibre (produced from eucalyptus, birch, poplar, etc.) [10]. A gradual move from softwood to hardwood can be observed. In 2013, 56% of bleached kraft pulp was produced with long-fibre wood (softwood), while the remaining 44% was produced with short-fibre wood (hardwood) (according to data from Ref. [6]). In 1980, the production capacity of bleached kraft pulp corresponded to 63% of softwood pulp. The entry into the market of non-traditional producing countries such as Brazil, Indonesia, Spain and Portugal, significantly increased the production of hardwood pulp. Eucalyptus bleached pulp production is rapidly increasing (from 8 million tons in 2003 to nearly 15 million in 2015), and eucalyptus wood is thus considered to be the most impor-

As kraft pulping is by far the most common process used these days, this chapter will focus

tant raw material of hardwood bleached market pulp in the world [11].

in the wastewaters generated in this process.

wood [7].

from FAO [6]).

122 Biological Wastewater Treatment and Resource Recovery

in modern mills [9].

#### *3.2.1. Sulphite pulping*

Sulphite process is very versatile, and covers the entire pH range, achieving high fibre flexibility in pulp yields and properties. The cooking process involves the use of aqueous sulphur dioxide (SO2 ) and a base: calcium, sodium, magnesium or ammonium. Sulphite pulping was developed in the second half of the nineteenth century and for several decades, the calcium acid sulphite process was the most common method. However, since 1950, the utilization of bases other than calcium has been a major development. The specific base used will determine the process's chemical and energy recovery system and water use. The use of the relatively cheap calcium base has become obsolete because the cooking chemicals cannot be recovered. Magnesium and sodium bases allow chemical recovery, and magnesium bases are currently the dominant choice in sulphite pulping process [7, 12].

#### *3.2.2. Kraft pulping*

In kraft pulping, white liquor, containing mainly active chemicals—sodium hydroxide and sodium sulphide—is used for cooking the chips at a high temperature (150–170°C) and pressure. Approximately, half of the wood composition degrades and dissolves during cooking. The spent cooking liquor (black liquor) contains reaction products of lignin and hemicelluloses, and is concentrated and burned in a recovery boiler that recovers the cooking chemicals and generates energy. The smelt is dissolved into water to form green liquor (mostly sodium carbonate and sodium sulphide), which then reacts with lime to convert the sodium carbonate into sodium hydroxide regenerating the white liquor. After cooking and washing, a brown pulp (brown stock pulp) is obtained. Printing, writing and tissue papers require the pulp to be bleached which removes the excess lignin and chromophores to produce a "white" pulp.

## **4. Background of pulp mill effluents: environmental fate and effects**

The pulp and paper industry consumes enormous amounts of water and natural resources and is also one of the largest effluents generators. Before the 1970s, wastewaters from the pulp and paper mills were normally discharged directly to the rivers or lakes, without any treatment or even a rough primary treatment. The high organic loads and solid content in the effluents affected the aquatic ecosystem in several ways such as localized damage to the benthic community, oxygen depletion in large areas and numerous changes in fish reproduction and physiology. In the 1980s, studies in Scandinavia, along the Baltic Coast and the Gulf of Bothnia, showed alterations in fish reproduction and increase of diseases and parasites [13, 14]. Studies conducted in USA and Canada in the beginning of the 1990s, under the Environmental Effects Monitoring (EEM) program [15, 16], revealed delayed sexual maturity, smaller gonads, changes in fish reproduction and depression in secondary sexual characteristics in species living downstream of pulp and paper mills discharges.

From the end of the 1970s until now, the main concern regarding effluents is the formation of chlorinated compounds in bleaching plants. In 1985, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) was discovered in the pulp mill effluents, which led to a general concern over the formation of chlorinated organic matter in chlorine bleaching. Consequently, the use of chlorine in the bleach plants gradually decreased until it was completely substituted with chlorine dioxide. In many countries, the environmental control authorities set strict restrictions on the discharges of chlorinated organics, measured as adsorbable organic halogen (AOX), into the aquatic environment. In 1992, the Swedish Environmental Protection Agency limited organochlorines emissions to 1.5 kg AOX/t of pulp and in 1995, Finland's official limit was set at 1.4 kg AOX/t of pulp [14].

Several authors reported that with the replacement of chlorine with chlorine dioxide, the effluent quality improved in AOX levels and the elimination of detectable amount of dioxins, polychlorinated compounds and chloroform [12, 13, 17].

The European Integrated Pollution and Prevention Control [12] has created reference documents (BREF) that set the Best Available Techniques (BAT) for several industrial sectors. The pulp and paper industry has a very defined set of operations to be especially applied in the new mills. Similarly, the International Finance Corporation [18] among others has defined directives that could be required to give financial support for the construction of new mills. For kraft pulp, the most important guidelines are listed in **Table 1**.


**Table 1.** Main BAT guidelines from IFC [18] and/or IPPC Bureau [12] regarding wastewater load minimization in bleached kraft pulp mills.

## **5. Mechanical pulping: wastewater characteristics**

carbonate into sodium hydroxide regenerating the white liquor. After cooking and washing, a brown pulp (brown stock pulp) is obtained. Printing, writing and tissue papers require the pulp to be bleached which removes the excess lignin and chromophores to produce a

The pulp and paper industry consumes enormous amounts of water and natural resources and is also one of the largest effluents generators. Before the 1970s, wastewaters from the pulp and paper mills were normally discharged directly to the rivers or lakes, without any treatment or even a rough primary treatment. The high organic loads and solid content in the effluents affected the aquatic ecosystem in several ways such as localized damage to the benthic community, oxygen depletion in large areas and numerous changes in fish reproduction and physiology. In the 1980s, studies in Scandinavia, along the Baltic Coast and the Gulf of Bothnia, showed alterations in fish reproduction and increase of diseases and parasites [13, 14]. Studies conducted in USA and Canada in the beginning of the 1990s, under the Environmental Effects Monitoring (EEM) program [15, 16], revealed delayed sexual maturity, smaller gonads, changes in fish reproduction and depression in secondary sexual characteristics in species living downstream of pulp and paper mills

From the end of the 1970s until now, the main concern regarding effluents is the formation of chlorinated compounds in bleaching plants. In 1985, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) was discovered in the pulp mill effluents, which led to a general concern over the formation of chlorinated organic matter in chlorine bleaching. Consequently, the use of chlorine in the bleach plants gradually decreased until it was completely substituted with chlorine dioxide. In many countries, the environmental control authorities set strict restrictions on the discharges of chlorinated organics, measured as adsorbable organic halogen (AOX), into the aquatic environment. In 1992, the Swedish Environmental Protection Agency limited organochlorines emissions to 1.5 kg AOX/t of pulp and in 1995, Finland's official limit was set at 1.4

Several authors reported that with the replacement of chlorine with chlorine dioxide, the effluent quality improved in AOX levels and the elimination of detectable amount of dioxins,

The European Integrated Pollution and Prevention Control [12] has created reference documents (BREF) that set the Best Available Techniques (BAT) for several industrial sectors. The pulp and paper industry has a very defined set of operations to be especially applied in the new mills. Similarly, the International Finance Corporation [18] among others has defined directives that could be required to give financial support for the construction of new mills.

**4. Background of pulp mill effluents: environmental fate and effects**

"white" pulp.

124 Biological Wastewater Treatment and Resource Recovery

discharges.

kg AOX/t of pulp [14].

polychlorinated compounds and chloroform [12, 13, 17].

For kraft pulp, the most important guidelines are listed in **Table 1**.

**Figure 4** shows a block diagram of the main part of the mechanical pulp production indicating the sources of emissions to the water from a pulp mill.

**Table 2** shows the specific water consumption and loads before wastewater treatment from the mechanical pulping [12].

**Figure 4.** Main unit operations of the mechanical pulping. Light brown arrows indicate wastewater sources.


BOD5: Biochemical Oxygen Demand; COD: Chemical Oxygen Demand; GW: groundwood pulping; PGW: pressurized groundwood pulping; RMP: refined mechanical pulping; TMP: thermomechanical pulping; ADt: air dry tone (10% water and 90% oven-dry pulp).

**Table 2.** Specific water consumption, organic and nutrient loads before wastewater treatment from the mechanical pulping. Data taken from Ref. [12].

## **6. Kraft pulping: wastewater characteristics**

#### **6.1. Process description and emissions to water**

A kraft pulp mill can be divided into four main parts: (1) raw material handling; (2) pulping line with an almost closed chemical and energy recovery system; (3) bleaching with an open water system and (4) the external wastewater treatment system. **Figure 5** shows the emissions sources to water from a kraft pulp mill.

**Table 3** shows the typical figures for the parameters in different sectors of a kraft pulp mill.

Data on current discharges to water (after wastewater treatment) expressed as loads based on available data from kraft pulp mills within the European Union are given in **Table 4**. **Figure 6** presents a comparison of the discharges to water of different existing mills with the performance of the new mills in South America that are processing eucalyptus wood and apply-

**Figure 5.** Main unit operations of kraft pulping. DOS: dissolved organic substances. Adapted from Ref. [12].


Flow in m3 /Adt, TSS, BOD, AOX, COD and Nitrogen in kg/ADt. Phosphorous in g/ADt.

**Table 3.** Sources of effluents and effluents loads from kraft pulp mill [12, 19]**.**


Flow in m3 /ADt, COD, BOD5 , AOX, TSS, nitrogen and phosphorous in kg/ADt.

1 Eucalyptus strands contain higher levels of phosphorus compared to other forest species used for pulp production. The average level discharged with the effluent is up to 0.12 kg total-P/ADt.

2 Emissions from eucalyptus pulp mills.

**6. Kraft pulping: wastewater characteristics**

A kraft pulp mill can be divided into four main parts: (1) raw material handling; (2) pulping line with an almost closed chemical and energy recovery system; (3) bleaching with an open water system and (4) the external wastewater treatment system. **Figure 5** shows the emissions

**Pulping process BOD5 (kg/ADt) COD (kg/ADt) Nitrogen (kg/ADt) Phosphorous (kg/ADt)**

BOD5: Biochemical Oxygen Demand; COD: Chemical Oxygen Demand; GW: groundwood pulping; PGW: pressurized groundwood pulping; RMP: refined mechanical pulping; TMP: thermomechanical pulping; ADt: air dry tone (10%

**Table 2.** Specific water consumption, organic and nutrient loads before wastewater treatment from the mechanical

GW 8.5–10 20–30 80–100 20–25 PGW 10–13 30–50 90–110 20–30 RMP 10–15 40–60 90–110 20–30 TMP 13–22 50–80 100–130 30–40

**Table 3** shows the typical figures for the parameters in different sectors of a kraft pulp mill. Data on current discharges to water (after wastewater treatment) expressed as loads based on available data from kraft pulp mills within the European Union are given in **Table 4**. **Figure 6** presents a comparison of the discharges to water of different existing mills with the performance of the new mills in South America that are processing eucalyptus wood and apply-

**Figure 5.** Main unit operations of kraft pulping. DOS: dissolved organic substances. Adapted from Ref. [12].

**6.1. Process description and emissions to water**

sources to water from a kraft pulp mill.

126 Biological Wastewater Treatment and Resource Recovery

water and 90% oven-dry pulp).

pulping. Data taken from Ref. [12].

3 Emissions from other hardwood (no eucalyptus) and softwood.

**Table 4.** Reported annual average discharges from kraft pulp mills within the EU [12]**.**

ing the Best Available Techniques (according to the European IPPC Bureau [12] and the IFC Guidelines [18]).

#### **6.2. Bleaching effluents**

Up to 85% of the total effluent volume is generated in the bleaching stage. Therefore, this part of the mill is broadly studied in order to minimize the effluent organic loads (especially the organochlorines loads) without impacting the pulp yield and brightness. Effluent loadings depend on the production process and the raw materials. The degree of delignification of the unbleached pulp, the bleaching process, the washing loss, type of wood, final brightness desired, chemical and water consumption and the degree of plant closure are important indicators of wastewater characteristics [12, 19]. To this end, kappa number is an important mill control parameter. The kappa number quantifies by a redox reaction to the amount of lignin (or the delignification degree) still in the pulp. The higher the kappa number, the higher the lignin content in the pulp. The low lignin amounts to be removed during bleaching, decreases the utilization of bleaching chemicals, which consequently reduces the load to the wastewater treatment. However, if the kappa number were to decrease too much during the cooking then the pulp yield and physical properties will be considerably low [10]. **Table 5** provides performance data of the different processes [12].

**Figure 6.** South America new mills performance compared with mills in North America and Europe (Data from EKONO and author personal sources). The vertical bars depicted in the graphs correspond to the 10th percentile to 90th percentile range. The column "All" corresponds to the average of the values reported.


**Table 5.** Kappa number currently achieved with different delignification technologies and comparison of the calculated effluent COD without considering the washing losses [12].

#### *6.2.1. Hardwood and softwood bleaching effluents*

The effluents from kraft pulp bleaching constitute varying quantities of organic and inorganic substances. The organic typically represents one-third of the dissolved material while the inorganics comprise two-thirds. The solid matter includes mainly fibres, pieces of fibres and the additives used in bleaching. The dissolved organic matter is composed of various species derived from the raw material and formed in the pulping and bleaching process (residual lignin, hemicelluloses and extractives) [19].

Wood material impact on the values of the effluents parameters can be assessed by comparing the figures for bleaching effluents derived from softwood and hardwood pulp. The former has higher COD and colour content than those of hardwood pulp. The compounds responsible for colour are lignin fragments of high molecular weight (HMW), which represents low biodegradability in the biological treatment [20]. Research has compared effluents from softwood and eucalyptus pulps [13, 20, 21] through AOX, COD, BOD5 and colour behaviour of the different kinds of pulp production (conventional bleached pulps and oxygen delignified bleached pulps). According to the findings, softwood and eucalyptus effluents have the same trend in AOX levels. For both conventional pulps, the AOX levels were higher than the corresponding oxygen delignified pulps. Furthermore, as it mentioned earlier, the total COD levels are dependent on the initial kappa numbers. The COD compositions of eucalyptus and softwood effluents are significantly different, where the effluents from the eucalyptus pulps are more biodegradable. The compounds forming the kappa number in softwood and hardwood (especially eucalyptus) differ as well: in softwood, the kappa number mainly representative of lignin, whereas in eucalyptus, the hexenuronic acids (HexA) are a large contributor [22, 23]. In this regard, the most common way to remove the hexenuronic acids is in the early bleaching stages through hot acid hydrolysis (A) and hot chlorine dioxide bleaching (DH) technologies [11, 22, 24].

#### *6.2.2. Chemical composition of the wastewater*

**Delignification technologies**

Conventional cooking

Conventional cooking + oxygen delignification

cooking

Extended cooking + oxygen delignification

Extended/modified

effluent COD without considering the washing losses [12].

range. The column "All" corresponds to the average of the values reported.

128 Biological Wastewater Treatment and Resource Recovery

**Kappa for hardwood Kappa for softwood Calculated COD load (kg/t) from the bleach plant**

14–22 30–35 28–44 60–70

**Figure 6.** South America new mills performance compared with mills in North America and Europe (Data from EKONO and author personal sources). The vertical bars depicted in the graphs correspond to the 10th percentile to 90th percentile

13–15 18–20 26–30 36–40

14–16 18–22 28–32 36–44

8–10 8–12 16–20 16–24

**Table 5.** Kappa number currently achieved with different delignification technologies and comparison of the calculated

**Hardwood Softwood**

The two main types of bleaching methods in use are elemental chlorine free (ECF), when no molecular or gaseous chlorine is dosed in the bleaching, and totally chlorine free (TCF) bleaching [12]. ECF is dominating the bleached chemical pulp market. In 2012, ECF pulp production reached approximately 93% of bleached kraft pulp's world market share. TCF production has declined a little over the last 10 years [25].

Owing to the differences between both the bleaching technologies and chemical composition of the bleaching effluents, it is necessary to study in order to predict and understand the environmental impact associated, and consequently to develop the most suitable treatment that decreases effluent loads and toxicity. A significant number of studies pertaining to the chemical composition of bleaching effluents have been published. Several authors have worked in identifying the chemical compounds in filtrates. More than 500 organic compounds have been identified in bleaching effluents so far. Most compounds identified in bleaching effluents are derived from lignin or other wood components, such as extractives or carbohydrates [26].

The most important difference, when comparing softwood effluents with the eucalyptus effluents, is the higher lignin content in the former and the hexenuronic acid content in the latter [20]. Lignin degradation products were commonly considered as the major precursors of chlorinated compounds. However, the presence of monochlorinated compounds derived from glucuronxylans were identified to be the major components of chlorine dioxide bleaching filtrates of eucalyptus kraft pulps [27, 28].

Other important compounds found in the effluents are wood-derived components: resin acids, fatty acids, phytosterols and retene. Lipophilic hardwood extractives consist of a complex mixture of compounds such as sterols, long chain aliphatic acids and alcohols, waxes, glycerides and sterol esters. If high amounts of these compounds are found in kraft mill effluent, their origin is frequently the spills of black liquor and soap or black liquor transported with the pulp [14, 29].

#### *6.2.3. Molecular weight distributions*

Several authors [14, 30, 31] have worked in determining the molecular weight distribution of the components in the effluents. The importance of determining the molecular weight distribution comes from the fact that significant removal in the biological treatment system is achieved from the low molecular weight (LMW) material. Evidence of this is the increment in the proportion of organic compounds with high molecular weight after biological treatment. Improvements in the removal of high molecular weight material would lead to greater efficiency and improve the effluent quality. Traditionally, the separation between low molecular weight (LMW) and high molecular weight (HMW) is done at 1000 Da. Bleach kraft mill effluents have an extended molecular weight distribution; from diverse kinds of monomeric compounds to large and complex molecules with molecular weights between 10,000 and 30,000 g/mol. The molecular weight distribution depends on the raw material and the bleaching process used. For example, the average molecular weight of organic matter in hardwood kraft pulp effluents is lower than the corresponding softwood effluents [14].

The molecular weight fractions in the bleaching filtrates of oxygen delignified eucalyptus pulps were studied. The HMW fraction contributed to approximately 40% of the total effluent load of COD both in softwood and hardwood ECF bleached pulps production, and about 30–40% to TCF bleached pulps effluents [30, 31]. Additionally, the most remarkable differences between softwood- and hardwood-derived effluents are in the aromatic region. The aromatic lignin-derived structures such as syringyl and guaiacyl units are not important structural elements in HMW effluent materials from ECF bleaching of oxygen delignified hardwood kraft pulps, but are important in softwood HMW effluents [31, 32]. Similarly, the results show that all HMW effluents contained carbohydrates. The carbohydrates found in the examined HMW could have had oligosaccharides, polysaccharides or both present in the effluent, either in dissolved or colloidal form. As can be expected, the HMW hardwood kraft pulps fraction contained more carbohydrates (mainly xylan) than the corresponding samples from softwood kraft pulps. Concerning the presence of carboxylic acids, the HMW samples showed high levels of these groups. They were formed due to the oxidation of lignin structures in the bleaching process [30–32].

Regarding the low molecular weight (LMW) compounds, it can be broadly classified into three main classes: acids, phenolic compounds and neutral compounds. The phenolic compounds and some of the acids are degradation products from lignin, while the resin acids, fatty acids, terpenes and sterols are residues of extractives presents in the raw material [14].

#### *6.2.4. ECF and TCF wastewaters treatability*

Lignin degradation products were commonly considered as the major precursors of chlorinated compounds. However, the presence of monochlorinated compounds derived from glucuronxylans were identified to be the major components of chlorine dioxide bleaching

Other important compounds found in the effluents are wood-derived components: resin acids, fatty acids, phytosterols and retene. Lipophilic hardwood extractives consist of a complex mixture of compounds such as sterols, long chain aliphatic acids and alcohols, waxes, glycerides and sterol esters. If high amounts of these compounds are found in kraft mill effluent, their origin is frequently the spills of black liquor and soap or black liquor transported

Several authors [14, 30, 31] have worked in determining the molecular weight distribution of the components in the effluents. The importance of determining the molecular weight distribution comes from the fact that significant removal in the biological treatment system is achieved from the low molecular weight (LMW) material. Evidence of this is the increment in the proportion of organic compounds with high molecular weight after biological treatment. Improvements in the removal of high molecular weight material would lead to greater efficiency and improve the effluent quality. Traditionally, the separation between low molecular weight (LMW) and high molecular weight (HMW) is done at 1000 Da. Bleach kraft mill effluents have an extended molecular weight distribution; from diverse kinds of monomeric compounds to large and complex molecules with molecular weights between 10,000 and 30,000 g/mol. The molecular weight distribution depends on the raw material and the bleaching process used. For example, the average molecular weight of organic matter in hardwood kraft

The molecular weight fractions in the bleaching filtrates of oxygen delignified eucalyptus pulps were studied. The HMW fraction contributed to approximately 40% of the total effluent load of COD both in softwood and hardwood ECF bleached pulps production, and about 30–40% to TCF bleached pulps effluents [30, 31]. Additionally, the most remarkable differences between softwood- and hardwood-derived effluents are in the aromatic region. The aromatic lignin-derived structures such as syringyl and guaiacyl units are not important structural elements in HMW effluent materials from ECF bleaching of oxygen delignified hardwood kraft pulps, but are important in softwood HMW effluents [31, 32]. Similarly, the results show that all HMW effluents contained carbohydrates. The carbohydrates found in the examined HMW could have had oligosaccharides, polysaccharides or both present in the effluent, either in dissolved or colloidal form. As can be expected, the HMW hardwood kraft pulps fraction contained more carbohydrates (mainly xylan) than the corresponding samples from softwood kraft pulps. Concerning the presence of carboxylic acids, the HMW samples showed high levels of these groups. They were formed due to the oxidation of lignin struc-

Regarding the low molecular weight (LMW) compounds, it can be broadly classified into three main classes: acids, phenolic compounds and neutral compounds. The phenolic compounds

pulp effluents is lower than the corresponding softwood effluents [14].

filtrates of eucalyptus kraft pulps [27, 28].

130 Biological Wastewater Treatment and Resource Recovery

with the pulp [14, 29].

*6.2.3. Molecular weight distributions*

tures in the bleaching process [30–32].

The biological treatment of the effluents from ECF and TCF is almost the same. There is a slight difference in the organic matter constitution among these bleaching effluents, but it is less than other parameters such as raw materials, effluents from the unbleached line, than the bleaching effluent itself.

TCF eucalyptus pulp produced an effluent with 3.5 times the BOD and twice the COD than ECF eucalyptus pulp effluent [30]. Similarly, TCF bleaching effluent had approximately twice the COD in softwood than the ECF effluents [33]. The larger amounts of COD and TOC in the TCF effluents can be explained because the bleaching reagents used in the TCF sequences (O3 , H2 O2 ) are less selective towards residual lignin than the ClO2 use in the ECF sequences. Bleaching of pulps with ozone is known to produce aldehyde and keto groups on carbohydrates, which are highly susceptible to oxidative degradation under alkaline conditions. An alkaline peroxide stage is used to further bleach ozone-treated pulps, resulting in an oxidative degradation of these carbohydrates and thus contributing to higher COD and TOC values in the TCF effluents. Moreover, the hardwood TCF effluents contained more carbohydrates (mainly xylan) than the ECF effluents. An explanation of these differences was that the process conditions in P-stage (long retention time under alkaline conditions) may favour dissolution of xylan from the pulp [30, 33].

However, while TCF effluent contains more dissolved organic matter, it is less coloured than ECF effluent, mainly because of the action of residual reagents (i.e. H<sup>2</sup> O2 ) in the TCF effluent. Normal values of colour at 525 nm in TCF effluents are 300 and 1300 C.U. in ECF effluents [31].

## **7. Kraft pulping: wastewater treatment**

The typical pulp mill wastewater treatment should include primary treatment (neutralization, screening or sedimentation), principally to remove suspended solids, and biological/ secondary treatment. The secondary treatment is mainly done to diminish the organic matter, which is removed by biological degradation, and is particularly useful for the removal of low molecular mass organic matter with a molecular weight of 800 Da or less. Some mills have tertiary treatment to further reduce toxicity, suspended solids, organics or colour [12, 13, 34].

Secondary biological treatment is applied in most types of pulp and paper mills. The most usual methods are activated sludge and aerated lagoons. Some variations of these systems include the use of filters and sequences reactors—Mobil Bed Bioreactor (MBBR) and Membrane Bioreactors (MBR). Sometimes anaerobic treatment is used followed by an aerobic biological stage [12, 18].

Aerated ponds and activated sludge methods are the most common treatment systems in pulp and paper industry. In an aerated pond, wastewater is treated through a combination of physical, biological and chemical processes. They have large residence times between 3 and 20 days, and consequently a large volume. They work with low microorganism concentration (low solids concentration) about 100–300 mg/L. These ponds use aeration devices to add oxygen to the wastewater (normally surface turbine aerators or bottom aerators) and mix the contents of the pond, thereby enhancing the microbial activity. However, due to low efficiency levels and the large surface required, the use of aerated lagoons has drastically diminished [12, 13, 34].

The largest secondary treatment system is activated sludge (60–75% of all the biological effluent treatment plants in pulp and paper industry use activated sludge systems); even in new plants. The advantages of the aerated activated sludge systems compared to the aerobic ponds are that they achieve high removal efficiencies, the process can be well controlled, requires less surface and the microorganisms are adapted to the receiving wastewater. The disadvantages are the high construction and operation costs (especially the energy cost of the aeration systems), the high rate of sludge production and the loss of efficiency due to bulking problems, and consequently, the need to add nutrients to avoid this problem. Sludge handling and nutrient dosage are additional to the energy cost, which is the major component contributing to the operational cost of the biological treatment of process effluents within the pulp and paper industry [34].

#### **7.1. Characteristics of activated sludge treatment**

Two main units of the activated sludge plant are the aeration basin and the sedimentation basin. In the aeration basin, the effluent is treated with a culture of microorganisms (the activated sludge), which is present in a high concentration. **Figure 7** shows a diagram of a pulp mill treatment with the activated sludge system. Activated sludge plants at kraft pulp mills have a retention time of about 15–48 h. The solids concentration in the activated sludge systems is typically 2000–6000 mg/L. The hydraulic residence time is 4–8 h for a conventional system and the cellular residence time (sludge age) is normally 5–15 days. Normal loads are between 0.05 and 0.1 kg BOD/kg sludge for extended aeration and 0.1–0.3 kg BOD/kg sludge for low load process. The common operating temperature is about 35–37°C and the dissolved oxygen (DO) concentration is 1.5–2.0 ppm. The nutrients concentration in relation to the organic matter is important in effluent treatment. Effluents from the wood processing industry generally have a BOD:N:P ratio of 100:(1–2):(0.15–0.3) and the addition of supplemental nutrients is normally required [13, 34].

The removal efficiencies reached vary according to the wastewater residence time and the operating conditions. Normal efficiencies figures are between 85 and 98% BOD<sup>5</sup> removal and 60–85% for COD removal. For AOX, the reduction is about 40–65%, 40–85% for phosphorus and 20–50% for nitrogen. The overall efficiency of TSS removal using primary and secondary treatment is about 85–90% [12].

#### **7.2. Aerobic treatability of the different effluent fractions**

The COD of treated effluent represents how effective a treatment technology is in its ability to remove the total organic material present in the influent. BOD measurements by themselves

physical, biological and chemical processes. They have large residence times between 3 and 20 days, and consequently a large volume. They work with low microorganism concentration (low solids concentration) about 100–300 mg/L. These ponds use aeration devices to add oxygen to the wastewater (normally surface turbine aerators or bottom aerators) and mix the contents of the pond, thereby enhancing the microbial activity. However, due to low efficiency levels and the large surface required, the use of aerated lagoons has drastically diminished [12, 13, 34].

The largest secondary treatment system is activated sludge (60–75% of all the biological effluent treatment plants in pulp and paper industry use activated sludge systems); even in new plants. The advantages of the aerated activated sludge systems compared to the aerobic ponds are that they achieve high removal efficiencies, the process can be well controlled, requires less surface and the microorganisms are adapted to the receiving wastewater. The disadvantages are the high construction and operation costs (especially the energy cost of the aeration systems), the high rate of sludge production and the loss of efficiency due to bulking problems, and consequently, the need to add nutrients to avoid this problem. Sludge handling and nutrient dosage are additional to the energy cost, which is the major component contributing to the operational cost of the biological treatment of process effluents within the

Two main units of the activated sludge plant are the aeration basin and the sedimentation basin. In the aeration basin, the effluent is treated with a culture of microorganisms (the activated sludge), which is present in a high concentration. **Figure 7** shows a diagram of a pulp mill treatment with the activated sludge system. Activated sludge plants at kraft pulp mills have a retention time of about 15–48 h. The solids concentration in the activated sludge systems is typically 2000–6000 mg/L. The hydraulic residence time is 4–8 h for a conventional system and the cellular residence time (sludge age) is normally 5–15 days. Normal loads are between 0.05 and 0.1 kg BOD/kg sludge for extended aeration and 0.1–0.3 kg BOD/kg sludge for low load process. The common operating temperature is about 35–37°C and the dissolved oxygen (DO) concentration is 1.5–2.0 ppm. The nutrients concentration in relation to the organic matter is important in effluent treatment. Effluents from the wood processing industry generally have a BOD:N:P ratio of 100:(1–2):(0.15–0.3) and the addition of supplemental

The removal efficiencies reached vary according to the wastewater residence time and the

60–85% for COD removal. For AOX, the reduction is about 40–65%, 40–85% for phosphorus and 20–50% for nitrogen. The overall efficiency of TSS removal using primary and secondary

The COD of treated effluent represents how effective a treatment technology is in its ability to remove the total organic material present in the influent. BOD measurements by themselves

removal and

operating conditions. Normal efficiencies figures are between 85 and 98% BOD<sup>5</sup>

**7.2. Aerobic treatability of the different effluent fractions**

pulp and paper industry [34].

132 Biological Wastewater Treatment and Resource Recovery

**7.1. Characteristics of activated sludge treatment**

nutrients is normally required [13, 34].

treatment is about 85–90% [12].

**Figure 7.** Diagram of a pulp mill treatment plant with activated sludge as biological treatment.

do not quantify the non-biodegradable or slowly biodegradable organic portion of the effluent. Moreover, studies seem to indicate that the residual colour in pulp mill effluents could be linked to the recalcitrant COD [35].

Recalcitrant organic matter is supposed to be partly responsible for long-term toxicity in receiving waters [21]. As discussed earlier, it is widely reported that the residual recalcitrant organic matter is composed predominantly by high molecular weight components, which are not metabolized due to its size. However, the contributions of high and low molecular weight fractions in bio-treated effluents are dissimilar [36]. In the LMW fraction, a large-scale removal of the chlorinated phenolic compounds, chlorinated resin acids and sterols occurs. In the HMW fraction, the carbohydrates are strongly affected; however, other compounds such as oxidized lignin were less affected [30].

Some findings are possible by comparing the high molecular weight (HMW) and low molecular weight (LMW) fractions of the acidic and alkaline filtrates post biological treatment [32]. In the alkaline filtrate, the COD and TOC in the HMW fraction increased after treatment. The same behaviour was observed with the AOX and lignin content in the acidic filtrate. This is attributable to the formation of soluble bacterial products or to the adsorption of the LMW into HMW matter [32, 35]. In the LMW filtrates, the COD/TOC decreased after biological treatment, as a result of the large removal of highly oxidized organic carbon. The colour increased in the HMW fractions of acid and alkaline filtrates. The biological treatment often leads to increased colour in ECF bleaching effluents due to the creation of new chromophores in the HMW fractions [13, 32].

#### **7.3. Bulking problems in the activated sludge systems**

Two critical operational aspects of an activated sludge plant are maintaining proper control of the dissolved oxygen (DO) concentration in the aeration tank and preserving a good settling sludge. Reduced settleablility results in poor plant performance, as it is difficult to maintain a low concentration of suspended solids in the plant effluent [13, 34]. Activated sludge plants that treat pulp and paper mill wastewaters seem to be particularly prone to this. There are several reasons for poor separation properties, such as filamentous bulking sludge, bulking due to excessive extracellular polymeric substances (EPS), production or formation of small flocs and dispersed biomass [37, 38]. In pulp mill wastewater, bulking is often due to the presence of filamentous bacteria. Common conditions that favour bulking are working at feeding loads ratios out of normal range, deficiencies in nitrogen and phosphorous species or in the level of DO [13]. In kinetic terms, the floc forming microorganisms have a competitive advantage at lower substrate concentrations because that allows the compounds to utilize oxygen and nutrients more efficiently than the not floc forming microorganisms [37].

The presence of filamentous bacteria was examined for two years in 15 French pulp, paper and board mills wastewater. The study of 25 bulking cases attributed the source in 10 cases to be COD hydraulic overloads, in 8 cases to deficient aeration and in 5 cases to nutrient deficiency [39].

### **8. Partial closure in water circuits**

The current market and environmental demands facing pulp and paper mills are the increased closure of the plant circuits and a further reduction or elimination of the wastes produced. The concept of a closed loop mill aims to eliminate discharges to the aquatic environment, recycle and reuse all possible solid and liquid process wastes, and reduce air emissions to the lowest possible quantity and toxicity. However, until today, no kraft mills are operating with complete closure and complete reutilization of the effluents. The most important problem experienced in mills that try to operate for long periods with zero discharge was corrosion caused by chlorides in a number of positions. Nevertheless, great progress has been made in minimizing impacts associated with pulp mill effluents. Water circulation closure methods include dry debarking, effective liquor spilling control, closed screening and washing, condensate stripping and other methods to minimize the loss of wood-derived organic matter. Extended and oxygen delignification can significantly reduce bleach plant effluent loads from kraft pulp mills. The bleach plant is the most important source of effluent within a pulp mill and the chlorinated effluents are more complicated to reutilize within the mill. For this reason, an important trend in bleaching development is to reduce volumes and decrease the effluent loads, especially of chlorinated compounds [40–42].

Up to now, a complete water closed circulation is not available; nevertheless, a partial closure of the water circuits is possible. This can be done segregating the acid and alkaline effluent streams and recirculating the liquids countercurrently from the last bleach stage through the sequence to the brown stock washer. The alkaline effluent could be used for washing the pulp in the unbleached part of the process.

## **9. Conclusions**

Some findings are possible by comparing the high molecular weight (HMW) and low molecular weight (LMW) fractions of the acidic and alkaline filtrates post biological treatment [32]. In the alkaline filtrate, the COD and TOC in the HMW fraction increased after treatment. The same behaviour was observed with the AOX and lignin content in the acidic filtrate. This is attributable to the formation of soluble bacterial products or to the adsorption of the LMW into HMW matter [32, 35]. In the LMW filtrates, the COD/TOC decreased after biological treatment, as a result of the large removal of highly oxidized organic carbon. The colour increased in the HMW fractions of acid and alkaline filtrates. The biological treatment often leads to increased colour in ECF bleaching effluents due to the creation of new chromophores

Two critical operational aspects of an activated sludge plant are maintaining proper control of the dissolved oxygen (DO) concentration in the aeration tank and preserving a good settling sludge. Reduced settleablility results in poor plant performance, as it is difficult to maintain a low concentration of suspended solids in the plant effluent [13, 34]. Activated sludge plants that treat pulp and paper mill wastewaters seem to be particularly prone to this. There are several reasons for poor separation properties, such as filamentous bulking sludge, bulking due to excessive extracellular polymeric substances (EPS), production or formation of small flocs and dispersed biomass [37, 38]. In pulp mill wastewater, bulking is often due to the presence of filamentous bacteria. Common conditions that favour bulking are working at feeding loads ratios out of normal range, deficiencies in nitrogen and phosphorous species or in the level of DO [13]. In kinetic terms, the floc forming microorganisms have a competitive advantage at lower substrate concentrations because that allows the compounds to utilize oxygen

The presence of filamentous bacteria was examined for two years in 15 French pulp, paper and board mills wastewater. The study of 25 bulking cases attributed the source in 10 cases to be COD hydraulic overloads, in 8 cases to deficient aeration and in 5 cases to nutrient deficiency [39].

The current market and environmental demands facing pulp and paper mills are the increased closure of the plant circuits and a further reduction or elimination of the wastes produced. The concept of a closed loop mill aims to eliminate discharges to the aquatic environment, recycle and reuse all possible solid and liquid process wastes, and reduce air emissions to the lowest possible quantity and toxicity. However, until today, no kraft mills are operating with complete closure and complete reutilization of the effluents. The most important problem experienced in mills that try to operate for long periods with zero discharge was corrosion caused by chlorides in a number of positions. Nevertheless, great progress has been made in minimizing impacts associated with pulp mill effluents. Water circulation closure methods include dry debarking, effective liquor spilling control, closed screening and washing, condensate stripping and other methods to minimize the loss of wood-derived organic mat-

and nutrients more efficiently than the not floc forming microorganisms [37].

in the HMW fractions [13, 32].

134 Biological Wastewater Treatment and Resource Recovery

**7.3. Bulking problems in the activated sludge systems**

**8. Partial closure in water circuits**

The pulp and paper industry has been considered a large consumer of wood, energy and water, and an important contributor of pollutant discharges to the environment (air, water courses and soil). However, the last decades have seen a lot of effort in creating solutions such as generating less pollutant wastewaters and reducing the amount and load of the emissions to the environment. The implementation of several measures like the dry debarking of wood, the introduction of extended cooking and oxygen delignification, the reuse of condensates, improvements in washing efficiency and especially the total substitution of chlorine, has brought a significant reduction in effluent flows and in the chlorinated and organic loads generated within the mill. In addition, the introduction of end-of-pipe secondary, and even tertiary, treatments have reduced large amounts of pollutant loads to the environment. However, the need for tertiary treatment is not yet well proven; while it purifies the effluent, the energy costs are high and even forms sludge.

Effluent characteristics are dependent on the production process and the raw materials. ECF eucalyptus pulp production is increasing appreciably but not much information on its effluents is available. The main difference between softwood and eucalyptus pulps is in the kappa number: the kappa number is mainly formed by lignin content in softwood pulp, and the Hexenuronic acids are important contributors to kappa number in eucalyptus pulp. Hence, the bleaching conditions for eucalyptus are less severe and consequently the effluents characteristics are different. Eucalyptus bleaching effluents have lower COD, AOX and colour content and higher biodegradability than the softwood effluents.

The environmental impact of effluent loads and the appropriate treatment can be determined by studying the chemical composition and molecular weight distribution of the bleaching effluents. The HMW in hardwood bleaching wastewaters constituted an important but not prevailing fraction of the wastewater composition (30–65% of the total). The hardwood HMW fraction is mainly composed of non-aromatic structural compounds.

Aerated activated sludge is the most common treatment system in pulp mills. BOD5 removals of 85–98% and COD removals of 60–85% are normally achieved with these systems. For AOX, the reduction is about 40–65%, 40–85% for phosphorus and 20–50% for nitrogen. Bulking problems are common in these systems mainly due to nitrogen deficiencies and phosphorous concentration or the level of DO.

Nowadays, plants that apply the best available technologies have their emissions controlled and present minimum environmental impact at the receiving waters.

The new developments are in the way to close even more the internal circuits in the plant, to reduce the flow discharged. Membrane technologies and similar technologies may be key in this regard.

## **Author details**

María Noel Cabrera

Address all correspondence to: ncabrera@fing.edu.uy

Universidad de la República, Engineering School, Chemical Engineering Institute, Forest Process Engineering, Montevideo, Uruguay

## **References**


[11] Colodette JL, Gomes CM, Rabelo M. Progess in eucalyptus kraft pulp bleaching. In: 2nd International Colloquium on Eucalyptus Pulp (2ICEP); Concepcion-Chile; 2005, pp. 1–18. Available from: http://www.eucalyptus.com.br/icep02/jorge\_colodette.pdf. [Accessed 2009 Oct 15]

problems are common in these systems mainly due to nitrogen deficiencies and phosphorous

Nowadays, plants that apply the best available technologies have their emissions controlled

The new developments are in the way to close even more the internal circuits in the plant, to reduce the flow discharged. Membrane technologies and similar technologies may be key in

Universidad de la República, Engineering School, Chemical Engineering Institute, Forest

[1] Diesen M. Economics of the pulp and paper industry. 1st Ed. Helsinki (Finland): Paperi

[2] FAO. Forest products yearbook, 1979–1990 [Internet]. Rome (Italy); 1992, p. 392. Available

[3] FAO. Forest products yearbook, 1991–1995 [Internet]. Rome (Italy); 1996. Available from:

[4] FAO. Forest products yearbook, 1996–2001 [Internet]. Rome (Italy); 2003. Available from:

[5] FAO. Forest products yearbook, 2002–2006 [Internet]. Rome (Italy); 2008. Available from:

[6] FAO. Forest products yearbook, 2009–2013 [Internet]. Rome (Italy); 2015. Available from:

[7] Smook GA. Handbook for pulp & paper technologists. 3rd ed. Tappi Press; Vancouver- -

[8] Sixta H, editor. Handbook of pulp. 1st Ed. Weinheim: WILEY-VCH Verlag GmbH &Co.

[9] Stenius P. Forest products chemistry. 1st Ed. Gullichsen J, Paulapuro H, editors. Fapet

[10] Gullichsen J. Fiber line operations. In: Gullichsen J, Fogelholm CJ, editors. Chemical

from: http://www.fao.org/forestry/statistics/80570/en/ [Accessed 2009 Oct 8]

http://www.fao.org/forestry/statistics/80570/en/ [Accessed 2009 Oct 15]

http://www.fao.org/forestry/statistics/80570/en/ [Accessed 2009 Sep 8]

http://www.fao.org/forestry/statistics/80570/en/ [Accessed 2010 Oct 15]

http://www.fao.org/forestry/statistics/80570/en/ [Accessed 2016 Jul 11]

KGaA; 2006, 1369 p. doi:10.1002/9783527619887.

pulping. Fapet Oy; Helsinki; 1999, p. 693.

and present minimum environmental impact at the receiving waters.

Address all correspondence to: ncabrera@fing.edu.uy

Process Engineering, Montevideo, Uruguay

ja Puu Oy; 2007, 186 p.

B.C.; 2003, 425 p.

Oy; Helsinki; 2000, 348 p. i

concentration or the level of DO.

136 Biological Wastewater Treatment and Resource Recovery

this regard.

**Author details**

María Noel Cabrera

**References**


[35] Konduru RR, Liss SN, Allen DG. Recalcitrant organics emerging from biological treatment of kraft mill effluents. Water Qual Res J Canada. 2001;**36**(4):737–57. doi:10.1139/ w02-006.

[23] Ventorim G, Oliveira KD, Colodette JL, Da Costa MM. Effect of pulp kappa number, lignin and hexenuronic acid contents on oxygen delignification performance. Sci For Sci.

[24] Ragnar M, Lindström ME. A comparison of emerging technologies. Pap ja Puu. 2004;

[25] Alliance for Environmental Technology. Trends in world bleached chemical pulp production: 1990–2012. AET Reports. 2013. Available from: http://www.aet.org/science\_of\_

[26] Mc Kague B, Carlberg G. Effluent characteristics and composition. In: Dence C, Reeve D, editors. Pulp bleaching-principles and practices. TAPPI Press; Atlanta- GA; 1996, pp.

[27] Freire CSR, Silvestre AJD, Pascoal Neto C. Carbohydrate-derived chlorinated compounds in ECF bleaching of hardwood pulps: Formation, degradation, and contribution to AOX in a bleached kraft pulp mill. Environ Sci Technol. 2003;**37**(4):811–4. doi:10.1021/

[28] Freire CSR, Silvestre AJD, Neto CP, Silva AMS, Evtuguin DV, Cavaleiro JAS. Easily degradable chlorinated compounds derived from glucuronoxylan in filtrates from chlorine dioxide bleaching of eucalyptus globulus kraft pulp. Holzforschung. 2003;**57**(1):81–

[29] Freire CSR, Silvestre AJD, Pascoal Neto C. Oxidized derivatives of lipophilic extractives formed during hardwood kraft pulp bleaching. Holzforschung. 2003;**57**(5):503–12.

[30] Dahlman O, Reimann A, Stromberg L, Mörck R. High molecular weight effluent materials from modern ECF and TCF bleaching. Tappi J. 1995;**78**(12):99–109. ISBN: 978-952-5216-30-1.

[31] Mounteer AH, Passos FML, Borges AC, Silva DO. Detecting structural and functional differences in activated sludge bacterial communities originating from laboratory treatment of elementally and totally chlorine-free bleaching effluents. Can J Microbiol.

[32] Souza L, Mounteer AH, Silva C., Dalvi L. A study on biological removal of recalcitrant organic matter in eucalyptus kraft pulp ECF bleaching filtrates. In: Tappi International Environmental Conference, CD-ROM. Atlanta, GA (USA): Tappi Press; 3–5 May 2003,

[33] Cates D, Eggert C, Yang J, Eriksson K. Comparison of effluents from TCF and ECF bleach-

[34] Hynninen P. Effluent treatment. In: Dahl O, editor. Environmental management and

ecf/eco\_risk/2013\_pulp.html [Accessed 2016 Aug 10].

2006;**71**:87–97.

**86**(1):39–44.

es0200847.

751–65. ISBN:0-89852-063-0.

138 Biological Wastewater Treatment and Resource Recovery

7. doi:10.1515/HF.2003.013.

doi:10.1515/HF.2003.075.

2002;**48**(3):245–55.

ing of kraft pulps. Tappi J. 1995;**78**(12):93–8.

control. 1st Ed. Paperi ja Puu Oy; Helsinki; 2008. pp. 86–116.

pp. 1–8.


#### **Molecular Biomonitoring of Microbial Communities in Tannery Wastewater Treatment Plant for the Removal of Retanning Chemicals Molecular Biomonitoring of Microbial Communities in Tannery Wastewater Treatment Plant for the Removal of Retanning Chemicals**

Adey Feleke Desta, Joyce Nzioki, Solomon Maina and Francesca Stomeo Adey Feleke Desta, Joyce Nzioki, Solomon Maina and Francesca Stomeo

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

http://dx.doi.org/10.5772/67349

#### **Abstract**

This chapter focuses on culture-independent characterization and monitoring of microbial communities in tannery wastewater treatment system, with special reference to the degradation of two xenobiotic chemicals used in retanning processes. Molecular survey of a tannery wastewater treatment system through metagenomic and metatranscriptomic approaches revealed a diverse microbial community in each component of the treatment system with high gene copies for enzymes involved in the degradation of cyclic aromatic compounds such as nitrotoluene. A combination of flow cytometry and molecular fingerprinting methods was used in a lab-scale reactor to monitor the dynamics of the microbes in the sludge and the fate of two retanning chemicals. The identified key microbial communities for the removal of the two xenobiotic chemicals belong to members of the group *Proteobacteria* and the phylum *Bacteroidetes*.

**Keywords:** tannery, retanning chemicals, bacteria, biomonitoring, wastewater, flow cytometry, fingerprinting

## **1. Introduction**

#### **1.1. The leather industry**

The leather industry in developing nations is a sector in continuous growth but leaving behind the toxic pollutants in the environment. The economy of Eastern African countries is predominantly agricultural where the livestock subsector plays a substantial role. Livestock

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

is an integral part of the national agricultural wealth of Eastern African countries serving as sources of power, meat, milk, egg, hides and skins, manure, and other products. Hides and skins, though by-products of animals, have been contributing greatly to the export earnings from the livestock sector since ancient times.

In most parts of sub-Saharan Africa and Asia, tanning is a family business, carried out in small- to medium-scale semi-mechanized units. Tanneries owned by different individuals are frequently grouped tightly in clusters which used to be nonresidential areas. Most of the tanning facilities are strategically located near to rivers and small streams so as to discharge their large amount of heavily polluted wastewater directly to these water bodies. Considering a case study in Ethiopia, the Awash River is used as inputs for small- and large-scale farms of fruits, vegetables, and sugarcane, yet experiencing a significant water quality deterioration. The discharge of properly untreated tannery effluent has caused severe pollution affecting surface and underground water resources, farms irrigated by such water, people working in the farms, and consumers of the farm products, not to mention the aquatic ecosystem.

#### **1.2. The leather manufacturing process**

The production of leather involves the whole process of converting raw hides or skins into useful commodities such as shoes and garments from the meat industry [1]. Hides and skins are processed to react with various chemical substances that prevent them from putrefaction to make them resistant to wetting and keep them supple and durable [2, 3]. It has been reported by Khan and colleagues [4] that about 130 chemicals are used in the entire process of leather production. The production process is generally divided into four main categories, namely, the beamhouse, tanning, retanning, and finishing processes. In this chapter, we will focus on pollutants of the retanning process and their fate during biological treatment processes.

#### **1.3. The retanning process: a closer look**

Retanning, also called post-tanning operation, involves neutralization and fat liquoring to improve the feel and handle of leather and provide frame retarding, water, and abrasion resistance properties [5, 6]. Retanning is carried out by employing various substances such as phenolic and naphthalene resins, melaminic resins, acrylic resins, and polymers [6].

#### *1.3.1. Melaminic resins as retanning agents*

Melaminic resins are condensation products from formaldehyde with amino and amido compounds, such as urea, melamine, and cyanamide (dicyandiamide). The amine resins are polymers synthesized by condensation of urea, formaldehyde, and melamine (2,4,6-triamino-1,3,5-triazine) [7]. The formaldehyde undergoes an addition reaction with amino group of urea or melamine with the formation of N-methylol groups. Urea-formaldehyde resins are synthesized and chemically modified by reaction with a sulfating agent to form a sulfonated soluble product. Regarding melamine, the methylol groups can react with amino or other methylol groups to form methylene or ether bridges based on the reaction scheme for melamine as depicted in **Figure 1**. These resins give light colored leathers with good resistance. Due to their availability, melaminic resins are among the widely used chemicals in industries processing leather to the retanning and finishing steps [7]. The trade name Retanal MD-80 refers to melamine-formaldehyde resin used in retanning of hides and skins.

**Figure 1.** Condensation of urea (a) and melamine (b) using formaldehyde (after Ref. [7]).

#### *1.3.2. Phenolic and naphthalene resins as retanning agents*

is an integral part of the national agricultural wealth of Eastern African countries serving as sources of power, meat, milk, egg, hides and skins, manure, and other products. Hides and skins, though by-products of animals, have been contributing greatly to the export earnings

In most parts of sub-Saharan Africa and Asia, tanning is a family business, carried out in small- to medium-scale semi-mechanized units. Tanneries owned by different individuals are frequently grouped tightly in clusters which used to be nonresidential areas. Most of the tanning facilities are strategically located near to rivers and small streams so as to discharge their large amount of heavily polluted wastewater directly to these water bodies. Considering a case study in Ethiopia, the Awash River is used as inputs for small- and large-scale farms of fruits, vegetables, and sugarcane, yet experiencing a significant water quality deterioration. The discharge of properly untreated tannery effluent has caused severe pollution affecting surface and underground water resources, farms irrigated by such water, people working in

the farms, and consumers of the farm products, not to mention the aquatic ecosystem.

The production of leather involves the whole process of converting raw hides or skins into useful commodities such as shoes and garments from the meat industry [1]. Hides and skins are processed to react with various chemical substances that prevent them from putrefaction to make them resistant to wetting and keep them supple and durable [2, 3]. It has been reported by Khan and colleagues [4] that about 130 chemicals are used in the entire process of leather production. The production process is generally divided into four main categories, namely, the beamhouse, tanning, retanning, and finishing processes. In this chapter, we will focus on pollutants of the retanning process and their fate during biological treatment processes.

Retanning, also called post-tanning operation, involves neutralization and fat liquoring to improve the feel and handle of leather and provide frame retarding, water, and abrasion resistance properties [5, 6]. Retanning is carried out by employing various substances such as

Melaminic resins are condensation products from formaldehyde with amino and amido compounds, such as urea, melamine, and cyanamide (dicyandiamide). The amine resins are polymers synthesized by condensation of urea, formaldehyde, and melamine (2,4,6-triamino-1,3,5-triazine) [7]. The formaldehyde undergoes an addition reaction with amino group of urea or melamine with the formation of N-methylol groups. Urea-formaldehyde resins are synthesized and chemically modified by reaction with a sulfating agent to form a sulfonated soluble product. Regarding melamine, the methylol groups can react with amino or other methylol groups to form methylene or ether bridges based on the reaction scheme for melamine as depicted in **Figure 1**. These resins give light colored leathers with good resistance.

phenolic and naphthalene resins, melaminic resins, acrylic resins, and polymers [6].

from the livestock sector since ancient times.

142 Biological Wastewater Treatment and Resource Recovery

**1.2. The leather manufacturing process**

**1.3. The retanning process: a closer look**

*1.3.1. Melaminic resins as retanning agents*

Phenolic and naphthalene resins are polymers synthesized using phenolic, naphthalene, and their derivatives condensed with urea and formaldehyde. The synthesis reaction which is patented by BASF in 1913 involves reaction of the basic phenolic and/or naphthalene constituents under acidic conditions which results in attachment of the aromatic compounds to one another with the aid of formaldehyde through methylene bridges. Then, they are adjusted to the optimum degree of condensation by making them binuclear or trinuclear and made water soluble by sulfonation or sulfomethylation which are finally adapted by buffering to meet the application requirements (**Figure 2**) [6, 8]. When used on chrome-tanned leather, they specifically impart it to a soft fullness and relaxed grain. These characteristics of mellowness and softness are very desirable in gloves, garment, and soft-type leathers [9].

**Figure 2.** (a) Basic constituents of phenolic and naphthalene resins, (b) structure of phenol formaldehyde condensate, and (c) structure of naphthalene formaldehyde condensate (after Refs. [6, 10]).


**Table1.**Characteristics of tannery wastewater based on studies from different countries and treatment systems.

#### **1.4. Characteristics of tannery wastewater**

The tanning process consumes high amount of water, estimated to be 34–56 m3 of water per ton of hide or skin processed [11]. Out of the total water consumed, 85% is discharged as a wastewater [12]. Interestingly, only 20% of the wet-salted hides/skins are converted into commercial leather, 25% becomes chromium-containing leather waste, and the remainder becomes non-tanned waste or is lost in wastewater as fat, soluble protein, and solid suspended pollutants [13]. Therefore, environmental pollution remains to be a serious problem in the leather sector.

The characteristics of the wastewater vary considerably from tannery to tannery depending on the size of the tannery, the chemicals used for the specific process, the amount of water used, and the type of final product produced by a tannery. The variations of effluent characteristics also occur through each working day in a tannery. According to Calheiros et al. [14], average COD and pH analyzed in 1 day were 2010 mg/l (±516) and 6.98 (±0.05), respectively, whereas 2068 mg/l (±446) and 7.93 (±0.08), respectively, in another day. **Table 1** summarizes the pollution load discharged from individual tannery processing operations.

Most of the studies on pollution load of tanneries do not include chemicals that are involved in the process after the tanning step. This is partly because the pollution load of the chemicals used in the retanning process is included in some parameters such as COD and TDS. The other reason is the absence of fast and cheap method to detect these specific chemicals. Reemtsma et al. [26] reported the presence of benzothiazoles in tannery wastewater in three forms, benzothiazole (BT), methylthiobenzothiazole (MTBT), and monobenzothiazole (MBT), with a dominance of MBT at a concentration of 3.3–6.9 μmol/L. These compounds have been detected in tannery wastewater samples by Fiehn et al. [27] in concentrations of 655 μg/L MBT, 10.5μg/L BT, and 39 μg/L of MTBT. A report by UNIDO [28] indicated that only 22% of all the chemicals used for post-tanning and finishing process is taken up and remained in/on the leather, whereas from the remaining waste chemicals (88%), 23% belongs to fat liquors and 20% to dyestuffs.

In this study, we explore the microbial community in the different components of a treatment plant and expressed genes for the target chemicals Basyntan and Retanal. In addition, we decipher the key microbial subcommunities responsible for the degradation of our target post-tanning chemicals.

## **2. Materials and methods**

### **2.1. Reactor setup and sampling**

**COD (mg/l)**

2250±565

3700 4800±350 1320–54,000

4100–6700

8000 2200 11123±563

2155 3114 2426 5650 **Table 1.**

–

–

–

–

The first column lists the parameters used in the different studies. Parameters with "–" means data are not available.

Characteristics of tannery wastewater based on studies from different countries and treatment systems.

19,755

5025

–

8.2–8.5

14,750

[25]

–

335

29.3

286

–

–

–

7.7

–

[24]

1126

33

83

55

18,884

1147

–

10.5

17,737

[23]

–

166

50.9

35.6

–

915

578

7.79

–

[22]

2983±259

122±8

32±6

630±67

–

–

–

10.8-0.1

6646±557

[21]

–

–

–

–

–

5300

1300

7.7

36,800

[20]

930

–

11

228

–

2004

1660

–

15,152

[19]

680–976

144–170

41,623

–

–

600–955

–

7.0–8.7

–

[18]

144 Biological Wastewater Treatment and Resource Recovery

840–18620

–

41–133

800–6480

–

220–1610

–

–

–

[17]

–

225±18

95±55

–

10,266±1460

2820±140

1505±90

7.06 ± 0.26

18,800–19700 [16]

1470

180

–

440

–

2690

1260

7.4

–

[15]

1000±88

–

**BOD (mg/l)**

**NH4(mg/l)**

**Cr (mg/l)** 0.027±0.075

–

–

92±36

–

6.14±1.1

–

[14]

**S2−(mg/l)**

**TS (mg/l)**

**SS (mg/l)**

**VSS (mg/l) pH**

**TDS (mg/l)**

**Reference**

The data shown in this chapter are from a study conducted on a pilot-scale biological wastewater treatment plant installed in the premises of a privately owned tannery in Modjo town, Ethiopia, 70 km south of the capital Addis Ababa. The system consists of two anaerobic reactors each with volume of 25 m3 : an aerobic reactor with a volume of 50 m3 and subsurface-flow constructed wetland vegetated with the perennial grass *Phragmites australis* (Cav.) (**Figure 3**). Performance of the treatment system was evaluated by taking samples of the influent and the treated effluent water and analyzing the different physicochemical parameters following the procedure in APHA [29].

**Figure 3.** Schematic presentation of the pilot tannery effluent treatment site comprising anaerobic-aerobic reactors integrated with constructed wetland system.

#### **2.2. Metagenomic and metatranscriptomic analyses**

Sludge and sediment samples were taken from the anaerobic, aerobic, and different parts of the constructed wetland. The extraction of DNA and RNA was carried out using Zymo ZR® kit for DNA and Zymo ZR® kit for RNA (Zymo Research, CA, USA), respectively. Shotgun sequencing of the metagenome was conducted by means of Illumina Nextera XT® protocol. Total RNA was sequenced following the Illumina TruSeq® RNA preparation protocol.

The quality of the generated DNA and RNA reads was checked using FastQC toolkit [30]. FASTX-Toolkit was used to dereplicate, screen for ambiguous reads, and trim based on the cutoff value of Phred score >20 [31]. Assembly of the trimmed DNA and RNA reads was performed using Velvet (v 1.1) [32] and Trinity (v 2014) [33], respectively. Ribosomal RNA was removed using the riboPicker software (v 0.4.1) [34]. Binning and normalization were performed using an in-house Perl script. Taxonomic identification was done using BLASTN for the metagenome contigs and BLASTX for the metatranscriptome against a local download of NCBI nonredundant GenBank database. A set of contigs from the metatranscriptomic dataset were analyzed for the frequency of various identified genes, and Blast2GO (v 1) [35] was employed for the annotation of the genes.

#### **2.3. Monitoring of microbial communities for the degradation of retanning chemicals**

A bench-scale sequencing batch reactor (SBR) mimicking the treatment system was set up to analyze the dynamics of microbial community and its functional significance for the removal of the various pollutants in the wastewater. The SBR was operated continuously in cycles of around 72 hours with the fill, react, settle, and draw cycles as depicted in **Figure 4**. A number of abiotic parameters including liquid chromatography-based analysis of the two retanning chemicals (Basyntan and Retanal) were measured at each batch throughout the entire running period. Similarly, flow cytometry-based quantification and sorting of sludge microbial community stained with DAPI were carried out using the MoFlo cell sorter (DakoCytomation, Fort Collins, CO). The sorted cells were processed for taxonomic identification of the different subcommunities using T-RFLP and clone library-based 16S rRNA sequence analysis described in Koch et al. [36]. Correlation analyses between the abiotic parameters and the gated subcommunities were done by Spearman's rankorder correlation coefficient using the program R (http://CRAN.R-project.org) Version 2.14.0.

**Figure 4.** Schematic presentation of the different phases of the lab-scale sequencing batch reactor.

## **3. Results and discussion**

treatment system was evaluated by taking samples of the influent and the treated effluent water and analyzing the different physicochemical parameters following the procedure in APHA [29].

Sludge and sediment samples were taken from the anaerobic, aerobic, and different parts of the constructed wetland. The extraction of DNA and RNA was carried out using Zymo ZR® kit for DNA and Zymo ZR® kit for RNA (Zymo Research, CA, USA), respectively. Shotgun sequencing of the metagenome was conducted by means of Illumina Nextera XT® protocol. Total RNA was sequenced following the Illumina TruSeq® RNA preparation protocol.

**Figure 3.** Schematic presentation of the pilot tannery effluent treatment site comprising anaerobic-aerobic reactors

The quality of the generated DNA and RNA reads was checked using FastQC toolkit [30]. FASTX-Toolkit was used to dereplicate, screen for ambiguous reads, and trim based on the cutoff value of Phred score >20 [31]. Assembly of the trimmed DNA and RNA reads was performed using Velvet (v 1.1) [32] and Trinity (v 2014) [33], respectively. Ribosomal RNA was removed using the riboPicker software (v 0.4.1) [34]. Binning and normalization were performed using an in-house Perl script. Taxonomic identification was done using BLASTN for the metagenome contigs and BLASTX for the metatranscriptome against a local download of NCBI nonredundant GenBank database. A set of contigs from the metatranscriptomic dataset were analyzed for the frequency of various identified genes, and Blast2GO (v 1) [35] was employed for the annotation of the genes.

**2.3. Monitoring of microbial communities for the degradation of retanning chemicals**

A bench-scale sequencing batch reactor (SBR) mimicking the treatment system was set up to analyze the dynamics of microbial community and its functional significance for the removal of the various pollutants in the wastewater. The SBR was operated continuously in cycles of around 72 hours with the fill, react, settle, and draw cycles as depicted in **Figure 4**. A number of abiotic parameters including liquid chromatography-based analysis of the two retanning chemicals (Basyntan and Retanal) were measured at each batch throughout the entire running period. Similarly, flow cytometry-based quantification and sorting of sludge microbial community stained with DAPI were carried out using the MoFlo cell sorter (DakoCytomation, Fort Collins, CO). The sorted cells

**2.2. Metagenomic and metatranscriptomic analyses**

integrated with constructed wetland system.

146 Biological Wastewater Treatment and Resource Recovery

#### **3.1. Performance of the treatment system**

Based on the physicochemical analysis, the untreated wastewater was characterized by its high concentration of sulfate, ammonia nitrogen, total suspended and dissolved solids, as well as high biological and chemical oxygen demands (BOD and COD). The high pH also indicated the alkalinity of the wastewater. Performance of the treatment system with regard to the removal of priority pollutants ranged between 70 and 99% (**Table 2**). The effluent parameters obtained for the COD, sulfate (SO4 2−), sulfide (S2−), nitrate (NO3 ), and ammonia nitrogen (NH<sup>3</sup> - N) were in line with the provisional emission limit values set for tannery effluents in Ethiopia which are 500 mg/l for COD, 1 g/l for SO4 2−, 1 mg/l for S2−, 20 mg/l for NO3 , and 30 mg/l for NH<sup>3</sup> .

#### **3.2. Metagenomic and metatranscriptomic analyses of tannery treatment system**

Shotgun metagenomic analysis of the pilot reactors revealed the presence of seven phyla in the anaerobic reactor and eight phyla in the aerobic and the constructed wetland areas. The most abundant bacterial phyla in the anaerobic and aerobic reactors belonged to phylum *Firmicutes* and *Proteobacteria*, respectively. In the wetland, members of the phyla *Proteobacteria*, *Chlorobi*, and *Chloroflexi* were dominant (**Figure 5**).

A closer look into the dominant phylum *Firmicutes* showed that the genera *Bacillus*, *Clostridium*, and *Tissierella* were relatively the most abundant genera in the anaerobic system; these microorganisms have been implicated in the degradation of aromatic hydrocarbons in other tannery wastewater treatment systems [37].


*Key*: TN, total nitrogen; TP, total phosphorous; TDS, total dissolved solids; TSS, total suspended solids; VSS, volatile suspended solids; total Cr, total chromium (Source: Desta et al. [37])

**Table 2.** Average characteristics of the influent and effluent wastewaters of the integrated treatment system at the time of sludge sampling (concentrations are in mg/l, except for pH).

**Figure 5.** Relative abundance of bacteria as classified in phylum level. Sample sites were classified based on the concentration of salt (measured as TDS) and qualitatively designated as high, medium, and low levels.

The phyla *Bacteroidetes*, *Cyanobacteria*, *Actinobacteria*, and *Chloroflexi* follow the next levels of abundance in the anaerobic reactor, with members identified in the degradation on both priority nutrients and synthetic aromatic compounds [36, 37].

**Parameter Influent Effluent % Removal**

*Key*: TN, total nitrogen; TP, total phosphorous; TDS, total dissolved solids; TSS, total suspended solids; VSS, volatile

**Table 2.** Average characteristics of the influent and effluent wastewaters of the integrated treatment system at the time

**Figure 5.** Relative abundance of bacteria as classified in phylum level. Sample sites were classified based on the

concentration of salt (measured as TDS) and qualitatively designated as high, medium, and low levels.

TN 245.25 ± 76 62.75 ± 14 74 SO4 800 ± 505 35 ± 61 96 TP 15.33 ± 1 4.23 ± 2 72 S2− 55.50 ± 6 4.91 ± 3 91 NO3 310 ± 203 40.25 ± 28 87 NO2 2.08 ± 3 0.03 99 NH<sup>3</sup> 287.70 ± 178 44.28 ± 26 85 COD 12,547.50 ± 3910 395 ± 139 97 BOD 4886.26 ± 266 308.91 ± 24 94 TDS 9470.50 ± 1335 2593.69 ± 344 73 TSS 1155 ± 203 92 ± 11 92 VSS 27,482.75 ± 197 2272.75 ± 724 92 Total Cr 27.25 ± 3 0.95 97

148 Biological Wastewater Treatment and Resource Recovery

pH 10.40 ± 0.3 7.66 ± 0.1

suspended solids; total Cr, total chromium (Source: Desta et al. [37])

of sludge sampling (concentrations are in mg/l, except for pH).

In the aerobic reactor, members of the phyla *Cyanobacteria* and *Deinococcus-Thermus* were the most abundant bacterial groups. The genus *Deinococcus* was more abundant in the aerobic reactor than in any other part of the treatment system. Members of the class *Betaproteobacteria* such as the genera *Burkholderia*, *Rhodocyclus*, and *Nitrosomonas* were identified from the aerobic system and are inferred to be involved in ammonia oxidation and aromatic compound degradation [36, 37].

Metatranscriptomic analysis of biological samples from the anaerobic reactor of the treatment system revealed the presence of genes coding enzymes involved in the degradation of nitrotoluene, chlorocyclohexane, toluene, and benzoate, apart from the enzymes for common anabolic and catabolic pathways (**Figure 6**). Relatively higher number of expressed genes were detected for nitrotoluene degradation coded for the enzymes DNT dehydrogenase (EC 1.2.99.2) and DHAT reductase (EC 1.8.99.3). These enzymes are implicated in the degradation of compounds such as nitrotoluene and related aromatic compounds.

**Figure 6.** Average contig coverage for 17 common metabolic gene anaerobic reactors of tannery WWTP. X-axis, The genes involved in the metabolic pathways, and Y-axis, the average contig coverage. Error bars represent the standard error.

#### **3.3. Dynamics and functional characterization of microbial communities**

Flow cytometric characterization of bacterial community in the sludge of the reactor followed up at bench-scale sequencing batch reactor (SBR) revealed the dynamics, succession, and shift of the microbial subcommunities during the course of reaction, with typical patterns in each batch. Starting from the first batch of the operation of the SBR, changes expressed as shift of clusters in the x- and y-axes were observed in each batch of reaction of the SBR, indicating increase in cell size and proliferation activity of the microbial communities over the whole running period of the SBR. Based on visual evaluation of the histograms of the dot plots, a gate template was created representing 30 clusters during the 14 batches of the reactor run (**Figure 7**). From each gate, cell abundance over the entire reaction period was evaluated.

**Figure 7.** Bacterial community dynamics of tannery wastewater running in sequencing batch reactor (SBR) for 45 days. The first dot plot (initial) refers to the bacterial community in the acclimatized sample used to seed the reactor at the beginning of the SBR. The gate template (top-left box) which is used as the basis for fingerprinting of the different cell types.

Correlation analysis of bacterial cell abundance in each gate with the 13 measured abiotic parameters revealed positive correlations (p<0.05) between removal of the retanning agents and bacterial groups in gates G6, G12, and G20. Considering the different UPLC-based peaks of Basyntan, highly positive correlation was found specifically between peak 1 (ΔB1) and peak 3 (ΔB3) of Basyntan and the cells in G21 and G23. The correlations between the rest of the retanning agents (ΔB2 and ΔR) with the cells in G21 and G23 were still found to be positive (p< 0.05), suggesting the possible role of the cells in G21 and G23 for the biodegradation of Retanal and all the components of Basyntan. In order to have a closer look at the clusters and identify the consistent members in the flow cytometric pattern from each batch of the SBR, eight of the 30 gated subcommunities, namely, G1, G2, G6, G12, G14, G16, G20, and G21 were sorted to analyze the composition and abundance of bacteria in each sorted gate. From all the sorted gates, eight bacterial families and classes belonging to the phyla *Proteobacteria*, *Bacteroidales*, and *Bacteroidetes* were identified using terminal restriction fragments (T-RFs). Out of the eight gated clusters, gate 14 (G14) contained the smallest portion of the sorted bacterial community with predominant members belonging to *Proteobacteria* (6%) and showed strong positive correlation (p< 0.01) with the degradation of Basyntan and Retanal (**Figure 8**).

The gates 16, 20, and 21 (G16, G20, and G21) which showed positive correlations with retanning chemicals degradation were dominated by members of the phylum *Bacteroidetes* constituting 13, 23, and 66%, respectively. *Rhodocyclaceae* (11%), *Brucellaceae* (10%), and unclassified *Proteobacteria* (8%) were the second, third, and fourth abundant groups identified in gate 20 (G20), respectively. The most abundant cells belonged to *Rhodocyclaceae* (48% and 22%). The second most abundant groups in this gate belonged to unclassified *Proteobacteria* (16 and 8%), followed by the family *Brucellaceae* (8%). The families *Caulobacteriaceae*, *Xanthomonadaceae*, and the phylum *Bacteroidetes* constituted a small proportion (15%) of the total community in the gate. The role of the identified bacterial groups in the degradation of the retanning agents is reflected by the positive correlation (p< 0.05) detected between cell abundances and removal of the retanning agents (**Figure 9**).

running period of the SBR. Based on visual evaluation of the histograms of the dot plots, a gate template was created representing 30 clusters during the 14 batches of the reactor run (**Figure 7**). From each gate, cell abundance over the entire reaction period was evaluated.

150 Biological Wastewater Treatment and Resource Recovery

Correlation analysis of bacterial cell abundance in each gate with the 13 measured abiotic parameters revealed positive correlations (p<0.05) between removal of the retanning agents and bacterial groups in gates G6, G12, and G20. Considering the different UPLC-based peaks of Basyntan, highly positive correlation was found specifically between peak 1 (ΔB1) and peak 3 (ΔB3) of Basyntan and the cells in G21 and G23. The correlations between the rest of the retanning agents (ΔB2 and ΔR) with the cells in G21 and G23 were still found to be positive (p< 0.05), suggesting the possible role of the cells in G21 and G23 for the biodegradation of Retanal and all the components of Basyntan. In order to have a closer look at the clusters and identify the consistent members in the flow cytometric pattern from each batch of the SBR, eight of the 30 gated subcommunities, namely, G1, G2, G6, G12, G14, G16, G20, and G21 were sorted to analyze the composition and abundance of bacteria in each sorted gate. From all the sorted gates, eight bacterial families and classes belonging to the phyla *Proteobacteria*, *Bacteroidales*, and *Bacteroidetes* were identified using terminal restriction fragments (T-RFs). Out of the eight gated clusters, gate 14 (G14) contained the smallest portion of the sorted bacterial community with predominant members belonging to *Proteobacteria* (6%) and showed strong positive correlation (p< 0.01) with the degradation of Basyntan and Retanal (**Figure 8**). The gates 16, 20, and 21 (G16, G20, and G21) which showed positive correlations with retanning chemicals degradation were dominated by members of the phylum *Bacteroidetes* constituting 13,

**Figure 7.** Bacterial community dynamics of tannery wastewater running in sequencing batch reactor (SBR) for 45 days. The first dot plot (initial) refers to the bacterial community in the acclimatized sample used to seed the reactor at the beginning of the SBR. The gate template (top-left box) which is used as the basis for fingerprinting of the different cell

types.

**Figure 8.** Correlation of 13 abiotic parameters with cell abundances in the 30 gates during the running period of the reactor (after Ref. [36]).

**Figure 9.** Taxonomic composition of the sorted gates associated with their role in the degradation of the two retanning agents Basyntan and Retanal (after Ref. [36]).

## **4. Conclusion**

The findings of this study provided a preliminary investigation on the biodegradability of two of the several types of xenobiotic compounds used in the tanning industry. It was possible to single out bacterial groups such as *Bacteroidetes* and *Proteobacteria*, with strong correlation with the complete degradation of some of the compounds in retanning chemicals.

Management of wastewater treatment plants (WWTPs) primarily focuses on process parameters and physicochemical (abiotic) properties of the wastewater before and after treatment. Stable performance of any biological wastewater treatment system can be achieved by understanding and manipulating the microbial communities residing in the system besides the management of the conventional process parameters and abiotic properties. Investigations of microorganisms responsible for efficient reduction of pollutants in various biological wastewater treatment plants have been conducted for many years. This study was successful in identifying bacterial groups involved in different nutrient removal processes from tannery wastewater such as sulfur oxidation, denitrification, and cyclic aromatic compound degradation. Moreover, this study is one of the few studies conducted in field-scale reactors that integrate different approaches to interpret the functional property of a biological treatment system.

## **Acknowledgements**

The authors like to acknowledge the Swedish International Development Cooperation Agency (SIDA), Addis Ababa University, and BIO-EARN (East African regional program and research network working for biotechnology, biosafety, and biotechnology policy development) for the financial support on the establishment of the WWTP. Biosciences eastern and central Africa, International Livestock Research Institute (BecA-ILRI Hub) and the German Academic Exchange Service (DAAD) are highly acknowledged for the financial and technical support of the various analyses of this study.

### **Author details**

Adey Feleke Desta1,\*, Joyce Nzioki<sup>2</sup> , Solomon Maina3 and Francesca Stomeo2

\*Address all correspondence to: adey.desta@gmail.com

1 Institute of Biotechnology, Addis Ababa University, Addis Ababa, Ethiopia

2 Biosciences eastern and central Africa-International Livestock Research Institute (BecA-ILRI) Hub, Nairobi, Kenya

3 School of Plant Biology, University of Western Australia, Perth, Western Australia, Australia

## **References**

**4. Conclusion**

152 Biological Wastewater Treatment and Resource Recovery

ment system.

**Acknowledgements**

**Author details**

support of the various analyses of this study.

\*Address all correspondence to: adey.desta@gmail.com

Adey Feleke Desta1,\*, Joyce Nzioki<sup>2</sup>

ILRI) Hub, Nairobi, Kenya

The findings of this study provided a preliminary investigation on the biodegradability of two of the several types of xenobiotic compounds used in the tanning industry. It was possible to single out bacterial groups such as *Bacteroidetes* and *Proteobacteria*, with strong correlation with the complete degradation of some of the compounds in retanning chemicals.

Management of wastewater treatment plants (WWTPs) primarily focuses on process parameters and physicochemical (abiotic) properties of the wastewater before and after treatment. Stable performance of any biological wastewater treatment system can be achieved by understanding and manipulating the microbial communities residing in the system besides the management of the conventional process parameters and abiotic properties. Investigations of microorganisms responsible for efficient reduction of pollutants in various biological wastewater treatment plants have been conducted for many years. This study was successful in identifying bacterial groups involved in different nutrient removal processes from tannery wastewater such as sulfur oxidation, denitrification, and cyclic aromatic compound degradation. Moreover, this study is one of the few studies conducted in field-scale reactors that integrate different approaches to interpret the functional property of a biological treat-

The authors like to acknowledge the Swedish International Development Cooperation Agency (SIDA), Addis Ababa University, and BIO-EARN (East African regional program and research network working for biotechnology, biosafety, and biotechnology policy development) for the financial support on the establishment of the WWTP. Biosciences eastern and central Africa, International Livestock Research Institute (BecA-ILRI Hub) and the German Academic Exchange Service (DAAD) are highly acknowledged for the financial and technical

, Solomon Maina3

2 Biosciences eastern and central Africa-International Livestock Research Institute (BecA-

3 School of Plant Biology, University of Western Australia, Perth, Western Australia, Australia

1 Institute of Biotechnology, Addis Ababa University, Addis Ababa, Ethiopia

and Francesca Stomeo2


[30] Andrews S. FastQC a quality control tool for high throughput sequence data [internet]. 2010. Available at: http://www.bioinformatics.babraham.ac.uk/projects/fastqc**/** [Accessed 2014-09-21]

[15] Apaydin O, Kurt U and Gonullu MT. An investigation on the tannery wastewater by

[16] Ganesh R, Balaji G and Ramaunjam RA. Biodegradation of tannery wastewater using sequencing batch reactor—respirometric assessment. Bioresour. Technol. 2006; **97**:

[17] Haydar S, Aziz JA and Ahmad MS. Biological treatment of tannery wastewater using

[18] Kongjao S, Damronglerd S and Hunsom M. Simultaneous removal of organic and inorganic pollutants in tannery wastewater using electrocoagulation technique. Korean J.

[19] Koteswari YN and Ramaniba R. The effect of tannery effluent on the colonization rate of

[20] Lefebvre O, Vasudevan N, Torrijos M, Thanasekaran K and Moletta R. Halophilic biological treatment of tannery soaks liquor in a sequencing batch reactor. Water Res. 2005;

[21] Leta S. Developing and optimizing processes for biological nitrogen removal from tannery wastewaters in Ethiopia [thesis]. Stockholm, Sweden: Royal Institute of Technology

[22] Orhon D, Ates E and Sözen S. Experimental evaluation of the nitrification kinetics for

[23] Ram B, Bajpai PK and Parwana HK. Kinetics of chrome-tannery effluent treatment by

[24] Szpyrkowicz L, Kaul SN and Neti RN. Tannery wastewater treatment by electro-oxidation coupled with a biological process. J. Appl. Electrochem. 2005; **35**: 381-390.

[25] Thanigavel M. Biodegradation of tannery effluent in fluidized bed bioreactor with low density biomass support [thesis]. Tamil Nadu, India: Annamalai University; 2004. [26] Reemtsma T, Fiehn O, Kalnowski G and Jekel M. Microbial transformations and biological effects of fungicide-derived benzothiazoles determined in industrial wastewater.

[27] Fiehn O, Wegener G and Jekel M. Capillary electrophoretic analysis of organic anions in tannery wastewater including high contents of chloride and sulfate. Int. J. Environ. Anal.

[28] UNIDO (United Nations Industrial Development Organization). Introduction to treatment of tannery effluents: What every tanner should know about effluent treatment.

[29] APHA (American Public Health Association). Standard Methods for the Examination of Water and Wastewater: 20th ed. Washington, DC, USA: American Public Health Association/American Water Works Association/Water Environment Federation; 1998.

activated sludge process. Pakistan J. Eng. Appl. Sci. 2007; **1**: 61-66.

plankets: a microcosm study. Turk. J. Biol. 2003; **27**: 163-170.

electrocoagulation. Global NEST J. 2009; **11**: 546-555.

1815-1821.

**39**: 1471-1480.

Chem. Eng. 2008; **25**: 703-709.

154 Biological Wastewater Treatment and Resource Recovery

(KTH); 2006. ISBN 91-7283-830-2. 77pp.

Environ. Sci. Technol. 1995; **29**: 478-485.

Chem. 1994; **69**: 257-271.

Vienna: UN Publication; 2011.

tannery wastewaters. Water S. A. 2000; **26**: 43-52.

the activated sludge system. Process Biochem. 1999; **35**: 255-265.


**Biofilm for Wastewater Treatment**

Provisional chapter

## **Application of Mixed Microbial Culture Biofilms for Manganese (II), Cobalt (II), and Chromium (VI) Biosorption by Horizontal Rotating Tubular Bioreactor** Application of Mixed Microbial Culture Biofilms for Manganese (II), Cobalt (II), and Chromium (VI)

Biosorption by Horizontal Rotating Tubular Bioreactor

Tonči Rezić, Iva Rezić, Michaela Zeiner and Božidar Šantek Tonči Rezić, Iva Rezić, Michaela Zeiner and

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

http://dx.doi.org/10.5772/66920

#### Abstract

Industrial wastewater contaminated with toxic heavy metals is a big ecological and environmental problem. Applying biological materials to effectively remove and recover heavy metals from contaminated wastewaters has gained importance as promising alternative to conventional treatment techniques. Thus, the objective of the presented paper is the investigation of the capability of microorganisms, isolated from polluted (metal-laden) soil, to biosorb toxic metals from aqueous solutions. Biosorption process for heavy metal removal was conducted in a new pilot scale horizontal rotating tubular bioreactor (HRTB). This bioreactor provides conditions for microorganism's growth in a form of suspended cells and biofilm. Biofilm is capable to protect microorganisms from interaction with toxic metals in the surrounding environment. Three metals were selected as model examples: cations of manganese and cobalt and hexavalent chromium (an oxyanion). Optimized bioreactor conditions, namely, medium inflow rate (F) and bioreactor rotation speed (n) for biofilm formation and metal removal were monitored, and under optimized bioreactor conditions, promising results were obtained.

Keywords: heavy metals, mixed microbial culture, biosorption, biofilm, horizontal rotating tubular bioreactor

## 1. Introduction

Heavy metal's wastewater pollution has always been a very serious problem because these elements are not biodegradable and can accumulate in living tissues causing serious health effects [1]. Heavy metals are introduced into the natural environment through many industrial

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

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.

processes including leather tanning, wood preservation, metal plating, mining operations, chloralkali, radiator manufacturing, smelting, alloy industries, storage batteries, and automobile manufacturing [2]. Since the early 1970s, there has been growing concern over the effect of heavy metals on humans and aquatic ecosystems [3]. The Agency for Toxic Substances and Disease Registry (ATSDR) classifies nickel as a human carcinogen based on its chronic and subchronic effects [4]. Iron and copper can cause stomach and intestinal distress, liver and kidney damage, and anemia. Zinc may cause anemia, damage the pancreas, and decrease levels of high-density lipoprotein (HDL) cholesterol [5].

There are many conventional methods (physical and chemical) for heavy metal removal, but in general, they result with much waste which is hard to treat. In addition, several widely applied processes such as ion exchange, membrane technologies, and adsorption are very expensive processes when used for large quantities of wastewater which contain low concentrations of heavy metals [6]. Heavy metals may be removed from water as an insoluble soil by the chemical process of precipitation, respectively. However, chemical treatment of heavy metals generates environmentally hazardous chemical byproducts. Additionally, chemical treatment requires constant adjustment of pH value to a narrow range for optimal heavy metal removal, thereby increasing the labor input and cost [7].

As an alternative, different biochemical methods can be applied because they do not destroy metals, but concentrate and immobilize them [8]. Biosorption is removal of metals and their complexes from samples by biological materials [9]. Bioadsorbens can efficiently remove heavy metals from solutions with low concentration; therefore, they are ideal adsorptive media for wastewaters with low concentrations of metal ions. Microbial metal accumulation has received much attention during recent years, due to the potential use of microorganisms for treatment of metal-polluted water or wastewater streams. Recently, several bacterial species have been identified to remove toxic heavy metals [10, 11]. Biosorption can be performed on live or dead microorganisms, as well as on their parts or extracellular products and microorganism aggregations on the surfaces in the structures called biofilm. Biofilm application in the biosorption showed great potential in the wastewater treatment systems. Different types of bioreactor systems such as trickling filters, fluidized or packed bed bioreactors, and thin layer or biodisc reactors were implemented for biofilm formation and wastewater treatment [12–14]. Horizontal rotating tubular bioreactor (HRTB) was designed as combination of a thin layer [15, 16] and biodisc reactor [17] with construction abilities for successful biofilm formation. Consequently, bioreactor interior is equipped with o-shaped partitional walls which provide area for biofilm formation. Wide investigation of HRTB mixing properties was previously done [18–22], and aerobic and anaerobic bioprocesses were successfully conducted. As a model of anaerobic bioprocess, fermentative glucose conversion was chosen [23]. Acetate removal with mixed microbial culture was selected as a model bioprocess for study of HRTB performance in aerobic condition [24]. As combination of aerobic and anaerobic bioprocesses, nitrification and denitrification were done in two consecutive steps in the same bioreactor vessel [25].

In this investigation HRTB was used for native mixed microbial biofilm formation and investigation of developed biofilm biosorption abilities. In the biosorption experiments, artificial wastewater with heavy metal ions Co(II), Cr(VI), and Mn(II) was applied as representative example of textile industry wastewaters. Observed results showed significant potential of developed mixed microbial culture biofilm to successfully remove toxic heavy metals in applied bioreactor.

## 2. Material and methods

processes including leather tanning, wood preservation, metal plating, mining operations, chloralkali, radiator manufacturing, smelting, alloy industries, storage batteries, and automobile manufacturing [2]. Since the early 1970s, there has been growing concern over the effect of heavy metals on humans and aquatic ecosystems [3]. The Agency for Toxic Substances and Disease Registry (ATSDR) classifies nickel as a human carcinogen based on its chronic and subchronic effects [4]. Iron and copper can cause stomach and intestinal distress, liver and kidney damage, and anemia. Zinc may cause anemia, damage the pancreas, and decrease

There are many conventional methods (physical and chemical) for heavy metal removal, but in general, they result with much waste which is hard to treat. In addition, several widely applied processes such as ion exchange, membrane technologies, and adsorption are very expensive processes when used for large quantities of wastewater which contain low concentrations of heavy metals [6]. Heavy metals may be removed from water as an insoluble soil by the chemical process of precipitation, respectively. However, chemical treatment of heavy metals generates environmentally hazardous chemical byproducts. Additionally, chemical treatment requires constant adjustment of pH value to a narrow range for optimal heavy metal removal,

As an alternative, different biochemical methods can be applied because they do not destroy metals, but concentrate and immobilize them [8]. Biosorption is removal of metals and their complexes from samples by biological materials [9]. Bioadsorbens can efficiently remove heavy metals from solutions with low concentration; therefore, they are ideal adsorptive media for wastewaters with low concentrations of metal ions. Microbial metal accumulation has received much attention during recent years, due to the potential use of microorganisms for treatment of metal-polluted water or wastewater streams. Recently, several bacterial species have been identified to remove toxic heavy metals [10, 11]. Biosorption can be performed on live or dead microorganisms, as well as on their parts or extracellular products and microorganism aggregations on the surfaces in the structures called biofilm. Biofilm application in the biosorption showed great potential in the wastewater treatment systems. Different types of bioreactor systems such as trickling filters, fluidized or packed bed bioreactors, and thin layer or biodisc reactors were implemented for biofilm formation and wastewater treatment [12–14]. Horizontal rotating tubular bioreactor (HRTB) was designed as combination of a thin layer [15, 16] and biodisc reactor [17] with construction abilities for successful biofilm formation. Consequently, bioreactor interior is equipped with o-shaped partitional walls which provide area for biofilm formation. Wide investigation of HRTB mixing properties was previously done [18–22], and aerobic and anaerobic bioprocesses were successfully conducted. As a model of anaerobic bioprocess, fermentative glucose conversion was chosen [23]. Acetate removal with mixed microbial culture was selected as a model bioprocess for study of HRTB performance in aerobic condition [24]. As combination of aerobic and anaerobic bioprocesses, nitrification and denitrifi-

cation were done in two consecutive steps in the same bioreactor vessel [25].

In this investigation HRTB was used for native mixed microbial biofilm formation and investigation of developed biofilm biosorption abilities. In the biosorption experiments, artificial wastewater with heavy metal ions Co(II), Cr(VI), and Mn(II) was applied as representative example of

levels of high-density lipoprotein (HDL) cholesterol [5].

160 Biological Wastewater Treatment and Resource Recovery

thereby increasing the labor input and cost [7].

#### 2.1. Microorganism, medium, and growth conditions

Mixed microbial culture was isolated from surface sediments sampled in the Kaštela bay industrial area located near town Split, at the Croatia Adriatic coast. Isolation was done from 5 g of soil samples. Samples were resuspended in Erlenmeyer flasks with different contents of heavy metals in feeding medium (Table 1) and cultivated 48 h at 23 1C. Rotation speed during cultivation was 150 rpm. After 48 h flat plates were inoculated with 1 mL sample from each flask. Medium content used for flat plate cultivation was the same as shown in Table 1 with 20 g/L of agar. Viable cells were determined as colony-forming units (CFU 1/mL). The number of colonies was counted after 48 h at 23 1C. Only medium 1 provided satisfied condition for microorganism colony forming. Therefore, this medium was used for cultivation in tank bioreactor and HRTB. In this research, the medium was sterilized at 121C for 20 min.


Table 1. Contents of feeding medium used during microorganism isolation from soil samples.

#### 2.2. Characteristics and experimental setup of the bioreactor

The HRTB is a stainless steel tube with 2.0 m length and 0.25 m diameter. O-ring–shaped partition walls (inner diameter 0.19 m) divide its interior in a 0.02 m long section. The liquid volume of the bioreactor was 15 L. In order to enable rotation of the entire reactor, the HRTB is horizontally placed on appropriate bearings. The aeration was performed via the central tube fixed in the bioreactor's axis. Improvement of the aeration was obtained by submerging the aeration tube on five positions along the HRTB. For all experimental works, the airflow rate was 152 L h<sup>1</sup> . In Figure 1 the sampling systems for broth and biofilm are shown, being places at 0.40 m intervals. On the cover of the sampling place, a flat plate (0.02 · 0.02 m) is fixed as device for biofilm thickness measurement.

Figure 1. Schematic diagram of HRTB and the inner structure of HRTB with O-ring-shaped partition walls.

Batch cultivation in a stirred tank bioreactor is used to obtain the suspended bacterial biomass (7.5 L) needed for inoculating the HRTB. The feeding process was started after 24 h at a rate of 1 L h<sup>1</sup> and a rotation speed of the HRTB of 10 min<sup>1</sup> . A stable biofilm in the bioreactor is available after 15 days, which is considered as ready to start the experiments with different parameter variations, such as medium inflow rate (0.5, 1.0, and 2.0 L h<sup>1</sup> ) and bioreactor rotation speed (5, 15, and 30 min<sup>1</sup> ). The dynamics of the bioprocess in HRTB was monitored by withdrawing the samples from five positions along the bioreactorlength after five residence times since the new set of process parameters was established. The bioreactorwas operated under a constantinfluent glucose concentration 10 g L<sup>1</sup> and metal ion concentration Co2+ = 0.125 g L<sup>1</sup> , Mn2+ = 0.125 g L<sup>1</sup> , and Cr6+ = 0.125 g L<sup>1</sup> .

Since it was known in previous studies that bioreactor rotation speed higher than 30 min<sup>1</sup> leads to intensive biofilm detachment [23, 25], no higher speed are tested in the current investigation. Experiments with varying bioreactor rotation speed are carried out prior to changes of medium inflow rate, since the latter have exhibited higher effects on the bioprocess dynamics and the biofilm stability [23, 24, 26, 27].

#### 2.3. Analytical methods

0.40 m intervals. On the cover of the sampling place, a flat plate (0.02 · 0.02 m) is fixed as device

Batch cultivation in a stirred tank bioreactor is used to obtain the suspended bacterial biomass (7.5 L) needed for inoculating the HRTB. The feeding process was started after 24 h at a rate

Figure 1. Schematic diagram of HRTB and the inner structure of HRTB with O-ring-shaped partition walls.

available after 15 days, which is considered as ready to start the experiments with different param-

samples from five positions along the bioreactorlength after five residence times since the new set of process parameters was established. The bioreactorwas operated under a constantinfluent glucose

Since it was known in previous studies that bioreactor rotation speed higher than 30 min<sup>1</sup> leads to intensive biofilm detachment [23, 25], no higher speed are tested in the current investigation. Experiments with varying bioreactor rotation speed are carried out prior to changes of medium inflow rate, since the latter have exhibited higher effects on the bioprocess

). The dynamics of the bioprocess in HRTB was monitored by withdrawing the

. A stable biofilm in the bioreactor is

) and bioreactor rotation speed

, Mn2+ = 0.125 g L<sup>1</sup>

, and

of 1 L h<sup>1</sup> and a rotation speed of the HRTB of 10 min<sup>1</sup>

(5, 15, and 30 min<sup>1</sup>

Cr6+ = 0.125 g L<sup>1</sup>

.

dynamics and the biofilm stability [23, 24, 26, 27].

eter variations, such as medium inflow rate (0.5, 1.0, and 2.0 L h<sup>1</sup>

concentration 10 g L<sup>1</sup> and metal ion concentration Co2+ = 0.125 g L<sup>1</sup>

for biofilm thickness measurement.

162 Biological Wastewater Treatment and Resource Recovery

Biomass concentration in suspension was determined by centrifuging the culture medium of 35 mL for 20 min at 4500 rpm (3629 g), washing twice with demineralized water and then drying at 105�C/48 h. Supernatants were used for determination of Co2+, Mn2+, and Cr6+ (UV-Vis spectrophotometrical method by Fries and Getrost) [28]. All determinations were done in triplicates.

Inductively coupled plasma-mass spectrometry (ICP-MS) was used to quantify the metals in biofilm and suspended biomass after acidic digestion. The spectrometer used had a GemCone nebulizer, a cyclone spray chamber, and a standard one-piece extended torch with a quartz injector tube. Each metal was quantified by measurements in triplicates at three different wavelengths. The biofilm samples were mineralized using a closed microwave digestion system. Each sample was digested with a mixture of 5 mL nitric acid, 1 mL hydrogen peroxide, and 1 mL double-distilled water. The digestion was performed in five steps—3 min at 250 W, 1 min without power, 4.5 min at 250 W, 6 min at 650 W, and 5 min at 400 W—followed by a ventilation time of 25 min.

The biofilm thickness was measured applying a modified Venkataraman and Ramanujam method [29]: graphite powder was used instead of chalk powder. The projector was replaced by a microscope with micrometric scale. In order to determine the mass of the biofilm, samples were collected from the inner surface of HRTB, suspended in demineralized water, and twice washed after centrifugation. Finally the biofilm samples were dried for 48 h at 105�C.

Suspended biomass sorption capacity (qx,L) was calculated as follows:

$$q\_{\mathbf{x},L} = \frac{m\_M}{m\_{\mathbf{x},L}}\tag{1}$$

where mM mass of metal ion (mg) and mx.L is the dry weight of suspended biomass (g).

#### 2.4. Mathematical model development

#### 2.4.1. Diffusion process

The diffusivity of metal ion in water was estimated using the Wilke-Chang equation [30]:

$$D\_{aq} = 7.4 \cdot 10^{-8} \frac{\left(\xi\_{aq} M\_{aq}\right)^{1/2} T}{V\_M \prescript{0.6}{}{\eta}\_{aq}} \tag{2}$$

where Daq is the diffusion coefficient of metal ion in water (m2 s �1 ), ξ is the metal ion connecting factor, VM is the metal ion molar volume (m3 mol�<sup>1</sup> ), η is the water dynamic viscosity (kg m�<sup>1</sup> s �1 ), T is the temperature (K), and Maq is the water molecular mass (kg mol�<sup>1</sup> ).

Metal ion relative diffusivity (fD) was computed from Horn-Morgenroth equation [31]:

$$f\_D = 1 - \frac{0.43c\_{\rm x,f}^{0.92}}{11.19 + 0.27c\_{\rm x,f}^{0.99}}\tag{3}$$

where cx,f is the biofilm density (kg m�<sup>3</sup> ).

Effective diffusion coefficient of metal ion in biofilm was calculated using this correlation [32]:

$$D\_{\mathfrak{ef},M} = f\_D D\_{\mathfrak{aq}} \tag{4}$$

where Def,M is the effective diffusion coefficient of metal ion in biofilm (m2 s �1 ).

Mass transport of all dissolved metal ions in biofilm follows Fick's second law of molecular diffusion:

$$D\_{cf,M} \frac{\partial^2 \mathfrak{c}\_{M,f}}{\partial \mathbf{z}^2} = \frac{\partial \mathfrak{c}\_{M,f}}{\partial t} \tag{5}$$

where cM,f is the concentration of metal ion in biofilm phase (kg m�<sup>3</sup> ), t is time (s), and z is biofilm depth (m).

#### 2.4.2. One-dimensional diffusion-bioadsorption model

In the dynamic equilibrium conditions, metal ion concentration in the biofilm is represented conceptually as functions of biofilm depth z as shown in Figure 2C. Concentrations of metal ion (cM,f) in biofilm phase are given by the second-order polynomial correlation:

$$\mathcal{L}\_{M,f}(z) = a\_0 + a\_1 z + a\_2 z^2 \tag{6}$$

where a0, a1, and a<sup>2</sup> are the second-order polynomial correlation coefficient and z is biofilm depth coordinate (m).

Metal ion concentrations in the bulk liquid phase (cM,L) represent as constant values for each ideal mixing segment (Figure 2B, C).

The biofilm zone is surrounded by the stagnant liquid layer of thickness Lg (Figure 2C). The mass transfer coefficient (km) in the stagnant liquid layer was estimated by the correlation [18, 19]:

$$k\_m = 0.664 \left( D\_{tb} / L\_k \right) Re\_N^{1/2} Sc^{1/3} \tag{7}$$

where LK is the wetted perimeter of bioreactor (0.254 m), Dtb = Daq is the diffusion coefficients of metal ions in water (m<sup>2</sup> h�<sup>1</sup> ), Sc is Schmidt number, and ReN is Reynolds rotation number. Schmidt number (Sc) was calculated from [33]

$$\mathcal{S}\mathcal{x} = \nu / D\_{tb} \tag{8}$$

where ν is kinematic viscosity (m<sup>2</sup> s �1 ).

Reynolds rotation number (ReN) of HRTB was calculated by following equation [18]:

Application of Mixed Microbial Culture Biofilms for Manganese (II), Cobalt (II), and Chromium (VI)… http://dx.doi.org/10.5772/66920 165

$$\text{Re}\_{\text{N}} = \frac{D\_{\text{LIP}} \pi \eta L\_{\text{K}} \rho}{2\eta} + \frac{D\_{\text{T}} \Pi \text{nL}\_{\text{K}} \rho}{2\eta} \tag{9}$$

where DUP is the inner diameter of partition wall in HRTB (m), LK is wetted perimeter of bioreactor (0.254 m), Π is the Ludolph's number (3.14159), D<sup>T</sup> is the bioreactor diameter (m), n is the bioreactor rotation speed (s�<sup>1</sup> ), ρ is the broth density (kg m�<sup>3</sup> ), and η is the dynamic viscosity of broth (kg m�<sup>1</sup> s �1 ).

<sup>f</sup> <sup>D</sup> <sup>¼</sup> <sup>1</sup> � <sup>0</sup>:43cx,<sup>f</sup>

).

where Def,M is the effective diffusion coefficient of metal ion in biofilm (m2 s

Def ,<sup>M</sup>

ion (cM,f) in biofilm phase are given by the second-order polynomial correlation:

where cM,f is the concentration of metal ion in biofilm phase (kg m�<sup>3</sup>

2.4.2. One-dimensional diffusion-bioadsorption model

where cx,f is the biofilm density (kg m�<sup>3</sup>

164 Biological Wastewater Treatment and Resource Recovery

diffusion:

biofilm depth (m).

depth coordinate (m).

[18, 19]:

ideal mixing segment (Figure 2B, C).

of metal ions in water (m<sup>2</sup> h�<sup>1</sup>

where ν is kinematic viscosity (m<sup>2</sup> s

Schmidt number (Sc) was calculated from [33]

0:92

<sup>0</sup>:<sup>99</sup> (3)

�1 ).

<sup>∂</sup><sup>t</sup> (5)

Sc<sup>1</sup>=<sup>3</sup> (7)

), t is time (s), and z is

Def ,<sup>M</sup> ¼ f <sup>D</sup>Daq (4)

cM,fð Þ¼ <sup>z</sup> <sup>a</sup><sup>0</sup> <sup>þ</sup> <sup>a</sup>1<sup>z</sup> <sup>þ</sup> <sup>a</sup>2z<sup>2</sup> (6)

11:19 þ 0:27cx,<sup>f</sup>

Effective diffusion coefficient of metal ion in biofilm was calculated using this correlation [32]:

Mass transport of all dissolved metal ions in biofilm follows Fick's second law of molecular

∂<sup>2</sup>cM,<sup>f</sup>

In the dynamic equilibrium conditions, metal ion concentration in the biofilm is represented conceptually as functions of biofilm depth z as shown in Figure 2C. Concentrations of metal

where a0, a1, and a<sup>2</sup> are the second-order polynomial correlation coefficient and z is biofilm

Metal ion concentrations in the bulk liquid phase (cM,L) represent as constant values for each

The biofilm zone is surrounded by the stagnant liquid layer of thickness Lg (Figure 2C). The mass transfer coefficient (km) in the stagnant liquid layer was estimated by the correlation

where LK is the wetted perimeter of bioreactor (0.254 m), Dtb = Daq is the diffusion coefficients

1=2

), Sc is Schmidt number, and ReN is Reynolds rotation number.

Sc ¼ ν=Dtb (8)

km ¼ 0:664ð Þ Dtb=Lk ReN

�1 ).

Reynolds rotation number (ReN) of HRTB was calculated by following equation [18]:

<sup>∂</sup>z<sup>2</sup> <sup>¼</sup> <sup>∂</sup>cM,<sup>f</sup>

Regarding to "spiral flow" mixing model [18, 19], based on the physical model which divided the bioreactor into ideally mixed compartments (Figure 2A), mass balances of the heavy metal ion for the first ideal mixing segment across the bulk liquid (Figure 2A, B) were

$$V\_{\perp}^{1,1}\frac{d\boldsymbol{c}\_{M,L}^{1,1}}{dt} = F\_{\mu}\boldsymbol{c}\_{M,L}^{0} + F\_{\sigma\tau}\boldsymbol{c}\_{M,L}^{1,Ni} + F\_{p}\boldsymbol{c}\_{M,L}^{2,1} - \left(F\_{\mu} + F\_{p}\right)\boldsymbol{c}\_{M,L}^{1,1} - F\_{\sigma\tau}\boldsymbol{c}\_{M,L}^{1,1} - V\_{\perp}^{1,1}\boldsymbol{r}\_{M,L}^{1,1} \tag{10}$$

where c 1,1 <sup>M</sup>,<sup>L</sup> is liquid section metal ion concentrations in the first segment (Ni = 1) of the first kaskade (Nl = 1) (kg m�<sup>3</sup> ), c<sup>0</sup> <sup>M</sup>,<sup>L</sup> is inflow metal ion concentration (kg m�<sup>3</sup> ), Fu is inflow (m3 h�<sup>1</sup> ), Fp is back flow (m<sup>3</sup> h�<sup>1</sup> ), Fcr is circulation flow (m<sup>3</sup> h�<sup>1</sup> ), rM,L 1,1 is liquid section reaction rate in the first segment (Ni = 1) of first kaskade (Nl = 1) (kg m�<sup>3</sup> h�<sup>1</sup> ), VL 1,1 is liquid section volume in the first segment (Ni = 1) of the first kaskade (Nl = 1) (m<sup>3</sup> ), and c 1,Ni <sup>M</sup>,<sup>L</sup> is liquid section metal ion concentrations in the Ni-segment of the first kaskade (kg m�<sup>3</sup> ).

First ideal mixing segments of all cascades were represented in the model without biofilm zone (Figure 2B). All other ideal mixing segments include biofilm zone (Figure 2C). Therefore, mass

Figure 2. Conceptual representation of metal biosorption in HRTB: (A) "spiral flow" mixing, (B) metal ion diffusion, and (C) biosorption reaction.

balances of the heavy metal ion were computed across the bulk liquid for the second ideal mixing segment (Figure 2A, C) as follows:

$$V\_{L}^{1,2}\frac{dc\_{M,L}^{1,2}}{dt} = F\_{cr}c\_{M,L}^{1,1} - F\_{cr}c\_{M,L}^{1,2} - \mathcal{S}^{1,2}k\_{\mathfrak{m}}\left(c\_{M,L}^{1,2} - c\_{M,f(Z=0)}^{1,2}\right) - V\_{L}^{1,2}r\_{M,L}^{1,2} \tag{11}$$

where c 1,2 <sup>M</sup>,<sup>L</sup> is liquid section metal ion concentrations in the second segment (Ni = 2) of the first kaskade (Nl = 1) (kg m�<sup>3</sup> ), c 1, 2 <sup>M</sup>,f Zð Þ <sup>¼</sup><sup>0</sup> is metal ion concentration in the second segment (Ni = 2) of the first kaskade (Nl = 1) on the biofilm surface (kg m�<sup>3</sup> ), VL 1,2 is liquid section volume in the second segment (Ni = 2) of the first kaskade (Nl = 1) (m<sup>3</sup> ), S1,2 is mass transfer surface in second ideal mixing segment (Ni = 2) of the first kaskade (Nl = 1) (m<sup>2</sup> ), and rM,L 1,2 is liquid section reaction rate in the second segment (Ni = 2) of the first kaskade (Nl = 1) (kg m�<sup>3</sup> h�<sup>1</sup> ).

Liquid section volume in the ideal mixing segment (VL Nl,Ni) was computed from the bioreactor liquid volume using the following equation:

$$V\_L^{Nl,Ni} = \frac{V\_L}{Nl \cdot Ni} \tag{12}$$

where VL Nl,Ni is liquid section volume in the ideal mixing segment (m3 ), VL is liquid volume in the HRTB (m3 ), Nl is the number of kaskades, and Ni is the number of ideal mixing segments.

Mass transfer surface in the ideal mixing segment (SNl,Ni) was computed from the inside bioreactor surface using the following equation:

$$\mathbf{S}^{Nl,Ni} = \frac{\mathbf{S}}{\mathbf{N}l \cdot \mathbf{N}i} \tag{13}$$

where SNl,Ni is mass transfer surface in the ideal mixing segment (m<sup>2</sup> ) and S is inside bioreactor surface (m<sup>2</sup> ).

Mass transport of all dissolved metal ions in biofilm is derived from Eq. (5) and equal to reaction rate (r 1, 2 <sup>M</sup>,<sup>f</sup> ):

$$D\_{\mathfrak{e}f,M} \frac{\partial^2 c\_{M,f}^{1,2}}{\partial z^2} = r\_{M,f}^{1,2} \tag{14}$$

The inner boundary conditions (at z = 0) at biofilm-liquid interface are given as

$$\mathbf{S}^{1,2}k\_m\left(\mathbf{c}\_{M,L}^{1,2} - \mathbf{c}\_{M,f(z=0)}^{1,2}\right) = \mathbf{S}^{1,2}D\_{\mathbf{c}f,M}\frac{dc\_{M,f}^{1,2}(z)}{dz}|\_{z=0} \tag{15}$$

The outer boundary conditions (at z=Lf 1,2) at biofilm-bioreactor interface are given as

$$0 = \frac{dc\_{M,f}^{1,2}(z)}{dz}\Big|\_{z=L\_f^{1,2}}\tag{16}$$

As mentioned before, concentrations of the metal ion in the biofilm are represented with

second-order polynomial correlation [Eq. (6)]. Assuming dynamic equilibrium conditions at time (t) model were derived from mass balances equation [Eqs. (11), (14), (15)] and secondorder polynomial correlation for metal ion concentration [Eq. (6)], taken across biofilm zone vertical to the biofilm surface [Eqs. (17)–(20) below]:

Bulk liquid section:

$$0 = F\_{cr}c\_{M,L}^{1,1} - F\_{cr}c\_{M,L}^{1,2} - S^{1,2}k\_m \left(c\_{M,L}^{1,2} - a\_0^{1,2}\right) - V\_{\perp}^{1,2}r\_{M,L}^{1,2} \tag{17}$$

Biofilm zone:

balances of the heavy metal ion were computed across the bulk liquid for the second ideal

km c 1, 2 <sup>M</sup>,<sup>L</sup> � c 1, 2 M,f Zð Þ ¼0 

<sup>M</sup>,<sup>L</sup> is liquid section metal ion concentrations in the second segment (Ni = 2) of the first

), Nl is the number of kaskades, and Ni is the number of ideal mixing segments.

1,2

Def ,<sup>M</sup>

dc<sup>1</sup>,<sup>2</sup> <sup>M</sup>,fð Þz

1,2) at biofilm-bioreactor interface are given as

<sup>¼</sup> <sup>S</sup><sup>1</sup>, <sup>2</sup>

 

Mass transfer surface in the ideal mixing segment (SNl,Ni) was computed from the inside

SNl,Ni <sup>¼</sup> <sup>S</sup>

Mass transport of all dissolved metal ions in biofilm is derived from Eq. (5) and equal to

∂<sup>2</sup>c 1,2 M,f <sup>∂</sup>z<sup>2</sup> <sup>¼</sup> <sup>r</sup>

Def ,<sup>M</sup>

The inner boundary conditions (at z = 0) at biofilm-liquid interface are given as

<sup>0</sup> <sup>¼</sup> dc<sup>1</sup>, <sup>2</sup>

<sup>M</sup>,fð Þz dz <sup>z</sup>¼L<sup>1</sup>,<sup>2</sup> f

As mentioned before, concentrations of the metal ion in the biofilm are represented with

<sup>M</sup>,f Zð Þ <sup>¼</sup><sup>0</sup> is metal ion concentration in the second segment (Ni = 2) of

), VL

� <sup>V</sup><sup>1</sup>, <sup>2</sup> <sup>L</sup> r 1, 2

1,2 is liquid section volume in the

), S1,2 is mass transfer surface in second

Nl,Ni) was computed from the bioreactor

), and rM,L

Nl � Ni (12)

Nl � Ni (13)

<sup>M</sup>,<sup>f</sup> (14)

dz <sup>j</sup><sup>z</sup>¼<sup>0</sup> (15)

(16)

<sup>M</sup>,<sup>L</sup> (11)

1,2 is liquid section

).

), VL is liquid volume in

) and S is inside bioreactor

1, 2 <sup>M</sup>,<sup>L</sup> � <sup>S</sup><sup>1</sup>,<sup>2</sup>

reaction rate in the second segment (Ni = 2) of the first kaskade (Nl = 1) (kg m�<sup>3</sup> h�<sup>1</sup>

VNl,Ni <sup>L</sup> <sup>¼</sup> VL

Nl,Ni is liquid section volume in the ideal mixing segment (m3

where SNl,Ni is mass transfer surface in the ideal mixing segment (m<sup>2</sup>

mixing segment (Figure 2A, C) as follows:

166 Biological Wastewater Treatment and Resource Recovery

dc<sup>1</sup>,<sup>2</sup> M,L dt <sup>¼</sup> Fcrc

1,1 <sup>M</sup>,<sup>L</sup> � Fcrc

), c 1, 2

the first kaskade (Nl = 1) on the biofilm surface (kg m�<sup>3</sup>

second segment (Ni = 2) of the first kaskade (Nl = 1) (m<sup>3</sup>

Liquid section volume in the ideal mixing segment (VL

liquid volume using the following equation:

bioreactor surface using the following equation:

S<sup>1</sup>,<sup>2</sup> km c 1, 2 <sup>M</sup>,<sup>L</sup> � c 1,2 M,f zð Þ ¼0 

The outer boundary conditions (at z=Lf

ideal mixing segment (Ni = 2) of the first kaskade (Nl = 1) (m<sup>2</sup>

V<sup>1</sup>, <sup>2</sup> L

where c

where VL

the HRTB (m3

surface (m<sup>2</sup>

reaction rate (r

).

1, 2 <sup>M</sup>,<sup>f</sup> ):

1,2

kaskade (Nl = 1) (kg m�<sup>3</sup>

$$1D\_{\text{cf},M}2a\_2^{1,2} = r\_{M,f}^{1,2} \tag{18}$$

The inner boundary conditions (at z = 0):

$$-\frac{k\_m}{D\_{\rm ef,M}}\left(c\_{M,L}^{1,2} - a\_0^{1,2}\right) = a\_1^{1,2}\tag{19}$$

The outer boundary conditions (at z=Lf 1,2):

$$a\_1^{1,2} = -2a\_2^{1,2} L\_f^{1,2} \tag{20}$$

where a 1, 2 <sup>0</sup> , a 1,2 <sup>1</sup> , and a 1,2 <sup>2</sup> are the second-order polynomial correlation coefficient in the second segment (Ni = 2) of the first kaskade (Nl = 1); Lf 1,2 biofilm thickness in the second segment (Ni = 2) of the first kaskade (Nl = 1) (m); VL 1,2 is liquid section volume in the second segment (Ni = 2) of the first kaskade (Nl = 1) (m<sup>3</sup> ); and rM,L 1,2 is liquid section reaction rate in the second segment (Ni = 2) of the first kaskade (Nl = 1) (kg m�<sup>3</sup> h�<sup>1</sup> ).

Adjusting mass balances and reaction rates for all ideal mixing segments according to Figure 2, system of the differential equations was developed for heavy metal ion concentration changes along HRTB.

#### 2.4.3. Bioadsorption kinetic model

Mass balance equations were coupled to the reaction rate terms in the liquid section (rM,L) and in the biofilm zone (rM,f) based on the Freundlich adsorption isotherm. Instead kinetic terms heavy metal removal was changed with bioadsorption model [34] [Eqs. (21), (22)]:

$$\mathfrak{q}\_{\mathbf{x},L} = K\_F (\mathfrak{c}\_{\mathbf{M},L})^{1/h} \tag{21}$$

$$\mathfrak{q}\_{\mathbf{x},f} = \mathcal{K}\_F \mathfrak{k}\_{\mathbf{M},f} \mathfrak{l}^{1/\hbar} \tag{22}$$

were qx,L is suspended biomass adsorption capacity (mg g�<sup>1</sup> ), qx,f is biofilm adsorption capacity (mg g�<sup>1</sup> ), and KF and h are Freundlich isotherm constant.

Bioadsorption model for biofilm zone was derived from Freundlich equation [Eq. (22)] and second-order polynomial correlation for metal ion biofilm concentration [Eq. (6)]:

$$q\_{x,f} = K\_F \left[ a\_0 + a\_1 z\_i + a\_2 (z\_i)^2 \right]^{1/h} \tag{23}$$

where zi is collocation point across biofilm zone parallel to the substratum surface.

Assuming the collocation point zi = Lf/b where b ∈ N(+) bioadsorption model are defined as follows:

$$q\_{x,f} = K\_F \left[ a\_0 + a\_1 \frac{L\_f}{b} + a\_2 \left(\frac{L\_f}{b}\right)^2 \right]^{1/h} \tag{24}$$

The kinetic model assumes that reaction rate is the function of biomass concentration in the liquid section (cx,L) and in the biofilm zone (cx,f) [Eqs. (25), (26) below]:

$$r\_{M,L} = \frac{c\_{x,L}q\_{x,L}}{\pi} \tag{25}$$

$$r\_{M,f} = \frac{c\_{x,f}q\_{x,f}}{\tau} \tag{26}$$

where rM,f is biofilm section reaction rate (kg m�<sup>3</sup> h�<sup>1</sup> ), rM,L is liquid section reaction rate (kg m�<sup>3</sup> h�<sup>1</sup> ), and τ is retention time (h).

#### 2.4.4. Numerical methods

The model equations were solved by personal computer using the "Wolfram Mathematica" program routine "NDSolve, FindRoot, FindMinimum, Fit," and orthogonal collocation methods [35–37] were applied for the inner biofilm concentration profiles representing.

#### 2.4.5. Initial parameter values

The model was initially simulated using kinetic parameters (KF and h) from previous studies [38] and mixing parameters (Nl, Ni, Fcr, and Fp) computed in this study (Table 1). Transport parameters include the mass transfer coefficient rate of metal ions (km), and the effective diffusion coefficient of metal ion in biofilm (Def,M) was estimated by Eqs. (7) and (2)–(4).

#### 2.4.6. Parameter optimization

The empirical equations developed from HRTB mixing modeling were used as a fitness function during mixing parameter optimization (Nl, Ni, Fcr, and Fp). Kinetic parameters (KF and h) were optimized computing variance between observed variables and simulated variables as

$$E\_n = \frac{1}{n\_u} \sum\_{i=1}^{n=n\_u} \frac{\left(c\_{n,\text{exp}}^i - c\_{n,\text{sim}}^i\right)^2}{c\_{n,\text{exp}}^i} \tag{27}$$

where ci <sup>n</sup>, exp is observed variables (kg m�<sup>3</sup> ), ci <sup>n</sup>,sim is simulated variables (kg m�<sup>3</sup> ), and nu is number of observations.

To determine dependence of parameter change on variance between observed variables and simulated variables (En), calculation were performed by polynomial regression with the "Wolfram Mathematica" routine "Fit." After this plug, optimization was preformed calculating global minimum variance between observed variables and simulated variables using routine "FindMinimum."

## 3. Results and discussion

qx,<sup>f</sup> ¼ KF a<sup>0</sup> þ a1zi þ a2ð Þ zi

Assuming the collocation point zi = Lf/b where b ∈ N(+) bioadsorption model are defined as

Lf <sup>b</sup> <sup>þ</sup> <sup>a</sup><sup>2</sup>

The kinetic model assumes that reaction rate is the function of biomass concentration in the

rM,<sup>L</sup> <sup>¼</sup> cx,Lqx,<sup>L</sup>

rM,<sup>f</sup> <sup>¼</sup> cx,<sup>f</sup> qx,<sup>f</sup>

The model equations were solved by personal computer using the "Wolfram Mathematica" program routine "NDSolve, FindRoot, FindMinimum, Fit," and orthogonal collocation methods

The model was initially simulated using kinetic parameters (KF and h) from previous studies [38] and mixing parameters (Nl, Ni, Fcr, and Fp) computed in this study (Table 1). Transport parameters include the mass transfer coefficient rate of metal ions (km), and the effective diffusion

The empirical equations developed from HRTB mixing modeling were used as a fitness function during mixing parameter optimization (Nl, Ni, Fcr, and Fp). Kinetic parameters (KF and h) were

> ci <sup>n</sup>, exp � ci

> > ci n, exp

n,sim � �<sup>2</sup>

<sup>n</sup>,sim is simulated variables (kg m�<sup>3</sup>

optimized computing variance between observed variables and simulated variables as

X<sup>i</sup>¼nu i¼1

), ci

[35–37] were applied for the inner biofilm concentration profiles representing.

coefficient of metal ion in biofilm (Def,M) was estimated by Eqs. (7) and (2)–(4).

En <sup>¼</sup> <sup>1</sup> nu

<sup>n</sup>, exp is observed variables (kg m�<sup>3</sup>

� �<sup>2</sup> " #<sup>1</sup>=<sup>h</sup>

Lf b

<sup>τ</sup> (25)

<sup>τ</sup> (26)

), rM,L is liquid section reaction rate

where zi is collocation point across biofilm zone parallel to the substratum surface.

qx,<sup>f</sup> ¼ KF a<sup>0</sup> þ a<sup>1</sup>

liquid section (cx,L) and in the biofilm zone (cx,f) [Eqs. (25), (26) below]:

where rM,f is biofilm section reaction rate (kg m�<sup>3</sup> h�<sup>1</sup>

), and τ is retention time (h).

168 Biological Wastewater Treatment and Resource Recovery

follows:

(kg m�<sup>3</sup> h�<sup>1</sup>

2.4.4. Numerical methods

2.4.5. Initial parameter values

2.4.6. Parameter optimization

number of observations.

where ci

<sup>2</sup> h i<sup>1</sup>=<sup>h</sup>

(23)

(24)

(27)

), and nu is

#### 3.1. Biofilm formation studies in HRTB

In this work the effect of process parameters (n and F) on the mixed microbial culture biofilm formation in HRTB was studied as a continuation of comprehensive research of mixing [18–20] and conduction of model bioprocesses in HRTB [23–27]. This investigation started with mixed microbial culture isolation from surface sediments highly contaminated with heavy metals [39–41].

Isolated mixed microbial culture was developed in HRTB as described in Section 2.2, whereby the culture first grew in suspension and then a biofilm was gradually established on the Oshaped rings and inner surface of bioreactor. Figure 3 represents O-shaped rings before and after biofilm formation.

Figure 3. O-shaped rings before (A) and after (B) biofilm formation in HRTB.

The biofilm obtained was used for the investigation of suspended biomass adsorption abilities and biofilm properties (thickness, density) by different combinations of process parameters. Changes of process parameters (n and F) during this investigation are presented in Figure 4.

A significant disturbance was observed at F = 2.0 L h<sup>1</sup> and n = 30 min<sup>1</sup> when biofilm detachment took place. Influence of biofilm detachment on suspended biomass concentration changes will be discussed in the next section.

#### 3.2. Suspended biomass concentration and biosorption capacity in HRTB

In the present study, biomass grew as suspended single cells, suspended cell clusters, and biofilm attached to the bioreactor inner surface. Table 2 shows the results of suspended biomass concentration in dependency of parameter variation: inflow rate (F = 0.5–2.0 L h<sup>1</sup> ) and bioreactor rotation speed (n = 5–30 min<sup>1</sup> ). The suspended biomass concentrations (cx.L) range from 0.95 to 1.07 g L<sup>1</sup> at inflow rate 0.5 L h<sup>1</sup> . The increase of the inflow rate to 1.0 and

Figure 4. Dynamics of process parameter changes [bioreactor rotation speed (n) and medium inflow rate (F)] during investigation in the HRTB.

2.0 L h<sup>1</sup> was related to the increase of suspended biomass concentrations (1.59–5.11 g L<sup>1</sup> ) as a consequence of biofilm detachment and erosion. Highest suspended biomass concentration was 5.11 g L<sup>1</sup> registrated as a consequence of more intensive biofilm detachment (release of larger biofilm parts) due to high inflow rate (F = 2.0 L h<sup>1</sup> ) and bioreactor rotation speed (n = 30 min<sup>1</sup> ). In this situation, considerable increase of metal ion concentrations was observed as a consequence of biomass washout from HRTB. Biofilm detachment (erosion and sloughing) is a complex process affected by hydrodynamic conditions together with morphological and physiological characteristics of the biofilm [8, 42]. Suspended biomass changes were also observed at inflow rates (1.0–2.0 L h<sup>1</sup> ) for all bioreactor rotation speed (5–30 min<sup>1</sup> ) as a consequence of biofilm erosion (continuous release of smaller biofilm parts) [43].

The suspended biomass biosorption capacity (qx.L) during heavy metal removal is presented in Table 3. The inflow rate had a more pronounced effect on the biosorption capacity than the bioreactor rotation speed. Nevertheless, highest bioreactor rotation speed (30 min<sup>1</sup> ) decreased thickness of stagnant liquid layer at the biomass surface and provided facilitate condition for metal ion adsorption. The increase of the inflow rate to 1.0 and 2.0 L h<sup>1</sup> was related to the increase of biomass biosorption capacity. Microbial biomass concentration and content have a significant effect on the biosorption capacity. Therefore, higher biomass biosorption capacity was observed for inflow rates 1.0 and 2.0 L h<sup>1</sup> where higher microbial biomass concentration and biofilm erosion were observed (Table 2). Biofilm structure and extracellular polysaccharide content increase possibility for metal ion accumulation. Molecule of extracellular polysaccharide has high molecular mass and enhanced capability for metal ion bonding [13, 42, 44–47]. Due to the biofilm detachment observed for F = 2.0 L h<sup>1</sup> and n = 30 min<sup>1</sup> and release of microbial biomass with high amount of biofilm, biosorption capacity reached highest value of 83.27 mg g<sup>1</sup> , respectively.

Biological and hydrodynamic factors (content of extracellular polymers and cell physiological and morphological state of same microbial species) have influence on the suspended biomass


Table 2. The suspended biomass concentration (cx.L) changes at different combinations of bioreactor process parameters (n and F) during heavy metal removal process.


Table 3. Metal ion sorption capacity (qx.L) changes at different combinations of bioreactor process parameters (n and F) during heavy metal removal process.

biosorption capacity. Situation is more complex in mixed culture where different microbiological content and cell distribution also influence biosorption capacity. In addition, hydrodynamic conditions have also influence on all previous denominate biological factors [8]. Therefore, on the basis of these results, it is clear that biological hydrodynamic conditions in HRTB have a significant effect on the suspended biomass concentration and biosorption capacity (Tables 2 and 3).

#### 3.3. Biofilm volumetric density and thickness along HRTB

2.0 L h<sup>1</sup> was related to the increase of suspended biomass concentrations (1.59–5.11 g L<sup>1</sup>

larger biofilm parts) due to high inflow rate (F = 2.0 L h<sup>1</sup>

were also observed at inflow rates (1.0–2.0 L h<sup>1</sup>

(n = 30 min<sup>1</sup>

investigation in the HRTB.

170 Biological Wastewater Treatment and Resource Recovery

83.27 mg g<sup>1</sup>

, respectively.

a consequence of biofilm detachment and erosion. Highest suspended biomass concentration was 5.11 g L<sup>1</sup> registrated as a consequence of more intensive biofilm detachment (release of

Figure 4. Dynamics of process parameter changes [bioreactor rotation speed (n) and medium inflow rate (F)] during

observed as a consequence of biomass washout from HRTB. Biofilm detachment (erosion and sloughing) is a complex process affected by hydrodynamic conditions together with morphological and physiological characteristics of the biofilm [8, 42]. Suspended biomass changes

The suspended biomass biosorption capacity (qx.L) during heavy metal removal is presented in Table 3. The inflow rate had a more pronounced effect on the biosorption capacity than the

thickness of stagnant liquid layer at the biomass surface and provided facilitate condition for metal ion adsorption. The increase of the inflow rate to 1.0 and 2.0 L h<sup>1</sup> was related to the increase of biomass biosorption capacity. Microbial biomass concentration and content have a significant effect on the biosorption capacity. Therefore, higher biomass biosorption capacity was observed for inflow rates 1.0 and 2.0 L h<sup>1</sup> where higher microbial biomass concentration and biofilm erosion were observed (Table 2). Biofilm structure and extracellular polysaccharide content increase possibility for metal ion accumulation. Molecule of extracellular polysaccharide has high molecular mass and enhanced capability for metal ion bonding [13, 42, 44–47]. Due to the biofilm detachment observed for F = 2.0 L h<sup>1</sup> and n = 30 min<sup>1</sup> and release of microbial biomass with high amount of biofilm, biosorption capacity reached highest value of

Biological and hydrodynamic factors (content of extracellular polymers and cell physiological and morphological state of same microbial species) have influence on the suspended biomass

as a consequence of biofilm erosion (continuous release of smaller biofilm parts) [43].

bioreactor rotation speed. Nevertheless, highest bioreactor rotation speed (30 min<sup>1</sup>

). In this situation, considerable increase of metal ion concentrations was

) as

)

) decreased

) and bioreactor rotation speed

) for all bioreactor rotation speed (5–30 min<sup>1</sup>

Since the sampling point at 75% of reactor length was also used for introducing the temperature sensor, the biofilm thickness could be measured only at four sampling sites. The differences in biofilm thickness given by changing medium inflow rate (F = 0.5–2.0 L h<sup>1</sup> ) and bioreactor rotation speed (n = 5–30 min<sup>1</sup> ) are presented in Table 4. The biofilm thickness was in the range of 0.23–1.43 mm that is thinner than the literature data for mixed culture biofilm but thicker than monomicrobial culture biofilm thickness measured in previous research [25].

The biofilm thickness in the bioreactor Lf was mainly stabile for inflow rates 0.5 and 1 L h<sup>1</sup> , and only smaller biofilm parts were observed in the liquid phase as a consequence of the biofilm erosion process. This tendency was maintained until the inflow rate became 2 L h<sup>1</sup> . Afterward, hydrodynamic conditions and high metal load inhibited biofilm growth and decreased biofilm thickness. The resultant accumulation of metal ions had an impact on the biofilm, its strength, and its density. In these conditions intensive detachment of the biofilm was observed. The increase of the inflow rate produces thinner biofilm with higher density. Therefore, the outer biofilm layers are more sensitive to the shear stress and abrasion than the


Table 4. Biofilm thickness changes (Lf) along HRTB at different medium inflow rates (F = 0.5–2.0 L h<sup>1</sup> ) and bioreactor rotation speed (n = 5–30 min<sup>1</sup> ) during heavy metal removal bioprocess.

inner biofilm layers. Moreover, outer biofilm layers can be released even at relatively small shear stress. After this, the detachment rate is considerably reduced [12, 47]. Thinner biofilms are less sensitive to process condition changes, which has positive influence on the process stability [44]. The impact of the biofilm detachment on the bioprocess was less pronounced from bioreactor inflow rate (Table 4).

Biofilm volumetric density (cx.f) for F = 2.0 L h<sup>1</sup> and n = 30 min<sup>1</sup> was measured at the inlet and the outlet of the HRTB. The HRTB is characterized by concentration gradient along bioreactor, so consequently higher volumetric biofilm density was observed at the inlet of HRTB (59.7 5.2 g L<sup>1</sup> ) than at the outlet of HRTB (39.3 4.4 g L<sup>1</sup> ). Similar results were observed during previous investigation of metal ion removal in HRTB [38].

The reason for this finding might be that the substrate concentrations for microorganism growth decrease with bioreactor length. Higher volumetric biofilm density was related to increase the biofilm sorption capacity. Both properties are influenced by structure and content of biofilm. Differences in extracellular polysaccharide content affect the gradient of the linkage strength between cell clusters inside the biofilm. While cells on the surface of the biofilm grow relatively fast and do not accumulate, cells inside the biofilm have lower growth rates and produce more extracellular polysaccharides [13, 42, 44–46]. The extracellular polysaccharides affect the microbial sorption capacity by their content and molecular size. The outer biofilm layer exhibits higher porosity, resulting in easier metal ion access to deeper layers. Additionally, high-volumetric-density biofilms have higher sorption capacity than the low-density biofilms that are characterized by the low content of extracellular polysaccharides [48].

#### 3.4. Biofilm application in removal of Co(II), Cr(IV), and Mn(II) from wastewater

After biofilm formation and characterization, investigation of biofilm sorption abilities in removal of Co(II), Cr(IV), and Mn(II) was done at different combinations of medium inflow rates and constant HRTB rotation speed. Results are presented as equilibrium metal ion concentration along HRTB in the liquid phase. Equilibrium metal ion concentration was reached after five residence time changes.

The metal ion concentrations along HRTB at different medium inflow rates (F = 0.5–2.0 L h<sup>1</sup> ) and constant bioreactor rotation speed (n = 15 min<sup>1</sup> ) are presented in Figure 5 (Co(II) concentration Figure 5A, Cr(VI) concentration Figure 5B, Mn(II) concentration Figure 5C). Points represent measured values, while simulated values are represented with curves. Metal ion concentration changes along HRTB were simulated using one-dimensional diffusionbiosorption model and optimized parameter values [38]. The inflow of all metal ion (Co(II), Cr(VI), and Mn(II)) concentration was 0.125 g L<sup>1</sup> , respectively. Lower metal ion concentrations were detected at a first measuring point in the bioreactor (located at the place of medium inflow, 0% LHRTB) because of medium dilution at this location in the HRTB.

Generally, increase in the inflow rate (F) caused increase of metal ion concentration along bioreactor. Higher inflow rate increased metal ion load in HRTB and concentration of metal ions in liquid phase. Metal ion concentration in biomass was in a dynamic equilibrium with metal ion concentration in the liquid phase. Biomass (solid phase) in bioreactor becomes saturated with metal ions and reaches maximum removal capacity. Consequence of biomass saturation is the decrease of metal ion concentration in the liquid phase (Figure 5).

inner biofilm layers. Moreover, outer biofilm layers can be released even at relatively small shear stress. After this, the detachment rate is considerably reduced [12, 47]. Thinner biofilms are less sensitive to process condition changes, which has positive influence on the process stability [44]. The impact of the biofilm detachment on the bioprocess was less pronounced

(0% LHRTB) (50% LHRTB) (100% LHRTB)

Biofilm volumetric density (cx.f) for F = 2.0 L h<sup>1</sup> and n = 30 min<sup>1</sup> was measured at the inlet and the outlet of the HRTB. The HRTB is characterized by concentration gradient along bioreactor, so consequently higher volumetric biofilm density was observed at the inlet of

) than at the outlet of HRTB (39.3 4.4 g L<sup>1</sup>

The reason for this finding might be that the substrate concentrations for microorganism growth decrease with bioreactor length. Higher volumetric biofilm density was related to increase the biofilm sorption capacity. Both properties are influenced by structure and content of biofilm. Differences in extracellular polysaccharide content affect the gradient of the linkage strength between cell clusters inside the biofilm. While cells on the surface of the biofilm grow relatively fast and do not accumulate, cells inside the biofilm have lower growth rates and produce more extracellular polysaccharides [13, 42, 44–46]. The extracellular polysaccharides affect the microbial sorption capacity by their content and molecular size. The outer biofilm layer exhibits higher porosity, resulting in easier metal ion access to deeper layers. Additionally, high-volumetric-density biofilms have higher sorption capacity than the low-density

biofilms that are characterized by the low content of extracellular polysaccharides [48].

3.4. Biofilm application in removal of Co(II), Cr(IV), and Mn(II) from wastewater

After biofilm formation and characterization, investigation of biofilm sorption abilities in removal of Co(II), Cr(IV), and Mn(II) was done at different combinations of medium inflow rates and constant HRTB rotation speed. Results are presented as equilibrium metal ion

observed during previous investigation of metal ion removal in HRTB [38].

) Lf (mm)

0.5 5 0.75 1.08 0.89

1.0 5 0.85 1.34 0.95

2.0 5 0.23 0.37 0.28

Table 4. Biofilm thickness changes (Lf) along HRTB at different medium inflow rates (F = 0.5–2.0 L h<sup>1</sup>

) during heavy metal removal bioprocess.

15 0.89 0.73 0.81 30 0.93 1.29 0.85

15 0.92 1.43 0.84 30 0.86 1.21 0.91

15 0.35 0.28 0.25 30 0.38 0.37 0.35

). Similar results were

) and bioreactor

from bioreactor inflow rate (Table 4).

HRTB (59.7 5.2 g L<sup>1</sup>

rotation speed (n = 5–30 min<sup>1</sup>

F (L h<sup>1</sup>

) n (min<sup>1</sup>

172 Biological Wastewater Treatment and Resource Recovery

As shown in previously performed hydrodynamic experiments in HRTB, medium flow in the bioreactor can be determined by plug-flow conditions [21]. These are attributed to the formation of temperature and/or concentration gradients along the reactor length [16]. Decrease in the metal ion concentration gradient along the bioreactor length in the second part of the HRTB (measurements points on 50% and 100% LHRTB) confirmed assumption of the plug-flow condition in HRTB (Figure 5). The highest metal ion concentration measured near the place of medium inflow (measurement points 0% and 25% LHRTB) inhibited biomass

Figure 5. Concentration of Co (A), Cr (B), and Mn (C) ion along the HRTB at different medium inflow rates F = 0.5 L h<sup>1</sup> (black dots, solid line), F = 1.0 L h<sup>1</sup> (dark gray dots, dashed line), F = 2.0 L h<sup>1</sup> (light gray dots, dot line), and constant bioreactor rotation speed (n = 15 min<sup>1</sup> ).

activity and produced a considerable deviations from plug-flow conditions. As was previously mentioned (in the Section 3.2), the biofilm biosorption is a complex process that is affected by hydrodynamic conditions as well as morphological and physiological characteristics of the biofilm [49, 50].

## 4. Conclusion

Microbial strains were isolated from heavy metal-contaminated surface sediments and selected due to their ability to grow in the presence of metal ions. The results obtained in this study proved technical feasibility of isolated strains to form biofilm in HRTB and to remove metal ions from contaminated water with concentrations up to 500 mg L<sup>1</sup> . The microbial removal ability was higher at lowest medium inflow rates of 0.5. When the inflow rate was in the range of 1.0–2.0 L h<sup>1</sup> , microbial removal ability was reduced.

Generally, the medium inflow rate had more pronounced effect on the bioprocess dynamics than bioreactor rotation speed. The biofilm biosorption capacity was reduced with decreased biofilm density. Similar trend shows suspended biomass biosorption capacity and suspended biomass concentration. The obtained results prove that HRTB can be successfully used for conducting the removal of heavy metals with isolated microbial strains.

## Acknowledgements

This chapter is supported by Croatian Science Foundation project UIP-2014-09-1534 and SPECH-LRM-9158 for which the authors are grateful. Any findings, conclusions, and remarks are of authors only and do not necessarily reflect the standings of Croatian Science Foundation.

## Author details

Tonči Rezić 1 \*, Iva Rezić 2 , Michaela Zeiner<sup>3</sup> and Božidar Šantek1

\*Address all correspondence to: trezic@pbf.hr

1 Department of Biochemical Engineering, Faculty of Food Technology and Biotechnology, University of Zagreb, Zagreb, Croatia

2 Department of Applied Chemistry, Faculty of Textile Technology, University of Zagreb, Zagreb, Croatia

3 Department of Chemistry, University of Natural Resources and Life Sciences, BOKU – Vienna, Wien, Austria

## References

activity and produced a considerable deviations from plug-flow conditions. As was previously mentioned (in the Section 3.2), the biofilm biosorption is a complex process that is affected by hydrodynamic conditions as well as morphological and physiological character-

Microbial strains were isolated from heavy metal-contaminated surface sediments and selected due to their ability to grow in the presence of metal ions. The results obtained in this study proved technical feasibility of isolated strains to form biofilm in HRTB and to remove

removal ability was higher at lowest medium inflow rates of 0.5. When the inflow rate was in

Generally, the medium inflow rate had more pronounced effect on the bioprocess dynamics than bioreactor rotation speed. The biofilm biosorption capacity was reduced with decreased biofilm density. Similar trend shows suspended biomass biosorption capacity and suspended biomass concentration. The obtained results prove that HRTB can be successfully used for

This chapter is supported by Croatian Science Foundation project UIP-2014-09-1534 and SPECH-LRM-9158 for which the authors are grateful. Any findings, conclusions, and remarks are of authors only and do not necessarily reflect the standings of Croatian Science Foundation.

, Michaela Zeiner<sup>3</sup> and Božidar Šantek1

1 Department of Biochemical Engineering, Faculty of Food Technology and Biotechnology,

2 Department of Applied Chemistry, Faculty of Textile Technology, University of Zagreb,

3 Department of Chemistry, University of Natural Resources and Life Sciences, BOKU –

, microbial removal ability was reduced.

. The microbial

metal ions from contaminated water with concentrations up to 500 mg L<sup>1</sup>

conducting the removal of heavy metals with isolated microbial strains.

istics of the biofilm [49, 50].

174 Biological Wastewater Treatment and Resource Recovery

the range of 1.0–2.0 L h<sup>1</sup>

Acknowledgements

Author details

1

\*, Iva Rezić

University of Zagreb, Zagreb, Croatia

2

\*Address all correspondence to: trezic@pbf.hr

Tonči Rezić

Zagreb, Croatia

Vienna, Wien, Austria

4. Conclusion


[33] Garrido JM, van Benthem W, van Loosdrecht MCM, Heijnen JJ. Influence of dissolved oxygen concentration on nitrite accumulation in a biofilm airlift suspension reactor. Biotechnol Bioeng. 1997;53:168–178.

[18] Šantek B, Horvat P, Novak S, Mayr B, Moser A, Marić V. Mathematical modeling of mixing in horizontal rotating tubular bioreactor: "simple flow" model. Bioproc Eng.

[19] Šantek B, Horvat P, Novak S, Mayr B, Moser A, Marić V. Mathematical modeling of mixing in horizontal rotating tubular bioreactor: "simple flow" model. Bioproc Eng.

[20] Šantek B, Horvat P, Novak S, Moser A, Marić V. Studies on mixing in a horizontal rotating tubular bioreactor. Part I: optimisation of adjustable parameters for "spiral flow" model.

[21] Šantek B, Horvat P, Novak S, Moser A, Marić V. Studies on mixing in a horizontal rotating tubular bioreactor. Part II: prediction systems for adjustable parameters of "spiral flow"

[22] Šantek B, Horvat P, Novak S, Moser A, Marić V. Studies on mixing in a horizontal rotating tubular bioreactor. Part III: influence of liquid level and distance between the partition walls on prediction systems for adjustable model parameters. Bioproc Eng. 1998;19:91–

[23] Ivančić M, Šantek B, Novak S, Marić V. Fermentative bioconversion in a horizontal

[24] Slavica A, Šantek B, Novak S, Marić V. Microbial acetate oxidation in horizontal rotating

[25] Rezić T, Šantek B, Novak S, Marić V. Heterotrophic cultivation of Paracoccus denitrificans in a horizontal rotating tubular bioreactor. World J Microbiol Biotechnol. 2007;23:987–996.

[26] Rezić T, Šantek B, Novak S, Marić V. Comparison between the heterotrophic cultivation of Paracoccus denitrificans in continuous stirred tank reactor and horizontal rotating tubu-

[27] Ivančić M, Šantek B, Novak S, Horvat P, Marić V. Bioprocess kinetics in a horizontal

[28] Fries J, Getrost H. Organic reagents for trace analysis. Merck Laboratory Press, Darm-

[29] Venkataraman R, Ramanujam TK. A study on microbiology of biological film layer in

[30] Wilke CR, Chang P. Correlation of diffusion coefficients in dilute solutions. AIChE J.

[31] Horn H, Morgenroth E. Transport of oxygen, sodium chloride and sodium nitrate in

rotating tubular bioreactor. Bioprocess Biosyst Eng. 2004;26:169–175.

rotating biological contactors. Bioproc Eng. 1998;18:181–186.

[32] Stewart PS. Diffusion in biofilms. J Bacteriol. 2003;185:1485–1491.

rotating tubular bioreactor. Proc Biochem. 2004;39:995–1000.

tubular bioreactor. J Biosci. 2004;29:169–177.

lar bioreactor. Proc Biochem. 2006;41:2024–2028.

biofilms. Chem Eng Sci. 2006;61:1347–1356.

1996;14:195–204.

1996;14:223–229.

102.

stadt; 1977.

1955;1:264–270.

Bioproc Eng. 1998;18:467–473.

176 Biological Wastewater Treatment and Resource Recovery

model. Bioproc Eng. 1998;19:19–28.


## **Electrocoagulative and Biological Treatment of** Provisional chapter

## **Laundry Wastewater** Electrocoagulative and Biological Treatment of

Laundry Wastewater

[47] Characklis WG. Biofilm processes. In: Characklis WG, Marshall KC (eds.) Biofilms, Wiley,

[48] Cloirec P, Andres Y, Faur-Brasquet C, Gerente C. Engineered biofilms for metal ion

[49] Quintelas C, Fonseca B, Silva B, Figueiredo H, Tavares T. Treatment of chromium(VI) solutions in a pilot-scale bioreactor through a biofilm of Arthrobacter viscosus supported

[50] Van Hullebusch ED, Zandvoort MH, Lens PNL. Metal immobilization by biofilms: mech-

anisms and analytical tools. Rev Environ Sci Bio Technol. 2003;2:9–33.

New York; 1990. pp. 93–130.

178 Biological Wastewater Treatment and Resource Recovery

removal. Rev Environ Sci Biotechnol. 2003;2:177–192.

on GAC. Bioresour Technol. 2009;100:220–226.

Terelle Ramcharan and Ajay Bissessur

Additional information is available at the end of the chapter Terelle Ramcharan and Ajay Bissessur

http://dx.doi.org/10.5772/67525 Additional information is available at the end of the chapter

#### Abstract

The greater demand for potable water, both locally and worldwide, has directed a huge interest amongst researchers to investigate the possibility of recycling and reusing wastewater from laundry run-offs. The advantage of using recycling wastewater from such sources is mainly due to the fact that these bulk volumes of wastewater are considered to be less chemically polluted in comparison to those discarded from industrial effluents and wastewater sources. Almost all laundry detergents contain surfactants, whose main function serves to remove dirt/soil from contaminated items. Thus, an analysis of the surfactant levels before and after a treatment process is important to confirm that the surfactant has in fact carried out its intended purpose. Electrocoagulative treatment of wastewater, a well-researched and well-documented clean-up process that involves the production of aluminium hydroxy species by oxidation of aluminium metal upon the application of a controlled voltage which adsorbs fine particulate matter and pollutants from the wastewater has been investigated as a clean-up application to the treatment of laundry wastewater. The use of a biological treatment process which entails treating the wastewater with aerobic bacterial specie specifically designed to degrade fats, lipids, protein, detergents and hydrocarbons has also been investigated.

Keywords: biological, biospinners, electrocoagulation, laundry wastewater, linear alkylbenzene sulfonates

## 1. Introduction

The composition of laundry detergents is generally complex due to the numerous factors that have to be taken into consideration to ensure fresh clean garments at the end of the wash process. Sodium dodecylbenzene sulfonate, more commonly known as SDS or linear alkylbenzene sulfonates (LAS),

© 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.

is the most abundant anionic surfactant utilised in laundry detergents due to its excellent performance in removing water insoluble substances such as greasy and oily stains. As a commercial commodity, LAS is sold as a sodium salt which contains a mixture of homologues that has between 10 and 14 linear carbon atoms with a phenyl group attached to the linear alkyl chain and the sulfonate anion as shown in Figure 1 [1–6].

Figure 1. Chemical structure of sodium dodecylbenzene sulfonate (SDS).

The rapid biodegradation of LAS compounds especially under aerobic conditions consumes a large amount of bio-available oxygen that significantly increases the chemical oxygen demand, thus negative impacting on the environment and organisms persisting within that system [4, 5]. Oxidation of LAS by oxygen results in the formation of sulfophenylcarboxylic acid (SPC) that comprises one of the main products of biodegradation [7–9].

## 2. Quantification of LAS by Ultraviolet-Visible spectrophotometry

Ultraviolet-Visible (UV-Vis) spectrophotometry is one of the commonly used techniques for the quantification of surfactants, whereby the method of determination of anionic surfactants entails the use of a cationic dye that complexes with the anionic surfactant through the mechanism of ion association as shown in Figure 2 [10, 11].

Valuable structural information by mass spectrometric (MS) detection often allows for the qualitative analysis of surfactants [12]. Analysis of ethoxylated surfactants using soft ionisation techniques such as electrospray ionisation (ESI) and atmospheric pressure chemical ionisation (APCI) determines analytes in a cationized molecular state [13]. The use of mass spectrometry and addition of a volatile reagent like ammonium salt, for example, ammonium acetate that suppresses the formation of alkali salts improves the accuracy of LAS determinations. The determination of non-ionic surfactants is possible via the application of positive or negative ionization modes for ESI and APCI, with the best response obtained using the positive ion mode [13].

Liquid chromatography-mass spectroscopy (LC-MS) is a powerful analytical technique, that is an applied qualitative detection method for non-ionic surfactants as reported by many researchers [12, 14–17]. In addition, a direct application of gas chromatography-mass spectroscopy (GC-MS) is used in the analysis of non-ionic surfactants; however, this method is limited in its

Figure 2. Ion association complex formed between LAS and methylene [10].

is the most abundant anionic surfactant utilised in laundry detergents due to its excellent performance in removing water insoluble substances such as greasy and oily stains. As a commercial commodity, LAS is sold as a sodium salt which contains a mixture of homologues that has between 10 and 14 linear carbon atoms with a phenyl group attached to the linear

The rapid biodegradation of LAS compounds especially under aerobic conditions consumes a large amount of bio-available oxygen that significantly increases the chemical oxygen demand, thus negative impacting on the environment and organisms persisting within that system [4, 5]. Oxidation of LAS by oxygen results in the formation of sulfophenylcarboxylic acid

Ultraviolet-Visible (UV-Vis) spectrophotometry is one of the commonly used techniques for the quantification of surfactants, whereby the method of determination of anionic surfactants entails the use of a cationic dye that complexes with the anionic surfactant through the mechanism of

Valuable structural information by mass spectrometric (MS) detection often allows for the qualitative analysis of surfactants [12]. Analysis of ethoxylated surfactants using soft ionisation techniques such as electrospray ionisation (ESI) and atmospheric pressure chemical ionisation (APCI) determines analytes in a cationized molecular state [13]. The use of mass spectrometry and addition of a volatile reagent like ammonium salt, for example, ammonium acetate that suppresses the formation of alkali salts improves the accuracy of LAS determinations. The determination of non-ionic surfactants is possible via the application of positive or negative ionization modes for ESI

Liquid chromatography-mass spectroscopy (LC-MS) is a powerful analytical technique, that is an applied qualitative detection method for non-ionic surfactants as reported by many researchers [12, 14–17]. In addition, a direct application of gas chromatography-mass spectroscopy (GC-MS) is used in the analysis of non-ionic surfactants; however, this method is limited in its

(SPC) that comprises one of the main products of biodegradation [7–9].

and APCI, with the best response obtained using the positive ion mode [13].

ion association as shown in Figure 2 [10, 11].

2. Quantification of LAS by Ultraviolet-Visible spectrophotometry

alkyl chain and the sulfonate anion as shown in Figure 1 [1–6].

180 Biological Wastewater Treatment and Resource Recovery

Figure 1. Chemical structure of sodium dodecylbenzene sulfonate (SDS).

application due to the derivatization requirement for long ethoxy chain containing surfactants [17]. The use of solid-phase extraction (SPE) and GC-MS for the direct analysis of APEs is carried out, whereby a graphitized carbon black SPE cartridge and use of methanol/dichloromethane solvent system was implemented [18]. The use of ethyl violet and acridine orange dyes has been reported by researchers for extraction of anionic surfactants [19, 20]. Specifically, toluene and benzene solvents have been used for extraction of LAS complexes, which is deemed less toxic than chloroform, and have therefore been reported as a recommended replacement to the methylene blue method [19, 20]. High-performance liquid chromatography (HPLC) is a commonly applied technique for LAS determination and detection which includes ultraviolet (UV), fluorescence (FL), diode-array detection (DAD) and mass spectroscopy (MS).

Another method for the analysis of LAS by an ion-pair SPE technique and HPLC has been developed [21]. Extraction of LAS using C8, C18 and multiwall carbon-nanotubes was investigated and samples were quantified by reversed-phase HPLC using a C8 column and UV detection with isocratic elution at a retention time of 15 min using a methanol/water mobile phase containing 5 mM sodium acetate [21]. Quantification of LAS in environmental samples by HPLC-FL has been developed which entails Soxhlet extraction of the sample with gradient elution, retention time of 22 min and application of mobile phases, which include acetonitrile, water, triethylamine and acetic acid [22, 23]. Quantification of LAS in sewage sludge samples using HPLC-FL with a C8 column with microwave-assisted extraction is used for sample preparation. A comparison of separation of LAS using HPLC-FL and HPLC-DAD showed no significant difference between the two sets of results and that usage of either a FL or DAD detector are applicable [24]. HPLC-MS is considered the most accurate method for determination of LAS as it permits for both a qualitative and quantitative analysis of LAS [3, 5]. GC-MS is less often used for analysis of LAS, as this method would require derivatization of LAS into a volatile compound [25]. Quantification of anionic surfactants and inorganic constituents' viz., phosphates, silicates and zeolite, has been analysed by Inductively coupled plasma-optical emission spectroscopy (ICP-OES) [26]. Specifically, alkylbenzene sulfonates and alkyl sulphates were determined due to their ability to precipitate upon addition of calcium ions [26]. Non-ionic surfactants that are used widely in domestic and

industrial detergents [27] are represented by two major classes, which include alcohol ethoxylates (AE) and alkylphenol ethoxylates (APEOs) [28]. The most common non-ionic surfactants used in detergents are octylphenol ethoxylate (OPEO) and nonylphenol ethoxylate (NPEO) as shown in Figure 3 [29, 30].

Figure 3. Structure of nonylphenol ethoxylate [31, 32].

The toxicity of the surfactant is dependent on the length of the ethoxy chain. A more toxic behaviour is known to be displayed by APEOs with a shorter ethoxy chain (typically <4) when compared to longer ethoxy chain length APEOs (typically >10) [18]. APEOs can be degraded under both anaerobic and aerobic conditions, thus leading to the biotransformation of APEO into lipophilic metabolites of APEO [33]. The most common degradation products of APEOs include nonylphenol, octylphenol and mono- and diethoxylated compounds of NPEO and OPEO [33, 34], which are deemed toxic and have been found to be persistent in the environment, thus causing endocrine disrupting effects amongst aquatic organisms [34–36]. Other contributing important ingredients found in laundry detergents include builders and antifoaming agents. A common zeolite-based builder, sodium aluminium silicate, is often used as a builder in laundry detergents to reduce water hardness, while polydimethylsiloxane acts as an anti-foaming reagent.

## 3. Application of biological and electrocoagulative treatment methods to laundry wastewater

The separation of the solid matrices from the liquid matrices forms the basis for treatment of wastewater, which is most commonly achieved through coagulation-flotation methods [37]. During coagulative processes, an alteration of the surface properties of the individual particles occurs and this permits transformation into larger particles [38]. Inorganic salts of aluminium, iron or calcium are commonly used in coagulation processes [39]. In the coagulation process, small particles may form which decrease the efficiency in removal of pollutants from the wastewater streams and for this reason, flocculent agents are commonly used in conjunction with coagulation agents [40]. The efficiency of coagulation is enhanced by an increase in flocculation through accumulation of particles into larger settleable masses [38]. Polymerbased flocculants are commonly used for this purpose as a result of their large surface area, hence enabling the particles to group and settle, thus facilitating easy removal of pollutants from the wastewater.

Biological treatments have been mainly applied to the treatment of industrial effluent wastewater. The advantages associated with biological treatment of wastewater include a decreased amount of toxic and harmful chemicals coupled with an easy to implement green process [41]. Waste from effluents is recycled into an organism-based biomass through biological treatment, and can be easily disposed of naturally into the environment [41]. Major disadvantages associated with biological treatment of wastewater include:

a. large space requirement for the storage of biological waste,

industrial detergents [27] are represented by two major classes, which include alcohol ethoxylates (AE) and alkylphenol ethoxylates (APEOs) [28]. The most common non-ionic surfactants used in detergents are octylphenol ethoxylate (OPEO) and nonylphenol ethoxylate (NPEO) as shown in

The toxicity of the surfactant is dependent on the length of the ethoxy chain. A more toxic behaviour is known to be displayed by APEOs with a shorter ethoxy chain (typically <4) when compared to longer ethoxy chain length APEOs (typically >10) [18]. APEOs can be degraded under both anaerobic and aerobic conditions, thus leading to the biotransformation of APEO into lipophilic metabolites of APEO [33]. The most common degradation products of APEOs include nonylphenol, octylphenol and mono- and diethoxylated compounds of NPEO and OPEO [33, 34], which are deemed toxic and have been found to be persistent in the environment, thus causing endocrine disrupting effects amongst aquatic organisms [34–36]. Other contributing important ingredients found in laundry detergents include builders and antifoaming agents. A common zeolite-based builder, sodium aluminium silicate, is often used as a builder in laundry detergents to reduce water hardness, while polydimethylsiloxane acts as

3. Application of biological and electrocoagulative treatment methods to

The separation of the solid matrices from the liquid matrices forms the basis for treatment of wastewater, which is most commonly achieved through coagulation-flotation methods [37]. During coagulative processes, an alteration of the surface properties of the individual particles occurs and this permits transformation into larger particles [38]. Inorganic salts of aluminium, iron or calcium are commonly used in coagulation processes [39]. In the coagulation process, small particles may form which decrease the efficiency in removal of pollutants from the wastewater streams and for this reason, flocculent agents are commonly used in conjunction with coagulation agents [40]. The efficiency of coagulation is enhanced by an increase in flocculation through accumulation of particles into larger settleable masses [38]. Polymerbased flocculants are commonly used for this purpose as a result of their large surface area,

Figure 3 [29, 30].

an anti-foaming reagent.

Figure 3. Structure of nonylphenol ethoxylate [31, 32].

182 Biological Wastewater Treatment and Resource Recovery

laundry wastewater


In the application of biological treatment of wastewater, addition of a specific strain of bacteria to the wastewater is the main thrust of the system that subsequently targets specific oxidation and degradation of pollutants.

Biological wastewater treatment is often seen as an environmental friendly method, as there are generally no added chemicals involved. Some of the major concerns with regard to biological treatment of wastewater include the longer time periods for treatment, a larger surface area required and the addition of specialised bacteria for the specific degradation of pollutants. Chan demonstrated a method for treatment of laundry effluent through a combination of biological and chemical treatment methods [41].

The laundry effluent was treated biologically prior chemical treatment. This treatment method permitted the production of high-quality water that could be used for activities such as flushing and cleaning which reduced the consumption of water by the launderette. The quality of the water was assessed by measuring the following parameters: pH, DO, SS, COD and total surfactant concentration.

Electrocoagulation is often implemented as the primary treatment for wastewater due to its efficient pollutant removal as well as its safe and environmental friendly nature. Electrocoagulation involves the dissolution of sacrificial anodes due to the application of electric current. Aluminium and iron are the most generic anodes used for this purpose.

$$\text{Al}(\text{s}) \rightarrow \text{Al}^{3+}(\text{aq}) + 3\text{e}^- \tag{1}$$

$$\text{3H}\_2\text{O}\left(\ell\right) + \text{3e}^- \rightarrow \text{H}\_2(\text{g}) + \text{3OH}^-(\text{aq}) \tag{2}$$

Eqs. (1) and (2) represent the reactions taking place at the anode and cathode, respectively. The resultant metal ion reacts with hydroxide in the wastewater to form various metal hydroxides.

$$\text{Al}^{3+}\text{(aq)} + 3\text{H}\_2\text{O}\text{ (\ell)}\text{ (aq)} \rightarrow \text{Al(OH)}\_3 + 3\text{H}^+\text{(aq)}\tag{3}$$

$$\text{Al(OH)}\_{3}\text{(aq)} + \text{OH}^{-}\text{(aq)} \rightarrow \text{Al(OH)}\_{4}^{-}\text{(aq)}\tag{4}$$

Eqs. (3) and (4) represent the generation of aluminium hydroxy species during electrocoagulation.

Treatment of wastewater by electrocoagulation is known to effectively remove heavy metals, minerals and dyes from wastewater streams, hence making it a good treatment method for laundry wastewater. A high removal efficiency of organic compounds is obtained due to the various mechanisms that occur in the electrocoagulation cell. The pollutants adsorb onto the different aluminium hydroxy species depending on the chemical structure of the pollutant.

The hydrogen gas produced at the cathode induces flotation of the hydroxy species, hence allowing for a quick and efficient removal of pollutants from the wastewater. Aside from the production of aluminium hydroxy species, other mechanisms in the electrocoagulation cell occur which increases the efficient removal of pollutants from the wastewater stream. Reactions at the surface of the cathode also remove carbonate salts, which is abundant in laundry wastewater.

$$\mathrm{HCO}^{3-}\mathrm{(aq)} + \mathrm{OH}^{-}\mathrm{(aq)} \rightarrow \mathrm{CO}\_{3}^{2-}\mathrm{(aq)} + \mathrm{H}\_{2}\mathrm{O}\,\mathrm{(l)}\tag{5}$$

$$\text{Ca}^{2+}(\text{aq}) + \text{CO}\_3^{2-}(\text{aq}) \to \text{CaCO}\_3(\text{s}) \tag{6}$$

$$\text{Mg}^{2+}\text{(aq)} + \text{COs}^{2-}\text{(aq)} \rightarrow \text{MgCO}\_3\text{(s)}\tag{7}$$

Eqs. (5)–(7) represent the removal of carbonate from the wastewater as salts of calcium and magnesium. Laundry wastewater is also known to contain chloride salts. Electrolysis generates molecular chlorine, which can lead to the formation of hypochlorous acid and hypochlorite ions as shown in Eqs. (8)–(10). These species contain a relatively high oxidative potential, which allows for further degradation of organic pollutants in the wastewater stream.

$$2\text{Cl}^-(\text{aq}) \rightarrow \text{Cl}\_2(\text{g}) + 2\text{e}^- \tag{8}$$

$$\text{Cl}\_2(\text{g}) + \text{H}\_2\text{O} \rightarrow \text{HOCl} \text{ (aq)} + \text{H}^+(\text{aq}) + \text{Cl}^-(\text{aq}) \tag{9}$$

$$\text{HClO(aq)} \rightarrow \text{ClO}^-(\text{aq}) + \text{H}^+(\text{aq}) \tag{10}$$

In research presented by many scientists, electrocoagulation is described as the treatment of laundry effluent [42–44]. Iron and aluminium electrodes are used for electrocoagulation; however, aluminium electrodes had a greater efficiency in removal of pollutants from the laundry wastewater. Some investigations applied an ultrasonic bath during electrocoagulation which had a profound effect on the efficiency of the removal [42].

Over time, the formation of an inhibiting film due to high voltages applied to the electrodes impacts negatively on the efficiency of electrocoagulation. The measured parameters of phosphorous levels, detergent, COD, turbidity and conductivity in the laundry wastewater before and after the process of electrocoagulation are good indicators of the effectiveness of the electrocoagulative process [42–44].

Electrocoagulation using aluminium electrodes, as shown in Figure 4, has been applied as a method for treatment of wastewater obtained from a textile industry aimed at the removal of dye substances from wastewater [45, 46]. This method has accounted for a 99% efficiency in removal of the dye substances, measured by determination of the COD before and after treatment [45, 46]. The removal of heavy metals such as nickel, copper, zinc and chromium from synthetic and industrial wastewater by electrocoagulation using aluminium electrodes

Figure 4. Illustration of an electrocoagulation cell adapted from Wang et al. [42].

Treatment of wastewater by electrocoagulation is known to effectively remove heavy metals, minerals and dyes from wastewater streams, hence making it a good treatment method for laundry wastewater. A high removal efficiency of organic compounds is obtained due to the various mechanisms that occur in the electrocoagulation cell. The pollutants adsorb onto the different aluminium hydroxy species depending on the chemical structure of the pollutant.

The hydrogen gas produced at the cathode induces flotation of the hydroxy species, hence allowing for a quick and efficient removal of pollutants from the wastewater. Aside from the production of aluminium hydroxy species, other mechanisms in the electrocoagulation cell occur which increases the efficient removal of pollutants from the wastewater stream. Reactions at the surface of the cathode also remove carbonate salts, which is abundant in laundry wastewater.

2�

2�

Eqs. (5)–(7) represent the removal of carbonate from the wastewater as salts of calcium and magnesium. Laundry wastewater is also known to contain chloride salts. Electrolysis generates molecular chlorine, which can lead to the formation of hypochlorous acid and hypochlorite ions as shown in Eqs. (8)–(10). These species contain a relatively high oxidative potential,

In research presented by many scientists, electrocoagulation is described as the treatment of laundry effluent [42–44]. Iron and aluminium electrodes are used for electrocoagulation; however, aluminium electrodes had a greater efficiency in removal of pollutants from the laundry wastewater. Some investigations applied an ultrasonic bath during electrocoagulation which

Over time, the formation of an inhibiting film due to high voltages applied to the electrodes impacts negatively on the efficiency of electrocoagulation. The measured parameters of phosphorous levels, detergent, COD, turbidity and conductivity in the laundry wastewater before and after the process of electrocoagulation are good indicators of the effectiveness of the

Electrocoagulation using aluminium electrodes, as shown in Figure 4, has been applied as a method for treatment of wastewater obtained from a textile industry aimed at the removal of dye substances from wastewater [45, 46]. This method has accounted for a 99% efficiency in removal of the dye substances, measured by determination of the COD before and after treatment [45, 46]. The removal of heavy metals such as nickel, copper, zinc and chromium from synthetic and industrial wastewater by electrocoagulation using aluminium electrodes

which allows for further degradation of organic pollutants in the wastewater stream.

<sup>2</sup>�ðaqÞþ H2O <sup>ð</sup>ℓ<sup>Þ</sup> (5)

ðaqÞ ! CaCO3ðsÞ (6)

ðaqÞ ! MgCO3ðsÞ (7)

2Cl�ðaqÞ ! Cl2ðgÞ þ 2e� (8)

HClOðaqÞ ! ClO�ðaqÞ þ HþðaqÞ (10)

Cl2ðgÞ þ H2O ! HOCl ðaqÞ þ HþðaqÞ þ Cl�ðaqÞ (9)

HCO<sup>3</sup>�ðaqÞþ OH�ðaqÞ ! CO3

Ca2þðaqÞ þ CO3

184 Biological Wastewater Treatment and Resource Recovery

Mg2þðaqÞ þ CO3

had a profound effect on the efficiency of the removal [42].

electrocoagulative process [42–44].

has been widely applied. An added advantage of electrocoagulation in addition to removal of heavy metals from wastewater stream also significantly decreased the COD [47]. In a research study by Ramcharan and Bissessur, a comparison of electrocoagulation and biological treatment of Laundry Wastewater (LWW) was reported [48]. The surfactant concentration, chemical oxygen demand and total dissolved solids were the general water guideline parameters used to assess the success of the treatment system. The wastewater was characterised after each wash and rinse cycle discharged from a domestic washing machine and are referred to as first wash cycle wastewater (W1), first rinse cycle wastewater (R1) and second rinse cycle wastewater (R2). The two major parameters, which influenced the above treatment methods, were the period allocated for treatment and the suitability of each treatment method to a variety of wastewater matrixes. The successful treatment of R1 and R2 was obtained using the biological method, while electrocoagulation was successful for W1, R1 and R2 (Figure 5). The sample matrix of W1 was not compatible for biological treatment, as the bacterium was not able to cultivate under such harsh conditions. Aeration of W1 proved to decrease the concentration of the surfactant because SDS is susceptible to degradation under oxidative conditions.

Degradation of the bacteria is imminent upon exposure to the strongly basic pH of the first wash laundry wastewater, which increased the organic content thereby increasing the COD in laundry wastewater from the first wash during biological treatment.

The dominance of the electrocoagulative treatment method over the biological method of LWW is further supported by the COD levels attained as shown in Figure 5. It is clearly evident that upon treatment of W1, a gradual increase in the COD levels occurs over a prolonged period of time. The highly alkaline nature of the wastewater induces breakdown of bacterial cells, thus implementing an increase in the organic content and thereby consequently

Figure 5. Decrease in surfactant concentration after application of (a) biological treatment and (b) electrocoagulative treatments to laundry wastewater from the first wash (W1), first rinse stage (R1) and second rinse stage (R2). Reproduced from Ref. [48].

causing an increase in the COD level of W1. However, a marked decrease in COD level occurred during the implementation of the electrocoagulative technique as shown in Figure 5. Finally, the persistence of LAS in solution is directly linked to the COD level. The effective removal of LAS by the electrocoagulative treatment caused a marked decrease in the organic content present; thus, a rapid decrease in the COD is observed especially for R2 in the initial onset (within the 5 minutes of implementation) of electrocoagulation as shown in Figure 6.

The dominance of the electrocoagulative treatment method over the biological method of LWW is further supported by the COD levels attained as shown in Figure 5. It is clearly evident that upon treatment of W1, a gradual increase in the COD levels occurs over a prolonged period of time. The highly alkaline nature of the wastewater induces breakdown of bacterial cells, thus implementing an increase in the organic content and thereby consequently

186 Biological Wastewater Treatment and Resource Recovery

Figure 5. Decrease in surfactant concentration after application of (a) biological treatment and (b) electrocoagulative treatments to laundry wastewater from the first wash (W1), first rinse stage (R1) and second rinse stage (R2). Reproduced from

Ref. [48].

Figure 6. COD Levels of laundry wastewater samples after first wash (W1), first rinse (R1) and second rinse (R2) cycles when subjected to (a) biological treatment and (b) electrocoagulation. Reproduced from Ref. [48].

TDS levels at the different wash and rinse cycles of LWW showed an increasing trend when treated biologically, while the electrocoagulation method of treatment for LWW showed a decrease in the TDS levels as shown in Figure 7. This is chiefly due to the quick polymeric generation of aluminium hydroxide species during electrocoagulation had allowed for adsorption of SDS in LWW whilst promoting effective TDS removal through settlement of the polymeric floc generated.

Figure 7. TDS levels in laundry wastewater for first wash (W1), first rinse (R1) and second rinse samples (R2) after (a) biological treatment and (b) electrocoagulation. Reproduced from Ref. [48].

Supporting kinetic data is pivotal when implementing pilot wastewater treatment systems. The adsorption kinetics is one of the important parameters used to assess sustainability of the treatment system. A kinetic study on the adsorption capacity of the aluminium hydroxy species was investigated by Ramcharan and Bissessur [48]. The Ho pseudo second-order expression was used to evaluate the adsorption capacity for surfactant removal in laundry wastewater from the first wash, first rinse cycle and second rinse cycle as shown in Eq. (11) below. A second-order reaction was observed from the plot of t/qt vs. t shown in Figure 8 with R2 values >0.99.

TDS levels at the different wash and rinse cycles of LWW showed an increasing trend when treated biologically, while the electrocoagulation method of treatment for LWW showed a decrease in the TDS levels as shown in Figure 7. This is chiefly due to the quick polymeric generation of aluminium hydroxide species during electrocoagulation had allowed for adsorption of SDS in LWW whilst promoting effective TDS removal through settlement of the

Figure 7. TDS levels in laundry wastewater for first wash (W1), first rinse (R1) and second rinse samples (R2) after

(a) biological treatment and (b) electrocoagulation. Reproduced from Ref. [48].

polymeric floc generated.

188 Biological Wastewater Treatment and Resource Recovery

$$\frac{\mathbf{t}}{\mathbf{q}\_{\text{t}}} = \frac{1}{\mathbf{k}\_{2}\mathbf{q}\_{\text{e}}^{2}} + \frac{\mathbf{t}}{\mathbf{q}\_{\text{e}}} \tag{11}$$

The percentage of efficiency of adsorption (% E) was based on calculations using Eq. (13) below, where C0 and C corresponds to the initial and specific concentration of the surfactant at time t. The values of the adsorption efficiency at equilibrium (qe) and rate of adsorption (k2) was based on calculations using Eqs. (13) and (14), respectively. The rate of adsorption of the surfactants is significantly lower for laundry wastewater discharged from the first wash as compared to laundry wastewater from the first and second rinses as shown in Table 1. It can be easily inferred that a reduced amount time is required for the treatment of laundry wastewater disposed after the first and second rinses.

Figure 8. A Plot of t/qt vs. t showing second-order reaction kinetics for the adsorption capacity of surfactant by aluminium hydroxy species. Reproduced from Ref. [48].


Table 1. The Lagergren parameters for adsorption of surfactants by aluminium hydroxy species.

$$\text{\textquotedblleft}\text{\textquotedblright}\text{\textquotedblright}\text{\textquotedblleft}=\frac{\text{\textquotedblleft}\_{0}-\text{\textquotedblleft}\_{0}}{\text{\textquotedblleft}\_{0}}\times 100\text{\textquotedblright}\tag{12}$$

$$\mathbf{q}\_{\mathbf{e}} = \frac{1}{\text{Slope}}\tag{13}$$

$$k\_2 = \frac{\text{Slope}^2}{\text{Intercept}}\tag{14}$$

#### 4. Conclusions

The application of electrocoagulative and biological treatment methods effectively decreased the amount of surfactant concentration in laundry wastewater after all rinsing stages. In comparison, the electrocoagulative technique was found to be a more efficient treatment method of the two due to its ability to reduce the levels of the surfactant, COD and TDS over a considerably shorter period of time and its ability to be applied to a wider range of wastewater samples. A modification to the electrocoagulation treatment process whereby the addition of Biospinners® was carried out and was found to further reduce the levels of the surfactant, COD and TDS within the same applied period of time. Modification due to addition of Biospinners was shown to increase aeration and surface area, and facilitated the removal of an overlaying film of aluminium hydroxy species formed on the electrodes. The adsorption of LAS by aluminium hydroxy species was found to take place at a lower rate for W1, in comparison to R1 and R2 as shown by the kinetics in this study. From this, it is evident that there is a need for isolated treatments of laundry wastewater W1, R1 and R2, thus ensuring a reduced period of treatment and also ensuring the total output cost of the treatment method is kept to a minimum.

#### Author details

Terelle Ramcharan and Ajay Bissessur\*

\*Address all correspondence to: bissessura@ukzn.ac.za

School of Chemistry and Physics, University of KwaZulu-Natal, Durban, South Africa

## References

% E <sup>¼</sup> <sup>C</sup><sup>0</sup> � <sup>C</sup> C0

Table 1. The Lagergren parameters for adsorption of surfactants by aluminium hydroxy species.

Lagergren parameter W1 R1 R2 Experimental qe 77.60 67.27 60.69 Calculated qe 77.52 68.97 57.47 R<sup>2</sup> 0.997 0.999 0.999 k2 8.53 · 10-4 2.53 · 10-3 2.21 · 10-3

qe <sup>¼</sup> <sup>1</sup>

k2 <sup>¼</sup> Slope<sup>2</sup>

The application of electrocoagulative and biological treatment methods effectively decreased the amount of surfactant concentration in laundry wastewater after all rinsing stages. In comparison, the electrocoagulative technique was found to be a more efficient treatment method of the two due to its ability to reduce the levels of the surfactant, COD and TDS over a considerably shorter period of time and its ability to be applied to a wider range of wastewater samples. A modification to the electrocoagulation treatment process whereby the addition of Biospinners® was carried out and was found to further reduce the levels of the surfactant, COD and TDS within the same applied period of time. Modification due to addition of Biospinners was shown to increase aeration and surface area, and facilitated the removal of an overlaying film of aluminium hydroxy species formed on the electrodes. The adsorption of LAS by aluminium hydroxy species was found to take place at a lower rate for W1, in comparison to R1 and R2 as shown by the kinetics in this study. From this, it is evident that there is a need for isolated treatments of laundry wastewater W1, R1 and R2, thus ensuring a reduced period of treatment and also

ensuring the total output cost of the treatment method is kept to a minimum.

School of Chemistry and Physics, University of KwaZulu-Natal, Durban, South Africa

4. Conclusions

Reproduced from Ref. [48].

190 Biological Wastewater Treatment and Resource Recovery

Author details

Terelle Ramcharan and Ajay Bissessur\*

\*Address all correspondence to: bissessura@ukzn.ac.za

· 100 (12)

Slope (13)

Intercept (14)


[25] McEvoy J., and Giger W., Determination of linear alkylbenzenesulfonates in sewage sludge by high-resolution gas chromatography/mass spectrometry. Environmental Science Technology, 1986. 20: pp. 376–383.

[13] Frömel T., and Knepper T. P., Mass spectrometry as an indispensable tool for studies of biodegradation of surfactants. TrAC Trends in Analytical Chemistry, 2008. 27(11): pp. 1091–

[14] Shao B., Hu J.-Y., and Yang M., Determination of nonylphenol ethoxylates in the aquatic environment by normal phase liquid chromatography–electrospray mass spectrometry. Journal

[15] Eadsforth C. V., Sherren A. J., Selby M. A., Toy R., Eckhoff W. S., McAvoy D. C., and Matthijs E., Monitoring of environmental fingerprints of alcohol ethoxylates in Europe and

[16] Andreu V., Ferrer E., Rubio J. L., Font G., and Picó Y., Quantitative determination of octylphenol, nonylphenol, alkylphenol ethoxylates and alcohol ethoxylates by pressurized liquid extraction and liquid chromatography–mass spectrometry in soils treated with sewage sludges.

[17] Koh Y. K. K., Chiu T. Y., Boobis A. R., Cartmell E., Pollard S. J. T., Scrimshaw M. D., and Lester J. N., A sensitive and robust method for the determination of alkylphenol polyethoxylates and their carboxylic acids and their transformation in a trickling filter wastewater treatment

[18] Earls A., and Reydellet I., Determination of specific alkylphenol ethoxylates in textiles. 2006. [cited 31 March 2014] Available from: http://www.governmentchemist.org.uk/ dm\_documents/Determination%20of%20specific%20alkylphenol%20ethoxylates%20in%

[19] Motomlzu S., Fujiwara F., Fujiwara A., and Toei K., Solvent extraction-spectrophotometric determination of anionic surfactants with ethyl violet. Analysis Chemistry, 1982. 54: pp. 392–397.

[20] Adak A., Pal A., and Bandyopadhyay M., Spectrophotometric determination of anionic surfactants in wastewater using acridine orange. Indian Journal of Chemical Technology, 2005.

[21] Guo P., Guan Z., Wang W., Chen B., and Huang Y., Determination of linear alkylbenzene sulfonates by ion-pair solid-phase extraction and high-performance liquid chromatography. Talanta,

[22] Villar M., Callejón M., Jiménez J. C., Alonso E., and Guiraúm A., New rapid methods for determination of total LAS in sewage sludge by high performance liquid chromatography (HPLC)

[23] Bengoechea C., and Cantarero A. S., Analysis of linear alkylbenzene sulfonate in waste water and sludge by high performance liquid chromatography: An exercise of validation. Journal of

[24] Villar M., Callejón M., Jiménez J. C., Alonso E., and Guiráum A., Optimization and validation of a new method for analysis of linear alkylbenzene sulfonates in sewage sludge by liquid chromatography after microwave-assisted extraction. Analytica Chimica Acta, 2007. 599(1): pp. 92–97.

and capillary electrophoresis (CE). Analytica Chimica Acta, 2009. 634: pp. 267–271.

Canada. Ecotoxicology and Environmental Safety, 2006. 64(1): pp. 14–29.

Science of the Total Environment, 2007. 378(1–2): pp. 124–129.

of Chromatography A, 2002. 950(1–2): pp. 167–174.

plant. Chemosphere, 2008. 73(4): pp. 551–556.

Surfactants and Detergents, 2009. 12(1): pp. 21–29.

20textiles\_6sQz8.pdf

12: pp. 145–148.

2011. 84: pp. 587–592.

1106.

192 Biological Wastewater Treatment and Resource Recovery


**Bioenergy for Resource Recovery**

[38] vanLoon G. W., and Duffy S. J., Enivirontal Chemistry, A Global Perpspective. 2005, Oxford

[39] Guida M., Mattei M., Della Rocca C., Melluso G., and Meriç S., Optimization of alumcoagulation/flocculation for COD and TSS removal from five municipal wastewater. Desalina-

[40] Lee C. S., Robinson J., and Chong M. F., A review on application of flocculants in wastewater treatment. Process Safety and Environmental Protection (2014), http://dx.doi.org/10.1016/j.

[41] Chan H., High performance achieved by microbes to separate laundry effluents resulting in producing high water quality in a compact area. Separation and Purification Technology,

[42] Wang C.-T., Chou W.-L., and Kuo Y.-M., Removal of COD from laundry wastewater by electrocoagulation/electroflotation. Journal of Hazardous Materials, 2009. 164(1): pp.

[43] Janpoor F., Torabian A., and Khatibikamal V., Treatment of laundry waste-water by electrocoagulation. Journal of Chemical Technology & Biotechnology, 2011. 86(8): pp. 1113–1120. [44] Önder E., Koparal A. S., and Öğütveren Ü. B., An alternative method for the removal of surfactants from water: Electrochemical coagulation. Separation and Purification Technology,

[45] Merzouk B., Yakoubi M., Zongo I., Leclerc J. P., Paternotte G., Pontvianne S., and Lapicque F., Effect of modification of textile wastewater composition on electrocoagulation effi-

[46] Chithra K., Thilakavathi R., Murugan A. A., and Marimuthu C., Treatment of textile effluent

[47] Dermentzis K., Christoforidis A., and Valsamidou E., Removal of nickel, copper, zinc and chromium from synthetic and industrial wastewater by electrocoagulation. International Journal

[48] Ramcharan T., and Bissessur A., Treatment of laundry wastewater by biological and electrocoagulation methods. Water Science & Technology, 2016. doi:10.2166/wst.2016.464.2016.

using sacrificial electrode. Modern Applied Science, 2008. 2: pp. 38–43.

University Press Inc., New York, p. 515.

tion, 2007. 211(1–3): pp. 113–127.

194 Biological Wastewater Treatment and Resource Recovery

psep.2014.04.010.

2012. 90: pp. 101–108.

2007. 52(3): pp. 527–532.

ciency. Desalination, 2011. 275(1–3): pp. 181–186.

of Environmental Sciences, 2011. 1(5): pp. 697–710.

81–86.

#### **Biohydrogen Production from Wastewaters Biohydrogen Production from Wastewaters**

Periyasamy Sivagurunathan, Periyasamy Sivagurunathan, Gopalakrishnan Kumar,

Gopalakrishnan Kumar, Arivalagan Pugazhendhi, Arivalagan Pugazhendhi, Guangyin Zhen,

Guangyin Zhen, Takuro Kobayashi and Kaiqin Xu Takuro Kobayashi and Kaiqin Xu

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

http://dx.doi.org/10.5772/65891

#### **Abstract**

Biohydrogen production technology is an emerging field for the advanced wastewater treatment with cogeneration of energy. Besides, hydrogen is an excellent candidate with high energy value (122 kJ/g) than other known carbon‐based fuels with no adverse effects to the environment as it releases only water vapor as the by‐products during the combustion. Biohydrogen production technology can be assisted through two major pathways: (a) light‐dependent reaction (biophotolysis and photofermentation) and (b) light‐independent reaction (dark fermentation and microbial electrohydrogenesis cells). The light‐dependent reaction can be catalyzed by photosynthetic bacteria, whereas the dark fermentation catalyzed by the heterotrophic bacterial group of facultative and obligate anaerobes. The wastewaters are a rich source of organic nutrients which supports the growth of hydrogen producers along with the disposal of waste and energy recovery. In the present chapter, the recent advancements on biohydrogen production technology from wastewaters with respect to the (a) inoculum development, (b) process optimization, (c) scale‐up and (d) the challenges and perspectives toward the improvement of this emerging technology for the wastewater treatment.

**Keywords:** biohydrogen, dark fermentation, wastewater

## **1. An overview of biohydrogen production**

The growing demand of the energy for daily life purposes urged us to seek an alternative and renewable energy carrier with less emission of the pollutants. Hydrogen is an essential and promising candidate for replacing the fossil fuels depletion and greenhouse gas emission

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.

reduction. When burning, it releases only water vapor as a by‐product with no adverse harmful gases such as NOx and SiO2 , and hence, it is considered as clean and carbon‐free energy carrier. The energy content of hydrogen is 122 kJ/g, which is 2.75‐fold greater than the existing hydrocarbon fuels makes an ideal energy carrier for various industrial, transportation and power generations.

Different types of hydrogen production are available such as fossil fuel by hydrocarbon reforming, coal gasification and partial oxidation which requires high temperature and pressure. The biologically adopted hydrogen production methods can be classified as (i) biophotolysis of water using algae/cyanobacteria, (ii) photodecomposition of organic compounds using photosynthetic bacteria, (iii) dark fermentative hydrogen production using strict anaerobic or facultative bacteria and (iv) microbial fuel cells (MFC). Each biological production method had distinct advantages and limitations. For example, the green algae/ cyanobacteria decomposes the water into gas (H2 ) and liquid (H2 O) in the presence of sunlight by photosynthesis pathway, whereas the slow growth of the algal cells and an inhibition of hydrogenase enzyme with the presence of traces of oxygen limit their application in large scale extent. The photosynthetic bacteria and dark fermentation bacteria share a similar metabolism for the breakdown of organic compounds for their energy and the liberation of energy [1, 2]. The photosynthetic bacteria use organic acids as a substrate and prone to the ammonium and oxygen toxicity, making it as unsuitable for commercial hydrogen production. In contrast, the dark fermentation degrades wide range of organic waste from complex lignocellulose, food waste and industrial wastewater to simpler monomers (sucrose, glucose). However, the chemical oxygen demand (COD) removal efficiency of the dark fermentation is relatively lower 33%, as it requires further treatment before discharge into the system. Moreover, the biomass growth rate and hydrogen production rate of the dark fermentation are comparatively higher than the other hydrogen production methods and make it as attractive candidate for industrial and commercial biohydrogen production [3]. Recently, the auxiliary methods for the hydrogen production from hydrogen effluent have been emerged through microbial fuel cell (MFC) or bioelectrochemical systems (BES) technology.

### **2. Hydrogen‐producing microorganisms**

**Table 1** displayed the microbial strains helpful for biohydrogen production through dark fermentation [4]. Hydrogen production during fermentation involves either facultative anaerobic bacteria or strict anaerobic bacteria. Facultative anaerobes are capable of growing in the absence of oxygen. The most common hydrogen*‐producing* facultative anaerobes are *Klebsiella pneumoniae* [5], *Escherichia coli* [6], *Enterobacter aerogenes* [7], *Rhodospirillum rubrum*, *Methanobacterium formiccium* [4]. Chookaew et al. [5] reported that *Klebsiella* sp. TR17 is able to produce biohydrogen from crude glycerol in an up‐flow anaerobic sludge blanket (UASB) reactor with highest HPR of 242.15 mmol H2 /L/d and HY of 44.27 mmol H2 /g glycerol. Besides, the *Klebsiella pneumoniae* produce valuable by‐products such as 1,3‐propanediol and 2,3‐butanediol [8]. Reungsang et al. [7] reported that the immobilized *E. aerogenes* ATCC


**Table 1.** Hydrogen production using pure cultures WW, wastewater; CMS, condensed molasses soluble.

13048 produced major soluble metabolite products (SMPs), such as ethanol, 1,3‐propanediol (1,3‐PD), formic acid and acetic acid.

#### **2.1. Facultative anaerobes**

reduction. When burning, it releases only water vapor as a by‐product with no adverse harm-

carrier. The energy content of hydrogen is 122 kJ/g, which is 2.75‐fold greater than the existing hydrocarbon fuels makes an ideal energy carrier for various industrial, transportation and

Different types of hydrogen production are available such as fossil fuel by hydrocarbon reforming, coal gasification and partial oxidation which requires high temperature and pressure. The biologically adopted hydrogen production methods can be classified as (i) biophotolysis of water using algae/cyanobacteria, (ii) photodecomposition of organic compounds using photosynthetic bacteria, (iii) dark fermentative hydrogen production using strict anaerobic or facultative bacteria and (iv) microbial fuel cells (MFC). Each biological production method had distinct advantages and limitations. For example, the green algae/

light by photosynthesis pathway, whereas the slow growth of the algal cells and an inhibition of hydrogenase enzyme with the presence of traces of oxygen limit their application in large scale extent. The photosynthetic bacteria and dark fermentation bacteria share a similar metabolism for the breakdown of organic compounds for their energy and the liberation of energy [1, 2]. The photosynthetic bacteria use organic acids as a substrate and prone to the ammonium and oxygen toxicity, making it as unsuitable for commercial hydrogen production. In contrast, the dark fermentation degrades wide range of organic waste from complex lignocellulose, food waste and industrial wastewater to simpler monomers (sucrose, glucose). However, the chemical oxygen demand (COD) removal efficiency of the dark fermentation is relatively lower 33%, as it requires further treatment before discharge into the system. Moreover, the biomass growth rate and hydrogen production rate of the dark fermentation are comparatively higher than the other hydrogen production methods and make it as attractive candidate for industrial and commercial biohydrogen production [3]. Recently, the auxiliary methods for the hydrogen production from hydrogen effluent have been emerged through microbial fuel cell (MFC) or bioelectrochemical systems (BES)

**Table 1** displayed the microbial strains helpful for biohydrogen production through dark fermentation [4]. Hydrogen production during fermentation involves either facultative anaerobic bacteria or strict anaerobic bacteria. Facultative anaerobes are capable of growing in the absence of oxygen. The most common hydrogen*‐producing* facultative anaerobes are *Klebsiella pneumoniae* [5], *Escherichia coli* [6], *Enterobacter aerogenes* [7], *Rhodospirillum rubrum*, *Methanobacterium formiccium* [4]. Chookaew et al. [5] reported that *Klebsiella* sp. TR17 is able to produce biohydrogen from crude glycerol in an up‐flow anaerobic sludge blanket

erol. Besides, the *Klebsiella pneumoniae* produce valuable by‐products such as 1,3‐propanediol and 2,3‐butanediol [8]. Reungsang et al. [7] reported that the immobilized *E. aerogenes* ATCC

/L/d and HY of 44.27 mmol H2

/g glyc-

, and hence, it is considered as clean and carbon‐free energy

) and liquid (H2

O) in the presence of sun-

ful gases such as NOx

power generations.

technology.

and SiO2

198 Biological Wastewater Treatment and Resource Recovery

cyanobacteria decomposes the water into gas (H2

**2. Hydrogen‐producing microorganisms**

(UASB) reactor with highest HPR of 242.15 mmol H2

Facultative anaerobes play important roles in H2 production by biological routes, as it can grow in the presence of oxygen, higher biomass growth rate and utilization of wide range of organic wastes. The widely studied facultative anaerobic model for hydrogen production is *E. coli* and *E. aerogenes*. Facultative anaerobes convert pyruvate to acetyl‐coA and formate with the catalysis of pyruvate formate‐lyase complex and then release H2 with formate hydrogen lyase. The maximum theoretical hydrogen yield is 2 mol of H2 per mole of glucose. The glucose metabolic pathway yields succinate, lactate, acetate, ethanol and formate, as fermentation end‐products. *Enterobacter* sp. have been widely used in various reactor configuration from batch to continuous mode operation. Several attempts like coculture of the facultative anaerobes with strict anaerobes have been assessed to improve the biohydrogen production. The coculture has advantages over pure culture due to the less maintenance, technical feasibility and faster substrate utilization rate. Sivagurunathan et al. [9] demonstrated that the addition of enriched mixed culture with *Enterobacter cloacae* enhanced the hydrogen production rate of 2.25 L/L‐d from beverage wastewater. In another report [6], immobilization of *E. coli* cells using sodium alginate increased the hydrogen production efficiency from fructose (1.17 mol/mol hexose) and beverage wastewater (1.65 mol/mol hexose), respectively.

#### **2.2. Mixed consortia**

The mixed consortia can be derived from a variety of different natural sources, such as sewage sludge, anaerobically digested sludge, compost, animal manure and contaminated soil (**Table 2**). Mixed culture contains different types of bacteria; it also contains methanogens or hydrogen‐consuming bacteria. Mixing also determines the local shear stress that the flow applies to microorganisms. Mixed culture can be obtained from aerobic or anaerobic sludge in wastewater treatment plants or compost piles or any other source of bacteria. Currently,


**Table 2.** Hydrogen production using mixed consortia.

researchers mainly focused two routes for microbial fermentative hydrogen production: one utilizes pure microbial strains and the other employs a mixed microbial consortium. Generally, the hydrogen‐producing efficiency and hydrogen yield of pure bacteria are lower than mixed consortia. Several investigators have focused on hydrogen production by microbial fermentation using a mixed microbial consortium, because of low‐cost organic substrates, high hydrogen yields and operated in non‐sterile conditions.

## **3. Process optimization for scale‐up**

Biohydrogen production is an emerging research area in the sustainable biofuel production via anaerobic fermentation technology. Though the hydrogen production from biological routes seems attractive over other commercial process, the operational conditions are essential to optimize in order to attain the maximum achievable hydrogen production rates and yields. A few important parameters on these aspects are as follows:


Biohydrogen production through mixed consortia is a complex bioprocess where the inoculum source, substrate type, environmental factors (pH, temperature and substrate concentration), nutrient availability and HRT can influence the metabolic reactions of hydrogen producers. Optimizing these factors is a paramount importance for enhancing the hydrogen production efficiency from organic wastes.

#### **3.1. Inoculum pretreatment**

researchers mainly focused two routes for microbial fermentative hydrogen production: one utilizes pure microbial strains and the other employs a mixed microbial consortium. Generally, the hydrogen‐producing efficiency and hydrogen yield of pure bacteria are lower than mixed consortia. Several investigators have focused on hydrogen production by microbial fermentation using a mixed microbial consortium, because of low‐cost organic substrates,

**(mol/mol hexose added)**

**References**

Biohydrogen production is an emerging research area in the sustainable biofuel production via anaerobic fermentation technology. Though the hydrogen production from biological routes seems attractive over other commercial process, the operational conditions are essential to optimize in order to attain the maximum achievable hydrogen production rates and

Biohydrogen production through mixed consortia is a complex bioprocess where the inoculum source, substrate type, environmental factors (pH, temperature and substrate concentration), nutrient availability and HRT can influence the metabolic reactions of hydrogen

high hydrogen yields and operated in non‐sterile conditions.

**Wastewater type Inoculum source Hydrogen yield (HY)** 

BWW EMC‐sewage sludge + pig slurry 1.95 mol/mol glu [37] BWW EMC + *E. coli* XL1 blue 260 mL/g COD [11] Sugar beet juice Anaerobic sludge 2.0 mol/mol glu [38] Distillery WW Anaerobic sludge 10.95 mmol/g COD [39] Dairy WW Anaerobic sludge 15.33 mmol/g COD [40] Cheese processing WW Mixed cultures 10.2 mM/g COD [41] Organic WW Soil 2.32 mol/mol [42] Herbal WW Slaughter house sludge 165 mL/g COD [43] CMS Anaerobic sludge 1.5 mol/mol [44] Brewery WW Anaerobic sludge 1.21 mol/mol [45] GWW Anaerobic sludge 0.75 mol/mol [46] WW, wastewater; BWW, beverage wastewater; CMS, condensed soluble molasses; GWW, glycerine wastewater.

yields. A few important parameters on these aspects are as follows:

**3. Process optimization for scale‐up**

**Table 2.** Hydrogen production using mixed consortia.

200 Biological Wastewater Treatment and Resource Recovery

(a) Inoculum pretreatment

(c) Nutrient availability (d) Hydraulic retention time

(b) pH

The active acidogenic hydrogen‐producing biocatalyst role is crucial, notably in a complex mixed culture microenvironment. In general, the hampering hydrogen yield from mixed consortia was observed due to (i) the competition of hydrogen‐consuming microbes and (ii) diversion of the metabolic flux toward non‐favorable hydrogen by‐products. The hydrogen consumers, such as lactic acid bacteria, methanogenes and sulfur‐reducing bacteria, not only act as a competitor for the hydrogen producers but also synthesize various by‐products, which affect the growth of hydrogen producers. For instance, the release of proteinaceous toxin (bacteriocins) by lactate‐producing bacteria acts as a suppressing factor for hydrogen production and microbial growth [10]. Thus, when the mixed culture is used as an inoculum source, pretreatment step acts as an important role in determining the efficiency of the hydrogen production from mixed consortia. **Table 3** showed the various pretreatment methods for enriching the hydrogen producers. The pretreatment step promotes the selective enrichment of hydrogen producers with a suppression of the hydrogenotrophic methanogenes and other hydrogen consumers. The suppression of the hydrogen consumers by pretreatment process allows the mixed consortia to produce the hydrogen as a major product. The fundamental basics relied with the pretreatment method are the physiological difference of the microorganisms. The spore‐forming hydrogen producers survive under the harsh pretreatment conditions, whereas the vegetative cells ruptured/killed during the pretreatment. Various pretreatment methods, such as heat shock, acid shock, alkali shock, chemical agents, load shock and oxygen shock, have been assessed for enriching the hydrogen producers from mixed consortia. Each pretreatment step has a significant impact on the suppression of the microbial populations and also the distribution of the microbial metabolism.

Among the various pretreatment methods, the heat‐shock [11] pretreatment has been widely accepted as a suitable method for preparing the hydrogen‐producing seed inocula, due to the relatively simple method for the suppression of the hydrogen consumers and selective enrich-


**Table 3.** Inoculum pretreatment method for enriching hydrogen production mixed consortia.

ment of the sporulating hydrogen‐producing bacteria such as *Clostridium* sp. The acid‐shock [12] and base‐shock [13] pretreatments suppress the methanogenic activity by the narrow selective growth pH range of the methanogenes (6–7.5), whereas the *Clostridium* populations survive in the harsh condition due to the spore‐forming capability. The chemical shock methods such as chloroform [14] and 2‐bromoethanesulfonic acid (BESA) [15] have a complex structure, analog to the methanogenic coenzyme, and it acts as a inhibitor for the methanogenes. This method facilitated the suppression of the methanogenes, whereas the other non‐spore‐forming hydrogen producers such as *Enterobacter* sp. can also survive with the presence of *Clostridium* sp., thus enhancing the substrate utilization and hydrogen yield. The load‐shock [16] treatment is directed by the exposure of the inoculum to a higher substrate concentration, and it leads to the surge in the pH with an accumulation of organic acids and inhibits the methanogenic populations.

Ren et al. [17] demonstrated that application of various pretreatment methods, such as acid, alkaline, heat‐shock and repeated aeration, can greatly affect the metabolic pathway and the microbial community distribution pattern. The dominant butyric acid‐mediated hydrogen metabolism was observed with heat‐shock and alkaline treatment, and mixed‐type fermentation pathway was observed with the acid pretreatment, whereas the ethanol‐type pathway was observed with repeated aeration treatment with a maximum hydrogen yield of 1.96 mol/ mol glucose. The microbial community characterized by denaturing gradient gel electrophoresis (DGGE) revealed that the changes in the composition of the microbial dynamics affect the hydrogen yield. The strain *Ethanoligenens harbinens* was detected under repeated aeration condition with an ethanol‐mediated pathway, and the hydrogen‐consuming propionic acid bacterium *Propionibacterium propionicus* was detected in acid treatment with low hydrogen productivity. The heat‐shock‐mediated mixed culture was dominated with *Clostridium* sp. which represents the butyric‐acid‐type metabolic pathway. Based on the evidence, the appropriate pretreatment method is essential for enriching the hydrogen‐producing bacterial populations and enhanced hydrogen production.

#### **3.2. pH**

pH is the key driven parameter affecting the cellular metabolism of hydrogen‐producing bacterial populations, since the prevalent end products of the bacterial metabolism vary with the changes in the medium pH. Based on the pH and the major end products formation, three metabolic pathways have been proposed (a) ethanol type (EtOH) (Eq. 1), (b) butyric type (HBu) (Eq. 2) and (c) propionic type (HPr) (Eq. 3). The former, HBu type, involved in the hydrogen‐generating reactions, whereas the latter, HPr type, involved in the hydrogen‐scavenging reactions. Hence, the elimination of the propionate formation is an essential step for the enhancement of hydrogen production.

$$\rm C\_{s}H\_{12}O\_{s} + 2H\_{2}O \rightarrow 2CH\_{3}CH\_{2}OH + 2HCO\_{3}^{-} + 2H\_{2} \tag{1}$$

*Δ G*<sup>0</sup> = − 235.0 kJ/mol

$$\rm C\_{s}H\_{12}O\_{s} \rightarrow CH\_{3}CH\_{2}CH\_{2}COOH + 2CO\_{2} + 2H\_{2} \tag{2}$$

$$
\Delta \text{G}\_{\text{o}}^{\cdot} = -254.0 \text{ kJ/mol}
$$

$$
\text{C}\_{\text{e}} \text{H}\_{\text{12}} \text{O}\_{\text{e}} + 2 \text{H}\_{\text{2}} \rightarrow 2 \text{CH}\_{3} \text{CH}\_{2} \text{COOH} + 2 \text{H}\_{2} \text{O} \tag{3}
$$

$$
\Delta \mathbf{G}\_{\mathrm{o}}^{\cdot} = -279.4 \text{ kJ/mol}
$$

pH affects the physiological conditions of the bacterial growth, metabolism and ions transport. Optimizing the pH is considering a key factor influenced the redox environment and the direction of electron flow toward the hydrogen formation. The experimental reports demonstrated that the optimal pH for the bacterial growth does not result in the elevated hydrogen production performances [3]. For the dark fermentative hydrogen fermentation, the optimal pH for efficient hydrogen production lied between 5.5 and 6.5 for various wastewaters and pure substrates [18]. In addition, the acidic pH induces the pyruvate transformation to volatile fatty acids (VFA) with concomitant hydrogen production, whereas the neutral pH facilitated the methanogenic pathway. Maintaining the acidogenic (5.5–6.5) pH is essential for controlling the methanogenic populations and efficient hydrogen production.

#### **3.3. Nutrients**

ment of the sporulating hydrogen‐producing bacteria such as *Clostridium* sp. The acid‐shock [12] and base‐shock [13] pretreatments suppress the methanogenic activity by the narrow selective growth pH range of the methanogenes (6–7.5), whereas the *Clostridium* populations survive in the harsh condition due to the spore‐forming capability. The chemical shock methods such as chloroform [14] and 2‐bromoethanesulfonic acid (BESA) [15] have a complex structure, analog to the methanogenic coenzyme, and it acts as a inhibitor for the methanogenes. This method facilitated the suppression of the methanogenes, whereas the other non‐spore‐forming hydrogen producers such as *Enterobacter* sp. can also survive with the presence of *Clostridium* sp., thus enhancing the substrate utilization and hydrogen yield. The load‐shock [16] treatment is directed by the exposure of the inoculum to a higher substrate concentration, and it leads to the surge in the pH with an accumulation of organic acids and

Ren et al. [17] demonstrated that application of various pretreatment methods, such as acid, alkaline, heat‐shock and repeated aeration, can greatly affect the metabolic pathway and the microbial community distribution pattern. The dominant butyric acid‐mediated hydrogen metabolism was observed with heat‐shock and alkaline treatment, and mixed‐type fermentation pathway was observed with the acid pretreatment, whereas the ethanol‐type pathway was observed with repeated aeration treatment with a maximum hydrogen yield of 1.96 mol/ mol glucose. The microbial community characterized by denaturing gradient gel electrophoresis (DGGE) revealed that the changes in the composition of the microbial dynamics affect the hydrogen yield. The strain *Ethanoligenens harbinens* was detected under repeated aeration condition with an ethanol‐mediated pathway, and the hydrogen‐consuming propionic acid bacterium *Propionibacterium propionicus* was detected in acid treatment with low hydrogen productivity. The heat‐shock‐mediated mixed culture was dominated with *Clostridium* sp. which represents the butyric‐acid‐type metabolic pathway. Based on the evidence, the appropriate pretreatment method is essential for enriching the hydrogen‐producing bacterial popu-

pH is the key driven parameter affecting the cellular metabolism of hydrogen‐producing bacterial populations, since the prevalent end products of the bacterial metabolism vary with the changes in the medium pH. Based on the pH and the major end products formation, three metabolic pathways have been proposed (a) ethanol type (EtOH) (Eq. 1), (b) butyric type (HBu) (Eq. 2) and (c) propionic type (HPr) (Eq. 3). The former, HBu type, involved in the hydrogen‐generating reactions, whereas the latter, HPr type, involved in the hydrogen‐scavenging reactions. Hence, the elimination of the propionate formation is an essential step for

 O → 2 CH3 CH2

*Δ G*<sup>0</sup> = − 235.0 kJ/mol

 OH + 2 HCO3

<sup>−</sup> + 2 H2 (1)

inhibits the methanogenic populations.

202 Biological Wastewater Treatment and Resource Recovery

lations and enhanced hydrogen production.

the enhancement of hydrogen production.

C<sup>6</sup> H12 O<sup>6</sup> + 2 H2

**3.2. pH**

The inorganic nutrient supplements, such as nitrogen (N), phosphorus (P) and iron (Fe), along with carbon (C) source, are important for microbial growth and improvement in the hydrogen production. The nutrient at proper concentration is beneficial for hydrogen production. For instance, Lin and Lay [19] explained that at a carbon/nitrogen (C/N) ratio of 47, the hydrogen yield from sucrose was 1.9 times higher than the control with a value of 4.8 mol/ mol substrate. In a pure culture thermotolerant *Kelbsiella* sp., the maximum hydrogen yield of 0.28 mol/mol glycerol was observed with 11.21 g/L glycerol, 2.84 g/L KH2 PO<sup>4</sup> and 5.66 g/L NH<sup>4</sup> Cl, respectively [20]. Wang et al. [21] mentioned that the hydrogen production efficiency of glucose (313.3 mL/g glucose) was improved with low supplementation of nitrate 0.1 g/L; however, increased concentration of nitrate over 0.1 g/L significantly affected the hydrogen yield and the substrate consumption rate. The drop in hydrogen production is attributed by the inhibition of nitrogenize activity by surplus ammonium ions [22, 23]. The iron (Fe) is an important element essential for the hydrogenase activity, which directs the metabolic pathway by stimulating the active site for the ferredoxin (Fd). The addition of iron supplement was shown to improve the hydrogen production. Gadhe et al. [24] demonstrated the effects of nano‐sized iron and nickel oxide nanoparticles by using dairy wastewater as a substrate, and it showed that an enhancement in hydrogen yield of 17.2 mmol/g COD is due to the enhanced activity of the ferredoxin oxidoreductase, ferredoxin and hydrogenase enzymes. Moreover, the optimal value for the Fe2+ concentration is varied with the type of substrates used. For instance, the optimal concentration reported by Liu and Shen [25] was 10 mg/L from starch, whereas palm oil mill effluent showed an optimal value of 257 mg/L [26].

#### **3.4. Hydraulic retention times**

The hydraulic retention time (HRT) is one of the key process control parameters influencing the continuous hydrogen production. HRT enables the better process control of the microorganisms that can regulate the metabolic pathway favorable for efficient hydrogen production. The long HRT permits the growth of hydrogen consumers mainly archaea, which is unsuitable for hydrogen production, whereas too low HRT leads to the washout of active biomass and deterioration of the reactor performances. The optimization of HRT is a paramount importance for the scale‐up, long‐term and sustainable hydrogen production. HRT controls the organic loading rate (OLR), substrate degradation and reaction kinetics. The organic wastes required long HRT, whereas the simple organics required short HRT [2]. The reported optimum HRT value for the wastewater ranges from 0.5 to 24 h. For example, the short HRT (0.5 h) provided the maximum hydrogen production rate of 14 L/L‐d from condensed soluble wastewater [27], whereas the long HRT (24 h) is required for efficient conversion of olive mill wastewater with a HPR of 7.0 L/L‐d [28]. The process parameters discussed above significantly influenced the hydrogen production; hence, careful assessment of each individual factor is important for stable hydrogen production.

## **4. Bioreactor design considerations for continuous hydrogen production**

Bioreactor configuration is a notable factor in dark fermentative hydrogen production, as it influences the contact between the organic waste and hydrogen producers, substrate utilization, biomass dilution rate, etc. According to the feeding regime, the biohydrogen production can be conducted in batch, semi‐continuous and continuous mode (**Table 4**). The batch mode operation is relatively simple and easier to control. Hence, the batch mode hydrogen reactors have been widely used to determine the feasibility of the organic waste feedstock and to optimize the environmental parameters such as pH, temperature, substrate concentration. In semi‐continuous mode operation, the organic substrate was operated in a sequencing batch which includes feeding, reaction, settle and decant stages [29]. The sequencing batch operation is recommended for a viscous substrate like a POME and solid organic biomass like food waste and lignocellulosic biomass, where the physical contact between the substrate and microorganisms is limited, and this reactor mode operation enables the better hydrolysis rate, avoids clogging in the pipes and retains the effective biomass concentration. In continuous mode operation, the continuous supply of nutrients and the removal of the pollutants occur simultaneously with the aid of peristaltic pumps.

Although various reactor models assessed, the continuous mode operation is preferred for bench‐scale and commercial‐scale applications. The widely investigated model for continuous mode operation is the CSTR type, wherein the substrates and feedstocks are well mixed inside the reactor with the aid of the mechanical rotor; however, the biomass washout usually occurred at lower HRT [27, 30]. In some cases, the biofilm formed inside the CSTR is resistance to the biomass washout and thereby enhancing the hydrogen production performances. Chu et al. [27] investigated the CSTR reactor model by using condensed soluble molasses as a substrate with suspended and immobilized cells as inoculum source. The hydrogen production


WW, wastewater; ASBR, anaerobic sequencing batch reactor; CSTR, continuously stirred tank reactor; MBR, membrane bioreactor; PBR, packed bed reactor; ICBR, immobilized cell bioreactor.

**Table 4.** Bioreactor types used in hydrogen production.

from immobilized cell was relatively lower with a maximum HPR of 7.6 L/L/d; however, the suspended cells operation provided the maxim HPR of 14.04 L/L/d, respectively. The observed variation is attributed by the washout of the active biomass in immobilized cells system (9.8 g volatile suspended solids (VSS)/L), poor mass transfer between the microbes and substrates and the increased lactic acid formation. On the other hand, the suspended cell system formed a hydrogen‐producing granule (HPG) inside the reactor, and thus, it retains the active biomass (12.30 g VSS/L) and less formation of the lactic acid. Sivagurunathan et al. [30] demonstrated that the hydrogen production from ICBR [31] was higher (55 L/L/d) than the suspended cells CSTR (37.56 L/L/d) operation. The superior performance of the ICBR is due to the formation of granular biomass at short HRT of 3 h with the presence of *Selenomonas* sp. and further maturation of granules with the presence of active hydrogen‐producing *Clostridium* Sp. The *Selenomonas* sp. act as a bio‐glue for the development of granules. Moreover, the energy content analysis of the beverage wastewater with immobilized cells system analysis showed that it has the capability of reducing the CO2 reduction efficiency of 2832 ton CO2 equivalent/year.

## **5. Conclusion**

**3.4. Hydraulic retention times**

204 Biological Wastewater Treatment and Resource Recovery

factor is important for stable hydrogen production.

simultaneously with the aid of peristaltic pumps.

The hydraulic retention time (HRT) is one of the key process control parameters influencing the continuous hydrogen production. HRT enables the better process control of the microorganisms that can regulate the metabolic pathway favorable for efficient hydrogen production. The long HRT permits the growth of hydrogen consumers mainly archaea, which is unsuitable for hydrogen production, whereas too low HRT leads to the washout of active biomass and deterioration of the reactor performances. The optimization of HRT is a paramount importance for the scale‐up, long‐term and sustainable hydrogen production. HRT controls the organic loading rate (OLR), substrate degradation and reaction kinetics. The organic wastes required long HRT, whereas the simple organics required short HRT [2]. The reported optimum HRT value for the wastewater ranges from 0.5 to 24 h. For example, the short HRT (0.5 h) provided the maximum hydrogen production rate of 14 L/L‐d from condensed soluble wastewater [27], whereas the long HRT (24 h) is required for efficient conversion of olive mill wastewater with a HPR of 7.0 L/L‐d [28]. The process parameters discussed above significantly influenced the hydrogen production; hence, careful assessment of each individual

**4. Bioreactor design considerations for continuous hydrogen production**

Bioreactor configuration is a notable factor in dark fermentative hydrogen production, as it influences the contact between the organic waste and hydrogen producers, substrate utilization, biomass dilution rate, etc. According to the feeding regime, the biohydrogen production can be conducted in batch, semi‐continuous and continuous mode (**Table 4**). The batch mode operation is relatively simple and easier to control. Hence, the batch mode hydrogen reactors have been widely used to determine the feasibility of the organic waste feedstock and to optimize the environmental parameters such as pH, temperature, substrate concentration. In semi‐continuous mode operation, the organic substrate was operated in a sequencing batch which includes feeding, reaction, settle and decant stages [29]. The sequencing batch operation is recommended for a viscous substrate like a POME and solid organic biomass like food waste and lignocellulosic biomass, where the physical contact between the substrate and microorganisms is limited, and this reactor mode operation enables the better hydrolysis rate, avoids clogging in the pipes and retains the effective biomass concentration. In continuous mode operation, the continuous supply of nutrients and the removal of the pollutants occur

Although various reactor models assessed, the continuous mode operation is preferred for bench‐scale and commercial‐scale applications. The widely investigated model for continuous mode operation is the CSTR type, wherein the substrates and feedstocks are well mixed inside the reactor with the aid of the mechanical rotor; however, the biomass washout usually occurred at lower HRT [27, 30]. In some cases, the biofilm formed inside the CSTR is resistance to the biomass washout and thereby enhancing the hydrogen production performances. Chu et al. [27] investigated the CSTR reactor model by using condensed soluble molasses as a substrate with suspended and immobilized cells as inoculum source. The hydrogen production

Biohydrogen production from industrial wastewaters seems to be appropriate and environmental benign option for future sustainable hydrogen economy with simultaneous energy recovery and waste disposal. Various studies revealed the hydrogen production potential of wastewaters. Among them, sugar‐rich wastewaters are the promising substrate for high‐efficient hydrogen production rates and yields, due to their easier degradation rate and higher substrate concentration. Other key challenges that rely on dark fermentative hydrogen production from organic wastes are the low substrate conversion efficiency, moderate‐to‐low hydrogen yield and residual organics in the effluents. In general, biohydrogen production is a primary step for wastewater treatment, in which a maximum 4 mol/mol glucose representing 33% of COD removal efficiency; nearly 70–80% of the residual organics remain untreated with the hydrogen‐producing effluent, thus seeks further disposal of the effluent in the wastewater streams. The post‐residual effluent has to be integrated with various two‐step processes, such as methane production, photofermentation, microbial electrolysis cells, bioplastics production and microalgae cultivation, for maximizing the energy recovery.

The author Dr. Periyasamy Sivagurunathan greatly acknowledged the financial assistance from Japan Society for Promotion of Science: JSPS ID: P15370 for this study.

## **Author details**

Periyasamy Sivagurunathan1 \*, Gopalakrishnan Kumar1 , Arivalagan Pugazhendhi2 , Guangyin Zhen<sup>3</sup> , Takuro Kobayashi1 and Kaiqin Xu1, 4\*

\*Address all correspondence to: contact2sivas12@gmail.com and joexu@nies.go.jp

1 Center for Materials Cycles and Waste Management Research, National Institute for Environmental Studies, Tsukuba, Japan

2 Department of Environmental Engineering, Daegu University, Gyeongsan, Gyeongbuk, Republic of Korea

3 School of Ecological and Environmental Sciences, East China Normal University, Shanghai, China

4 School of Environmental Science and Engineering, Shanghai Jiao Tong University, Shanghai, China

## **References**


[5] Chookaew T, O‐Thong S, Prasertsan P. Biohydrogen production from crude glycerol by immobilized *Klebsiella* sp. TR17 in a UASB reactor and bacterial quantification under non‐sterile conditions. International Journal of Hydrogen Energy. 2014;39:9580–7.

hydrogen yield and residual organics in the effluents. In general, biohydrogen production is a primary step for wastewater treatment, in which a maximum 4 mol/mol glucose representing 33% of COD removal efficiency; nearly 70–80% of the residual organics remain untreated with the hydrogen‐producing effluent, thus seeks further disposal of the effluent in the wastewater streams. The post‐residual effluent has to be integrated with various two‐step processes, such as methane production, photofermentation, microbial electrolysis cells, bioplastics produc-

The author Dr. Periyasamy Sivagurunathan greatly acknowledged the financial assistance

1 Center for Materials Cycles and Waste Management Research, National Institute for

2 Department of Environmental Engineering, Daegu University, Gyeongsan, Gyeongbuk,

3 School of Ecological and Environmental Sciences, East China Normal University, Shanghai,

4 School of Environmental Science and Engineering, Shanghai Jiao Tong University, Shanghai,

[1] Singh L, Wahid ZA. Methods for enhancing bio‐hydrogen production from biological process: A review. Journal of Industrial and Engineering Chemistry. 2015;21:70–80. [2] Sivagurunathan P, Kumar G, Bakonyi P, Kim S‐H, Kobayashi T, Xu KQ, et al. A critical review on issues and overcoming strategies for the enhancement of dark fermentative hydrogen production in continuous systems. International Journal of Hydrogen Energy.

[3] Ghimire A, Frunzo L, Pirozzi F, Trably E, Escudie R, Lens PNL, et al. A review on dark fermentative biohydrogen production from organic biomass: Process parameters and

[4] Elsharnouby O, Hafez H, Nakhla G, El Naggar MH. A critical literature review on biohydrogen production by pure cultures. International Journal of Hydrogen Energy.

use of by‐products. Applied Energy. 2015;144:73–95.

, Arivalagan Pugazhendhi2

, Guangyin

tion and microalgae cultivation, for maximizing the energy recovery.

and Kaiqin Xu1, 4\*

**Author details**

Republic of Korea

Zhen<sup>3</sup>

China

China

**References**

2016;41:3820–36.

2013;38:4945–66.

Periyasamy Sivagurunathan1

, Takuro Kobayashi1

Environmental Studies, Tsukuba, Japan

206 Biological Wastewater Treatment and Resource Recovery

from Japan Society for Promotion of Science: JSPS ID: P15370 for this study.

\*, Gopalakrishnan Kumar1

\*Address all correspondence to: contact2sivas12@gmail.com and joexu@nies.go.jp


[30] Sivagurunathan P, Lin C‐Y. Enhanced biohydrogen production from beverage wastewater: process performance during various hydraulic retention times and their microbial insights. RSC Advances. 2016;6:4160–9.

[17] Ren N‐Q, Guo W‐Q, Wang X‐J, Xiang W‐S, Liu B‐F, Wang X‐Z, et al. Effects of different pretreatment methods on fermentation types and dominant bacteria for hydrogen pro-

[18] Lin C‐Y, Lay C‐H, Sen B, Chu C‐Y, Kumar G, Chen C‐C, et al. Fermentative hydrogen production from wastewaters: A review and prognosis. International Journal of

[19] Lin CY, Lay CH. Effects of carbonate and phosphate concentrations on hydrogen production using anaerobic sewage sludge microflora. International Journal of Hydrogen

[20] Chookaew T, O‐Thong S, Prasertsan P. Statistical optimization of medium components affecting simultaneous fermentative hydrogen and ethanol production from crude glycerol by thermotolerant *Klebsiella* sp. TR17. International Journal of Hydrogen Energy.

[21] Wang B, Wan W, Wang J. Effects of nitrate concentration on biological hydrogen production by mixed cultures. Frontiers of Environmental Science & Engineering in China.

[22] Wang B, Wan W, Wang J. Effect of ammonia concentration on fermentative hydrogen

[23] Redwood MD, Macaskie LE. A two‐stage, two‐organism process for biohydrogen from

[24] Gadhe A, Sonawane SS, Varma MN. Enhancement effect of hematite and nickel nanoparticles on biohydrogen production from dairy wastewater. International Journal

[25] Liu G, Shen J. Effects of culture and medium conditions on hydrogen production from starch using anaerobic bacteria. Journal of Bioscience and Bioengineering. 2004;98:251–6.

[26] O‐Thong S, Prasertsan P, Intrasungkha N, Dhamwichukorn S, Birkeland N‐K. Optimization of simultaneous thermophilic fermentative hydrogen production and COD reduction from palm oil mill effluent by Thermoanaerobacterium‐rich sludge.

[27] Chu CY, Wu SY, Hsieh PC, Lin CY. Biohydrogen production from immobilized cells and suspended sludge systems with condensed molasses fermentation solubles.

[28] Scoma A, Bertin L, Fava F. Effect of hydraulic retention time on biohydrogen and volatile fatty acids production during acidogenic digestion of dephenolized olive mill wastewa-

[29] Badiei M, Jahim JM, Anuar N, Sheikh Abdullah SR. Effect of hydraulic retention time on biohydrogen production from palm oil mill effluent in anaerobic sequencing batch reac-

production by mixed cultures. Bioresource Technology. 2009;100:1211–3.

glucose. International Journal of Hydrogen Energy. 2006;31:1514–21.

International Journal of Hydrogen Energy. 2008;33:1221–31.

International Journal of Hydrogen Energy. 2011;36:14078–85.

tor. International Journal of Hydrogen Energy. 2011;36:5912–9.

duction. International Journal of Hydrogen Energy. 2008;33:4318–24.

Hydrogen Energy. 2012;37:15632–42.

208 Biological Wastewater Treatment and Resource Recovery

of Hydrogen Energy. 2015;40:4502–11.

ters. Biomass and Bioenergy. 2013;48:51–8.

Energy. 2004;29:275–81.

2014;39:751–60.

2009;3:380–6.


**Provisional chapter**

## **Valorization of Glucose-Based Wastewater Through Production of Hydrogen, Volatile Fatty Acids and Alcohols Valorization of Glucose-Based Wastewater Through Production of Hydrogen, Volatile Fatty Acids and Alcohols**

Eduardo Lucena Cavalcante de Amorim, Leandro Takano Sader, Lucas Rodrigues Ramos and Edson Luiz Silva Leandro Takano Sader, Lucas Rodrigues Ramos and Edson Luiz Silva Additional information is available at the end of the chapter

Additional information is available at the end of the chapter

Eduardo Lucena Cavalcante de Amorim,

http://dx.doi.org/10.5772/67101

#### **Abstract**

[43] Sivaramakrishna D, Sreekanth D, Sivaramakrishnan M, Sathish Kumar B, Himabindu V, Narasu ML. Effect of system optimizing conditions on biohydrogen production from herbal wastewater by slaughterhouse sludge. International Journal of Hydrogen Energy.

[44] Wu J‐H, Lin C‐Y. Biohydrogen production by mesophilic fermentation of food wastewa-

[45] Shi X‐Y, Jin D‐W, Sun Q‐Y, Li W‐W. Optimization of conditions for hydrogen production from brewery wastewater by anaerobic sludge using desirability function approach.

[46] Fernandes BS, Peixoto G, Albrecht FR, Saavedra del Aguila NK, Zaiat M. Potential to produce biohydrogen from various wastewaters. Energy for Sustainable Development.

[47] Yin Y, Hu J, Wang J. Enriching hydrogen‐producing bacteria from digested sludge by different pretreatment methods. International Journal of Hydrogen Energy.

[48] Kim MS, Lee DY, Kim DH. Continuous hydrogen production from tofu processing waste using anaerobic mixed microflora under thermophilic conditions. International

[49] Kongjan P, O‐Thong S, Angelidaki I. Hydrogen and methane production from desugared molasses using a two‐stage thermophilic anaerobic process. Engineering in Life

[50] Chu C‐Y, Hastuti ZD, Dewi EL, Purwanto WW, Priyanto U. Enhancing strategy on renewable hydrogen production in a continuous bioreactor with packed biofilter from

sugary wastewater. International Journal of Hydrogen Energy. 2016;41:4404–12.

ter. Water Science and Technology. 2004;49:223–8.

Journal of Hydrogen Energy. 2011;36:8712–8.

Renewable Energy. 2010;35:1493–8.

2014;39:7526–33.

210 Biological Wastewater Treatment and Resource Recovery

2010;14:143–8.

2014;39:13550–6.

Sciences. 2013;13:118–25.

The production of hydrogen in an anaerobic fluidized bed reactor (AFBR) was evaluated under different organic loading rates (OLRs) with the addition of 1 g L−1 sodium bicarbonate for pH control. Expanded clay was used as the support material for microbial attachment. Two AFBRs were operated with glucose concentrations of 10 and 25 g L−1 and a hydraulic retention time (HRT) decreasing from 8 to 1 h at a controlled temperature of 30°C. A linear correlation was observed between the hydrogen production rate (HPR) and the OLR, except for the reactor operated with 25 g L−1 glucose. The maximum HPR of 1.58 L h−1 L−1 was obtained with an HRT of 1 h, and the maximum H2 yield of 1.32 mol H2 mol−1 glucose was obtained with an HRT of 2 h, in the reactor operated with 10 g L−1 glucose.

**Keywords:** hydrogen production, anaerobic fluidized bed reactor, substrate concentration, hydraulic retention time, organic loading rate

## **1. Introduction**

The acidogenic fermentation of wastewater or biowaste for H2 production has attracted great global interest because it is a cheap and simple technology that produces clean energy from renewable sources while reducing pollutants [1, 2].

According to Reddy et al. [3], one of the major drawbacks of using organic wastes is that only 30–40% of the substrate is used to H2 production and 60–70% is converted to several other metabolites. However, some metabolites are commercially attractive, such as acetic

© 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, 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.

acid, butyric acid, propionic acid, lactic acid, succinic acid, 1,3-propanediol, ethanol, methanol, etc. [4, 5].

H2 production has been carried out with a variety of organic wastes, in which the source of carbonaceous organic material is based on glucose, sucrose, starch, xylose, cheese-processing wastewater, tapioca-processing wastewater, and sugarcane vinasse [6–9].

The fermentation process for the production of H2 in anaerobic reactors is greatly influenced by several factors, such as the type of wastewater, the inoculum, the type of reactor, the nutritional requirements, the temperature, and the pH [10–12].

For practical engineering, industrial H2 production requires continuous or semicontinuous production processes. Several types of reactors have been studied to effectively generate H2 . Reactors for continuous H2 production include suspended biomass reactors, e.g., continuous stirred tank reactors (CSTRs) [13–15] and anaerobic sequencing bed reactors (ASBRs) [16], and biofilm reactors such as anaerobic packed bed reactors (APBRs) [17] and anaerobic fluidized bed reactors (AFBRs) [6–9, 18]. The advantages and disadvantages of different reactor types vary. Biofilm reactors can overcome the drawbacks of suspended biomass reactors by decoupling the biomass retention time from HRT, thus increasing the biomass concentration in the reactor. The hydraulic mixing regime is usually more turbulent in AFBRs than in APBRs, which improves mass transfer and treatment efficiencies because bed fluidization favors contact between the biofilm and substrate [19–21].

Hydrogen production is a microbial-mediated process dependent on several parameters that can affect the performance. Some of these are the inoculum source, pH, substrate concentration, accessible nutrients, HRT, and temperature [11, 21]. Their control in appropriate range can enrich the microbial community with hydrogen producers, eliminate hydrogen consumers, shift the metabolism to favor hydrogen production, increase substrate conversion efficiency, and increase the overall process potential [1, 10, 11, 21]. The organic loading rate (OLR; influent substrate concentration/HRT) is a parameter that evaluates the simultaneous effects of influent substrate concentrations and HRTs when synthetic or real wastewaters are used to produce H<sup>2</sup> in anaerobic reactors [13–18, 22–26]. Previous studies in our research group observed hydrogen production with glucose concentrations of 2000 mg L−1 [27–29], 4000 mg L−1 [6, 30] and 5000 mg L−1 [31]. Increasing glucose concentration to 10 g L−1 and 25 g L−1 can determine the range where hydrogen-producing acidogenesis shifts to solventogenesis. Therefore, the present study examines the effect of both OLR and alkalinity supplementation on H<sup>2</sup> production in AFBRs with influent glucose concentrations of 10 g L−1 (OLRs of 30–240 kg COD m−3 day−1) and 25 g L−1 (OLRs of 75–600 kg COD m−3 day−1).

## **2. Materials and methods**

#### **2.1. Anaerobic fluidized bed reactors and feed composition**

A schematic diagram of the two identical jacketed AFBRs used in this study is presented in **Figure 1**. The reactors were constructed with a transparent acrylic tube, within 5.3 cm of internal diameter and 190 cm of height, and filled with expanded clay (diameter = 2.8–3.3 mm, density = 1.5 g cm−3). Each AFBR was equipped with a water jacket that recirculated heated water from a thermostatic bath to maintain the temperature at 30°C. The AFBRs were fed with synthetic wastewater containing glucose (10 and 25 g L−1) as the main carbon source supplemented with the following nutrients: SeO2 , 0.07 mg L−1; CoCl2 ·2H2 O, 0.08 mg L−1; FeCl3 ·6H<sup>2</sup> O, 0.5 mg L−1; NiSO4 ·6H<sup>2</sup> O, 1 mg L−1; FeSO4 ·7H2 O, 5 mg L−1; K2 HPO4 , 21.7 mg L−1; Na2 HPO4 ·2H2 O, 33.4 mg L−1: CaCl2 ·6H<sup>2</sup> O, 47 mg L−1; KH2 PO4 , 85 mg L−1; and CO(NH2 ) 2 N2 O, 125 mg L−1. In order to control the pH of the reactors at 5.0–5.5, hydrochloric acid (10 M) and sodium bicarbonate (1 g L−1) were also used [6].

**Figure 1.** Schematic description of the AFBR.

acid, butyric acid, propionic acid, lactic acid, succinic acid, 1,3-propanediol, ethanol, metha-

by several factors, such as the type of wastewater, the inoculum, the type of reactor, the nutri-

production processes. Several types of reactors have been studied to effectively generate H2

stirred tank reactors (CSTRs) [13–15] and anaerobic sequencing bed reactors (ASBRs) [16], and biofilm reactors such as anaerobic packed bed reactors (APBRs) [17] and anaerobic fluidized bed reactors (AFBRs) [6–9, 18]. The advantages and disadvantages of different reactor types vary. Biofilm reactors can overcome the drawbacks of suspended biomass reactors by decoupling the biomass retention time from HRT, thus increasing the biomass concentration in the reactor. The hydraulic mixing regime is usually more turbulent in AFBRs than in APBRs, which improves mass transfer and treatment efficiencies because bed fluidization

Hydrogen production is a microbial-mediated process dependent on several parameters that can affect the performance. Some of these are the inoculum source, pH, substrate concentration, accessible nutrients, HRT, and temperature [11, 21]. Their control in appropriate range can enrich the microbial community with hydrogen producers, eliminate hydrogen consumers, shift the metabolism to favor hydrogen production, increase substrate conversion efficiency, and increase the overall process potential [1, 10, 11, 21]. The organic loading rate (OLR; influent substrate concentration/HRT) is a parameter that evaluates the simultaneous effects of influent substrate concentrations and HRTs when synthetic or real wastewaters are used

observed hydrogen production with glucose concentrations of 2000 mg L−1 [27–29], 4000 mg L−1 [6, 30] and 5000 mg L−1 [31]. Increasing glucose concentration to 10 g L−1 and 25 g L−1 can determine the range where hydrogen-producing acidogenesis shifts to solventogenesis. Therefore, the present study examines the effect of both OLR and alkalinity supplementation

A schematic diagram of the two identical jacketed AFBRs used in this study is presented in **Figure 1**. The reactors were constructed with a transparent acrylic tube, within 5.3 cm of

production in AFBRs with influent glucose concentrations of 10 g L−1 (OLRs of 30–240

in anaerobic reactors [13–18, 22–26]. Previous studies in our research group

wastewater, tapioca-processing wastewater, and sugarcane vinasse [6–9].

The fermentation process for the production of H2

For practical engineering, industrial H2

212 Biological Wastewater Treatment and Resource Recovery

Reactors for continuous H2

to produce H<sup>2</sup>

**2. Materials and methods**

on H<sup>2</sup>

tional requirements, the temperature, and the pH [10–12].

favors contact between the biofilm and substrate [19–21].

kg COD m−3 day−1) and 25 g L−1 (OLRs of 75–600 kg COD m−3 day−1).

**2.1. Anaerobic fluidized bed reactors and feed composition**

 production has been carried out with a variety of organic wastes, in which the source of carbonaceous organic material is based on glucose, sucrose, starch, xylose, cheese-processing

in anaerobic reactors is greatly influenced

.

production requires continuous or semicontinuous

production include suspended biomass reactors, e.g., continuous

nol, etc. [4, 5].

H2

#### **2.2. Heat treatment of inoculum, AFBR setup and operation conditions**

The AFBRs were inoculated with sludge from an upflow anaerobic sludge blanket (UASB) reactor treating swine wastewater effluent. The sludge was heat treated for 10 min at 90°C according to the methodology of Kim et al. [25] in order to eliminate hydrogen consumers and select for endospore producers. The reactors were inoculated at a rate of 10% of the sludge feed volume.

The total liquid flow rate into the AFBRs was maintained at 128 L h−1 (expansion = 30%). This flow rate produced a superficial velocity 1.30 times greater than the minimum fluidization velocity. At first, in order to activate the H2 -producing biomass, the two AFBRs were operated in batch mode for 48 h while periodically recording the substrate consumption by microorganisms. When the activation period was over, the reactors were operated in continuous mode with an HRT of 8 h, which was then decreased stepwise to 6 h, 4 h, 2 h, and 1 h. The composition of the gaseous products (H2 and CO2 ) and soluble metabolites (volatile organic acids and alcohols) produced during fermentative H2 production was monitored as a function of time.

To facilitate discussion of the results and to identify the reactors, each reactor was named according to the influent glucose concentration: the reactor operated with 10 g L−1 glucose was named "R10," and the reactor operated with 25 g L−1 glucose was named "R25."

#### **2.3. Chemical analyses**

The GOD-PAP enzymatic method [32] was used to determine the glucose concentrations. Total solids (TS), volatile suspended solids (VSS), total volatile solids (TVS), and chemical oxygen demand (COD) analyses were performed according to Standard Methods for the Examination of Water and Wastewater [33].

A gas chromatograph (GC-2010, Shimadzu, Tokyo, Japan) equipped with a thermal conductivity detector (TCD) was used to determine the biogas composition. Argon was used as the carrier gas with a Carboxen 1010 PLOT column (30 m long × 0.53 mm internal diameter). A gas chromatograph (GC-2010, Shimadzu, Tokyo, Japan) equipped with a flame ionization detector (FID) was used to determine volatile organic acids and alcohols. The GC used a COMBI-PAL headspace sample introduction system (AOC 5000 model) and HP-INNOWAX column (30 m long × 0.25 mm internal diameter × 0.25 mm film thickness) [32].

A gas meter (type TG1; Ritter Inc., Germany) was used to measure the amount of H2 generated.

## **3. Results and discussion**

#### **3.1. Effect of OLR on H<sup>2</sup> production**

**Figure 2** presents the variation in pH effluent as a function of OLR for the two AFBRs used in this study. The pH remained stable throughout the system operation within the operating range of acidogenic anaerobic systems, i.e., between 3.7 in Barros et al. [6], 3.4 and 3.6 in R10, and 3.3 and 3.5 in R25. The influent pH remained between 5.2 and 5.9 in Barros et al. [6], 4.8 and 5.6 in R10, and 5.5 and 5.9 in R25 (**Figure 2**).

**Figure 3** presents the variation in glucose conversion as a function of OLR for the AFBRs used in this study. To estimate glucose consumption during fermentation, glucose levels were measured in the fermentation medium (**Figure 3**). Glucose consumption by microorganisms was recorded at all OLR intervals in both AFBRs. The data indicate that glucose conversion decreased with the increase of OLR at all concentrations. For reactor R10, when OLR was increased from 30–120 kg COD m−3 day−1, glucose conversion decreased from 57 to 36%, but when OLR increased to 240 kg COD m−3 day−1, glucose conversion increased to 41%. For reactor R25, when OLR increased from 75 to 600 kg COD m−3 day−1, glucose conversion decreased from 36 to 20%.

 **Figure 2.** pH effluent as a function of the OLR for the AFBRs.

The total liquid flow rate into the AFBRs was maintained at 128 L h−1 (expansion = 30%). This flow rate produced a superficial velocity 1.30 times greater than the minimum fluidization

in batch mode for 48 h while periodically recording the substrate consumption by microorganisms. When the activation period was over, the reactors were operated in continuous mode with an HRT of 8 h, which was then decreased stepwise to 6 h, 4 h, 2 h, and 1 h. The

and CO2

To facilitate discussion of the results and to identify the reactors, each reactor was named according to the influent glucose concentration: the reactor operated with 10 g L−1 glucose was

The GOD-PAP enzymatic method [32] was used to determine the glucose concentrations. Total solids (TS), volatile suspended solids (VSS), total volatile solids (TVS), and chemical oxygen demand (COD) analyses were performed according to Standard Methods for the

A gas chromatograph (GC-2010, Shimadzu, Tokyo, Japan) equipped with a thermal conductivity detector (TCD) was used to determine the biogas composition. Argon was used as the carrier gas with a Carboxen 1010 PLOT column (30 m long × 0.53 mm internal diameter). A gas chromatograph (GC-2010, Shimadzu, Tokyo, Japan) equipped with a flame ionization detector (FID) was used to determine volatile organic acids and alcohols. The GC used a COMBI-PAL headspace sample introduction system (AOC 5000 model) and HP-INNOWAX column (30 m long × 0.25 mm internal diameter × 0.25 mm film thick-

**Figure 2** presents the variation in pH effluent as a function of OLR for the two AFBRs used in this study. The pH remained stable throughout the system operation within the operating range of acidogenic anaerobic systems, i.e., between 3.7 in Barros et al. [6], 3.4 and 3.6 in R10, and 3.3 and 3.5 in R25. The influent pH remained between 5.2 and 5.9 in Barros et al. [6],

**Figure 3** presents the variation in glucose conversion as a function of OLR for the AFBRs used in this study. To estimate glucose consumption during fermentation, glucose levels were measured in the fermentation medium (**Figure 3**). Glucose consumption by microorganisms was recorded

A gas meter (type TG1; Ritter Inc., Germany) was used to measure the amount of H2

 **production**

4.8 and 5.6 in R10, and 5.5 and 5.9 in R25 (**Figure 2**).

named "R10," and the reactor operated with 25 g L−1 glucose was named "R25."


) and soluble metabolites (volatile organic

production was monitored as a function

generated.

velocity. At first, in order to activate the H2

214 Biological Wastewater Treatment and Resource Recovery

composition of the gaseous products (H2

Examination of Water and Wastewater [33].

of time.

ness) [32].

**3. Results and discussion**

**3.1. Effect of OLR on H<sup>2</sup>**

**2.3. Chemical analyses**

acids and alcohols) produced during fermentative H2

 **Figure 3.** Glucose conversion as a function of the OLR for the AFBRs.

**Figure 4** presents the variation in the hydrogen production rate (HPR) as a function of OLR for the two AFBRs used in this study. Similar to the results of Barros et al. [6] for an AFBR with expanded clay as the support material, an influent glucose concentration of 4 g L−1, and alkalinity supplementation (values presented in **Figure 2**), the HPR values for R10 increased linearly from 0.12 to 1.58 L h−1 L−1 when OLR increased from 30 to 240 kg COD m−3. By contrast, a linear relationship between HPR and OLR was not observed in R25 for OLR ranging from 75 to 600 kg COD m−3. The maximum HPR values were 1.58 and 0.84 L h−1 L−1 for reactors R10 and R25, respectively.

 **Figure 4.** HPR as a function of the OLR for the AFBRs.

**Figure 5** presents the variation in HY as a function of OLR for the two AFBRs used in this study. The HY values increased with increasing OLR in both reactors. For reactor R10, when OLR was increased from 30 to 120 kg COD m−3 day−1, HY increased significantly from 0.48 to 1.32 mol H2 mol−1 glucose, but when OLR increased to 240 kg COD m−3 day−1, HY decreased to 1.04 mol H2 mol−1 glucose. For reactor R25, when OLR increased from 75 to 300 kg COD m−3 day−1, the increase in HY was less significant, i.e., from 0.44 to 0.63 mol H2 mol−1 glucose, but when OLR increased to 600 kg COD m−3 day−1, the yield decreased to 0.56 mol H2 mol−1 glucose.

**Figure 6** presents the variation in H2 content as a function of OLR for the two AFBRs used in this study. In reactors R10 and R25, the behavior of the H2 content also varied according to changes in OLR. The hydrogen content of the biogas increased with increasing OLR in both reactors, with a higher H2 content for HRT 1 h (240 and 600 kg COD m−3 day−1, respectively). The H<sup>2</sup> content ranged from 8 to 58% for R10 and 10 to 57% for R25.

The glucose conversion, HPR, HY, and H2 content of the reactors are consistent with the results of several studies conducted using AFBRs [6, 18, 27, 28, 30–32, 34, 35].

Valorization of Glucose-Based Wastewater Through Production of Hydrogen, Volatile Fatty Acids and Alcohols http://dx.doi.org/10.5772/67101 217

 **Figure 5.** HY as a function of the OLR for the AFBRs.

**Figure 4** presents the variation in the hydrogen production rate (HPR) as a function of OLR for the two AFBRs used in this study. Similar to the results of Barros et al. [6] for an AFBR with expanded clay as the support material, an influent glucose concentration of 4 g L−1, and alkalinity supplementation (values presented in **Figure 2**), the HPR values for R10 increased linearly from 0.12 to 1.58 L h−1 L−1 when OLR increased from 30 to 240 kg COD m−3. By contrast, a linear relationship between HPR and OLR was not observed in R25 for OLR ranging from 75 to 600 kg COD m−3. The maximum HPR values were 1.58 and 0.84 L h−1 L−1 for reactors

**Figure 5** presents the variation in HY as a function of OLR for the two AFBRs used in this study. The HY values increased with increasing OLR in both reactors. For reactor R10, when OLR was increased from 30 to 120 kg COD m−3 day−1, HY increased significantly from 0.48 to

changes in OLR. The hydrogen content of the biogas increased with increasing OLR in both

day−1, the increase in HY was less significant, i.e., from 0.44 to 0.63 mol H2

content ranged from 8 to 58% for R10 and 10 to 57% for R25.

results of several studies conducted using AFBRs [6, 18, 27, 28, 30–32, 34, 35].

this study. In reactors R10 and R25, the behavior of the H2

when OLR increased to 600 kg COD m−3 day−1, the yield decreased to 0.56 mol H2

mol−1 glucose, but when OLR increased to 240 kg COD m−3 day−1, HY decreased

mol−1 glucose. For reactor R25, when OLR increased from 75 to 300 kg COD m−3

content as a function of OLR for the two AFBRs used in

content of the reactors are consistent with the

content for HRT 1 h (240 and 600 kg COD m−3 day−1, respectively).

mol−1 glucose, but

content also varied according to

mol−1 glucose.

R10 and R25, respectively.

216 Biological Wastewater Treatment and Resource Recovery

1.32 mol H2

The H<sup>2</sup>

to 1.04 mol H2

**Figure 6** presents the variation in H2

 **Figure 4.** HPR as a function of the OLR for the AFBRs.

The glucose conversion, HPR, HY, and H2

reactors, with a higher H2

 **Figure 6.** H<sup>2</sup> content as a function of the OLR for the AFBRs.

**Table 1** compares studies that evaluated OLR and HY. Studies that observed a decrease in HY with increasing OLR used an OLR range of 6–833.3 kg COD m−3 day−1 and reported HYs of 4.26–0.81 mol H2 .mol−1 substrate. By contrast, studies that observed an increase in HY with increasing OLR worked with an OLR range of 13.5–480 kg COD m−3 day−1 and reported HYs of 0.94–2.49 mol H2 mol−1 substrate.


**Table 1.** Comparison of the studies that varied the OLR by changing the substrate concentration.

According to Kraemer and Bagley [26], the reason for the variations of H2 yield at lower or higher OLRs is unknown. High OLR values may reduce the production of H2 by (1) increasing inhibition by volatile fatty acids (VFAs) with increasing OLR, (2) decreasing thermodynamic regulation due to lower dissolved H2 concentrations at lower OLRs, (3) affecting acetogenic activity, and (4) increase CO2 inhibition by increasing the concentration of dissolved CO2 .

Inhibition by VFAs at high OLR values appears to be a valid explanation. The ability of added external VFA to reduce or inhibit the production of H2 in mixed-culture and continuous-flow systems has been studied, and there is consensus that butyrate increases higher inhibition than the acetate [18, 24, 40].

H2 production was also assessed with or without the addition of sodium bicarbonate as an alkalizing agent. The effect of the alkalizing agent on pH was important for controlling the hydrogen content and CO2 in the system. The high HY in the absence of a buffering agent can be attributed to the pH range of the reactor and the CO2 concentrations produced at steady bicarbonate concentrations [41–44].

#### **3.2. Soluble microbial products**

**Table 2** presents the distribution of soluble microbial products (SMPs) with increasing glucose concentration and increasing OLRs in the AFBRs. The molar fractions of acetic and butyric acid were the largest by percentage. Barros et al. [6] for an AFBR with expanded clay as the support material, an influent glucose concentration of 4 g L−1, and alkalinity supplementation (values presented in **Table 2**) observed a descending order of products of acetate (32.99–46.81%), butyrate (37.30–41.49%), ethanol (10.18–22.95%), and propionate (1.26–4.90%). In our reactor R10, the products in descending order were ethanol (45.54–71.54%), acetate (27.11–50.63%), butyrate (2.91–31.03%) and methanol (0.00–14.41%). In reactor R25, the products in descending order were ethanol (48.00–71.54%), acetate (12.05–37.43%), butyrate (01.02–29.09%), and methanol (0.00–14.41%) (**Table 2**).


HAc acetate, HBu butyrate, HPr propionate, EtOH ethanol, MetOH methanol, TVFA total volatile fatty acids, TVFA HAc + HBu + HPr, SMP TVFA + EtOH + MetOH, HAc/SMP molar acetate-to-SMP ratio, HBu/SMP molar butyrate-to-SMP ratio, HPr/SMP molar propionate-to-SMP ratio, EtOH/SMP molar ethanol-to-SMP ratio, MetOH/SMP molar methanolto-SMP ratio, HAc/HBu molar acetate-to-butyrate ratio

**Table 2.** Effect of glucose concentration and OLR on the SMP distribution in the AFBRs.

According to Kraemer and Bagley [26], the reason for the variations of H2

**Table 1.** Comparison of the studies that varied the OLR by changing the substrate concentration.

**Study Substrate OLR (kg m−3 d−1) HY (mol H<sup>2</sup>**

Yu et al. [36] Rice winery 168 432 1.89 1.79 Van Ginkel and Logan [24] Glucose 25.6 76.8 2.20 2.00 Van Ginkel and Logan [37] Glucose 6 24 2.80 2.20 Kyazze et al. [15] Sucrose 22.4 112.2 1.65 0.81 Lin et al. [38] Sucrose 34.7 833.3 4.26 2.31 Davila-Vasquez et al. [39] Cheese whey 54 138.6 2.4 1.0

Lin et al. [18] Sucrose 13.5 107.9 1.69 2.49

Zhang et al. [35] Glucose 60 480 0.94 1.19 Shida et al. [27] Glucose 6 48 1.84 2.29 Perna et al. [17] Cheese whey 22 37 0.5 0.67

higher OLRs is unknown. High OLR values may reduce the production of H2

regulation due to lower dissolved H2

external VFA to reduce or inhibit the production of H2

be attributed to the pH range of the reactor and the CO2

activity, and (4) increase CO2

Adapted from Kraemer and Bagley [26].

Lower OLR improves H2

Higher OLR improves H2

production

218 Biological Wastewater Treatment and Resource Recovery

production

than the acetate [18, 24, 40].

hydrogen content and CO2

bicarbonate concentrations [41–44].

**3.2. Soluble microbial products**

H2

inhibition by volatile fatty acids (VFAs) with increasing OLR, (2) decreasing thermodynamic

Inhibition by VFAs at high OLR values appears to be a valid explanation. The ability of added

systems has been studied, and there is consensus that butyrate increases higher inhibition

 production was also assessed with or without the addition of sodium bicarbonate as an alkalizing agent. The effect of the alkalizing agent on pH was important for controlling the

**Table 2** presents the distribution of soluble microbial products (SMPs) with increasing glucose concentration and increasing OLRs in the AFBRs. The molar fractions of acetic and butyric acid were the largest by percentage. Barros et al. [6] for an AFBR with expanded clay as the support material, an influent glucose concentration of 4 g L−1, and alkalinity supplementation (values

concentrations at lower OLRs, (3) affecting acetogenic

in mixed-culture and continuous-flow

concentrations produced at steady

inhibition by increasing the concentration of dissolved CO2

Sucrose 20 160 1.34 2.17

in the system. The high HY in the absence of a buffering agent can

yield at lower or

 **mol−1 substrate)**

**Low High Low OLR High OLR**

by (1) increasing

.

Previous studies employing conditions similar to those used in the present study observed the production of similar metabolites, although differences in the distributions of the metabolites were observed [6, 18, 27, 28, 30–32, 34, 35].

The reactors R10 and R25 produced higher amounts of solvents, such as MetOH and EtOH in the R25 reactor. The higher EtOH concentrations observed in R10 and R25 are similar to the results of Wu et al. [34]. However, our recent studies [6, 27, 29] that used the same medium composition, inoculum, and support material have significantly different results. Barros et al. [6] with an influent glucose concentration of 4 g L−1, and alkalinity supplementation, observed ethanol percentages lower than 22.95% at the beginning of the operation and  subsequently decreased and stabilized to 11%. EtOH production is considered unfavorable for hydrogen metabolite production because no H2 is consumed or produced (Eq. (1)):

$$\rm C\_{\rm 6}H\_{12}O\_{\rm 6} \rightarrow 2\ CH\_3CH\_2OH + 2CO\_2 \tag{1}$$

Propionate was only detected during the operation of the reactor containing 25 g L−1, with maximum concentration of 1.20 mM in the OLR of 100 kg COD m−3 day−1. Propionic acid production was not observed in AFBRs with influent glucose concentration of 2 g L−1 [27, 29]. Zhang et al. [35] suggested that the absence of propionic acid may be due to inhibition of the activity of the bacteria that form this acid under low pH conditions; these bacteria may be sensitive to both low HRTs and high OLRs. Moreover, the absence of propionic acid production ensures greater production of hydrogen due to the lower consumption of H2 for forming propionate (Eq. (2)):

$$\rm C\_8H\_{12}O\_8 + 2H\_2 \rightarrow CH\_3CH\_2COOH + 2H\_2O \tag{2}$$

Both HAc and HBu are soluble metabolites favoring H2 production because these products are generated during H<sup>2</sup> production (Eqs. (3) and (4)):

$$\rm C\_6H\_{12}O\_6 + 2H\_2O \rightarrow 2\cdot CH\_3COOH + 2CO\_2 + 4H\_2 \tag{3}$$

$$\rm C\_6H\_{12}O\_6 \rightarrow \rm CH\_3CH\_2CH\_2COOH + 2CO\_2 + 2H\_2 \tag{4}$$

Previous studies have observed that H2 production increases with the molar ratio of HAc/HBu [45, 46]. **Table 2** presents the variation of the HAc/HBu ratio in R10 and R25. Barros et al. [6] for an influent glucose concentration of 4 g L−1, and alkalinity supplementation, observed the best proportion of soluble metabolites and therefore a higher yield of hydrogen, with molar ratios of HAc/HBu ranging from 0.81 to 1.21 for OLRs varied 12–96 kg COD m−3 day−1, respectively, but decreasing to 1.08 for an OLR of 96 kg COD m−3 day−1. In our R25, similar behavior of Barros et al. [6] were obtained, but in R10 HAc/HBu ratio decreased from 17.42 to 0.87 when the OLRs increased from 30 to 240 kg COD m−3 day−1.

According to Hafez et al. [45], when OLR increased from 6.5 to 103 g COD L−1 day−1, acetate and butyrate were the main liquid products, with trace concentrations of ethanol and no detectable lactate, whereas in the OLR range of 154–206 g COD L−1 day−1, the concentrations of propionate, isovalerate, valerate, and ethanol increased markedly. The steady-state average molar ratios of acetate/butyrate were 2.3, 2.3, 2.0, and 2.2 for OLRs of 6.5, 25.7, 51.4, and 103 g COD L−1 day−1, respectively, but decreased to 1.1 for OLRs of 154 and 206 g COD L−1 day−1.

According to Prakasham et al. [47], at lower substrate conditions with the limitation of substrate concentration, increasing glucose concentration progressively increases H2 production because of effective metabolism and further H2 production process. However, higher concentrations can also negatively impact H2 production. When the H2 yield observed value reduced because the glucose concentration was above the optimum value, a limited glucose utilization occurred, or a shift in the metabolic pathway from the acidogenic phase to a solventogenic phase took place.

Hydrogen and CO2 were the only gaseous metabolites during all stages of the experiment. NO CH4 was detected in the biogas from either reactor. The combination of heat treatment of the inoculum and operation under acidogenic pH conditions inhibited the methanogenic activity responsible for the consumption of hydrogen in the system. Furthermore, the results in the literature suggest that manipulating some operational parameters such as the HRT contributes to the elimination of methanogenic archaea in the reactors.

According to Chen et al. [48], these microorganisms fail to thrive in part because the maximum specific growth rate of methanogenic archaea (μmaximum = 0.0167 h−1) is significantly lower than that of acidogenic microorganisms (μmaximum = 0.083 h−1). Thus, methanogenic microorganisms are unable to reproduce or remain in equilibrium under these conditions, resulting in their removal from the reactor.

#### **3.3. COD removal and carbon balance**

subsequently decreased and stabilized to 11%. EtOH production is considered unfavorable

Propionate was only detected during the operation of the reactor containing 25 g L−1, with maximum concentration of 1.20 mM in the OLR of 100 kg COD m−3 day−1. Propionic acid production was not observed in AFBRs with influent glucose concentration of 2 g L−1 [27, 29]. Zhang et al. [35] suggested that the absence of propionic acid may be due to inhibition of the activity of the bacteria that form this acid under low pH conditions; these bacteria may be sensitive to both low HRTs and high OLRs. Moreover, the absence of propionic acid produc-

C6 H <sup>12</sup> O<sup>6</sup> → 2 CH <sup>3</sup> CH <sup>2</sup>

tion ensures greater production of hydrogen due to the lower consumption of H2

production (Eqs. (3) and (4)):

+ 2  H <sup>2</sup>

C6 H <sup>12</sup> O<sup>6</sup> → CH <sup>3</sup> CH <sup>2</sup> CH <sup>2</sup>

+ 2  H <sup>2</sup> → CH <sup>3</sup> CH <sup>2</sup>

 O → 2 CH <sup>3</sup>

[45, 46]. **Table 2** presents the variation of the HAc/HBu ratio in R10 and R25. Barros et al. [6] for an influent glucose concentration of 4 g L−1, and alkalinity supplementation, observed the best proportion of soluble metabolites and therefore a higher yield of hydrogen, with molar ratios of HAc/HBu ranging from 0.81 to 1.21 for OLRs varied 12–96 kg COD m−3 day−1, respectively, but decreasing to 1.08 for an OLR of 96 kg COD m−3 day−1. In our R25, similar behavior of Barros et al. [6] were obtained, but in R10 HAc/HBu ratio decreased from 17.42 to 0.87 when

According to Hafez et al. [45], when OLR increased from 6.5 to 103 g COD L−1 day−1, acetate and butyrate were the main liquid products, with trace concentrations of ethanol and no detectable lactate, whereas in the OLR range of 154–206 g COD L−1 day−1, the concentrations of propionate, isovalerate, valerate, and ethanol increased markedly. The steady-state average molar ratios of acetate/butyrate were 2.3, 2.3, 2.0, and 2.2 for OLRs of 6.5, 25.7, 51.4, and 103 g COD L−1 day−1, respectively, but decreased to 1.1 for OLRs of 154 and 206 g COD L−1 day−1. According to Prakasham et al. [47], at lower substrate conditions with the limitation of sub-

production. When the H2

strate concentration, increasing glucose concentration progressively increases H2

 COOH + 2  H <sup>2</sup>

 COOH + 2 CO2

 COOH + 2 CO2

production increases with the molar ratio of HAc/HBu

production process. However, higher concen-

yield observed value reduced

is consumed or produced (Eq. (1)):

 OH + 2 CO2 (1)

for forming

production

 O (2)

+ 4  H <sup>2</sup> (3)

+ 2  H <sup>2</sup> (4)

production because these products

for hydrogen metabolite production because no H2

220 Biological Wastewater Treatment and Resource Recovery

Both HAc and HBu are soluble metabolites favoring H2

C6 H <sup>12</sup> O<sup>6</sup>

the OLRs increased from 30 to 240 kg COD m−3 day−1.

because of effective metabolism and further H2

trations can also negatively impact H2

propionate (Eq. (2)):

are generated during H<sup>2</sup>

C6 H <sup>12</sup> O<sup>6</sup>

Previous studies have observed that H2

The carbon balance in the reactors can be calculated by Eq. (5) according to Gavala et al. [49]. The comparison between measured and calculated COD concentrations for each steady state is also presented. The COD calculations were performed as the following: the products (CODproducts) and the glucose (CODglucose) COD concentrations were calculated according to Eqs. (5) and (6), respectively. The CODresidual was calculated after subtraction of the sum of the CODproducts and CODglucose from the CODmeasured (Eq. (3)).The CODothers corresponds to the nonidentified metabolic products during glucose fermentation:

$$\text{COD}\_{\text{products}} = a.\left(\frac{mmolHAc}{1}\right).64\frac{mg\text{COD}}{mmolHAc} + b.\left(\frac{mmolHBu}{1}\right).160\frac{mg\text{COD}}{mmolHBu}\left(\frac{mmolHAc}{1}\right)$$

$$+c.\left(\frac{mmolHPr}{1}\right).112\frac{mg\text{COD}}{mmolPr} + d.\left(\frac{mmolMetOH}{1}\right).48\frac{mg\text{COD}}{mmolMetOH}$$

$$+e.\left(\frac{mmolEtOH}{1}\right).96\frac{mg\text{COD}}{mmolEtOH}\tag{5}$$

where a, b, c, d, and e are the measured concentrations of the acetic acid, butyric acid, propionic acid, methanol, and ethanol, respectively.

$$\text{COD}\_{\text{glucose}} = f. \left(\frac{\text{mgGlucose}}{1}\right) \frac{192 \text{ mg COD}}{180 \text{ mg}} \tag{6}$$

where f is the measured concentration of glucose.

The difference between CODmeasured and COD based on SMP may be attributed to the presence of other soluble metabolites that were not detected, e.g., lactic acid and formic acid, because the chromatographic method of headspace extraction used in this study only detects alcohols and volatile acids.

This difference was calculated based on Eq. (7):

$$\text{COD}\_{\text{others}} = \text{COD}\_{\text{massud}} - \left( \text{COD}\_{\text{products}} + \text{COD}\_{\text{glucose}} \right) \tag{7}$$

**Table 3** presents influent and effluent COD values and standard deviations as well as efficiencies for all reactors. Influent COD represents glucose added to the wastewater and carbonaceous matter present in urea. Effluent COD corresponds to the carbonaceous matter in the effluent that was oxidized. Carbonaceous matter present in the effluent consists of nonconsumed glucose; soluble metabolites, e.g., organic acids, solvents, and other intermediary compounds; and biomass detached from the support medium.


**Table 3.** Influent COD, effluent COD, and COD removal in AFBRs.

The theoretical effluent COD was calculated based on stoichiometric relationships for oxidation of glucose, acetic acid, butyric acid, propionic acid, biomass, ethanol, and methanol to estimate the carbon balance. Theoretical COD values for the remaining glucose, soluble metabolites, and biomass as well as the difference between the theoretical total COD and the COD measured for all reactors are presented in **Table 4**.

In the reactor operated by Barros et al. [6], this difference varied between 12 and 350 mg L−1, which corresponded to a variation of 0.34 and 9.19%. The reactor R10 showed a difference ranging from 91 to 301 mg L−1 (variation of 1.05 and 3.28%), whereas in the reactor R25, the difference varied between 17 and 1026 mg L−1 (variation of 0.07 and 4.62%).Those differences


may be accredited to the presence of other metabolites such as lactic acid and formic acid that were not detected, probably due to the chromatographic method performed (headspace extraction), considering that this method can only detect volatile acids and alcohols.

**Table 4.** Theoretical COD values of soluble metabolites, biomass COD, and effluent COD measured in AFBRs.

The largest variation between COD measured in the effluent and the theoretical COD (corresponding to glucose, soluble metabolites, and biomass in the effluent) was 9.19% based on the results obtained from the carbon balance. However, according to Standard Methods [33], the determination of metabolites and COD produces errors of close to 10%. For that reason, this variation may be attributed to the margin of error of the determination methods used.

## **4. Conclusions**

This difference was calculated based on Eq. (7):

222 Biological Wastewater Treatment and Resource Recovery

compounds; and biomass detached from the support medium.

Barros et al. [6] 12 4216 ± 210 3788 ± 153 10 ± 6

R10 30 11,298 ± 954 8617 ± 457 24 ± 5

R25 75 26,126 ± 1024 20,202 ± 978 23 ± 3

 *COD others* = *CODmeasured* — (*CODproducts* + *CODglucose*) (7) **Table 3** presents influent and effluent COD values and standard deviations as well as efficiencies for all reactors. Influent COD represents glucose added to the wastewater and carbonaceous matter present in urea. Effluent COD corresponds to the carbonaceous matter in the effluent that was oxidized. Carbonaceous matter present in the effluent consists of nonconsumed glucose; soluble metabolites, e.g., organic acids, solvents, and other intermediary

**OLR (kg COD m−3 day−1) Influent COD (mg L−1) Effluent COD (mg L−1) COD removal (%)**

 4140 ± 206 3349 ± 146 19 ± 9 4139 ± 270 3718 ± 165 10 ± 4 4487 ± 220 3805 ± 191 15 ± 2 4312 ± 226 3680 ± 136 15 ± 4

 10,439 ± 843 9056 ± 419 13 ± 6 10,693 ± 977 8639 ± 433 19 ± 3 10,175 ± 799 8589 ± 447 16 ± 2 10,969 ± 901 8705 ± 512 21 ± 2

 26,447 ± 1201 22,352 ± 883 15 ± 2 27,285 ± 1392 22,207 ± 791 19 ± 2 26,116 ± 1273 23,502 ± 943 10 ± 1 28,216 ± 1321 25,242 ± 967 11 ± 2

The theoretical effluent COD was calculated based on stoichiometric relationships for oxidation of glucose, acetic acid, butyric acid, propionic acid, biomass, ethanol, and methanol to estimate the carbon balance. Theoretical COD values for the remaining glucose, soluble metabolites, and biomass as well as the difference between the theoretical total COD and the

In the reactor operated by Barros et al. [6], this difference varied between 12 and 350 mg L−1, which corresponded to a variation of 0.34 and 9.19%. The reactor R10 showed a difference ranging from 91 to 301 mg L−1 (variation of 1.05 and 3.28%), whereas in the reactor R25, the difference varied between 17 and 1026 mg L−1 (variation of 0.07 and 4.62%).Those differences

COD measured for all reactors are presented in **Table 4**.

**Table 3.** Influent COD, effluent COD, and COD removal in AFBRs.

Satisfactory performance for H2 production was observed in the anaerobic fluidized bed reactor containing 10 g L−1 glucose. However, in the reactor containing 25 g L−1 glucose, the yield was limited.

The HPR had a linear increase with OLR, with the exception of reactor operated with 25 g L−1 glucose. The maximum HPR was 1.58 L h−1 L−1 obtained in the reactor with 10 g L−1 glucose for OLR of 240 kg COD m−3 day−1 (HRT = 1 h). The maximum HY was 1.32 mol H2 mol−1 glucose obtained in the reactor with 10 g L−1 glucose for HRT 2 h (OLR = 240 kg COD m−3 day−1).

The H<sup>2</sup> production with addition of sodium bicarbonate was important to control the pH and CO2 system. The reactors operated at high glucose concentrations (10 and 25 g L−1) showed higher proportions of solvents.

## **Author details**

Eduardo Lucena Cavalcante de Amorim1, 2, Leandro Takano Sader3 , Lucas Rodrigues Ramos3 and Edson Luiz Silva3 \*

\*Address all correspondence to: edsilva@ufscar.br

1 Technology Center, Federal University of Alagoas, Maceió, AL, Brazil

2 Department of Hydraulics and Sanitation, University of São Paulo, São Carlos, SP, Brazil

3 Department of Chemical Engineering, Federal University of São Carlos, São Carlos, SP, Brazil

## **References**


[7] Rosa PRF, Santos SC, Silva EL. Different ratios of carbon sources in the fermentation of cheese whey and glucose as substrates for hydrogen production and ethanol production in continuous reactors. International Journal of Hydrogen Energy. 2014;**39**:1288–1296. DOI: 10.1016/j.ijhydene.2013.11.011.

OLR of 240 kg COD m−3 day−1 (HRT = 1 h). The maximum HY was 1.32 mol H2

Eduardo Lucena Cavalcante de Amorim1, 2, Leandro Takano Sader3

1 Technology Center, Federal University of Alagoas, Maceió, AL, Brazil

The H<sup>2</sup>

higher proportions of solvents.

224 Biological Wastewater Treatment and Resource Recovery

\*

\*Address all correspondence to: edsilva@ufscar.br

7626. DOI: 10.1016/j.ijhydene.2013.09.157.

ijhydene.2010.01.108.

**Author details**

**References**

and Edson Luiz Silva3

CO2

obtained in the reactor with 10 g L−1 glucose for HRT 2 h (OLR = 240 kg COD m−3 day−1).

2 Department of Hydraulics and Sanitation, University of São Paulo, São Carlos, SP, Brazil

3 Department of Chemical Engineering, Federal University of São Carlos, São Carlos, SP, Brazil

[1] Bartacek J, Zabranska J, Lens PNL. Developments and constraints in fermentative hydrogen production. Biofuels, Bioproducts Biorefining. 2007;**1**:201–214. DOI: 10.1002/bbb. [2] Mohan SV. Harnessing of biohydrogen from wastewater treatment using mixed fermentative consortia: Process evaluation towards optimization. International Journal of

[3] Reddy MV, Amulya K, Rohit MV, Sarma PN, Mohan SV. Valorization of fatty acid waste for bioplastics production using Bacillus tequilensis: Integration with dark-fermentative hydrogen production process. International Journal of Hydrogen Energy. 2014;**39**:7616–

[4] Sarma SJ, Brar SK, Bihan YL, Buelna G. Liquid waste from bio-hydrogen production—a commercially attractive alternative for phosphate solubilizing bio-fertilizer. International Journal of Hydrogen Energy. 2013;**38**:8704–8707. DOI: 10.1016/j.ijhydene.2013.05.032 [5] Sarma SJ, Pachapur V, Brar SK, Bihan YL, Buelna G. Hydrogen biorefinery: Potential utilization of the liquid waste from fermentative hydrogen production. Renewable and

Sustainable Energy Reviews. 2015;**50**:942–951. DOI: 10.1016/j.rser.2015.04.191.

[6] Barros AR, Amorim ELC, Reis CM, Shida GM, Silva EL. Biohydrogen production in anaerobic fluidized bed reactors: Effect of support material and hydraulic retention time. International Journal of Hydrogen Energy. 2010;**35**:3379–3388. DOI: 10.1016/j.

Hydrogen Energy. 2009;**34**:7460–7474. DOI: 10.1016/j.ijhydene.2009.05.062.

production with addition of sodium bicarbonate was important to control the pH and

system. The reactors operated at high glucose concentrations (10 and 25 g L−1) showed

mol−1 glucose

, Lucas Rodrigues Ramos3


used to treat cheese whey. International Journal of Hydrogen Energy. 2013;38:54–62. DOI: 10.1016/j.ijhydene.2012.10.022.


[29] Amorim ELC, Barros AR, Damianovic MHRZ, Silva EL. Anaerobic fluidized bed reactor with expanded clay as support for hydrogen production through dark fermentation of glucose. International Journal of Hydrogen Energy. 2009;**34**:783–790. DOI: 10.1016/j. ijhydene.2008.11.007.

used to treat cheese whey. International Journal of Hydrogen Energy. 2013;38:54–62. DOI:

[18] Lin CN, Wu SY, Chang, JS. Fermentative hydrogen production with a draft tube fluidized bed reactor containing silicone-gel-immobilized anaerobic sludge. International Journal of Hydrogen Energy. 2006;**31**:2200–2210. DOI: 10.1016/j.ijhydene.2006.05.012.

[19] Jung KW, Kim DH, Kim SH, Shin HS. Bioreactor design for continuous dark fermentative hydrogen production. Bioresource Technology. 2011;**102**:8612–8620. DOI: 10.1016/j.

[20] Show K, Lee D, Chang J. Bioreactor and process design for biohydrogen production. Bioresource Technology. 2011;**102**:8524–8533. DOI: 10.1016/j.biortech.2011.04.055.

[21] Barca C, Soric A, Ranava D, Giudici-Orticoni MT, Ferrasse JH. Anaerobic biofilm reactors for dark fermentative hydrogen production from wastewater: A review. Bioresource

[22] Mohammadi P, Ibrahim S, Annuar MSM, Ghafari S, Vikineswary S, Zinatizadeh AA. Influences of environmental and operational factors on dark fermentative hydrogen production: A review. Clean—Soil, Air, Water. 2012;**40**:1297–1305. DOI: 10.1002/

[23] Tawfik A, Salem A. The effect of organic loading rate on bio-hydrogen production from pre-treated rice straw waste via mesophilic up-flow anaerobic reactor. Bioresource

[24] Van Ginkel SW, Logan BE. Inhibition of biohydrogen production by undissociated acetic and butyric acids. Environmental Science and Technology. 2005;**39**:9351–9356. DOI:

[25] Kim S, Han S, Shin H. Effect of substrate concentration on hydrogen production and 16S rDNA-based analysis of the microbial community in a continuous fermenter. Process

[26] Kraemer JT, Bagley DM. Improving the yield from fermentative hydrogen production.

[27] Shida GM, Barros AR, Reis CM, Amorim ELC, Damianovic MHRZ, Silva EL. Long-term stability of hydrogen and organic acids production in an anaerobic fluidized-bed reactor using heat treated anaerobic sludge inoculum. International Journal of Hydrogen

[28] Shida GM, Sader LT, Amorim ELC, Sakamoto IK, Maintinguer SI, Saavedra NK, Varesche MBA, Silva EL. Performance and composition of bacterial communities in anaerobic fluidized bed reactors for hydrogen production: Effects of organic loading rate and alkalinity. International Journal of Hydrogen Energy. 2012;**37**:16925–16934. DOI: 10.1016/j.

Technology. 2015;**185**:386–398. DOI: 10.1016/j.biortech.2015.02.063.

Technology. 2012;**107**:186–190. DOI: 10.1016/j.biortech.2011.11.086.

Biochemistry. 2006;**41**:199–2007. DOI: 10.1016/j.procbio.2005.06.013.

Energy. 2009;**34**:3679–3688. DOI: 10.1016/j.ijhydene.2009.02.076.

Biotechnology Letters. 2007;**29**:685–695. DOI: 10.1007/s10529-006-9299-9.

10.1016/j.ijhydene.2012.10.022.

226 Biological Wastewater Treatment and Resource Recovery

biortech.2011.03.056.

clen.201100007.

10.1021/es0510515.

ijhydene.2012.08.140.


## **Production of Biogas and Performance Evaluation of Ultrasonic Membrane Anaerobic System (UMAS) for Palm Oil Mill Effluent Treatment (POME) Production of Biogas and Performance Evaluation of Ultrasonic Membrane Anaerobic System (UMAS) for Palm Oil Mill Effluent Treatment (POME)**

Abdurahman Hamid Nour and Azhari Hamid Nour Azhari Hamid Nour

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

http://dx.doi.org/10.5772/67602

Abdurahman Hamid Nour and

#### **Abstract**

[41] Leite JAC, Fernandes BS, Pozzi E, Barboza M, Zaiat M. Application of an anaerobic packed-bed bioreactor for the production of hydrogen and organic acids. International Journal of Hydrogen Energy. 2008;**33**:579–586. DOI: 10.1016/j.ijhydene.2007.10.009.

[42] Valdez-Vazquez I, Poggi-Varaldo HM. Alkalinity and high total solids affecting H2 production from organic solid waste by anaerobic consortia. International Journal of

[43] Choi J, Ahn Y. Biohydrogen fermentation from sucrose and piggery waste with high levels of bicarbonate alkalinity. Energies. 2015;**8**:1716–1729. DOI: 10.3390/en8031716. [44] Silva AJ, Pozzi E, Foresti E, Zaiat M. The influence of the buffering capacity on the production of organic acids and alcohols from wastewater in anaerobic reactor. Applied Biochemistry and Biotechnology. 2015;**175**:2258–2265. DOI: 10.1007/s12010-014-1424-y.

[45] Hafez H, Nakhla G, El Naggar MH, Elbeshbishy E, Baghchehsaraee B. Effect of organic loading rate on a novel hydrogen bioreactor. International Journal of Hydrogen Energy.

[46] Wang JL, Wan W. The effect of substrate concentration on biohydrogen production by using kinetic models. Science in China Series B-Chemistry. 2008;**51**:1110–1117. DOI:

[47] Prakasham RS, Brahmaiah P, Satish T, SambasivaRao KRS. Fermentative biohydrogen production by mixed anaerobic consortia: Impact of glucose to xylose ratio. International Journal of Hydrogen Energy. 2010;**34**:9354–9361. DOI: 10.1016/j/ijhydene.2009.09.104.

[48] Chen CC, Lin CY, Chang JS. Kinetics of hydrogen production with continuous anaerobic cultures utilizing sucrose as the limiting substrate. Applied Microbiology and

[49] Gavala HN, Skiadas IV, Ahring BK. Biological hydrogen production in suspended and attached growth anaerobic reactor systems. International Journal of Hydrogen Energy.

2010;**35**:81–92. DOI: 10.1016/j.ijhydene.2009.10.051.

Biotechnology. 2001;**57**:56–64. DOI: 10.1007/s002530100747.

2006;**31**:1164–1175. DOI: 10.1016/j/ijhydene.2005.09.009.

10.1007/s11426-008-0104-6.

228 Biological Wastewater Treatment and Resource Recovery

Hydrogen Energy. 2009;**34**:3639–3646. DOI: 10.1016/j.ijhydene.2009.02.039.

This study proposes a new approach for integrated technology of ultrasonic and membrane for a palm oil mill effluent treatment. This study evaluated the performance of the new design of ultrasonic membrane anaerobic system (UMAS) when a palm oil mill effluent (POME) introduces this approach. To fit kinetic study, six steady states were investigated and the results have shown that the mixed liquor volatile suspended solids (MLVSSs) range from 10,400 to 17,350 mg/l while the mixed liquor suspended solids (MLSSs) range from 13,800 to 22,600 mg/l. Three kinetic models of Monod, Contois, and Chen and Hashimoto were used to evaluate the integrated system at organic loading rates ranging from 1 to 15 kg COD/m3 /day. The percentage efficiency of COD removal was from 92.8 to 98.3%, and hydraulic retention time (HRT) was from 500.8 to 8.6 days. The influent COD concentrations of the POME ranged from 70,400 to 90,200 mg/l.The integrated technology of UMAS is a more attractive one as it avoids membrane fouling problems.

**Keywords:** membrane, ultrasonic, POME, methane, CO2 , UMAS

## **1. Introduction**

The palm oil industry has grown tremendously in the recent years and accounted for the largest percentage of oil and fats production in the world in 2011. Over the last few decades, the palm oil industry has been growing rapidly. Palm oil has risen to become the most produced and consumed vegetable oil in the world, widely used in food, cosmetic and hygienic

© 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, 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.

products due to its affordable price, efficient production and high oxidative stability [1]. Palm oil is the most produced vegetable oil in the world with a global production of almost 60 million tons and a global vegetable oil market share of more than 35% by weight in 2015 as reported by Hansen et al. [2] and MPOB [3]. The industry continues to generate huge revenues for the producing countries. Accordingly, it is not surprising that the oil palm industry is expected to grow further in the coming years as shown in **Figure 1**.

Over the long term, global palm oil demand shows an increasing trend as an expanding global population gives rise to increased consumption of palm-oil based products world consumption of palm oil [5]. [6] Stated that palm oil industries have been significantly contributing towards the economic growth and increase standard of living among the South East Asian countries. Nowadays, the global production and demand for palm oil are increasing rapidly where the plantations are spreading across Asia, Africa and Latin America. The five leading palm oil producing countries are Indonesia, Malaysia, Thailand, Colombia and Nigeria [7] as shown in **Figure 2**.

The development of palm oil industry in Malaysia has turned into a phenomenal in which the area of plantation expanded from year to year. The country is experiencing a robust development in new oil palm plantations and palm oil mills. This commodity plays a significant role in the Malaysia economic growth [8]. Throughout the year, Malaysia is blessed with favorable weather conditions, which are advantageous for palm oil cultivation [9]. Thus, it is not surprising that the highest yields have been obtained from palms grown in this region, which is far from its natural habitat. Besides, the Malaysian palm oil industry has grown to become a very important agriculture-based industry, where the country is today one of the world's leading producer and exporter of palm oil.

**Figure 1.** Global consumption of palm oil from 1995/1996 to 2014/2015 (USDA, 2016) [4].

**Figure 2.** Palm oil production by country [10, 11].

products due to its affordable price, efficient production and high oxidative stability [1]. Palm oil is the most produced vegetable oil in the world with a global production of almost 60 million tons and a global vegetable oil market share of more than 35% by weight in 2015 as reported by Hansen et al. [2] and MPOB [3]. The industry continues to generate huge revenues for the producing countries. Accordingly, it is not surprising that the oil palm industry

Over the long term, global palm oil demand shows an increasing trend as an expanding global population gives rise to increased consumption of palm-oil based products world consumption of palm oil [5]. [6] Stated that palm oil industries have been significantly contributing towards the economic growth and increase standard of living among the South East Asian countries. Nowadays, the global production and demand for palm oil are increasing rapidly where the plantations are spreading across Asia, Africa and Latin America. The five leading palm oil producing countries are Indonesia, Malaysia, Thailand, Colombia and Nigeria [7] as

The development of palm oil industry in Malaysia has turned into a phenomenal in which the area of plantation expanded from year to year. The country is experiencing a robust development in new oil palm plantations and palm oil mills. This commodity plays a significant role in the Malaysia economic growth [8]. Throughout the year, Malaysia is blessed with favorable weather conditions, which are advantageous for palm oil cultivation [9]. Thus, it is not surprising that the highest yields have been obtained from palms grown in this region, which is far from its natural habitat. Besides, the Malaysian palm oil industry has grown to become a very important agriculture-based industry, where the country is today one of the world's

**Palm Oil**

1995/962005/062006/072007/082008/092009/102010/112011/122012/132013/142014/15

**Year**

**Figure 1.** Global consumption of palm oil from 1995/1996 to 2014/2015 (USDA, 2016) [4].

is expected to grow further in the coming years as shown in **Figure 1**.

shown in **Figure 2**.

0

10

20

30

40

**World Consumption in Million Metric** 

**Tons**

50

60

70

leading producer and exporter of palm oil.

230 Biological Wastewater Treatment and Resource Recovery

**Figure 3** depicts the statistics production of palm oil superseded soybean oil from 13% in 1990 to 28% of total oil and fats production in 2011. This is because oil palm has higher annual oil yield per hectare than other oil seeds crops including soybean [11] and palm oil has a relatively lower price as compared to the major alternative vegetable oils [12]. POME is highly polluted wastewater if not treated properly; it causes a lot of environment issues. POME is a colloidal suspension of 95–96% water, 0.6–0.7% oil and 4–5% total solids including 2–4% suspended solids originating from mixture of a sterilizer condensate, separator sludge and hydrocyclone wastewater [13]. The conventional treatment technology of POME employed in most of the palm oil mills in Malaysia is the ponding system of biological treatment [14–16]. However, coping with the increasing production in most palm oil mills, the undersized biological treatment system is unable to cope with the increased volume of POME [17]. Thus, proper POME treatment is urgently needed to ensure the sustainable economic growth of palm oil industry in Malaysia besides protecting the environment. Several researchers have proposed other biological treatments.

The treatment system includes aerated lagoon system [18], conventional anaerobic digester [19], anaerobic contact process [20], upflow anaerobic sludge blanket (UASB) reactor [17, 19], close tank digester [21], trickling filter, aerobic lagoon system [18], aerobic rotating biological contactor [19] and evaporation process [13].

The main objective of this study was to evaluate the performance and kinetics of the new designed ultrasonic membrane anaerobic system (UMAS) in the treatment of palm oil mill effluent (POME) based on three models [22–24]. **Table 1** shows mathematical expressions for specifics substrate utilization rate for three kinetic models (Monod, Contois, and Chen and Hashimoto).

**Figure 3.** World oil and fat production in 1990 and 2011 [3–5].


**Table 1.** Mathematical expressions of specifics substrate utilization rates for known kinetic models.

#### **1.1. Mechanisms of anaerobic digestion**

In anaerobic degrading of POME, biogas is formed when microorganisms, especially bacteria, degrade organic material in the absence of oxygen. Biogas consists of 50–75% methane (CH4 ), 25–45% carbon dioxide (CO2 ) and small amounts of other gases [25–27]. A simplified schematic representation of anaerobic degradation of organic matter is given in **Figure 4**. The AD process can be subdivided into the following four phases, each requires its own characteristic group of microorganisms.

The sequence of reactions involved in the mechanisms of AD is hydrolysis, acidogenesis, acetogenesis and methanogenesis [28]. Hydrolysis is conversion of nonsoluble biopolymers to soluble organic compounds. Acidogenesis is summarized as a conversion of soluble organic compounds to volatile fatty acids (VFA) and CO2 while acetogenesis is the conversion of VFAs to acetate and H2 [29]. Methanogenesis represents conversion of acetate and CO2 plus H2 to methane and carbon dioxide gas.

Production of Biogas and Performance Evaluation of Ultrasonic Membrane Anaerobic System... http://dx.doi.org/10.5772/67602 233

**Figure 4.** Process stages of anaerobic digestion [30].

## **2. Materials and methods**

#### **2.1. Raw POME wastewater preparation**

The raw POME was collected from a near local palm oil mill in Lebah Hillier, Kuantan, Malaysia. The raw POME was stored in a cold room at 4°C before use. Different dilutions of POME were prepared using tap water. The pH of the feed was adjusted to 7.0 using a 6 N NaOH solution.

#### **2.2. UMAS bioreactor operation and experimental setup**

A laboratory scale, with an effective 200-L UMAS reactor (**Figure 5**), was used in this study. The UMAS reactor consists of a cross-flow ultrafiltration membrane apparatus, a centrifugal pump and an anaerobic reactor. The total volume of the reactor was 200 L, and the working volume was 150 L. Six multifrequency ultrasonic transducers, operated at 25 KHz, are bonded to two sides of the tank chamber and connected to a Crest Genesis Generator (250 W, 25 KHz; Crest Ultrasonic, Trenton, NJ, USA). The maximum operating pressure on the membrane was 55 bars at 70 WC, and the pH ranged from 2 to 12.

#### **2.3. Analytical methods**

),

plus H2

to

**1.1. Mechanisms of anaerobic digestion**

**Figure 3.** World oil and fat production in 1990 and 2011 [3–5].

232 Biological Wastewater Treatment and Resource Recovery

**Kinetic Model Equation 1 Equation 2**

*<sup>U</sup>* <sup>=</sup> *<sup>U</sup>*max <sup>×</sup> *<sup>S</sup>* \_\_\_\_\_\_\_\_\_ *Y*(*B* × *X* + *S* )

*<sup>U</sup>* <sup>=</sup> *<sup>μ</sup>*max <sup>×</sup> *<sup>S</sup>* \_\_\_\_\_\_\_\_\_\_\_\_\_ *<sup>Y</sup> <sup>K</sup> So* <sup>+</sup> (1 <sup>−</sup> *<sup>K</sup>* ) *<sup>S</sup> <sup>Y</sup>*

*ks* <sup>+</sup> *<sup>S</sup>* \_\_1

compounds to volatile fatty acids (VFA) and CO2

25–45% carbon dioxide (CO2

Monod *<sup>U</sup>* <sup>=</sup> \_\_\_\_ *<sup>k</sup> <sup>S</sup>*

Contois

Chen & Hashimoto

group of microorganisms.

methane and carbon dioxide gas.

to acetate and H2

In anaerobic degrading of POME, biogas is formed when microorganisms, especially bacteria, degrade organic material in the absence of oxygen. Biogas consists of 50–75% methane (CH4

*<sup>U</sup>* <sup>=</sup> *<sup>K</sup>*\_\_*<sup>s</sup> K*( \_\_1 *<sup>S</sup>* ) +\_\_1 *k*

> *<sup>μ</sup>*max <sup>×</sup> *<sup>S</sup>* <sup>+</sup> *<sup>Y</sup>*(1 <sup>+</sup> *<sup>a</sup>* ) \_\_\_\_\_\_ *μ*max

*<sup>μ</sup>*max *<sup>S</sup>* <sup>+</sup> *<sup>Y</sup>*(1 <sup>−</sup> *<sup>K</sup>* ) \_\_\_\_\_\_ *μ* max

\_\_1 *<sup>U</sup>* <sup>=</sup> \_\_\_\_\_\_\_ *<sup>a</sup>* <sup>×</sup> *<sup>X</sup>*

\_\_1 *<sup>U</sup>* <sup>=</sup> *<sup>Y</sup> <sup>K</sup> <sup>S</sup>* \_\_\_\_\_*<sup>o</sup>*

**Table 1.** Mathematical expressions of specifics substrate utilization rates for known kinetic models.

matic representation of anaerobic degradation of organic matter is given in **Figure 4**. The AD process can be subdivided into the following four phases, each requires its own characteristic

The sequence of reactions involved in the mechanisms of AD is hydrolysis, acidogenesis, acetogenesis and methanogenesis [28]. Hydrolysis is conversion of nonsoluble biopolymers to soluble organic compounds. Acidogenesis is summarized as a conversion of soluble organic

[29]. Methanogenesis represents conversion of acetate and CO2

) and small amounts of other gases [25–27]. A simplified sche-

while acetogenesis is the conversion of VFAs

[22]

[23]

[24]

The following parameters were analyzed: COD, BOD, pH, VSS and TSS.

Methane gas was determined by gas chromatography with a stainless steel column (200 × 0.3 cm) packed with active carbon (30–60 mesh) using thermal conductivity detection. For TSS, VSS, volatile fatty acids and alkalinity were determined according to the standard methods [31]. The COD was measured using a Hach colorimetric digestion method (Method # 8000, Hach Company, and Loveland, CO, USA). The MLSS and MLVSS were determined by drying the sample at 105 and 550 ± 50°C.

**Figure 5.** Experimental setup.

#### **2.4. Bioreactor operation**

The steady-state performance of ultrasonic membrane anaerobic system (UMAS) was evaluated under different influent COD concentrations (70,400–90,200 mg/l), hydraulic retention time (HRTs) (500.8–14.7 days) and OLR of 1.5–9.0 kg COD/m3 /day (**Table 2**). In this study, the system was considered to have achieved steady state when the operating and control parameters were within ±10% of the average value. The produced biogas contained only CO2 and CH4 , so the addition of sodium hydroxide solution (NaOH) to absorb CO2 effectively isolated methane gas (CH4 ). **Table 2** depicts the results of the application of three known substrate utilization models.


**Table 2.** Summary of results (SS: steady state).

## **3. Results and discussion**

**2.4. Bioreactor operation**

**Figure 5.** Experimental setup.

234 Biological Wastewater Treatment and Resource Recovery

(CH4

CH4

OLR, kg COD/m3

SUR, kg COD/m3

**Table 2.** Summary of results (SS: steady state).

(HRTs) (500.8–14.7 days) and OLR of 1.5–9.0 kg COD/m3

addition of sodium hydroxide solution (NaOH) to absorb CO2

The steady-state performance of ultrasonic membrane anaerobic system (UMAS) was evaluated under different influent COD concentrations (70,400–90,200 mg/l), hydraulic retention time

was considered to have achieved steady state when the operating and control parameters were

). **Table 2** depicts the results of the application of three known substrate utilization models.

within ±10% of the average value. The produced biogas contained only CO2

**Steady state (SS) 1 2 3 4 5 6** COD feed, mg/L 70,400 73,478 76,200 83,570 86,700 90,200 COD permeate, mg/L 1197 1617 3048 3343 4508 6494 Gas production (L/day) 290 310 340 400 480 540 Total gas yield, L/g COD/day 0.48 0.53 0.58 0.67 0.78 0.81 % Methane 81 78.5 75.6 73.8 68.6 64.6

 yield, l/g COD/day 0.39 0.54 0.57 0.60 0.64 0.70 MLSS, mg/L 13,800 12,400 13,400 14,800 17,648 22,600 MLVSS, mg/L 10,269 10,751 11,765 13,320 15,530 20,159 % VSS 74.41 86.70 87.80 90.00 88.00 89.20 HRT, day 500.8 60.6 22.6 14.7 11.20 8.6 SRT, day 300 250 180 30.5 20.30 15.80

SSUR, kg COD/kg VSS/day 0.164 0.195 0.252 0.263 0.294 0.314

Percent COD removal (UMAS) **98.3 97.8 96 96.0 94.8 92.8**

/day 1.0 3.5 6.0 8.5 11.0 15

/day 0.023 0.724 2.225 4.576 5.685 7.347

/day (**Table 2**). In this study, the system

and CH4

effectively isolated methane gas

, so the

#### **3.1. The performance of ultrasonic membrane anaerobic system (UMAS)**

The operating conditions for the ultrasonic membrane anaerobic system (UMAS) over the 500 day experimental setup are given in **Table 2**. The performance evaluation of the integrated ultrasonic membrane anaerobic system (UMAS) was generated at different influent COD concentrations and hydraulic retention times (HRTs). **Table 3** depicted and summarized the kinetic coefficients. For the system results at influent COD concentrations from 70,400 to 90,200 mg/l and pH (6.7–7.8), UMAS was performed well. The mixed liquor volatile suspended solids (MLVSSs) for the first steady state were 10,400 mg/l, whereas the mixed liquor suspended solids (MLSSs) were 13,800 mg/l, equivalent to 75.36% of the MLSS. This low result can be explained due the palm oil mill effluent wastewater contains very high suspended solids.

The volatile suspended solid (VSS) fraction in the reactor at sixth steady state was increased to 89.20%. Results have shown that the long solid retention time (SRT) of UMAS facilitated the decomposition of the suspended solids and their subsequent conversion to methane (CH4 ); these findings found by Nagano et al. [32] and Abdurahman et al. [33]. At organic loading rate, OLR of 15 kg COD/m3 /day, the system registered the highest influent of COD 90,200 mg/l at this stage; the UMAS achieved 92.8% COD removal. **Figures 6**–**8** shown that UMAS can be applied and treat POME efficiently. Among the three models applied, the Monod and Chen


**Table 3.** Summary of the three known substrate utilization models application.

**Figure 6.** The Monod model.

and Hashimoto models performed better, shown that UMAS reactor performance should consider organic loading rates. These two models suggested that the predicted permeate COD concentration (S) is a function of influent COD concentration (S<sup>o</sup> ).

The percentage removal of COD by UMAS at various HRTs was shown in **Figure 9**. It was observed that COD removal efficiency increased as HRT increased from 8.6 to 500.8 days and it was in the range of 92.8–98.3%. It was found that this value higher than the 85% COD removal is observed for POME wastewater treatment using anaerobic fluidized bed reactors [34] and the 91.7–94.2% removal is observed for palm oil mill effluent wastewater treatment using membrane anaerobic system [35], and the 93.6–97.5% removal is observed for POME treatment using membrane anaerobic system [33]. Interestingly, it was found that there is no much difference in COD removal efficiency between HRTs of 500.8 days (98.3%) and 14.7 days

**Figure 7.** The Contois model.

**Figure 8.** The Chen and Hashimoto model.

and Hashimoto models performed better, shown that UMAS reactor performance should consider organic loading rates. These two models suggested that the predicted permeate

The percentage removal of COD by UMAS at various HRTs was shown in **Figure 9**. It was observed that COD removal efficiency increased as HRT increased from 8.6 to 500.8 days and it was in the range of 92.8–98.3%. It was found that this value higher than the 85% COD removal is observed for POME wastewater treatment using anaerobic fluidized bed reactors [34] and the 91.7–94.2% removal is observed for palm oil mill effluent wastewater treatment using membrane anaerobic system [35], and the 93.6–97.5% removal is observed for POME treatment using membrane anaerobic system [33]. Interestingly, it was found that there is no much difference in COD removal efficiency between HRTs of 500.8 days (98.3%) and 14.7 days

).

COD concentration (S) is a function of influent COD concentration (S<sup>o</sup>

**Figure 6.** The Monod model.

236 Biological Wastewater Treatment and Resource Recovery

**Figure 7.** The Contois model.

(96.0%). On the other hand, the COD removal efficiency has declined at shorter hydraulic retention time; at HRT of 8.6 days, the COD removal efficiency was reduced to 92.8%. **Table 2** results show that UMAS result might because of grown of volatile fatty acids inside the reactor. Usually, the hydraulic retention times were mainly effected by the ultrafiltration (UF) membrane influx rates, which directly determined the volume of influent (POME) that can be fed to the reactor.

#### **3.2. Evaluation of UMAS biokinetic coefficients**

The evaluated biokinetic coefficients based on COD basis by UMAS were analyzed as shown in **Table 2**.

**Figure 9.** COD removal efficiency of UMAS under steady-state conditions with various hydraulic retention times.

The kinetic coefficients were calculated and summarized in **Table 3**. The growth yield coefficient, Y, value ranges from 0.32 to 0.68 gm VSS/gm COD, specific microorganic decay rate, b, and maximum substrate utilization rate, K, ranges from 0.350 to 0.374 COD/g VSS.day. **Figure 10** depicts the relationship between the substrate utilization rates (SUR) and the specific substrate utilization rate for COD with various hydraulic retention times. The HRTs range from 8.6 to 500.8 days. The biokinetic coefficients of growth yield, Y, and specific microorganic decay rate, b, were calculated from the slope and intercept as shown in **Figures 11** and **12**. The evaluated maximum specific biomass growth rates, μmax, range from 0.248 to 0.474 day−1.

**Figure 10.** The specific substrate utilization rate for COD with various hydraulic retention times.

**Figure 11.** Evaluation of the growth yield, Y, and the specific biomass decay rate, b.

**Figure 12.** Evaluation of the maximum specific substrate utilization and the saturation constant, K.

#### **4. Production of methane (CH4 ) and carbon dioxide (CO2 ) gases**

A semicontinuous operation was conducted to verify the performance of the integrated ultrasonic membrane anaerobic system (UMAS) throughout a different hydraulic retention times (HRTs) and influent COD concentrations. In this study, the influent COD concentration was increased from 70,400 to 90,200 mg/l (for the six steady states). **Figure 13** illustrates the gas production rate and the methane content of the biogas. It was clear that the methane CH4 yield decreased with increasing OLRs. Methane gas contents were varied from 64.6 to 81%, and the methane yield was varied from 0.39 to 0.70 CH4 /g COD/day. The decreased CH4 yield with increasing OLR was also

**Figure 13.** Gas production and methane content.

The kinetic coefficients were calculated and summarized in **Table 3**. The growth yield coefficient, Y, value ranges from 0.32 to 0.68 gm VSS/gm COD, specific microorganic decay rate, b, and maximum substrate utilization rate, K, ranges from 0.350 to 0.374 COD/g VSS.day. **Figure 10** depicts the relationship between the substrate utilization rates (SUR) and the specific substrate utilization rate for COD with various hydraulic retention times. The HRTs range from 8.6 to 500.8 days. The biokinetic coefficients of growth yield, Y, and specific microorganic decay rate, b, were calculated from the slope and intercept as shown in **Figures 11** and **12**. The evaluated maximum specific biomass growth rates, μmax, range from 0.248 to

**Figure 11.** Evaluation of the growth yield, Y, and the specific biomass decay rate, b.

**Figure 10.** The specific substrate utilization rate for COD with various hydraulic retention times.

0.474 day−1.

238 Biological Wastewater Treatment and Resource Recovery

noted in many previous studies [36–40]. One of the reasons might be that shorter HRT of the system contributed to more active methanogens that were washed out during the removal of effluent. The gas production has increased from 290 to 540 L per day during the study. Biogas production increased with increasing OLRs from 0.48 l/g COD/day at 1.0 kg COD/m3 /day to 0.81 l/g COD/ day at 15 kg COD/m3 /day. These findings are in line with the results obtained from Refs. [41–43].

## **5. Conclusions**

The kinetic performance of newly designed ultrasonic membrane anaerobic system (UMAS) was evaluated in the treatment of palm oil mill effluent (POME).

The steady-state performance of ultrasonic membrane anaerobic system (UMAS) was evaluated under different influent COD concentrations (70,400–90,200 mg/l), hydraulic retention times (HRTs) (500.8–14.7 days) and OLR of 1.5–9.0 kg COD/m3 /day.

Among the three models applied, the Monod and Chen and Hashimoto models performed better, shown that UMAS reactor performance should consider organic loading rates. These two models suggested that the predicted permeate COD concentration (S) is a function of influent COD concentration (S<sup>o</sup> ).

It was observed that COD removal efficiency increased as HRT increased from 8.6 to 500.8 days, and it was in the range of 92.8–98.3%. The evaluated maximum specific biomass growth rates, μmax, range from 0.248 to 0.474 day−1.

It was found that the methane CH4 yield decreased with increasing OLRs. Methane gas contents were varied from 64.6 to 81%, and the methane yield was varied from 0.39 to 0.70 CH4 /g COD/day.

## **Author details**

Abdurahman Hamid Nour<sup>1</sup> \* and Azhari Hamid Nour2

\*Address all correspondence to: nour2000\_99@yahoo.com

1 Faculty of Chemical and Natural Resources Engineering, University Malaysia Pahang-UMP, Pekan, Pahang, Malaysia

2 Faculty of Pure and Applied Sciences International University of Africa, Khartoum, Sudan

## **References**

[1] Jundika, C. K., Sachin, V. J., Saad, A., Agus, P. S., Arun, S. M. (2016). Advances in biofuel production from oil palm and palm oil processing wastes: a review. *Biofuel Research Journal, 9,* 332–346.

[2] Hansen, S. B., Padfield, R., Syayuti, K., Evers, S., Zakariah, Z., Mastura, S. (2015). Trends in global palm oil sustainability research. *Journal of Cleaner Production, 100*, 140–149.

noted in many previous studies [36–40]. One of the reasons might be that shorter HRT of the system contributed to more active methanogens that were washed out during the removal of effluent. The gas production has increased from 290 to 540 L per day during the study. Biogas production

The kinetic performance of newly designed ultrasonic membrane anaerobic system (UMAS)

The steady-state performance of ultrasonic membrane anaerobic system (UMAS) was evaluated under different influent COD concentrations (70,400–90,200 mg/l), hydraulic retention

Among the three models applied, the Monod and Chen and Hashimoto models performed better, shown that UMAS reactor performance should consider organic loading rates. These two models suggested that the predicted permeate COD concentration (S) is a function of

It was observed that COD removal efficiency increased as HRT increased from 8.6 to 500.8 days, and it was in the range of 92.8–98.3%. The evaluated maximum specific biomass

contents were varied from 64.6 to 81%, and the methane yield was varied from 0.39 to

1 Faculty of Chemical and Natural Resources Engineering, University Malaysia Pahang-UMP,

2 Faculty of Pure and Applied Sciences International University of Africa, Khartoum, Sudan

[1] Jundika, C. K., Sachin, V. J., Saad, A., Agus, P. S., Arun, S. M. (2016). Advances in biofuel production from oil palm and palm oil processing wastes: a review. *Biofuel Research* 

\* and Azhari Hamid Nour2

/day. These findings are in line with the results obtained from Refs. [41–43].

/day.

yield decreased with increasing OLRs. Methane gas

/day to 0.81 l/g COD/

increased with increasing OLRs from 0.48 l/g COD/day at 1.0 kg COD/m3

was evaluated in the treatment of palm oil mill effluent (POME).

times (HRTs) (500.8–14.7 days) and OLR of 1.5–9.0 kg COD/m3

).

growth rates, μmax, range from 0.248 to 0.474 day−1.

\*Address all correspondence to: nour2000\_99@yahoo.com

day at 15 kg COD/m3

240 Biological Wastewater Treatment and Resource Recovery

**5. Conclusions**

influent COD concentration (S<sup>o</sup>

/g COD/day.

Abdurahman Hamid Nour<sup>1</sup>

Pekan, Pahang, Malaysia

*Journal, 9,* 332–346.

**References**

0.70 CH4

**Author details**

It was found that the methane CH4


[31] APHA. (2005). Standard Methods for the Examination of Water and Wastewater. American Public Health Association, New York.

[17] Ismail, S. (2005). Membrane separation technology for palm oil mill effluent (POME) treatment: an integrated approach. Ph.D. thesis, Universiti Sains Malaysia. 149-206. [18] Thani, M. I., Hussin, R., Ibrahim, W. W. R., Sulaiman, M. S. (1999). Industrial Processes and the Environment: Crude Palm Oil Industry. Handbook No. 3, Department of

[19] Najafpour, G. D., Zinatizadeh, A. A. L., Mohamed, A. R., Isa, M. H., Nasrollahzadeh, H. (2006). High-rate anaerobic digestion of palm oil mill effluent in an up-flow sludgefixed

[20] Mahabot, S., Harun, A. R. (1986). Effluent treatment system for FELDA's palm oil mill.

[21] Oswal, N., Sarma, P. M., Zinjarde, S. S, Pant, A. (2002). Palm oil mill effluent treatment

[22] Monod, J. (1949). Growth of bacteria cultures. *Annual Review of Microbiology*, *3*, 371–394. [23] Contois, D. E. (1959). Kinetics of bacteria growth: relationship between population density and space growth rate of continuous cultures. *The Journal of General and Applied* 

[24] Chen, Y. R., Hashimoto, A. G. (1980). Substrate utilization kinetic model for biological

[25] Rahayu, A. S., Karsiwulan, D., Yuwono, H., Trisnawati, I., Mulyasari, S., Rahardjo, S., Hokermin, S., Paramita, V. (2015). Handbook POME-to-Biogas Project Development in Indonesia. In: B. Castermans, et al. Eds. 2nd ed., Winrock International, United States of

[26] Youngsukkasem, S., Barghi, H., Rakshit, S. K., Taherzadeh, M. J. (2013). Rapid biogas production by compact multi-layer membrane bioreactor, efficiency of synthetic poly-

[27] Yunus, A., Zahira, Y., Parul, A., Kamaruzzaman, S. (2015). Production of biogas and performance evaluation of existing treatment processes in palm oil mill effluent (POME).

[28] Gerardi, M. H. (2003). The Microbiology of Anaerobic Digester. Wiley-Interscience, New

[29] Mes, T. Z. D., Stams, A. J. M., Reith, J. H., Zeeman, G. (2003). Methane production by anaerobic digestion of wastewater and solid wastes. In: J. H. Reith*,* et al. (Eds.), *Biomethane and Biohydrogen: Status and Perspectives of Biological Methane and Hydrogen Production*.

[30] Girard, M., Palacios, J. H., Belzile, M., Godbout, S., Pelletier, F. (2013). Biodegradation in Animal Manure Management. In: Dr. R. Chamy (Ed.) *Biodegradation – Engineering and Technology* Retrieved from http://www.intechopen.com/books/biodegradation-engineering-and-technology/biodegradation-in-animal-manure-management. doi:10.5772/56151

Sustainable, Sanitation and Water Management, Switzerland, pp. 58–94.

treatment processes. Biotechnology and Bioengineering, *22*, 2081–2095.

Environment, Kuala Lumpur, pp. 7–54.

242 Biological Wastewater Treatment and Resource Recovery

film bioreactor. *Process Biochemistry*, *41*, 370–379.

Technical Service Department, Kuala Lumpur.

meric membranes. *Energies, 6*, 6211–6224.

*Renewable and Sustainable Energy Reviews, 42*, 1260–1278.

*Microbiology*, *21*, 40–50.

America. 8-19.

Jersey, pp. 51–57.

by a tropical marine yeast. *Bioresource Technology*, *85*, 35–37.


## *Edited by Robina Farooq and Zaki Ahmad*

Biological treatment of wastewater is a low-cost solution for remediation of wastewater. This book focuses on the bioremediation of wastewater, its management, monitoring, role of biofilms on wastewater treatment and energy recovery. It emphasizes on organic, inorganic and micropollutants entering into the environment after conventional wastewater treatment facilities of industrial, agricultural and domestic wastewaters. The occurrence of persistent pollutants poses deleterious effects on human and environmental health. Simple solution for recovery of energy as well as water during biological treatment of wastewater is a viable option. This book provides necessary knowledge and experimental studies on emerging bioremediation processes for reducing water, air and soil pollution.

Photo by Alejandra Estrada / iStock

Biological Wastewater Treatment and Resource Recovery

Biological Wastewater

Treatment and Resource

Recovery

*Edited by Robina Farooq and Zaki Ahmad*