Wastewater Engineering: Advanced Wastewater Treatment Systems

© IJSR Publications

Wastewater Engineering: Advanced

Wastewater Treatment Systems

ii

Wastewater Engineering: Advanced Wastewater

Treatment Systems

ISSN: 2322-4657

DOI 10.12983/1-2014-03-01

© IJSR Publications, Penang, Malaysia,

This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of

the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation,

broadcasting, reproduction on microfilms or in any other physical way, and transmission or information

storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology

now known or hereafter developed. Exempted from this legal reservation are brief excerpts in connection

with reviews or scholarly analysis or material supplied specifically for the purpose of being entered and

executed on a computer system, for exclusive use by the purchaser of the work. Duplication of this

publication or parts thereof is permitted only under the provisions of the Copyright Law of the Publisher’s

location, in its current version, and permission for use must always be obtained from IJSR Publications.

Permissions for use may be obtained through RightsLink at the Copyright Clearance Center. Violations

are liable to prosecution under the respective Copyright Law. The use of general descriptive names,

registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence

of a specific statement, that such names are exempt from the relevant protective laws and regulations and

therefore free for general use. While the advice and information in this book are believed to be true and

accurate at the date of publication, neither the authors nor the editors nor the publisher can accept any

legal responsibility for any errors or omissions that may be made. The publisher makes no warranty,

express or implied, with respect to the material contained herein.

IJSR Publications; www.ijsrpub.com

http://www.ijsrpub.com/

iii

PREFACE

As the global population grows and many developing countries modernize, the importance of

water supply and wastewater treatment becomes a much greater factor in the welfare of nations.

Clearly, in today’s world the competition for water resources coupled with the unfortunate

commingling of wastewater discharges with freshwater supplies creates additional pressure on

treatment systems. Recently, researchers focus on wastewater treatment by difference methods

with minimal cost and maximum efficiency.

This volume of the Wastewater Engineering: Advanced Wastewater Treatment Systems is a

selection of topics related to physical-chemical and biological processes with an emphasis on

their industrial applications. It gives an overview of various aspects in wastewater treatments

methods including topics such as biological, bioremediation, electrochemical, membrane and

physical-chemical applications. Experts in the area of environmental sciences from diverse

institutions worldwide have contributed to this book, which should prove to be useful to

students, teachers, and researchers in the disciplines of wastewater engineering, chemical

engineering, environmental engineering, and biotechnology. We gratefully acknowledge the

cooperation and support of all the contributing authors.

Hamidi Abdul Aziz

Amin Mojiri

Professor, School of Civil Engineering,

Engineering Campus, Universiti Sains Malaysia,

[email protected]

Research Assistant, School of Civil Engineering,

Engineering Campus, Universiti Sains Malaysia,

[email protected]

mailto:[email protected]


iv

TABLE OF CONTENTS

PREFACE ............................................................................................................................... iii

TABLE OF CONTENTS ........................................................................................................ iv

CHAPTER 1: INTRODUCTION OF PRELIMINARY AND SECONDARY

TREATMENTS ...................................................................................................................... 1

1.1 Introduction of preliminary and Secondary Treatments; Z. Amirossadat .............. 2

CHAPTER 2: WASTEWATER BY TREATMENT BY PHYSICAL-CHEMICAL

TECHNOLOGIES ............................................................................................................... 05

2.1 Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent

Prepared From Waste Material; N. Azmi, J.K. Bashir, S. Sethupathi, C.A. Ng ......... 06

2.2 Removal of Colour from Synthetic Dye Wastewater Using Adsorbent Prepared from

Psyllium Husk; I. Dahlana and S.M.O. Tayeh ............................................................. 15

2.3 COD and BOD Removal from Textile Wastewater Using Naturally Prepared

Adsorbents and Their Activation forms Using Sulphuric Acid; Patel and Vashi ........ 31

2.4 Fenton oxidation for the Treatment of Liquid Waste with High COD and

Anionic/Non-ionic Surfactants; M. Collivignarelli, S. Sorlini, A. Abbà, M. Sordi ..... 41

2.5 Ultrasound Irradiation on the Treatment of Aromatic Compounds in Wastewater

W.L. Peng, G. Xinxin, M.J.K. Bashir .......................................................................... 48

CHAPTER 3: WASTEWATER TREATMENT BY BIOLOGICAL METHODS ........ 62

3.1 Wastewater Treatment by Biological Methods; A. Dadrasnia, N. Shahsavari and

C.U. Emenike ............................................................................................................... 63

3.2 Biological Treatment of Recycled Paper Mill Wastewater Using Modified Anaerobic

Inclining-Baffled (MAIB) Bioreactor; H.M. Zwaina and I. Dahlan ............................ 71

3.3 Augmentation of Biological Nitrogen Removal via Optimization of Support Media

Size and Aeration Strategy in Moving Bed Sequencing Batch Reactor; J.Wei Lim, M.J.K.

Bashir, S.L. Ng, S. Sethupathi, L.P. Wong. ................................................................. 87

v

CHAPTER 4: ELECTROCHEMICAL METHODS ........................................................ 96

4.1 Electrochemical Oxidation Process Contribution in Remediating Complicated

Wastewaters; M.J. K. Bashir, J.W. Lim, S.Q. Aziz, S.S.A. Amr ................................ 97

CHAPTER 5: WASTEWATER TREATMENT BY BIOREMEDIATION

TECHNOLOGIES ............................................................................................................. 107

5.1 Wastewater Treatment by Bioremediation Methods; A.N. Amenaghawon and K.O.

Obahiagbon ................................................................................................................ 108

5.2 Supplementation of Novel Solid Carbon Source Prepared from Dried Attached-

Growth Biomass for Bioremediation of Wastewater Containing Nitrogen; J.W. Lim,

M.J.K. Bashir, C.A. Ng, X. Guo. ............................................................................... 125

CHAPTER 6: WASTEWATER TREATMENT BY MEMBRANE TECHNIQUES . 136

6.1 Supported Liquid Membrane in wastewater Treatment; T.T. Teng, A. Talebi, and G.

Muthuraman ............................................................................................................... 137

6.2 Role of Emulsion Liquid Membrane (ELM) in Separation Processes; T.T. Teng, M.

Soniya, G. Muthuraman and A. Talebi ...................................................................... 149

6.3 Bulk Liquid Membrane and its Applications in Wastewater Treatment; T.T. Teng, S.

Elumalai, G. Muthuraman and A. Talebi ................................................................... 158

6.4 Challenges in Fabricating Suitable Membrane for Water Treatment Application;

L.Y. Wong, C.A. Ng, MJ.K. Bashir, T.L. Chew ....................................................... 171

6.5 Removal of Copper from Aqueous Solution by Using Thermo-responsive Polymer

Hydrogel as Copper Chelator via Polymer-enhanced Ultrafiltration (PEUF); J.J. Chen,

A.L. Ahmad and B.S. Ooi .......................................................................................... 183

CHAPTER 7: WASTEWATER TREATMENT BY PHYTOREMEDIATION

TECHNOLOGIES ............................................................................................................. 193

7.1 Wastewater Treatment by Phytoremediation Methods; H. Farraji ..................... 194

CHAPTER 8: LANDFILL LECHATE TREATMENT TECHNIQUES...................... 207

8.1 Municipal Landfill Leachate Treatment Techniques: An Overview; S.Q. Aziz, H.A.

Aziz, M.J.K. Bashir, A. Mojiri .................................................................................. 208

vi

CHAPTER 9: APPLICATION OF OPTIMIZATION IN TREATMENT ................... 225

9.1 Application of Optimization in Wastewater Treatment; Y.L. Lim, Y.C. Ho, A.F.M.

Alkarkhi ..................................................................................................................... 226

Wastewater Engineering: Advanced Wastewater Treatment Systems

Available online at http://www.ijsrpub.com/books

©2014 IJSRPUB

1


Treatment Systems

Chapter 1: Introduction of preliminary and

Secondary Treatments



©2014 IJSRPUB

2

Introduction of preliminary and Secondary Treatment

Zahra Amirossadat

Isfahan (Khorasgan) Branch, Islamic Azad University, Isfahan, Iran

Abstract. Recently, the amounts of wastewater are sharply increasing and the kinds of pollutants are also varied as the world

wide industry is being developed incessantly. With respect to both the quantity and composition, the textile processing

wastewater is recorded as the most polluted source among all industrial sectors. This chapter explained the preliminary and

secondary treatment of wastewater.

Keywords: Preliminary treatment, Secondary treatment, Wastewater

1. INTRODUCTION

Recently, the amounts of wastewater are sharply

increasing and the kinds of pollutants are also varied

as the world wide industry is being developed

incessantly. With respect to both the quantity and

composition, the textile processing wastewater is

recorded as the most polluted source among all

industrial sectors (Chang et al., 2009). At wastewater

treatment plants, wastewater is treated before it is

allowed to be returned to the environment, lakes, or

streams.

Discharge criteria required the installation of

facilities that performed what is now called primary

treatment of wastewater. This involved using screens

and sedimentation tanks to remove most of the

materials in the wastewater that float or settle. As

subsequent discharge criteria were tightened,

secondary treatment became necessary. Secondary

treatment is accomplished by bringing together waste,

bacteria and oxygen in trickling filters or the activated

sludge process. Bacteria are used to consume the

organic parts of the wastewater. Facilities and their

designers are now considering and installing tertiary

treatment facilities to comply with the latest

regulatory and permit parameters. These advanced

treatment processes go beyond conventional

secondary treatment and include the removal of

recalcitrant organic compounds, as well as excess

nutrients such as nitrogen and phosphorus (Coppen,

2004). Conventional wastewater treatment consists of

a combination of physical, chemical, and biological

processes and operations to remove solids, organic

matter and, sometimes, nutrients from wastewater.

General terms used to describe different degrees of

treatment, in order of increasing treatment level, are

preliminary, primary, secondary, and tertiary and/or

advanced wastewater treatment. In some countries,

disinfection to remove pathogens sometimes follows

the last treatment step.

2. Primary treatment

The objective of primary treatment is the removal of

settleable organic and inorganic solids by

sedimentation, and the removal of materials that will

float (scum) by skimming. Approximately 25 to 50%

of the incoming biochemical oxygen demand (BOD5),

50 to 70% of the total suspended solids (SS), and 65%

of the oil and grease are removed during primary

treatment. Some organic nitrogen, organic

phosphorus, and heavy metals associated with solids

are also removed during primary sedimentation but

colloidal and dissolved constituents are not affected.

The effluent from primary sedimentation units is

referred to as primary effluent. In many industrialized

countries, primary treatment is the minimum level of

reapplication treatment required for wastewater

irrigation. It may be considered sufficient treatment if

the wastewater is used to irrigate crops that are not

consumed by humans or to irrigate orchards,

vineyards, and some processed food crops. However,

to prevent potential nuisance conditions in storage or

flow-equalizing reservoirs, some form of secondary

treatment is normally required in these countries, even

in the case of non-food crop irrigation. It may be

possible to use at least a portion of primary effluent

for irrigation if off-line storage is provided. Primary

sedimentation tanks or clarifiers may be round or

rectangular basins, typically 3 to 5 m deep, with

hydraulic retention time between 2 and 3 hours.

Settled solids (primary sludge) are normally removed

from the bottom of tanks by sludge rakes that scrape

the sludge to a central well from which it is pumped to

sludge processing units. Scum is swept across the tank

surface by water jets or mechanical means from which

it is also pumped to sludge processing units.

(http://www.fao.org/docrep/t0551e/t0551e05.htm).

Primary treatment involves:

http://www.fao.org/docrep/t0551e/t0551e05.htm

Amirossadat

Advanced Wastewater Treatment

3

1. Screening- to remove large objects, such as

stones or sticks that could plug lines or block tank

inlets.

2. Grit chamber- slows down the flow to allow grit

to fall out

3. Sedimentation tank (settling tank or clarifier)-

settleable solids settle out and are pumped away,

while oils float to the top and are skimmed off

(http://www.sd1.org/resourcehandler.aspx?id=28).

2. Secondary treatment

Secondary wastewater treatment is the second stage of

wastewater treatment that takes place after the

primary treatment process. The process consists of

removing or reducing contaminants or growths that

are left in the wastewater from the primary treatment

process. Usually biological treatment is used to treat

wastewater in this step because it is the most effective

type of treatment on bacteria, or contaminant, growth.

Secondary treatment processes can remove up to 90

percent of the organic matter in wastewater by using

biological treatment processes. The two most common

conventional methods used to achieve secondary

treatment are attached growth processes and

suspended growth processes

(http://www.water.siemens.com/en/applications/waste

water_treatment/secondary-treatment).

Secondary treatment typically utilizes biological

treatment processes, in which microorganisms convert

nonsettleable solids to settleable solids. Sedimentation

typically follows, allowing the settleable solids to

settle out. Three options include:

1. Activated Sludge- The most common option

uses microorganisms in the treatment process to break

down organic material with aeration and agitation,

then allows solids to settle out. Bacteria-containing

“activated sludge” is continually recirculated back to

the aeration basin to increase the rate of organic

decomposition.

2. Trickling Filters- These are beds of coarse

media (often stones or plastic) 3-10 ft. deep.

Wastewater is sprayed into the air (aeration), then

allowed to trickle through the media. Microorganisms

attached to and growing on the media, break down

organic material in the wastewater. Trickling filters

drain at the bottom; the wastewater is collected and

then undergoes sedimentation.

3. Lagoons- These are slow, cheap, and relatively

inefficient, but can be used for various types of

wastewater. They rely on the interaction of sunlight,

algae, microorganisms, and oxygen (sometimes

aerated).

After primary and secondary treatment, municipal

wastewater is usually disinfected using chlorine (or

other disinfecting compounds, or occasionally ozone

or ultraviolet light). An increasing number of

wastewater facilities also employ tertiary treatment,

often using advanced treatment methods.

Tertiary treatment may include processes to

remove nutrients such as nitrogen and phosphorus,

and carbon adsorption to remove chemicals. These

processes can be physical, biological, or chemical.

Settled solids (sludge) from primary treatment and

secondary treatment settling tanks are given further

treatment and undergo several options for disposal

(http://www.sd1.org/resourcehandler.aspx?id=28).

The objective of secondary treatment is the further

treatment of the effluent from primary treatment to

remove the residual organics and suspended solids. In

most cases, secondary treatment follows primary

treatment and involves the removal of biodegradable

dissolved and colloidal organic matter using aerobic

biological treatment processes. Aerobic biological

treatment (see Box) is performed in the presence of

oxygen by aerobic microorganisms (principally

bacteria) that metabolize the organic matter in the

wastewater, thereby producing more microorganisms

and inorganic end-products (principally CO2, NH3,

and H2O). Several aerobic biological processes are

used for secondary treatment differing primarily in the

manner in which oxygen is supplied to the

microorganisms and in the rate at which organisms

metabolize the organic matter.

High-rate biological processes are characterized by

relatively small reactor volumes and high

concentrations of microorganisms compared with low

rate processes. Consequently, the growth rate of new

organisms is much greater in high-rate systems

because of the well-controlled environment. The

microorganisms must be separated from the treated

wastewater by sedimentation to produce clarified

secondary effluent. The sedimentation tanks used in

secondary treatment, often referred to as secondary

clarifiers, operate in the same basic manner as the

primary clarifiers described previously. The biological

solids removed during secondary sedimentation,

called secondary or biological sludge, are normally

combined with primary sludge for sludge processing.

Common high-rate processes include the activated

sludge processes, trickling filters or biofilters,

oxidation ditches, and rotating biological contactors

(RBC). A combination of two of these processes in

series (e.g., biofilter followed by activated sludge) is

sometimes used to treat municipal wastewater

containing a high concentration of organic material

from industrial sources.

(a) Activated Sludge

In the activated sludge process, the dispersed-growth

reactor is an aeration tank or basin containing a

http://www.sd1.org/resourcehandler.aspx?id=28

http://www.water.siemens.com/en/applications/wastewater_treatment/secondary-treatment

http://www.water.siemens.com/en/applications/wastewater_treatment/secondary-treatment

http://www.sd1.org/resourcehandler.aspx?id=28


Chapter 1: Introduction of preliminary and Secondary Treatments

4

suspension of the wastewater and microorganisms, the

mixed liquor. The contents of the aeration tank are

mixed vigorously by aeration devices which also

supply oxygen to the biological suspension. Aeration

devices commonly used include submerged diffusers

that release compressed air and mechanical surface

aerators that introduce air by agitating the liquid

surface. Hydraulic retention time in the aeration tanks

usually ranges from 3 to 8 hours but can be higher

with high BOD5 wastewaters. Following the aeration

step, the microorganisms are separated from the liquid

by sedimentation and the clarified liquid is secondary

effluent. A portion of the biological sludge is recycled

to the aeration basin to maintain a high mixed-liquor

suspended solids (MLSS) level. The remainder is

removed from the process and sent to sludge

processing to maintain a relatively constant

concentration of microorganisms in the system.

Several variations of the basic activated sludge

process, such as extended aeration and oxidation

ditches, are in common use, but the principles are

similar.

(b) Trickling Filters

A trickling filter or biofilter consists of a basin or

tower filled with support media such as stones, plastic

shapes, or wooden slats. Wastewater is applied

intermittently, or sometimes continuously, over the

media. Microorganisms become attached to the media

and form a biological layer or fixed film. Organic

matter in the wastewater diffuses into the film, where

it is metabolized. Oxygen is normally supplied to the

film by the natural flow of air either up or down

through the media, depending on the relative

temperatures of the wastewater and ambient air.

Forced air can also be supplied by blowers but this is

rarely necessary. The thickness of the biofilm

increases as new organisms grow. Periodically,

portions of the film 'slough off the media. The

sloughed material is separated from the liquid in a

secondary clarifier and discharged to sludge

processing. Clarified liquid from the secondary

clarifier is the secondary effluent and a portion is

often recycled to the biofilter to improve hydraulic

distribution of the wastewater over the filter.

(c) Rotating Biological Contactors

Rotating biological contactors (RBCs) are fixed-film

reactors similar to biofilters in that organisms are

attached to support media. In the case of the RBC, the

support media are slowly rotating discs that are

partially submerged in flowing wastewater in the

reactor. Oxygen is supplied to the attached biofilm

from the air when the film is out of the water and from

the liquid when submerged, since oxygen is

transferred to the wastewater by surface turbulence

created by the discs' rotation. Sloughed pieces of

biofilm are removed in the same manner described for

biofilters

(http://www.fao.org/docrep/t0551e/t0551e05.htm).

REFERENCES

Chang W, Tran H, Park D, Zhang R, Ahn D (2009).

Ammonium nitrogen removal characteristics

of zeolite media in a Biological Aerated Filter

(BAF) for the treatment of textile wastewater.

Journal of Industrial and Engineering

Chemistry, 15: 524-528.

Cppen J (2004) Advanced Wastewater Treatment

Systems. Courses ENG4111 and 4112

Research Project, University of Southern

Queensland, Faculty of Engineering and

Surveyin.

http://www.fao.org/docrep/t0551e/t0551e05.htm



©2014 IJSRPUB

5


Treatment Systems

Chapter 2: Wastewater Treatment by Physical-

Chemical Technologies



©2014 IJSRPUB

6

Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent

Prepared From Waste Material

Nurshazwani Binti Azmi, Mohammed J.K. Bashir*, Sumathi Sethupathi, Choon-Aun Ng

Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul

Rahman, 31900 Kampar, Perak, Malaysia

*Corresponding Author: [email protected]; Tel: 605-4688888 ext: 4559; Fax: 605-4667449

Abstract. Landfill leachate has become the subject of recent research interest as it is a strongly polluted wastewater. The

produced leachate is one of the most important drawbacks of municipal solid waste disposal (MSW) in sanitary landfill.

Adsorption by activated carbon (AC) appears to have considerable potential in landfill leachate treatment due to the simplicity

design, superior removal of organic compound and less land area required. However, the high demand for AC is a major

problem due to limited carbon based substances such as coal, wood and lignite. Therefore, waste material seems to be a good

option as an alternatives source of AC. Consequently, this paper focuses on effectiveness of using AC in landfill leachate

treatment and highlighted the recent development treating landfill leachate using adsorbent prepared from waste material.

Keywords: Landfill Leachate, Treatment, Adsorbent

1. INTRODUCTION

As the exponential population and social civilization

growth, together with the developments of industries

and technologies, rapid generation of MSW has

becomes a global environmental problem (Saeed et

al., 2009). There are many options available for MSW

disposal such as sanitary landfill, open dump,

incineration, composting, grinding, hog feeding,

milling, and anaerobic digestion (Aziz et al., 2010).

Sanitary landfill is the most common MSW disposal

method due to such advantages as the simple disposal

procedure, low cost, and landscape-restoring effect on

holes from mineral workings (Bashir et al., 2010).

However, the production of highly contaminated

leachate is a major drawback of this method (Aziz et

al., 2010).

Landfill leachate is defined as any contaminated

liquid effluent percolating through deposited waste

and emitted within a landfill or dump site through

external sources (Taulis, 2005). In a more precise

definition, it is a soluble organic and mineral

compound formed when water infiltrates into the

refuse layers, which extracts a series of contaminants

and instigates a complex interplay between the

hydrological and biogeochemical reactions that acts as

a mass transfer mechanisms for producing of moisture

content sufficiently high to start the liquid flow (Aziz

et al., 2004). As shown in Figure 1, leachate

generation induced by the gravity force, precipitation,

irrigation, surface runoff, rainfall, snowmelt,

recirculation, liquid waste co-disposal, refuse

decomposition, groundwater intrusion and initial

moisture content present within the landfills

(Achankeng, 2004). As the consequences, leachate

may contain high concentration of organic matter

(biodegradable and non-biodegradable), ammonia

nitrogen, heavy metals, chlorinated organic and

inorganic salts (Renou et al., 2008). Without an

appropriate treatment, landfill leachate could be a

potential source of surface and groundwater

contamination, as it could seep into soils and subsoil,

causing severe pollutions to receiving water body

(Oman and Junestedt, 2008).

Typically, leachate characteristics and

compositions depends on various factors such as

waste composition, age of landfill, site hydrogeology,

specific climate conditions, moisture routing through

the landfill, and the landfill design and operation

(Ghafari et al., 2010). Age of landfill site is one of the

main variables that affect the leachate characteristics

(Bashir et al., 2012), where the concentration of

leachates parameters changes with the age of the

leachate. Young acidogenic landfill leachate

commonly characterized by high biochemical oxygen

demands(BOD) and chemical oxygen demands(COD)

, high concentration of ammonium nitrogen followed

by low pH value as low as pH 4 (Wu et al., 2001). The

degradation of biological matter by microorganism

lead to generation of Volatile Fatty Acid(VFA) that

lead to low ph value and high BOD/COD ratio. On

the contrary, aged landfill (i.e. >10 years old)

produces mature (stabilized) leachate that contains

bio-refractory compounds such as humic acid (HA)

and fulvic acids (FA), with BOD5/COD ratio less than

0.1 (Alvariz-Vazqurez et al., 2004) as illustrated in

Table 1.

Biological treatment of landfill leachates have been

shown to be very effective in removing organic matter

in early stages (Berruetta and Castrillon, 1992) with

high BOD/COD ratio. As the BOD/COD ratio

decrease with the passage of time (Rodriguez et al.,

2000), the biodegradable organic content of leachate

reduced where biological treatment no longer

effective due to the presence of refractory organic

matter and physico-chemical processes may become


http://www.sciencedirect.com/science/article/pii/S0011916410003516#bib1

Azmi et al.

Recent Development in Landfill Leachate Treatment Using Low Cost Adsorbent Prepared From Waste Material

7

one of the appropriate options for stabilized landfill

leachate. Various physico-chemical process have been

practiced for old landfill leachate treatments such as

adsorption (Halim et al., 2012), ion-exchange (Bashir

et al., 2012), Fenton reaction (Mohajeri et al., 2011),

coagulation/flocculation (Ghafari et al., 2010),

electrochemical oxidation process (Marco et al., 2013)

ozonation (Salem et al., 2013), and air stripping

(Bloor and Banks, 2005). Among all process,

adsorption technology is one of the most applicable

and simple methods.

Adsorption is defined as a mass transfer process by

which a substance is transferred from liquid or gas

phases to the solid surface of adsorbent and form

attachment via physical or chemical interactions. The

material providing the solid surface is called the

adsorbent and material removed from the liquid phase

is called as adsorbate. AC demonstrated significant

adsorption efficiency in gas and liquid phases due to

its high micropore volume, large specific surface area,

favorable pore size distribution, thermal stability, and

capability for rapid adsorption and low acid/base

reactivity (Li et al., 2009). The unique adsorptive

properties of AC, makes it as one of the best filtration

media in the world. However, high manufacturing

cost and expensive carbonaceous material for

producing high quality AC (Mohan and Pittman,

2006) lead to limitation of this application for landfill

leachate treatment especially in developing countries.

Thus, the use of non-conventional material such as

agriculture waste and industrial by-product that are

locally available can be chemically modified and

utilized as a low carbon adsorbent (Babel and

Kurniawan, 2003). Several studies have been

conducted by using AC for various types of waste

water. Consequently, the present work reviews and

evaluates the recent published works focuses on

landfill leachate treatment using adsorbent prepared

from the waste materials.

Fig. 1: Leachate Formation (Agamuthu, 2001)


Chapter 2: Wastewater Treatment by Physical-Chemical Technologies

8

Table 1: Classification of landfill leachate (Alvarez-Vazquez et al., 2004) Parameters Young Intermediate Stabilized

Age (years) <5 5-10 >10

pH <6-5 6-5-7-5 >7.5

COD (mg L-1) >10 000 4000-10 000 <4000

BOD5/COD 0.5-1 0.1-0.5 <0.1

TOC/COD <0.3 0.3-0.5 >0.5

NH-N <400 NA >400

Heavy metal (mg L-1) Low to

medium

Low Low

Organic compound 80%

VFA

5-30%

VFA+ HA+ FA

HA+FA

Biodegradability Important Medium Low

Kjeldahl Nitrogen (g L-1) 0.1-0.2 NA NA

2. PREPARATION OF ACTIVATED CARBON

In general, the process for manufacturing of AC

involves two steps, carbonization of the raw material

followed by the activation of the carbonized product.

The activation process can be carried out in two ways,

either by physical activation (PhA) or chemical

activation (Cha). Pretreatment of the raw material

normally conducted before carbonization process.

Few preliminary stages involves such as crushing,

milling and sieving for appropriate particle size are

important for subsequent handling of the raw material

(Alslaibi et al., 2013). During carbonization, the raw

material will undergo thermal decomposition in inert

atmosphere through gasification by Nitrogen gas.

Carbonization process can be carried out using tubular

furnaces, reactors, muffle furnace and more recently

in a glass reactor placed in a modified microwave

oven (Foo and Hameed, 2011). This process known as

pyrolysis where the non-carbon elements such

hydrogen, nitrogen and oxygen will be released

leaving a rigid carbon skeleton with a rudimentary

pore structure. Based on previous studies, pyrolysis of

lignocellulosic material such as olive stones, coconut

shells or olive shells will remove most of non-carbon

elements in the form of char, tar (oil) and gas where

the proportion depends on the parameter during

pyrolysis such as temperature of pyrolysis, nitrogen

flow rate and heating rate. According to Mohammed

et al., (2010), flash pyrolysis will give high liquid

production, while slow heating rate recommended for

high yields of the char residue.

Physical activation is normally made by

carbonization followed by activation in partial or

control gasification at high temperature (Rodriguez-

Reiniso and Molina-Sabio, 1992). Activation

completed through gasification using oxidizing agent

(Table 2) or mixture of it at temperature ranging from

700-1100 o

C to develop the porosity (micropores and

mesopores) of the carbonized material (Rodrigues-

Reinoso and Molina-Sabio, 1992). CO2 can develop

narrow micropores, while steam will widen the initial

micropores. Accordingly, CO2 produced larger

volume and narrower micropores (Mohamed et al.,

2010) while steam created a larger volume of

mesopores and micropores.

Meanwhile in ch(A), carbonization and activation

can be run simultaneously, where the raw material

will be impregnated with chemical agent (Table 2)

followed by conventional heating at moderate

temperature between 400-800oC (Demiral and

Gunduzoglu, 2010). The chemical agent acts as

deactivating agent where it inhibits formation of tar

and develops AC porosity via dehydration and

degradation. The pore distribution and surface area is

determined by the impregnation ratio between the

mass of precursor and the chemical agent chosen. The

resulting AC produced is then washed with distilled

water or mild acid in order to remove residual

chemicals from the material. Nowadays, chemical

activation is widely applied for the activation because

of its lower activation temperature and higher product

yield compared with the physical one (Guo and

Rockstraw, 2007). Based on the literature reviewed, it

was found that alkaline hydroxides (KOH and NaOH)

(Lillo-Rodenas et al., 2007) can be used to prepare the

activated carbon, which gave high specific surface

area in the range of 2318–3500 m2/g (Tseng, 2006).

http://www.sciencedirect.com/science/article/pii/S026387621000184X#bib10



Azmi et al.


9

Table 2: Common activating agent for different activation method in preparation of AC (Alslaibi et al., 2013)

Some additional studies combine the physical

activation with chemical activation where it is known

as physicochemical activation. In general,

physicochemical activation is performed by changing

the activation atmosphere of the chemical activation

by a gasification atmosphere (steam, CO2) at higher

temperatures or the chemical activation is carried out

directly under the presence of a gasifying agent. The

combination of both types of carbon activation makes

ACs with textural and chemical properties which are

different from those obtained by any of the activations

alone. For example, steam reduces the occurrence of

heteroatoms into the carbon structures. Also, the

combination of oxidizing reagents in the liquid phase

(nitric or sulfuric acids) with gasification agents

improves the development of porosity on the final

carbons. In addition, continued development of

activated carbon lead to the new application of

microwave heating in preparing activated carbons.

Recently, the usage of microwave heating as an

alternative to conventional heating in ch(A) receives

considerable attention in many research works.

Microwave synthesis is an alternative technique that

overcomes the problems of conventional fast firing

because microwave synthesis is a non-contact

technique where the heat is transferred to the product

via electromagnetic waves, and large amounts of heat

can be transferred to the interior of the material,

minimizing the effects of differential synthesis (Jones

et al., 2002). This leads to uniform heating, rapid

temperature rise and saving energy (Zhang et al.,

2011). In addition, the short period of treatment time,

reduced extra processing cost by reduction in energy

consumption (Xin-Hui et al., 2011) and gas

consumption within the process (Foo and Hameed,

2009). Besides, previous study illustrate relatively

higher surface areas developed during AC preparation

by microwave heating compared to conventional

heating for the same precursor (Alsalibi et al., 2013).

Since carbonaceous materials are a good microwave

absorber, microwave assisted thermal process turned

to interesting alternative approach in preparation of

AC. The experimental apparatus used in ch(A) by

microwave heating in the laboratory scale is shown in

Figure 2.

3. PERFORMANCE OF AC

Performance of AC is largely depends on two

parameters e.g. the surface area and the pore structure

(Figueiredo et al., 1999). The larger surface area will

contribute to higher adsorption capacity. Although a

microporous AC is generally desired for adsorption

purposes, the presence of mesopores is also valuable

for the adsorption of large molecules or where a faster

adsorption rate is required (Huang et al., 2011).

Meanwhile, the macropores have larger pore volume

and act as access pores that provide a passageway to

the particles. The pore volume limits the size of

particles that can be access while the surface area

limits the amount of material that can be adsorbed.

According to the International Union of Pure and

Applied Chemistry (IUPAC), porosity within the

adsorbent has been categorized according to width.

The largest macropores (access pores) generally have

width between 50nm-100nm, the mesopores (transport

pores), between 2-50nm wide and the smallest pores,

micropores (high adsorption properties) have width <

2nm. Figure 3 shows the three types of AC pore

structure: micropores,mesopores and macropores.



10

Fig. 2: A schematic diagram of the experimental apparatus used in activation process (Deng et al., 2009)

4. RECENT DEVELOPMENTS IN LANDFILL

LEACHATE TREATMENT VIA ADSORBENTS

FROM WASTE MATERIALS

In the recent years, adsorption technology has been

applied in landfill leachate treatment by using low

cost adsorbents prepared from waste materials,

particularly agricultural waste, as precursors for AC

production. As adsorption by AC offer number of

advantages such as not getting affected by toxic

compound, superior removal of organic contaminants

(Weber, 1978), high degree of porosity and very well

developed surface area (Sevilla et al., 2011), it has

turned to a powerful standard technique for removing

toxic and non-biodegrable metal ion (Foo and

Hameed, 2011). The effectiveness of adsorption

process by physical and chemical activation method is

presented in Table 3.

Study conducted by Foo et al., (2013a) showed

good removal of NH3-N and orthophosphate by

79.63% and 85.06% via adsorption, where

adsorptive removal increased as adsorbent dosage

increased from 0.5g/200ml to 5.0g/200ml. The

experimental data revealed the feasibility of sugarcane

bagasse as developed adsorbent for leachate treatment

with micropore volume 0.515cm3/g and micropore

surface area 659.25 mg/cm2. In another study

conducted by Foo et al., (2013b), the potential of the

banana front as the precursor in AC production was

attempted. The experimental data showed that, with

increasing of adsorbent dosage, from 0.5g/200ml to

3.5g/200ml and from 0.5g/200ml to 4.0mg/200ml at

30oC, the adsorptive removal of Boron and Iron

increased. The greatest adsorptive removal of boron

and total Iron is by 97.45% and 95.14% with

maximum monolayer adsorption capacity at 11.09 and

26.15 mg/g. Meanwhile, tamarind fruit seed derived

granular activated carbon showed higher adsorptive

removal of colour (91.22%) while COD (79.93%).

The optimum adsorbent dosage in this experiment is

6.0g/200ml at temperature 30oC. The adsorption of

COD and color onto TSAC was best described by the

Langmuir isotherm model, with a monolayer

adsorption capacity of 64.93 mg/g and 168.57 Pt–

Co/g, respectively. The results indicated that the

tamarind fruit seed derived granular activated carbon

adsorption process was sufficient for removing these

targeted contaminants, to reach to the industrial

discharge limit of COD (500 mg/L) permitted by the

World Health Organization (WHO).Kamarudin et al.,

(2012) presented the physical activation of durian

peels for AC preparation, with 41.98% and 39.86%

removal of color and COD with optimum activation

temperature at 800oC, activation time 2.1H and CO2

flow rate 68.68 mg/L. In another study conducted by

Ching et al., (2011), impregnated coffee ground with

Azmi et al.


11

H2SO4 by chemical activation via microwave heating,

they found that the optimum impregnation ratio was

0.5 and 2.5 with a 10g AC dosage at pH 8.1, with

removal of iron by 77% and PO4-P by 84%. In the

case of rice husk, Kalderis et al., (2008) observed that

the micropore volume of Rice Husk Activated Carbon

produced was 0.42cm3/g with chemical activation by

ZnCl2 via microwave heating. The percentage removal

of COD was by 60% and colour by 70% with

optimum AC dosage at 30g/L.

Fig. 3: Diagram of Activated Carbon (Source: Sushrut Chemicals, 2013)

Table 3: Application of adsorbent prepared from various waste materials in treating landfill leachate



12

Based on the previous experimental works, waste

material especially agriculture waste has high

potential to be converted to AC and then utilized in

treating hazardous landfill leachate. The percentage of

COD, colour, and heavy metals such as Iron,

orthophosphate and Boron reduced and meet the

requirement set by the officially authorized prior

discharge to the surface water. However, only small

numbers of research works conducted for landfill

leachate treatment by using waste material as

adsorbent. Thus, more studies should be conducted to

understand the process of low-cost landfill leachate

treatment through adsorption process by AC.

5. OPPORTUNITIES AND CHALLENGES

Recently, many research works have been done to

highlight the potential use of waste material as a

precursor for landfill leachate treatment. Higher

availability with lower cost makes waste material such

as agriculture waste and industrial waste as an

alternative choice of precursor. Furthermore, with

favorable characteristic as an adsorbent (high porosity

and large surface area), waste materials are

comparable with the typical precursor (coal, wood,

lignite).

Although the precursor selection depends on their

availability, cost and purity, but the manufacturing

process and the application of the product are also

important considerations (Yavuz et al., 2010). In

addition, low social acceptability and lack of

economical affordability, especially in the

management of the treatment such as treatment

facilities, chemical, labor, energy consumption,

transportation, collection and maintenance are among

the main key drivers deciding its flexibility, reliability

and suitable manner.

Thus, corrective and transparent policies, mandates

and standards which governing the collection,

transportation, disposal prevention, recycling, reuse,

monitoring, designing and supervision of solid waste

management should to be pointed out and well-

planned. Besides, the sound of professional in order to

creates environmental awareness for public

participation, adequate financial provisions,

engineering and operating standards, responsibilities

sharing, product stewardship, staff capacities

upgrading formal procedures redressing, regular

opinion survey, site rehabilitation and aftercare

maintenance need to be properly assigned and

counteracted (Bernache, 2003).

6. CONCLUSION

Landfill leachate treatment by adsorption process

using AC has a potential to be practice in sanitary

landfill. Together with simplicity design, less land

area required, alongside with the recycling of waste

material as a low- cost precursor reduces the numbers

of MSW and resolves the landfill leachate problem.

However, full cooperation between communities,

private sectors, local government and states are

required in order to apply this valuable technology

towards sustainable environment.

ACKNOWLEDGEMENTS

The authors are grateful for the financial support

provided by the Universiti Tunku Abdul Rahman

(UTAR) through grant No:

IPSR/RMC/UTARRF/2012-C2/M03 and

IPSR/RMC/UTARRF/2013-C2/T02.

REFERENCES

Achankeng E (2004). Sustainability in municipal solid

waste management in Bamenda and Yaounde,

Cameroon, PhD thesis of Geographical and

Environmental Studies University of Adelaide,

Australia.

Agamuthu P (2001). Solid Waste: Principles and

Management: with Malaysian Case Studies,

Insitute of Biological Sciences, University of

Malaya.

Alvarez-Vazquez H, Jefferson B, Judd SJ

(2004).Membrane bioreactors vs conventional

biological treatment of landfill leachate: a brief

review. Journal of Chemical Technology and

Biotechnology, 79 (10): 1043–1049.

Alslaibi TM, Abustan I, Ahmad MA, Foul AA (2012).

Production of Activated Carbon from

agriculture byproduct via conventional

microwave heating : a review. J. Chem.

Technol. Biotechnol., 88(7): 1183-1190

Aziz HA, Adlan MN, Zahari MSM, Alias S (2004).

Removal of ammoniacal nitrogen (N-NH3)

from municipal solid waste leachate by using

activated carbon and limestone. Waste Manage.

Res, 22: 371–375.

Aziz SQ, Aziz HA, Yusoff MS, Bashir MJK, Umar M

(2010). Leachate characterization in semi-

aerobic and anaerobic sanitary landfills: A

comparative study. Journal of Environmental

Management, 9: 2608-2614.

Babel S, Kurniawan TA (2003). A research study on

Cr (VI) removal from contaminated wastewater

using natural zeolite. J. Ion Exch., 14: 289–292

Bernache G (2003). The environmental impact of

municipal waste management: the case of

Guadalajara metro area, Resour. Conserv.

Recycl, 39: 223–237.

Azmi et al.


13

Beruetta J, Castrillon L (1992). Anaerobic treatment

of leachates in UASB reactors. J. Chem.

Technol. Biotecnol, 54: 33-7.

Bashir MJK, Aziz HA, Yusoff MS, Adlan MN (2010).

Application of response surface methodology

(RSM) for optimization of nitrogen removal

from semi-aerobic leachate using ion exchange

resin. Desalination, 254: 154-161.

Bashir MJK, Aziz HA, Yusoff MS, Aziz SQ (2012).

Investigation of color and COD eliminations

from mature semi-aerobic landfill leachate

using anion-exchange resin: Equilibrium and

kinetic study. Environ. Eng. Sci., 29: 297–305

Blorr MC, Banks CJ (2005). Acute and sublethal

toxicity of lanffill leachate towards aquatic

macro-intervebrates:Demonstarating the

remediation potential of air-stripping.

Environmental International, 32(8): 1114-1122

Chandra TC, Mirna MM, Sunarso J, Sudaryanto Y,

Ismadji S (2009). Activated carbon from durian

shell: preparation and characterization. J.

Taiwan Inst. Chem. Eng, 40: 457–462

Ching SL, Yusoff MS, Aziz HA, Umar M (2011).

Influence of impregnation ratio on coffe ground

activated Carbon as landfill leachate adsorbent

for removal of total iron and orthophosphate.

Desalination, 279: 225-234.

Demiral H, Gunduzoglu G (2010). Removal of nitrate

from aqueous solutions by activated carbon

prepared from sugar beet bagasse. Bioresource

Technol, 101: 1675–1680

Deng H, Yang L, Tao G and Dai J (2009). Preparation

and characterization of activated carbon from

cotton stalk by microwave assisted chemical

activation - application in methylene blue

adsorption from aqueous solution. J. Hazard.

Matter., 166: 1514–1521

Foo KY, Hameed BH (2009). An overview of landfill

leachate treatment via activated carbon

adsorption process. Journal of Hazardous

Material, 171: 54-60.

Foo KY, Hameed BH (2011). The environmental

applications of activated carbon/zeolite

composit materials. Adv. Colloid Interface Sci,

162:22-28

Foo KY, Lee LK, Hameed BH (2013a). Preparation of

activated carbon from sugarcane by microwave

assisted activation for the remediation of semi-

aerobic landfill leachte. Bioresource

Technology, 134: 166-172

Foo KY, Lee LK, Hameed BH (2013b). Preparation

of banana frond activated carbon by microwave

induced activation for the removal of boron and

total iron from landfill leachate. Chemical

Engineering Journal, 223: 604-610

Foo KY, Lee LK, Hameed BH (2013c). Batch

adsorption of semi-anaerobic landfill leachate

by granular activated carbon prepared by

microwave heating. Chemical Engineering

Journal, 222: 259-264

Ghafari G, Aziz HA, Bashir MJK (2010). The use of

poly-aluminum chloride and alum for the

treatment of partially stabilized leachate: A

comparative study. Desalination, 257: 110–116.

Guo Y, Rockstraw RA (2007). Physicochemical

properties of carbons from pecan shell by

phosphoric acid activation.Bioresource

Technol., 98: 1513–1521

Halim A, Aziz HA, Johari MAM, Ariffin KS, Bashir

MJK (2012). Semiaerobic landfill leachate

treatment using carbon minerals composite

adsorbent, Environ. Eng. Sci., 29:306–312

Hameed BH (2009). Spent tea leaves: a new non-

conventional and low-cost adsorbent for

removal of basic dye from aqueous solutions, J.

Hazard. Matter, 161: 753–759

Hejazifar M, Azizian S, Sarikhani H, Li Q, Zhao D (

2011).Microwave assisted preparation of

efficient activated carbon from grapevine

rhytidome for the removal of methyl violet

from aqueous solution, Journal of Analytical

and Applied Pyrolysis,92:258–266.

Huang L, Sun Y, Wang W, Yue Q, Yang T (2011).

Comparative study on characterization

ofactivated carbons prepared by microwave and

conventional heating methods and application

in removal of oxytetracycline (OTC).Chemical

Engineering Journal, 171: 1446–1453.

Jones DA, Lelyveld TP, MavrofidisSD, Kingman SW,

Miles NJ (2002). Microwave heating

applications in environmental engineering: a

review, Resources, Conservation and recycling

34: 75–90.

Jusoh AB, Cheng WH, Low WM, Nora’ aini A, Noor

MJMM (2005).Study on the removal of iron

and manganese in groundwater by granular

activated carbon, Desalination 182:347–353.

Kalderis DS, Dimitricous S, Koutaulakis, P.

Paraskeva, E. Diamadopolous, E.Otal, J.

Olivares del Valle, C. Ferrarez-Pereira (2008).

Adsorption of polluting substances on

Activated Carbon prepared from rice husk and

sugarcane bagasse. 144: 42-50

Kamaruddin MA, Yusof MS, Ahmad MA (2012).

Treatment of seme-aerobic landfill leachate

using during-peel based activated carbon

adsorption, optimization of preparation

condition. IJEE, 3: 223-236)

Lillo-Rodenas MA, Marco-Lozar JP, Cazorla-Amoros

D, Linares-Solano A (2007). Activated carbons

prepared by pyrolysis of mixtures of carbon



14

precursor/alkaline hydroxide J. Anal. Appl.

Pyrol, 80: 166–174

Li W, Peng J, Zhang L, Yang K, Xia H, Zhang S, Guo

SH (2009). Preparation of activated carbon

from coconut shell chars in pilot-scale

microwave heating equipment at 60 Kw. Waste

Management, 29:756–760.

Marco P, Carlos A, Martinez-Huitle (2013). Role of

electrode materials for the anodic oxidation of a

real landfill leachate-Comparison between Ti-

Ru-Sn ternary oxide,PbO2 and Boron doped

diamond anode.Chemosphere, 90(4): 1455-

1460

Mohajeri S, Aziz HA, Zahed MA, Mohajeri L, Bashir

MJK, Aziz SQ, Adlan MN, Isa MH (2011).

Multiple responses analysis and modeling of

Fenton process for treatment of high strength

landfill leachate. Water Sci. Technol, 64:1652–

1660.

Mohamed AR, Mohammadi M., Darzi, GN (2010).

Preparation of carbon molecular sieve from

lignocellulosic biomass: A review. Renewable

and Sustainable Energy Reviews, 14(6):1591-

1599

Mohan D.,Pittman CUJ (2006). Activated carbon and

low cost adsorbents for remediation of tri- and

hexavalent chromium from water. Journal of

Hazardous Materials, B137:762-811.

Oman CB, Junestedt C (2008). Chemical

characterization of landfill leachate-400

parameter and compound.Waste Management,

28(10): 1876-1891

Renou S, Givaudan JG, Poulain S, Dirassauyan F,

Moulin P (2008). Landfill Leachate Treatment:

Review & Opportunity. J Hazard Mater, 150:

468-493.

Rodriguez IJ, Castrillon PL, Maranon ME., Sastre AH

(2000). Biomethanisation of municipal solid

waste in a pilot plant. Water Res., 4(2):447-54.

Rodríguez-Reinoso F, Molina-Sabio, M. (1992).

Activated carbons from lignocellulosic

materials by chemical and/or physical

activation: an overview. Carbon, 30(7): 1111-

1118.

Saeed MO, Hassan MN, Mujeebu MA (2009).

Developement of municipal solid waste

generation and recyclable components rate of

Kuala Lumpur : Perspective study. Waste

Mangement, 9: 2209.

Abu Amr S.S, Aziz HA, Adlan MN Bashir MJK

(2013). Pretreatment of stabilize leachate using

ozone/persulfate oxidation. Chem. Eng. Journsl,

221: 492-499

Sevilla M, Fuertes AB, Mokaya R (2011). Preparation

and hydrogen storage capacity of highly porous

activated carbon materials derived from

polythiophene. Int. J. Hydrog. Energy, 36:

15658–15663

Taulis M (2005). Metal contaminants in leachate from

sanitary landfills', Metal Contaminants in New

Zealand: Sources, Treatments, and Effects on

Ecology and Human Health, pp. 119-38.

Tseng RL (2006). Mesopore control of high surface

area NaOH-activated carbon. J.

Coll.Sci.Imp.U.Tok, 303:494-502

Wu JJ, Wu CC, Ma HW, Chang CC, Masten SJ

(2001). Treatment of landfill leachate by ozone-

based advanced oxidation process. In:

Proceeding of the 15th IOA world congress,

London,UK: 339-34.

Xin-hui D, Srinivasakannan C, Jin-hui P, Li-bo Z,

Zheng-yong Z (2011). Comparison of activated

carbon prepared from Jatropha hull by

conventional heating and microwave heating.

Biomass Bio-energy, 35:3920–3926.

Yavuz R, Akyildiz H, Karatepe N, C Etinkaya E

(2010). Influence of preparation conditions on

porous structures of olive stone activated by

H3PO4. Fuel Process Technol., 91:80–87.

Zhang G, Zhang L, Deng Hand Sun P (2011).

Preparation and characterization of sodium

carboxymethyl cellulose from cotton stalk using

microwave heating. J. Chem. Technol.

Biotechnol, 86:584–589.



©2014 IJSRPUB

15

Removal of Colour from Synthetic Dye Wastewater Using Adsorbent Prepared from

Psyllium Husk

Irvan Dahlan1*

, Somaia M.O. Tayeh2

1School of Chemical Engineering, Universiti Sains Malaysia, Engineering Campus, Seri Ampangan, 14300 Nibong Tebal,

Pulau Pinang, Malaysia. 2School of Civil Engineering, Universiti Sains Malaysia, Engineering Campus, Seri Ampangan, 14300 Nibong Tebal, Pulau

Pinang, Malaysia.

*Corresponding Author: [email protected]

Abstract. Color removal from textile effluents has been given much attention in the last few years by the adsorption process

using low cost adsorbents. In this study, four different adsorbents were prepared from psyllium husk (i.e. PH/CFO-Al,

PH/CFO-Ac, QPH and NPH) for removal of direct blue dye (DB71) from synthetic wastewater. From the preliminary study,

PH/CFO-Ac adsorbent gave the highest color removal efficiency. PH/CFO-Ac adsorbent was prepared from psyllium husk and

CoFe2O4 by a facile refluxing route in acidic solution. PH/CFO-Ac adsorbent was selected to be used for further batch studies

to examine the effect of various experimental parameters, i.e. contact time, adsorbent amount, shaking rate, initial dye

concentration, pH and temperature. The best conditions for dye removal using PH/CFO-Ac adsorbent were obtained at pH 9.0,

temperature of 30°C, shaking rate of 150 rpm and contact time of 2 hours. The adsorption kinetics was found to follow pseudo-

second-order kinetic model. The experimental data fitted well with the Langmuir model with a monolayer adsorption capacity

of 188.7mg/g. It was also found that prepared and spent PH/CFO-Ac adsorbents have a homogenous particle size distribution.

In addition, the surface morphology of the spent PH/CFO-Ac adsorbents had more compact structures with small granular

particles attached on the surface.

Keywords: Colour removal, Synthetic Dye Wastewater, Psyllium Husk

1. INTRODUCTION

Due to the increase in the world population and

development of industrial applications, environmental

pollution problem became very important, especially

wastewater pollution problem. Communities produce

both liquid and solid wastes. The liquid waste, i.e.

wastewater, is essentially the water supply of the

community after it has been used in a variety of

applications. Wastewater handling, disposal &

treatment are serious worldwide problem. Many

industrial and agricultural activities use water in an

excessive way. However, it is now well known that

the fresh water resources are limited and fragile, so

they must be protected. Discharge of sanitary

wastewater, industrial effluent and agricultural field’s

runoff can be the main source of freshwater pollution.

This causes many diseases for human, and it is known

that 70-80% of illness in developing countries is

related to water contamination, particularly for

children and women (WHO/UNICEF, 2000).

Textile industries consume large volumes of water

and chemicals for wet processing of textiles. The

chemical reagents used are very diverse in chemical

composition, ranging from inorganic compound to

polymers and organic compound (Correia et al.,

1994). The color is an evident indicator of water

pollution by the dyes. Industrial dye effluents are

visible even at concentrations lower than 1 mg/l.

Moreover, some dyes and their degradation products

are carcinogenic (Ahn et al., 1999). Also, some dyes

are harmful to aquatic life in rivers where they are

discharged. Since, dye can reduce light penetration

into the water thereby decreasing the efficiency of

photosynthesis in aquatic plants and hence having

adverse impact on their growth (Che Ani, 2004).

Textile wastewater is generally high in both color

and organic content. Effluents discharged from dyeing

industries are highly colored and they can be toxic to

aquatic life in receiving waters (Lee et al., 1999,

Kadirvelu et al., 2003). Color removal from textile

effluents has been given much attention in the last few

years, not only because of its potential toxicity, but

mainly due to its visibility problems (Morais et al.,

1999). The total dye consumption of the textile

industry worldwide is in excess of 107 kg/year, and an

estimated 90% of this ends up on fabrics.

Consequently, 1000 tonnes/year or more of dyes are

discharged into waste streams by the textile industry

worldwide (Ahmad et al., 2007).

Development of the appropriate techniques for

treatment of dye wastewater is important for the

protection of natural water. To eliminate dyes from

aqueous colored effluents and reduce their ecological

consequences, several biological and chemical

techniques have been proposed: anaerobic/aerobic

degradation (Ahmed et al., 2007),

coagulation/flocculation (Papić et al., 2000) and also

oxidative/reductive chemical and photochemical

processes (Lucas and Peres, 2006). Due to relatively

high operating costs and low removal efficiencies

using the above-mentioned processes, textile, pulp and

Dahlan and Tayeh

Removal of Colour from Synthetic Dye Wastewater Using Adsorbent Prepared from Psyllium Husk

16

paper industries seldom apply these to treat their

effluents.

Among several chemical and physical methods, the

adsorption has been found to be superior to other

techniques in water reuse methodology because of its

capability for adsorbing a broad range of different

types of adsorbates efficiently, and simplicity of

design. Many researchers researched for cheaper

substitutes, which are relatively inexpensive, and are

at the same time endowed with reasonable adsorptive

capacity. These studies include the use of coal, fly

ash, activated clay, palm-fruit bunch, Bagasse pith,

cellulose-based waste, peat, bentonite, slag and fly

ash, rice husk, activated sludge, etc (Ahmad et al.,

2007).

Psyllium husk has not been investigated as

adsorbent for color removal from dye solutions. This

research studied the adsorption for color removal from

synthetic dye wastewater using an adsorbent prepared

from an inexpensive and readily available material,

i.e. psyllium husk. Also, the study aims to achieve the

following measureable objectives:

1- To prepare and characterize adsorbent from

psyllium husk using quaternized and magnetic

methods.

2- To investigate the ability of the best psyllium

husk adsorbent for removal of color from synthetic

dye under various operating conditions (initial dye

concentration, amount of sorbent, shaking rate,

contact time, pH, and temperature).

3- To determine the kinetic behavior and

isotherms for the adsorption process of color onto

psyllium husk adsorbent.

2. MATERIALS AND METHODS

2.1. Materials

The raw psyllium husk (Fig. 1) was obtained directly

from a store in Parit Buntar, Penang, and used in the

preparation of the adsorbent for this study. Direct blue

dye (DB71) was provided by Sigma – Aldrich, Co.,

and used without further purification. The dye was

used as the adsorbate in the batch experimental study.

The chemicals used in this study are Na2CO3 and

Ferric nitrate [Fe(NO3)3·9H2O] which is provided by

Bendosen Laboratory Chemicals, whereas NaOH,

H2SO4 and Cobalt nitrate [Co(NO3)2·6H2O] are

provided by Qrëc, Bright Chem. Sdn Bhd. In addition,

N-(3-chloro-2-hydroxy-propyl)-trimethylammonium

chloride was provided by Aldrich Chemistry. All

chemicals were used without any purification process.

Fig. 1: Raw psyllium husk (PH).

2.2. Preparation of Adsorbent

2.2.1. Preparation of raw psyllium husk

The raw psyllium husk was dried in the oven at 100oC

for 24 hours, and then ground by using domestic

blender to pass through a 1mm-sieve (Laboratory Test

Sieve). The husk obtained was kept in a closed plastic

and labeled as raw psyllium husk (RPH).

2.2.2. Preparation of Quaternized Adsorbent

About 20g of the raw psyllium husk was treated with

250 ml of 1% w/v Na2CO3 solution (Fig. 2a). The

mixture was shaken for 45 min at room temperature

(Fig. 2b). It was subsequently rinsed several times

with distilled water (Fig. 2c) and dried at 60°C (Fig.

2d). The husk obtained was labeled as natural

psyllium husk (NPH). A portion of NPH was

quaternized according to the method reported by Low

and Lee (1997). The natural psyllium husk (10g) in

12.5 ml of 5 M NaOH solution was left at room

temperature for 30 min (Fig. 2e). At the end of

incubation period 10 ml of 4 M N - (3-chloro-2-

hydroxy-propyl) - trimethylammonium chloride was

added to the mixture (Fig. 2f). It was thoroughly left

in the oven at 60-70°C for 4 hours (Fig. 2g) with

intermittent stirring. The reaction mixture was then

rinsed several times with water and finally with



17

distilled water (Fig. 2h). The product after drying was

labeled as quaternized psyllium husk (QPH).

2.2.3. Preparation of psyllium husk/CoFe2O4

adsorbent (PH/CFO)

Psyllium husk/CoFe2O4 adsorbent was synthesized by

a facile refluxing route in alkaline solution as reported

by Ai et al., (2010) with some modifications. In a

typical procedure, 8g of raw psyllium husk was added

into a 150mL alkaline solution containing 3.4g NaOH,

and stirred at room temperature for 30 minutes (Fig.

3a) to get the psyllium husk suspension. The

suspension was then maintained at 100°C to keep

boiling state (Fig. 3b). A 50mL metal nitrate aqueous

solution was prepared by dissolving Fe(NO3)3·9H2O

(5.4944 g) and Co(NO3)2·6H2O (1.9790 g) in distilled

water (Fig. 3c). The solution was poured as quickly as

possible into the above boiling suspension (Fig. 3d).

The mixture solution was then refluxed at 100°C for 2

h (Fig. 3e). By a simple magnetic procedure, the

resulting product was separated from water, dried at

80°C for 12 h (Fig. 3f), and the product was labeled as

PH/CFO-Al. This preparation method was repeated

twice, the second time was by replacing the 150 mL

alkaline solution NaOH by 150ml acidic solution

H2SO4 (2N ) to obtain PH/CFO-Ac.

Table 1: Summary of batch studies

Preliminary Batch study Main Batch study

Using four types of adsorbent prepared from

psyllium husk, to select the best adsorbent with the

highest efficiency to be used in the main study, the

effect of two parameters were studied:

a- Effect of adsorbent amount (using 1g

and 3g adsorbent)

b- Effect of initial dye concentration (using

5mg/ L, 15mg/L)

Main experiments for synthetic wastewater

a. Effect of contact time using 1hr, 2hr, 3hr, 4hr and 5hr

b. Effect of adsorbent amount using 0.1g, 0.3g, 0.6g, 1g

and 1.5g

c. Effect of shaking rate using 100rpm, 150rpm,

200rpm, 250rpm and 350rpm.

d. Effect of initial dye concentration using 15, 30, 60,

100 and 200mg/L

e. Effect of pH using pH 3, 4.5, 6, 7.5, 9 and10.5

f. Effect of temperature using 30, 40, 50 and 60oC

2.4. Isotherm and Kinetic studies

For the isotherm and kinetic studies, one batch

experiment was conducted using 1000ml of synthetic

wastewater with initial dye concentration of 200mg/L

and pH 9. About 1.0g of adsorbent was added to the

sample, it was agitated at the agitation rate of 150

rpm. Samples were taken at different time intervals (0

– 60 min), and the final concentration of the dye was

measured at each time t. At time t = 0 and

equilibrium, the dye concentration was measured and

the amount of adsorption at equilibrium, qe (mg/g)

was calculated using Equation

where Co and Ce (mg/L) are the liquid-phase

concentrations of sample at initial and equilibrium,

respectively. W (g) is the mass of sorbent used and V

(L) is the volume of the solution. The removal

efficiency of dye can be calculated from Equation 2

where Ct is the dye concentration at time t.

Adsorption isotherm is fundamentally essential to

explain how solutes interact with sorbents, and is

critical in optimizing the use of sorbents. In the

present study, the equilibrium isotherms were

analyzed using the Langmuir and Freundlich,

isotherms. The linear form of Langmuir isotherm is

given in Equation 3 (Keleşoğlu, 2007) and Freundlich

isotherm is given in Equation 4 (Schwarzenbach et

al., 2003),

where the constant Qo signifies the adsorption

capacity (mg/g) and b is related with the energy of the

adsorption (L/mg), KF and n are Freundlich constants.

To investigate the adsorption mechanism, pseudo-

first-order and pseudo-second-order kinetic models

were tested to find the best fitted model for the

experimental data. The pseudo-first-order equation is

given by Equation 5.

where k1 is the pseudo-first-order rate constant

(min−1

), qe and qt are the amounts of dye adsorbed

(mg/g ) at equilibrium and at time t (min). The

pseudo-second-order model can be expressed in

Equation 6 (Ho and McKay, 1998):

Dahlan and Tayeh


18

where k2 (g mg

−1 min

−1 ) is the rate constant of the

pseudo-second-order adsorption.

2.5. Analysis and Characterization

The initial concentration of dye (direct blue 71 dye

solution) and samples after sorption treatment were

measured using DR 2800 spectrophotometer

according to the Platinum – Cobalt Standard Method

(Adopted from Standard Methods for the Examination

of Water and wastewater and NCASI, Technical

Bulletin No. 253, Dec. 1971). All the concentrations

of sample analysis were conducted in triplicate to

increase the precision of the results, and only the

average value was reported throughout this analysis.

Selected magnetic sorbents, before and after the

adsorption process, were characterized using

Mastersizer 2000 in order to study the particle size

distribution for the adsorbent. To obtain the surface

morphologies of the selected adsorbent, before and

after the adsorption process, the scanning electron

microscopy (SEM) examinations were performed with

5 kV of accelerating voltage using Leo Supra 35VP-

24-58 microscope.

3. RESULTS AND DISCUSSIONS

3.1. Preliminary Batch Study

As described before, four types of adsorbents were

prepared to select the best adsorbent for the main

batch experimental studies. Figure 4(a) shows the

results for the adsorbent selection based on dye

concentration, and Figure 4(b) shows the results for

the adsorbent selection based on adsorbent amount. It

was noticed from Figure 4(a) that the PH/CFO-Ac

which is the magnetic adsorbent prepared from

psyllium husk treated with acidic solution (H2SO4) has

the best removal efficiency with 79% color removal

from the 5mg/L initial DB71 dye concentration, and

88% color removal from the 15mg/L initial dye

concentration, while the other kinds of adsorbent

(PH/CFO-Al, QPH and NPH), gave negative

efficiency due to the color of dye became more

darker.

It was noticed also from Figure 4 (b) that the

PH/CFO-Ac, has the best removal efficiency, with

88.6%% color removal when 1g adsorbent was used

and 73% color removal when 3g adsorbent was used,

while the other three kinds of adsorbent gave negative

efficiency due to the color of dye became more

darker. Based on the results of the preliminary study,

PH/CFO-Ac was selected to be used for the main

study.

3.2. Batch Experiments for Synthetic Wastewater

3.2.1. Effect of contact time

One of the most important parameters affecting the

adsorption process is the contact time. It is supposed

that the color removal efficiency will increase by the

increasing of the contact time, until reaching a time at

which no significant amount of dye is removed; this

time is called the equilibrium time. Figure 5 shows the

effect of contact time on the adsorption of DB71 using

PH/CFO-Ac adsorbent

It was shown in Figure 5 that the removal

efficiency of PH/CFO-Ac increased rapidly from

90.8% to 94.03 % when the contact time was

increased from 1 to 2 hours. After that, no significant

increase was noticed in the removal efficiency with

increasing the contact time. This is probably due to

larger surface area of the PH/CFO-Ac adsorbent being

available at the beginning of the adsorption process,

so the adsorption rate was fast. As the surface

adsorption sites become exhausted, the uptake rate is

controlled by the rate at which the adsorbate is

transported from the exterior to the interior sites of the

adsorbent particles (Ai et al., 2010, Kumar et al.,

2010).



19

Fig. 2: Preparation of quaternized adsorbent

Dahlan and Tayeh


20

Fig. 3: Preparation of psyllium husk/CoFe2O4 adsorbent (PH/CFO)



21

Fig. 4: (a) Adorbent selection based on dye concentration, and (b) adsorbent selection based on adsorbent amount.

Fig. 5: Effect of contact time on the adsorption of DB71 from synthetic wastewater

3.2.2. Effect of adsorbent amount

Figure 6 shows the plot of amount of PH/CFO-Ac

adsorbent against percentage removal efficiency of

DB71. It was observed that the removal efficiency is

varied with varying amount of adsorbent and it

decreased with increasing the amount of adsorbent.

The removal efficiency decreased from 94.78% to

81.1% for an increase in adsorbent amount from

0.1g/100ml to 1.5g/100ml. The decrease in removal

efficiency with increasing adsorbent amount may be

due to the concentration gradient between solute

concentration in the solution and the solute

concentration in the surface of the adsorbent. Thus

Dahlan and Tayeh


22

with increasing adsorbent amount, the amount of dye

adsorbed onto unit weight of adsorbent gets reduced,

thus causing a decrease in adsorption capacity and

removal efficiency with increasing adsorbent amount

(Vadivelan and Kumar, 2005).

Fig. 6: Effect of adsorbent amount on the adsorption of DB71 from synthetic wastewater

3.2.3. Effect of shaking rate

Figure 7 shows the effect of shaking rate on the

adsorption of DB71. The removal efficiency of DB71

was 98.74% at low shaking rate (100 rpm) and rose to

99.36% as the shaking rate was increased to 150 rpm.

This effect can be attributed to the decrease in

boundary layer thickness around the adsorbent

particles being a result of increasing the degree of

mixing (Gupta et al., 2011). The removal efficiency

decreased again when the shaking rate was increased

to 350 rpm. This might be due to higher boundary

layer resistance to mass transfer in the bulk (Dahlan

and Razali, 2011).

Fig. 7: Effect of shaking rate on the adsorption of DB71 from synthetic wastewater

3.2.4. Effect of initial dye concentration

The effect of initial dye concentration in the range of

15 to 200 mg/l on adsorption was investigated and the

results are shown in Figure 8. It is evident from figure

that the percentage dye removal increased with the

increasing in initial concentration of dye. The increase

in initial dye concentration enhances the reaction

between dye and PH/CFO-Ac. The percentage dye

removal was found to be 95.11% for 15 mg/l of initial

concentration, and it increased to 99.54% when the

initial dye concentration was increased to 200mg/l.



23

Fig. 8: Effect of initial dye concentration on the adsorption of DB71 from synthetic wastewater

3.2.5. Effect of pH

The pH of the solution plays an important role in the

whole adsorption process. To determine the optimum

pH for the adsorption of DB71 using PH/CFO-Ac, the

effect of solution pH was investigated in the range of

3–10.5 and the results are shown in Figure 9. The

removal efficiency of DB71 at pH less than 7 is

relatively lower than at pH more than 7. It increased

up from 87.8% to 91.5% when pH was increased from

3 to 9. Solution pH may affect both aqueous chemistry

and surface binding-sites of the adsorbent. The

decrease of adsorption at pH less than 7 can be

explained by the fact that at this acidic pH, H+ may

compete with dye ions for the adsorption sites of

adsorbent, thereby inhibiting the adsorption of dye

(Hameed and El-Khaiary, 2008).

Fig. 9: Effect of pH on the adsorption of DB71 from synthetic wastewater

3.2.6. Effect of temperature

The effect of temperature on the adsorption process

was investigated by carrying out the adsorption

experiments using PROTECH Shaker Incubator with

different degrees of temperature ranging from 30 -

60oC. Figure 10 shows that the highest removal

efficiency (92.04%) was obtained at temperature

30oC, and decreased slowly until temperature 50

oC,

after that it decreased rapidly to 76.36% when the

temperature was increased to 60oC. Similar results

were obtained by various authors for the adsorption of

dyes on various adsorbents (Ho and McKay, 1998;

Chern and Wu, 2001; Chiou and Li, 2002; Hamdaoui,

2006). This can be explained by the exothermic

spontaneity of the adsorption process and by the

weakening of bonds between dye molecules and

active sites of adsorbents at high temperatures (Amin,

2009).

Dahlan and Tayeh


24

Fig. 10: Effect of temperature on the adsorption of DB71 from synthetic wastewater

3.3. Adsorption kinetics

Figure 11 shows that the adsorption capacity of

PH/CFO-Ac increased rapidly in the initial stages of

the experiment and it reached equilibrium at 40

minutes although the data were measured for 2 hours.

The results indicate that, at the beginning the dye ions

were adsorbed by the exterior surface of PH/CFO-Ac,

so the adsorption rate was fast. When the adsorptions

of the exterior surface reached saturation, the dye

exerted onto the pores and was adsorbed by the

interior surface of the adsorbent. The time profile of

dye uptake is a single, smooth and continuous curve

leading to saturation, suggesting also the possible

monolayer coverage of dye on the surface of the

PH/CFO-Ac A similar phenomenon was observed

from previous study for the adsorption of direct blue

71 dye on palm ash adsorbent and the equilibrium

time was 1 hour (Sengil, 2003).

Fig. 11: The variation of adsorption capacity of DB71 onto PH/CFO-Ac with adsorption time, 1L solution, 1g adsorbent,

200mg/L dye concentrations, pH 9, at 30 ◦C

To investigate the adsorption kinetic of DB71 on

the surface of PH/CFO-Ac, pseudo-first-order and

pseudo-second-order kinetic models were tested. The

plots of pseudo-first-order and pseudo-second-order

kinetic models are shown in Figure 12 and Figure 13,

respectively. Kinetic constants obtained by linear

regression for the two models are listed in Table 2.

The correlation coefficient for the pseudo-first-order

model is relatively good (R2 = 0.85), however, the

calculated qe (qe,cal ) obtained from this equation does

not give reasonable value (Table 2), which is much

lower compared with qe obtained from the

experimental data (qe,exp ).



25

Fig.12: Pseudo-first-order kinetic plot for adsorption of DB71 onto PH/CFO-Ac (C0 = 200mg/L, pH 9)

Fig. 13: Pseudo-second-order kinetic plot for adsorption of DB71 onto PH/CFO-Ac (C0 = 200mg/L, pH 9)

Table 2: Kinetic parameters for adsorption of DB71 onto PH/CFO-Ac

q e,exp

(mg g -1)

Pseudo-first-order Pseudo-second-order

K1 (min-1) q e,cal

(mg g -1)

R2 K2

(g mg-1 min-1)

q e,cal

(mg g -1)

R2

196.48 0.032 3.0 0.85 0.031 196 0.997

This result suggests that the adsorption process

does not follow the pseudo-first-order kinetic model,

which is similar to the result reported for adsorption

of direct blue 71 onto palm ash adsorbent (Ahmad et

al., 2007). In many cases the pseudo-first-order

equation of Lagergren does not fit well to the whole

range of contact time and is generally applicable over

the initial stage of the adsorption processes (Ho and

McKay, 1999, Hameed and Hakimi, 2008). For the

pseudo-second-order kinetic model, the R2 value is

0.997 and the (qe,cal =196mg/g) agrees very well with

the (qe,exp =196.48mg/g) value, which indicates that

the adsorption of DB71 onto PH/CFO-Ac follows a

pseudo-second-order kinetic model.

3.4. Adsorption isotherm

The adsorption isotherm is the most important

information which indicates how the adsorbate

molecules distribute between the liquid phase and the

Y= -0.013X + 0.545

R2 = 0.85

Y = 0.00509 X + 0.00083

R2 = 0.997

Dahlan and Tayeh


26

solid phase when the adsorption process reaches an

equilibrium state. To optimize the design of an

adsorption system for the adsorption of adsorbates, it

is important to establish the most appropriate

correlation for the equilibrium curves. In this study,

the equilibrium isotherms were examined using the

Langmuir and Freundlich isotherms.

In the linear form of Langmuir isotherm, a plot of

Ce/qe versus Ce (Fig. 14) yields a straight line with

slope 1/Q0 and intercepts 1/Q0b. The essential

characteristics of the Langmuir isotherm can be

expressed in terms of dimensionless constant

separation factor RL given by (Mahmoodi et al., 2011)

(7)

where b is the Langmuir constant and Co is the

highest initial dye concentration (mg/L). According to

the value of RL the isotherm shape may be interpreted

as shown in Table 3.

Fig. 14: Linearized Langmuir isotherm

Table 3: Adsorption Types based on RL value (Mahmoodi et al., 2011)

Value of RL Type of adsorption

RL > 1.0

RL = 1.0

0 < RL < 1.0

RL = 0

Unfavourable

Linear

Favourable

Irreversible

The value of RL calculated (0.00066) was in the

range between 0 and 1 which indicate that the

adsorption is favorable at operation conditions

studied. Table 4 summarizes Langmuir constants and

computed maximum adsorption capacity Q0 of DB71

onto the PH/CFO-Ac.

Table 4: Langmuir and Freundlich isotherm model constants and correlation coefficients for adsorption of DB71 onto

PH/CFO-Ac

Langmuir

isotherm

b (L/mg) Qo (mg/g) R

L R

2

7.57 188.7 0.00066 0.997

Freundlich isotherm KF Qo (mg/g) n R

2

203 177.7 40 0.991

In addition, the Freundlich isotherm is an empirical

equation based on a heterogeneous surface. A plot of

ln qe versus ln C

e (Fig. 15) enables the constant K

F

and exponent n to be determined. KF can be defined as

adsorption of distribution coefficient and represents

the quantity of dye adsorbed onto adsorbent for an

equilibrium concentration. The slope 1/n is a measure

of adsorption intensity or surface heterogeneity. These

values together with the correlation coefficient are

presented in Table 3.3. Based on the values of

correlation coefficients (R2

) and Qo shown in Table

3.3, the adsorption isotherm with PH/CFO-Ac can be

described by Langmuir equation. The Langmuir

equation yields a better fit of the experimental data

than Freundlich equation.



27

Fig. 15: Linearized Freundlich isotherm

3.5. Characterization of magnetic PH/CFO-Ac

adsorbent

3.5.1. Particle Size Distribution analysis

Figure 16 shows the particle size distribution of

prepared adsorbent, spent adsorbent with high

removal efficiency and spent adsorbent with low

removal efficiency, respectively. The maximum

volume percentage corresponds to the particle size of

approximately 1000µm, 900µm and 600µm for

prepared adsorbent, spent adsorbent with low removal

and spent adsorbent with high removal, respectively.

The average volume of the particle size distribution

range of spent PH/CFO-Ac adsorbents is higher than

the prepared PH/CFO-Ac adsorbent, which means that

higher particle size distributions range are formed

after DB71 adsorption onto PH/CFO-Ac adsorbent. It

was also found that the specific surface area of the

prepared adsorbent was much higher (0.034 m2/g)

than that of the spent adsorbent with low removal and

spent adsorbent with high removal (0.0158 and 0.0112

m2/g, respectively). The lower specific surface area of

spent PH/CFO-Ac adsorbents could be due to the

agglomerating of smaller particles (which is most

probably the dye particles that cover the surface of the

adsorbent) during the adsorption process. It was also

shown from the figure that prepared and spent

PH/CFO-Ac adsorbents have a unimodal particle size

distribution, which indicates a homogeneous particle

size for the prepared and the spent adsorbents.

Dahlan and Tayeh


28

Fig. 16: Particle size distribution of prepared PH/CFO-Ac adsorbent, spent adsorbent with high removal efficiency, and spent

adsorbent with low removal efficiency.

3.6.2. Surface morphology analysis (SEM)

The morphology of the raw psyllium husk, prepared

and spent PH/CFO-Ac adsorbent is shown in Figure

17. Surface analysis indicated that the surface

morphology of the raw psyllium husk (Figure 17a)

consists of lumps of uneven shapes particles. After the

preparation of PH/CFO-Ac adsorbent, it can be

observed from Figure 17b that the prepared adsorbent

consists of irregular rough particles that were

unevenly scattered together which might be due to the

CoFe2O4 particles deposited on the surface of raw

psyllium husk. After the adsorption of DB71, the

surface of the PH/CFO-Ac adsorbents was found to

have more compact structures with small granular

particles attached on the surface (Figure 17c) which is

most probably the dye particles that covered the

external surface of PH/CFO-Ac adsorbents.

4. CONCLUSIONS

This study showed that PH/CFO-Ac adsorbent was

successfully synthesized by a facile one-step refluxing

route could be used as an effective adsorbent for the

removal of direct blue dye (DB71) from synthetic

wastewater. All experimental parameters such as

contact time, adsorbent amount, shaking rate, initial

dye concentration, pH and temperature affected the

adsorption of dye from synthetic wastewater. The

equilibrium of adsorption of dye onto PH/CFO-Ac

adsorbent was suitably described by the Langmuir

models. The process of adsorption was relatively rapid

and was best described by the pseudo-second-order

kinetic model. The results obtained in this study

shows that PH/CFO-Ac adsorbent could be used as an

effective adsorbent for the removal of direct blue dye

from synthetic wastewater and this has never been

reported in the literature.



29

Fig. 17: SEM micrographs of (a) raw psyllium husk, (b) prepared PH/CFO-Ac adsorbent and (c) spent PH/CFO-Ac adsorbent

ACKNOWLEDGMENT

The authors wish to acknowledge the financial

support from the Universiti Sains Malaysia (Short

Term Grant A/C. 60310014 and Incentive Grant).

REFERENCES

Ahmad A, Hameed B, Aziz N (2007). Adsorption of

direct dye on palm ash: Kinetic and equilibrium

modeling. Journal of Hazardous Materials, 141:

70-76.

Ahn DH, Chang WS, Yoon TI (1999). Dyestuff

wastewater treatment using chemical oxidation,

physical adsorption and fixed bed biofilm

process. Process Biochemistry, 34: 429-439.

Ai L, Huang H, Chen Z, Wei X, Jiang J (2010).

Activated carbon/CoFe2O4 composites: Facile

synthesis, magnetic performance and their

potential application for the removal of

malachite green from water. Chemical

Engineering Journal, 156: 243-249.

Amin NK (2009). Removal of direct blue-106 dye

from aqueous solution using new activated

carbons developed from pomegranate peel:

Adsorption equilibrium and kinetics. Journal of

Hazardous Materials, 165: 52-62.

Che Ani Y (2004). Adsorption Studies Of Dyes Using

Clay-based And Activated Carbon Adsorbents.

MscThesis.Universiti Sains Malaysia, Penang.

Chern, JM, Wu CY (2001). Desorption of dye from

activated carbon beds: effects of temperature,

pH, and alcohol. water research, 35: 4159-4165.

Chiou MS, Li HY (2002). Equilibrium and kinetic

modeling of adsorption of reactive dye on

Dahlan and Tayeh


30

cross-linked chitosan beads. Journal of


Correia VM, Stephenson T, Judd SJ (1994).

Characterisation of textile wastewaters‐a

review. Environmental Technology, 15: 917-

929.

Dahlan I, Razali MHM (2011). Waste-Derived

Siliceous Materials as a Novel Sorbent for

Removal of Ni 2+ from Aqueous Solutions.

Water, Air, & Soil Pollution, 1-7.

Gupta V, Agarwa S, Saleh TA (2011). Chromium

removal combining the magnetic properties of

iron oxide with adsorption properties of carbon

nanotubes. Water Research, 45: 2207-2212.

Hamdaoui O (2006). Batch study of liquid-phase

adsorption of methylene blue using cedar

sawdust and crushed brick. Journal of


Hameed B, El-Khaiary M (2008). Malachite green

adsorption by rattan sawdust: Isotherm, kinetic

and mechanism modeling. Journal of


Hameed B, Hakimi H (2008). Utilization of durian

(Durio zibethinus Murray) peel as low cost

sorbent for the removal of acid dye from

aqueous solutions. Biochemical Engineering

Journal, 39: 338-343.

Ho Y, Mckay G (1998). Sorption of dye from aqueous

solution by peat. Chemical Engineering Journal,

70: 115-124.

Ho Y, Mckay G (1999). The sorption of lead (II) ions

on peat. Water Research, 33: 578-584.

Kadirvelu K, Kavipriya M, Karthika C, Radhika M,

Vennilamani N, Pattabhi S (2003). Utilization

of various agricultural wastes for activated

carbon preparation and application for the

removal of dyes and metal ions from aqueous

solutions. Bioresource Technology, 87: 129-

132.

Kelesoglu S (2007). Comparative Adsorption Studies

of Heavy Metal Ions on Chitin and Chitosan

Biopolymers.Msc Thesis. İzmir Institute of

Technology, İzmir.

Kumar PS, Gayathri R, Arunkumar RP (2010).

Adsorption of Fe (III) ions from aqueous

solution by Bengal Gram Husk powder:

equilibrium isotherms and kinetic approach.

Electronic Journal of Environmental,

Agricultural and Food Chemistry, 9: 1047-

1058.

Lee C, Low K, Gan P (1999). Removal of some

organic dyes by acid-treated spent bleaching

earth. Environmental Technology, 20: 99-104.

Low KS, Lee CK (1997). Quaternized rice husk as

sorbent for reactive dyes. Bioresource

Technology, 61: 121-125.

Lucas MS, Peres JA (2006). Decolorization of the azo

dye Reactive Black 5 by Fenton and photo-

Fenton oxidation. Dyes and Pigments, 71: 236-

244.

Mahmoodi NM, Hayati B, Arami M, Lan C (2011).

Adsorption of textile dyes on Pine Cone from

colored wastewater: Kinetic, equilibrium and

thermodynamic studies. Desalination, 268: 117-

125.

Morais L, Freitas O, Goncalves E, Vasconcelos L,

Gonzalez BC (1999). Reactive dyes removal

from wastewaters by adsorption on eucalyptus

bark: variables that define the process. Water

Research, 33: 979-988.

Papic S, Koprivanac N, Bozic AL (2000). Removal of

reactive dyes from wastewater using Fe (III)

coagulant. Coloration Technology, 116: 352-

358.

Schwarzenbach RP, Gschwend PM, Imboden DM,

Wiley J (2003). Environmental organic

chemistry, Wiley Online Library.

Sengil IA (2003). Adsorption of reactive dyes on

calcined alunite from aqueous solutions.

Journal of Hazardous Materials, 98: 211-224.

Vadivelan V, Kumar KV (2005). Equilibrium,

kinetics, mechanism, and process design for the

sorption of methylene blue onto rice husk.

Journal of Colloid and Interface Science, 286:

90-100.

WHO/UNICEF (2000). Global Water Supply and

Sanitation Assessment Report 2000. WHO,

Geneva.



©2014 IJSRPUB

31

COD and BOD Removal from Textile Wastewater Using Naturally Prepared

Adsorbents and Their Activation Forms Using Sulphuric Acid

Himanshu Patel*, R.T. Vashi

Department of Chemistry, Navyug Science College, Rander Road, Surat – 395009, Gujarat, INDIA


Abstract. Comparative adsorption studies of naturally prepared adsorbents (Neem leaf powder, Gauva leaf powder and

Tamarind seed powder) and their activation forms using sulphuric acid were conducted for removal COD and BOD from

textile wastewater. Process parameters like adsorbent dose, conduct duration, temperature and pH were examined in this study.

The adsorption data were analyzed using Freundlich, Langmuir, Dubinin-Raduskevich, Flory-Huggins, Redlich-Peterson, Sips,

Toth and Khan model to understand adsorption mechanism. Activated Neem leaf powder was most suitable than investigated

adsorbents. The maximum adsorption capacities were found to be 87.58, 85.65 and 81.25 mg/g for NLPs, GLPs and TSPs

respectively for COD and 84.51, 81.45 and 75.54 mg/g for NLPs, GLPs and TSPs respectively for BOD.

Keywords: Textile wastewater, Naturally prepared adsorbents and their activated form, Process Parameters, Adsorption

equilibrium isotherms

1. INTRODUCTION

Increasing public pressure and administrative concern

to arrest further deterioration of the environment may

result in more stringent standards. The objective of

wastewater treatment has expanded considerably from

simple nuisance control to include public health,

environmental, aesthetic and ecological

considerations. The day is not far off when industries

will have to acquire a nonpolluting or zero polluting

status to meet future environmental regulations. The

problem is more severe for the textile industry

because of its dynamic nature. Constantly changing

process lines result in diverse and complex wastes that

are ever changing in constitution and color (Venkata

Mohan et al., 1999). The textile industry is very

chemical-intensive; wastewater from textile-

processing contains huge residues from different

textile dyeing and finishing operations. Of particular

concern are dyestuffs, which are often major sources

of heavy metals, salt, adsorbable organic halogens and

color in dyehouse effluent, sizing agents, which have

high BOD and COD levels; and anionic/non-ionic

surfactants, most of which are still of poorly

biodegradable nature (Alaton et al., 2006).

There are many methods for the removal of

contaminations from wastewater, such as membrane

process, ion exchange, biological degradation and

adsorption using various kinds of adsorbent.

Adsorption process is proven to be an effective

process for the removal of various pollutants from its

aqueous solutions because adsorption process can

remove pollutant in wide range of concentrations

(Budyanto et al., 2008). The adsorption process has

not been used extensively in wastewater treatment, but

demands for a better quality of treated wastewater

effluent, including toxicity reduction, have led to an

intensive examination and use of the process of

adsorption. The relative advantages of adsorption over

other conventional advanced treatment methods are:

(1) it can remove both organic as well as inorganic

constituents even at very low concentrations, (2) it is

relatively easy and safe to operate, (3) both batch and

continuous equipment can be used, (4) no sludge

formation, and (5) the adsorbent can be regenerated

and used again. Moreover the process is economical

because it requires low capital cost and there are

abundant low-cost materials available which can be

used as adsorbents (Mohanty, et al., 2006). Low cost

adsorbents especially made from natural sources like

plant root, leaf, seed, peel, etc are being investigated.

Ulmus leaves and their ash, rubber (Hevea

brasiliensis) seed shell, Platanus orientalis leaves,

Rich husk, Sunflower stalks and Neem leaf powder

have been used for wastewater treatment and

adsorption has emerged as a cost-effective and

efficient alternative for the removal of hazardous

contaminations including dyes from low strength

wastewaters. Also, some acids were also utilized for

activation of adsorbents (Patel and Vashi, 2011).

Previously we had utilized activated neem leaf

powder (Patel and Vashi, 2012), activated guava leaf

powder (Patel and Vashi, 2011) and tamarind seed

powder (Patel and Vashi, 2010) for removal of dyes

from its aqueous solution and provided outstanding

results. There are very few research literatures

available for removal of COD and BOD from real

wastewater as per author’s best knowledge. Present

investigation involved the preparation of naturally

prepared adsorbents (and its activation form) and

treatment of textile wastewater using these adsorbents

for removal of COD and BOD. Effect of adsorbent

dose, contact duration, temperature and pH were

studied. Freundlich, Langmuir, Dubinin-Raduskevich,

Flory-Huggins, Redlich-Peterson, Sips, Toth and

Khan models were tested for their applicability.


Patel and Vashi

COD and BOD Removal from Textile Wastewater Using Naturally Prepared Adsorbents and Their Activation Forms

Using Sulphuric Acid

32

2. METHODS AND MATERIALS

2.1. Adsorbent

The Neem (scientific name: Azadirachta indica)

belongs to the meliaceae family and is native to Indian

sub-continent. The Guava (Psidium Guajava; Family:

Myrtaceae) tree are easily available in Indian region.

The mature leaves of plant (Neem and Guava) used in

the present investigation are collected from the

available trees near Navyug Science College, Gujarat.

The mature leaves of plant washed thrice with water

to remove dust and water soluble impurities and were

dried until the leaves become crisp. The dried leaves

were crushed and powdered and further washed with

distilled water till the washings were free from color

and turbidity. Then this powder was dried in an oven

at 60 ± 2 °C and placed in desiccator for the

adsorption studies, thus natural adsorbent prepared.

The Tamarind (Tamarindus indica), a family of

Fabaceae, has used for preparation of medicines for

internal and external applications and as condiment in

many dishes. Tamarind fruit seed, collected from

nearby Navyug Science College, Gujarat, a waste

product of tamarind pulp, are washed, dried and

pulverized. This powder was washed with distilled

water till the washings were free from color and

turbidity and thereafter, dried in the oven at 60 ºC. For

activation of adsorbent, each adsorbent was stirred

with excess amount of 0.1 N sulphuric acid.

Thereafter, it washed with de-ionized water to remove

untreated acid dried in an oven at 60 ± 2 °C.

2.2. Experimental Details

The textile wastewater samples were withdrawn from

Pandesara, GIDC, Gujarat, India. Combine

wastewater samples were collected bimonthly for

three times in sampling bottles and placed in ice box

to preserve the characteristics of wastewater and were

analyzed as per standard method (APHA, 1992). For

removal of COD and BOD, batch experiments of

textile wastewater were carried out as per table 1.

Briefly, each adsorbent i.e. NLP and a-NLP (NLPs),

GLP and a-GLP (GLPs) and TSP and a-TSP (a-TSP)

was added to wastewater samples and the mixture was

stirred at 400 rpm. Each adsorbent was kept in contact

till equilibrium state was attained. The required pH of

system was maintained by using 0.1 N HCl or 0.1 N

NaOH during experiment. All chemicals used were of

analytical reagent grade and purchased from

Qualigens, India. The important physico-chemical

characteristics i.e. COD and BOD were determined

before and after treatment using standard methods.

Table 1: Experimental Details for treatment of wastewater using adsorbents

Effect of System Adsorption Dose (g/l) Temperature (K) Contact Duration (min) pH

Effect of adsorption dose 1, 2, 4, 6, 8, 10, 12, 14,

16, 18 and 20 300 180 7

Effect of contact duration 5.0 300 30, 60, 90, 120, 150,

180, 210 and 240 7

Effect of temperature 5.0 298, 303, 308, 313, 318,

323 and 328 180 7

Effect of pH 5.0 300 180 3, 5, 7, 9 and 11

2.3. Batch adsorption model

The equilibrium sorption isotherm is fundamentally

important in the design of sorption systems.

Equilibrium studies in sorption give the capacity of

the sorbent. It is described by sorption isotherm

characterized by certain constants whose values

express the surface properties and affinity of the

sorbent. (samsun clone). Different equilibrium models

were available to predict the sorption behavior. The

most commonly available models are Freundlich,

Langmuir, Dubinin-Raduskevich, Flory-Huggins,

Redlich-Peterson, Sips, Toth and Khan models.

2.3.1. Freundlich Isotherm

The Freundlich expression is an empirical equation

based on sorption on a heterogeneous surface and

effectively on multilayer and is expressed by the

following equation.

qe = KFCe1/n

or log qe = log KF + 1/n log Ce

Where, qe and Ce is the amount of adsorbed

adsorbate per unit weight of adsorbent and

unadsorbed adsorbate concentration in solution at

equilibrium, respectively and KF (L/mg) and n are

Freundlich constant characteristics of the system,

which are determined from the log qe vs. log Ce. If the

value of exponent n was greater than 1 (n >1) then the

adsorption represent favorable adsorption

2.3.2. Langmuir Isotherm

This monolayer adsorption isotherm is very useful for

predicting adsorption capacities and also interpreting

into mass transfer relationship. The isotherm can be

written as follows:



33

qe = kLCe / (1 + aLCe) or (Ce/qe) = (1/ KL) + (aL/KL)

Ce

The constant KL (L/g) is the Langmuir equilibrium

constant and the KL/aL gives the theoretical monolayer

saturation capacity, Qmax (mg/g). These Langmuir

parameters were obtained from the linear correlations

between the values of Ce/qe and Ce. Generally, the

Langmuir equation, applies to the cases of adsorption

on completely homogeneous surfaces where

interactions between adsorbed molecules are

negligible (Ozacar and Sengil, 2005).

2.3.3. Dubinin-Raduskevich Isotherm

This isotherm can be used to describe adsorption on

both homogenous and heterogeneous surfaces. The

Dubinin–Radushkevich equation has the following

form:

qe = qme-βε2

or In qe = In qm - βε2

Where qm (mg/g) is the Dubinin–Radushkevich

monolayer capacity a constant related to sorption

energy, and ε is the Polanyi potential which is related

to the equilibrium concentration as follows:

ε = RT In (1 + 1/Ce)

Where R is the gas constant (8.314 J/mol K) and T

is the absolute temperature. The constant β gives the

mean free energy, E, of adsorption per molecule of the

adsorbate when it is transferred to the surface of the

solid from infinity in the solution and can be

computed using the relationship.

E = 1/ (2β) ½

The magnitude of E is useful for estimating the

mechanism of the adsorption reaction. In the case of E

< 8 kJ/mol, physical forces may affect the adsorption.

If E is in the range of 8–16 kJ/mol, adsorption is

governed by ion exchange mechanism, while for the

values of E > 16 kJ/mol, adsorption may be dominated

by particle diffusion.

2.3.4. Flory-Huggins Isotherm

The Flory-Huggins model accounts for the degree of

surface coverage characteristics of adsorbate on the

adsorbent and the linear form of the Flory-Huggins

equation is expressed as:

ln (q/Ce) = ln(KFH) + aFH ln(1- qe)

Where q is the surface coverage of the adsorbent

by adsorbate. KFH (mg/g) and aFH are the Flory-

Huggins constants. These constants can be obtained

from the plot of log (q/Ce) versus log (I-qe)

Furthermore the equilibrium constant KFH,

obtained from the Flory-Huggins isotherm model is

used to compute the Gibbs free energy for the

adsorption process. The Gibbs free energy is related to

equilibrium constant as follows:

ΔG0 = - RT/ In KFH

Where R is universal gas constant 8.314 J/K/mol,

T is absolute temperature (K) and KFH is equilibrium

constant from Flory-Huggins isotherm equation

(Israel et al., 2010).

2.3.5. Redlich-Peterson Isotherm

The Redlich-Peterson isotherm model is widely used

as a compromise between the Langmuir and

Freundlich systems, since it combines elements from

both the Langmuir and Freundlich equations, where

the mechanism of adsorption is a hybrid one and does

not follow ideal monolayer adsorption. The Redlich-

Peterson isotherm has a linear dependence on

concentration in the numerator and an exponential

function in the denominator. It approaches the

Freundlich model at high concentration and is in

accordance with the low concentration limit of the

Langmuir equation. It can be applied either in

homogenous or heterogeneous systems due to the high

versatility of the equation. This model incorporates

three parameters into an empirical isotherm. It can be

described as follows,

ln[(KRPCe /qe) – 1] = βRP ln(Ce) + ln(aRP)

Where, qe (mg/g), is the solid-phase adsorbate

concentration at equilibrium, Ce (mg/l) is the liquid-

phase adsorbate concentration at equilibrium, KRP

(1/mg) and aRP (Lβmg

–β), are the Redlich-Peterson

isotherm constants and βRP is an exponent, which lies

between 1 and 0. If ‘βRP’ is equal to one, the equation

modifies to Langmuir model, and if ‘βRP’ is equal to

zero then the equation changes to Henry’s law

equation (Quintelas et al., 2008).

2.3.6. Sip Isotherm

Sips isotherm is a combined form of Langmuir and

Freundlich expressions deduced for predicting the

heterogeneous adsorption systems and circumventing

the limitation of the rising adsorbate concentration

associated with Freundlich isotherm model. At low

adsorbate concentrations, it reduces to Freundlich

isotherm; while at high concentrations, it predicts a

monolayer adsorption capacity characteristic of the

Langmuir isotherm. As a general rule, the equation

parameters are governed mainly by the operating

conditions such as the alteration of pH, temperature

and concentration. The linear form of Sip isotherm

model is represented as

βs ln(Ce) = - ln(Ks/qe) + ln(as)

Where Ks is the total number of binding sites

(mg/g), as the median association constant (L/mg) and

1/n is the heterogeneity factor. If the value for βs is

Patel and Vashi



34

less than one, it indicates that it is heterogeneous

adsorbents, while values closer to or even one

indicates that the adsorbent has relatively more

homogeneous binding sites.

2.3.7. Toth Isotherm

Toth isotherm model, is another empirical equation

developed to improve Langmuir isotherm fittings

(experimental data), and useful in describing

heterogeneous adsorption systems, which satisfying

both low and high-end boundary of the concentration.

Its correlation presupposes an asymmetrical quasi-

Gaussian energy distribution, with most of its sites has

adsorption energy lower than the peak (maximum) or

means value

qe = (qm bT Ce) / [1 + (bTCe) 1/nT

] nT

Where, qm (mg/g) is the Khan adsorption capacity,

and bT (1/mg) is the Toth model constant and nT the

Toth model exponent (Foo and Hameed, 2010).

3. RESULTS AND DISCUSSION

3.1. Characterization of Textile Wastewater

It was reported that water requirement vary from 61 to

646 liters per kg of cloth processed with an average

value of 235 liters per kg. These values correspond to

12 - 130 liter per meter of cloth with an average value

of 47 liters per kg. The pH of the investigated textile

mill effluent ranges from 9.24 to 7.64 having average

value of 8.2. The color of the wastewater is brownish

yellow having color unit 291.3 Hazen. The average

value of COD, BOD, Total Dissolved Solid (TDS),

Hardness, Alkalinity, Fluoride, Chloride and Sulphate

were 1556.2, 924.1, 6845.3, 1086.7, 670.7, 874.3,

475.3 and 877.1 ppm respectively. Also, average

value of Electrical Conductivity and Sulphate were

found to be 7417.0 µs/cm. High value of

contaminations are presented due to usage of

chemicals such as starches, dextrin, gums, glucose,

waxes, pectin, alcohol, fatty acids, acetic acid, soap,

detergents, sodium hydroxide, carbonates, sulfides,

sulfites, chlorides, dyes, pigments, carboxymethyl

cellulose, gelatin, peroxides, silicones, flourcarbons,

resins, etc. in wet processes of textile mill. Same types

of contaminations were found, while investigating

textile wastewater (Fanchiang et al., 2006, Pathe et al.,

2005, Martins et al., 2006).

3.2. Effect of Adsorbent Dose

Figure 1, 2 and 3 shows the effect of different doses

(Ceq vs. Adsorbent dose) of naturally prepared

adsorbents i.e. NLPs, GLPs and TSPs respectively

maintaining temperature of 300 K, contact duration of

180 and pH 7. It can be seen that there is a large

reduction in the COD content when NLPs was used,

from an initial value of 1625.8 ppm to zero and 110.2

ppm at a dosage of 20 g/L of a-NLP and NLP

respectively. BOD had an initial value of 1002.4 to

11.2 and 51.5 ppm by 20 g/L of a-NLP and NLP

respectively. It can be also seen that there is a large

reduction in the COD content when GLPs is used,

from an initial value of 1625.8 to 258.4 and 411.2

ppm at a dosage of 20 g/L of a-GLP and GLP

respectively. BOD had an initial value of 1002.4 to

201.2 and 212.2 ppm when 20 g/L of a-GLP and GLP

was used respectively. It can be seen that there is a

large reduction in the COD content using TSPs, from

an initial value of 1625.8 to 411.2 and 312.2 ppm at a

dosage of 20 g/L of TSP and a-TSP respectively.

BOD had an initial value of 1002.4 to 280.5 and 251.2

ppm by 20 g/L of TSP and a-TSP respectively.

Fig. 1: Effect of adsorbent dose for percentage COD and BOD removal using NLPs



35

Fig. 2: Effect of adsorbent dose for percentage COD and BOD removal using GLPs

Fig. 3: Effect of adsorbent dose for percentage COD and BOD removal using TSPs

From all the adsorbents, a-NLP was found to be

more effective than all investigated adsorbents for

removal of COD and BOD. Also, it was easily

appeared that value of Ceq was continuously

decreasing with increasing adsorbent dose upto 18.0

g/L. The increase in adsorption with increase in

adsorbent may be attributed due to the reason of

increased adsorbent surface and availability of more

adsorption sites. The reason behind the phenomenon

may be speculated to be due to the interference

between binding sites at higher concentrations in

solution with respect to available binding sites. Then

after straight line indicated the equilibrium was

attained at dose of 18.0 g/l for all adsorbents

investigated. It can explain that further increase in the

dose of adsorbent did not affect the uptake capacity

because of the unavailability of adsorbate sites due to

saturation (Li et al., 2008).

3.3. Effect of Contact Duration

Table 2 reveals effect of different contact durations

(30 to 240 minutes) on the COD, BOD and color

removal using 5.0 g/L dosage of NLPs, GLPs and

TSPs at constant temperature (300 K), pH (7) and

agitator speed (400 rpm). Percentage removal of COD

was found to be 30.3 to 63.6, 19.6 to 63.4 and 9.5 to

41.5 using a-NLP, a-GLP and a-TSP respectively.

Also, percentage removal of COD was found to be

20.1 to 57.2, 11.2 to 47.5 and 13.4 to 56.2 using NLP,

GLP and TSP respectively. Percentage removal of

BOD was found to be 26.1 to 57.4, 14.6 to 49.7 and

5.2 to 32.1 using a-NLP, a-GLP and a-TSP

respectively. Also, percentage removal of BOD was

found to be 15.2 to 46.2, 9.2 to 42.1 and 10.1 to 40.1

using NLP, GLP and TSP respectively. These

removals were found at contact duration of 30 to 240

min respectively.

In adsorption process, a knowledge regarding the

contact time required to achieve the equilibrium is

highly important. Table 2 clearly indicates a rapid

increase in the amount of adsorption with increase in

time initially, gradually leading to equilibrium.

Although at higher contact time, the rate of adsorption

decreased and a saturation stage was attained due to

the accumulation of the adsorption sites. This decline

is due to decrease in total adsorbent surface area and

increased diffusion pathway. It can also elucidate that

Patel and Vashi



36

mechanism of solute transfer to the solid includes

diffusion through the fluid film around the adsorbent

particle and diffusion through the pores to the internal

adsorption sites. Initially the concentration gradient

between the film and the solid surface is large, and

hence the transfer of solute onto the solid surface is

faster. That is why it takes lesser time to attain

percentage removal of COD and BOD. As time

increases, intraparticle diffusion becomes

predominant. Hence solute takes more time to transfer

from solid surface to internal adsorption sites through

the pores (Hameed et al., 2008, Gulipalli et al., 2011).

Table 2: Effect of contact duration for percentage COD and BOD removal using naturally prepared adsorbents

Adsorbent % Removal Contact Duration (min)

30 60 90 120 150 180 210 240

a-NLP COD 30.3 39.9 50.4 54.1 57.2 60.1 63.6 63.6

BOD 26.1 31.2 39.0 49.5 53.6 55.7 57.4 57.4

NLP COD 20.1 30.2 41.2 45.2 51.0 55.2 57.2 57.2

BOD 15.2 20.5 28.5 40.5 44.2 45.5 46.2 46.2

a-GLP COD 19.6 32.3 44.3 48.9 56.5 58.5 63.4 63.4

BOD 14.6 28.6 34.5 40.1 45.1 48.1 49.7 49.7

GLP COD 11.2 20.2 30.2 35.5 41.2 45.5 47.4 47.4

BOD 9.2 15.4 20.4 30.2 34.5 40.1 42.1 42.1

a-TSP COD 9.5 14.5 19.5 28.9 33.3 39.5 41.5 41.5

BOD 5.2 10.2 16.5 20.5 27.5 30.2 32.1 32.1

TSP COD 13.4 20.9 30.5 39.7 50.4 53.1 56.2 56.2

BOD 10.1 18.8 24.7 31.2 36.7 38.0 40.1 40.1

3.4. Effect of Temperature

The effect of variation in temperature i.e. 298, 303,

308, 313, 318, 323 and 328 K on the adsorption by a-

NLP, a-GLP and TSP at constant contact duration,

with respect to COD, BOD and color of wastewater is

depicted in Table 3. It can be seen that the increase in

temperature leads to linear decrease in percent

removal in three cases. Also, straight line after

temperature of 323 K indicating equilibrium

accomplished at temperature of 323 K. The values of

percent removal of COD were seen to increase from

39.1 to 75.6 % by a-NLP, 20.15 to 65.6 by NLP

whereas from 30.3 to 72.3 % by a-GLP, 20.1 to 61.2

% using GLP and from 15.2 to 47.7 % by a-TSP and

also, 24.3 and 63.0 % by TSP at temperature of 298 to

328 K respectively. The values of percent removal of

BOD were seen to increase from 34.7 to 63.2 % by a-

NLP, 20.25 to 55.5 by NLP whereas from 29.0 to 68.0

% by a-GLP, 18.2 to 60.0 % using GLP and from 7.2

to 47.7 % by a-TSP and also, 18.7 and 58.1 % by TSP

at temperature of 298 to 328 K respectively. The

adsorption capacity increases with the increasing

temperature, indicating that the adsorption is an

endothermic process. This may be a result of an

increase in the mobility of chemicals (contributing

COD and BOD) with increasing temperature. An

increasing number of molecules may also acquire

sufficient energy to undergo an interaction with active

sites at the surface. Furthermore, the increasing

temperature may produce a swelling effect within the

internal structure of adsorbents (NLPs, GLPs and

TSPs), enabling to penetrate further (Malkoc et al.,

2007).

3.5. Effect of pH

The effect of pH (3 to 11) is represented in Table 4

using NLPs, GLPs and TSPs (dose: 5.0 g/L) for

removal of COD, BOD and color from textile

wastewater at constant contact duration of 180 min,

temperature of 300 K and agitator speed of 400 rpm.

The percentage removal was increasing, as pH of

system increases. The highest percentage removal of

COD was found to be 64.8, 55.5, 57.4, 47.5, 50.2 and

56.5 by a-NLP, NLP, a-GLP, GLP, a-TSP and TSP

respectively at pH 11. The highest percentage removal

of BOD was found to be 60.0, 54.2, 48.7, 40.2, 41.2

and 44.7 by a-NLP, NLP, a-GLP, GLP, a-TSP and

TSP respectively at pH 11. Though, the factor pH is

an important parameter in adsorption studies, but due

to presence of various elements in dyeing mill

wastewater such as starches, dextrin, gums, glucose,

waxes, pectin, alcohol, fatty acids, acetic acid, soap,

detergents, sodium hydroxide, carbonates, sulfides,

sulfites, chlorides, dyes, pigments, carboxymethyl

cellulose, gelatin, peroxides, silicones, flourcarbons,

resins; the moderate removal of COD, BOD and color

was found with change of factor, pH (O’Neill et al.,

1999).



37

Table 3: Effect of temperature for percentage COD and BOD removal using naturally prepared adsorbents

Adsorbent

% Removal

Temperature (K)

298 303 308 313 318 328 328

a-NLP COD 39.1 48.6 56.0 62.2 70.1 75.6 76.0

BOD 34.7 41.4 48.0 52.6 60.1 63.0 63.2

NLP COD 20.1 32.2 45.5 50.5 61.2 65.6 65.6

BOD 20.2 32.2 35.6 42.5 51.2 55.5 55.5

a-GLP COD 30.3 39.1 54.6 62.2 68.3 72.3 73.0

BOD 29.0 35.0 45.4 52.1 61.6 67.6 68.0

GLP COD 20.0 28.5 44.2 52.1 57.2 61.2 61.2

BOD 18.2 25.5 34.4 41.1 54.2 60.0 60.0

a-TSP COD 15.2 20.2 26.5 37.3 44.5 47.5 47.5

BOD 7.8 12.5 25.5 35.8 43.6 47.7 47.7

TSP COD 24.3 37.9 44.9 52.2 56.5 62.5 63.0

BOD 18.7 25.9 38.6 45.9 51.6 58.1 58.1

Table 4: Effect of temperature for percentage COD and BOD removal using naturally prepared adsorbents

Adsorbent % Removal pH

3 5 7 9 11

a-NLP COD 14.8 23.6 39.0 56.5 64.8

BOD 9.7 17.6 32.1 50.1 60.0

NLP COD 10.0 15.5 27.4 45.5 55.5

BOD 5.2 10.5 20.5 41.4 54.2

a-GLP COD 7.5 14.8 26.5 42.1 57.4

BOD 3.2 10.2 20.2 34.7 48.7

GLP COD 5.5 10.2 15.5 30.2 47.5

BOD 1.2 5.5 11.1 25.5 40.2

a-TSP COD 4.5 12.5 24.5 41.5 50.2

BOD 3.5 15.2 20.5 31.5 41.4

TSP COD 8.0 24.2 34.8 50.4 56.5

BOD 5.7 21.7 31.7 40.0 44.7

3.6. Adsorption Model

Table 5 described Freundlich and Langmuir

parameters for COD and BOD removal using

naturally prepared adsorbents i.e. NLPs, GLPs and

TSPs. Freundlich adsorption capacity, KF was found

to be 79.39 and 41.64 L/mg using a-NLP and 45.12

and 40.25 L/mg using NLP for COD and BOD

respectively. Freundlich adsorption capacity was

found to be 28.69 and 12.13 L/mg using a-GLP and

20.45 and 15.18 L/mg using GLP for COD and BOD

respectively. Freundlich adsorption capacity was

found to be 20.45 and 15.18 L/mg using a-TSP and

23.07 and 11.71 L/mg using TSP for COD and BOD

respectively. Langmuir adsorption capacity, Qmax for

COD and BOD was obtained 87.58 and 84.51 mg/g

respectively using a-NLP and 80.45 and 79.58 mg/g

respectively using NLP. Langmuir adsorption capacity

for COD and BOD was obtained 85.65 and 81.45

mg/g respectively using a-GLP and 82.41 and 78.57

mg/g respectively using GLP. Langmuir adsorption

capacity for COD and BOD was obtained 75.14 and

70.45 mg/g respectively using a-TSP and 81.25 and

75.54 mg/g respectively using TSP. Freundlich and

Langmuir adsorption capacities indicate the good

sorbing capacity of the naturally prepared adsorbents.

Also, it shows that a-NLP is more preferable

adsorbent then other investigated adsorbents.

Activation of Neem Leaf Powder and Guava Leaf

Powder using sulphuric acid were found more

proficient than regular NLP and GLP. Regular

Tamarind Seed Powder (TSP) was found to be more

efficient than activated TSP using sulphuric acid. This

shows that H2SO4 is effective in creating well-

developed pores on the surface of NLP and GLP with

large surface area and porous structure. But in the case

of TSP, there was no change in surface of TSP while

activation using sulphuric acid. Further, value of

exponent n was greater than 1 (n > 1), so, the

adsorption represents favorable adsorption.

Dubinin–Radushkevich and Flory–Huggins


naturally prepared adsorbents were depicted in Table

6. Maximum Dubinin–Radushkevich monolayer

capacity, qm was found to be 18.45, 15.27 and 10.65

mg/g for NLPs, GLPs and TSPs respectively for COD

and 13.27, 12.17 and 8.54 mg/g for NLPs, GLPs and

TSPs respectively for BOD. All these value are close

to that obtained from the Langmuir isotherm model.

The calculate energy was found to be in the range of

1.2412 - 0.5258 KJ/mol, which confirm physical

adsorption reaction. Physisorption processes have

adsorption energies less than 8 KJ/mol and the energy

in the range of 1.2412 - 0.5258 KJ/mol for removal of

Patel and Vashi



38

COD and BOD using naturally prepared adsorbents

indicates that the sorption process is physisorption and

the positive value of energy (E), of sorption indicates

that the sorption process is endothermic and that

higher solution temperature will favor the sorption

process. The Flory–Huggins adsorption capacity, kFM

was calculated and found to be 0.0555, 0.0450 and

0.0252 mg/g for NLPs, GLPs and TSPs respectively

for COD and 0.0452, 0.0402 and 0.0301 mg/g for

NLPs, GLPs and TSPs respectively for BOD. The

ΔG0 for this study was computed in the range of -

9.1909 to -13.8271 kJ/mol. The negative value of ΔG0

indicates that the sorption process is spontaneous in

nature.

Table 5: Freundlich and Langmuir parameters for COD and BOD removal using naturally prepared adsorbents

Adsorbent Removal Freundlich Parameters Langmuir Parameters

KF (L/mg) n Qmax (mg/g) KL (L/mg)

a-NLP COD 79.39 0.1377 87.58 0.049

BOD 41.64 0.2262 84.51 0.037

NLP COD 45.12 0.1379 80.45 0.045

BOD 40.25 0.1919 79.58 0.012

a-GLP COD 28.69 0.4065 85.65 0.12

BOD 12.13 0.1447 81.45 0.14

GLP COD 20.45 0.3175 82.41 0.21

BOD 15.18 0.2809 78.57 0.54

a-TSP COD 17.4 0.2353 75.14 0.14

BOD 7.45 0.1212 70.45 0.024

TSP COD 23.07 0.1570 81.25 0.17

BOD 11.71 0.1350 75.54 0.016

Table 6: Dubinin–Radushkevich and Flory-Huggins parameters for COD and BOD removal using naturally prepared

adsorbents

Adsorbent Particular Dubinin-Radushkevich Parameters Flory-Huggins Parameters

qm (mg/g) Β (mmol2/J2) E (kJ/mol) nFH kFH (mg/g) ΔG0 (kJ/mol)

a-NLP COD 18.45 4.613 1.2412 -0.699 0.0552 -6.9707

BOD 13.27 10.332 0.9574 -0.657 0.0452 -7.2395

NLP COD 15.21 15.485 0.8747 -0.547 0.0212 -6.1091

BOD 8.27 22.258 0.8547 -0.512 0.0392 -6.7909

a-GLP COD 15.27 0.3547 0.8258 -0.431 0.0450 -10.2111

BOD 12.17 0.3948 0.6847 -0.379 0.0402 -11.2551

GLP COD 12.41 0.5158 0.5258 -0.3954 0.0166 -9.1909

BOD 7.55 0.7414 0.6868 -0.354 0.0305 -10.7929

a-TSP COD 8.65 0.3470 0.7265 -0.407 0.0185 -9.4405

BOD 5.87 0.5081 0.8584 -0.317 0.0205 -12.6157

TSP COD 10.65 0.6587 0.7457 -0.375 0.0252 -10.2332

BOD 8.54 0.8714 0.6585 -0.298 0.0301 -13.8271

Table 7 depicted Redlish-Peterson and Sips


naturally prepared adsorbents, in which an exponent

of Redlish-Peterson, βRP not far from 1, indicate that

the Langmuir isotherm best-fits the isotherm data for

removal of COD and BOD using naturally prepared

adsorbents. The Sips adsorption capacity, qm was

calculated and found to be 19.7, 17.9 and 15.5 mg/g

for NLPs, GLPs and TSPs respectively for COD and

18.5, 17.9 and 10.0 mg/g for NLPs, GLPs and TSPs

respectively for BOD. Similar to the Redlich–Peterson

model constants, the same trends were observed for

the Sips model constants. The exponent, n values were

close to unity. This means that sorption data obtained

in this study is more suited to Langmuir form than

Freundlich.

Table 8 described the Toth and Khan Isotherm


naturally prepared adsorbents. Maximum Toth

adsorption capacity, qm was found to be 14.25, 10.47

and 6.58 mg/g for NLPs, GLPs and TSPs respectively

for COD and 7.87, 5.84 and 4.58 mg/g for NLPs,

GLPs and TSPs respectively for BOD. Further, It is

obvious that for Toth model exponent, nT = 1, this

isotherm reduces to the Langmuir sorption isotherm

equation. Maximum Khan adsorption capacity, qm was

found to be 17.34, 11.44 and 11.14 mg/g for NLPs,

GLPs and TSPs respectively for COD and 15.82,

10.78 and 10.47 mg/g for NLPs, GLPs and TSPs

respectively for BOD.



39

Table 7: Redlish-Peterson and Sips parameters for COD and BOD removal using naturally prepared adsorbents

Adsorbent Particular

Redlish-Peterson Parameters Sips Parameters

KRP

(1/mg) aRP (Lβmg–β) βRP qm (mg/g) Ks (L/mg) n

a-NLP COD 2.40 0.234 0.874 19.7 1030.2 0.734

BOD 3.18 7.489 0.907 18.5 9217.4 0.847

NLP COD 1.25 45.12 0.915 14.4 9845.5 0.745

BOD 1.02 0.245 0.995 10.8 5545.5 0.824

a-GLP COD 1.23 61.67 0.940 17.9 8462.5 0.885

BOD 0.31 0.032 1.007 15.2 1251.4 0.755

GLP COD 0.915 75.45 0.954 10.5 4578.5 0.705

BOD 0.29 0.025 0.54 7.4 1020.2 0.821

a-TSP COD 0.83 149.8 0.991 12.1 1656.5 0.927

BOD 0.24 0.004 0.96 7.2 768.5 0.835

TSP COD 0.71 158.5 0.542 15.5 754.5 0.812

BOD 0.12 0.0011 0.547 10.0 625.5 0.714

Table 8: Toth and Khan parameters for COD and BOD removal using naturally prepared adsorbents

Adsorbent Particular Toth Isotherm Khan Isotherm

qm (mg/g) bT (1/mg) nT qm (mg/g) bK (L/mg) aK

a-NLP COD 14.25 0.025 0.854 17.34 0.245 0.057

BOD 7.87 0.058 0.754 15.82 0.547 0.047

NLP COD 11.24 0.047 0.687 14.14 0.657 0.068

BOD 4.58 0.052 0.847 10.25 0.847 0.084

a-GLP COD 10.47 0.047 0.856 11.44 0.574 0.064

BOD 5.84 0.065 0.884 10.78 0.632 0.057

GLP COD 8.66 0.043 0.785 7.45 0.478 0.068

BOD 4.58 0.045 0.847 15.45 0.558 0.062

a-TSP COD 5.58 0.042 0.748 8.45 0.874 0.041

BOD 3.85 0.028 0.789 7.45 0.658 0.038

TSP COD 6.58 0.045 0.840 11.14 0.782 0.057

BOD 4.58 0.036 0.705 10.47 0.687 0.047

4. CONCLUSION

This study provides characterization of textile

wastewater and, simple and effective adsorption

process for COD and BOD removal using naturally

prepared adsorbents (NLP, GLP and TSP) and their

activation forms using sulfuric acid. It is observed

from the present investigation that the textile

wastewater contains various types of pollutants, which

are not easily removed. Batch adsorptive treatment

was carried out using various parameters like

adsorbent dose, conduct duration, temperature and

pH, in whixch activation of Neem Leaf Powder and

Guava Leaf Powder using sulphuric acid was found

more proficient than investigated regular NLPs and

GLPs. Regular Tamarind Seed Powder (TSP) was

found to be more efficient than activated TSP using

sulphuric acid. The adsorption equilibrium isotherms,

viz. Freundlich, Langmuir, Dubinin-Raduskevich,

Flory-Huggins, Redlich-Peterson, Sips, Toth and

Khan isotherm ware analyzed, in which maximum

adsorption capacities of linear equation Langmuir

isotherm was calculated and found to be 87.58, 85.65

and 81.25 mg/g for NLPs, GLPs and TSPs

respectively for COD and 84.51, 81.45 and 75.54

mg/g for NLPs, GLPs and TSPs respectively for

BOD. Also, Langmuir isotherm best-fits the isotherm

data for removal of COD and BOD using naturally

prepared adsorbents. This sorption process is

physisorption and endothermic and that higher

solution temperature will favor the sorption process,

derived from calculated free energy, E. The negative

value of ΔG0 indicates that the sorption process is

spontaneous in nature.

REFERENCES

Alaton IA, Insel G, Eremektar G, Babuna FG, Orhon

D (2006). Effect of textile auxiliaries on the

biodegradation of dyehouse effluent in

activated sludge, Chemosphere, 62: 1549-1557.

APHA (1992). Standard Methods for the examination

of water and wastewater, 18th ed. American

Public Health Association, Washington, DC,

USA.

Budyanto S, Soedjono S, Irawaty W, Indraswati N

(2008). Studies of Adsorption Equilibria and

Kinetics of Amoxicillin from Simulated

Wastewater using Activated Carbon and

Natural Bentonite, Journal of Environmental

Protection Science, 2: 72-80.

Fanchiang J, Tseng D, Guo G, Chen H (2009).

Ozonation of complex industrial park

wastewater: effects on the change of wastewater

Patel and Vashi



40

characteristics, Journal of Chemical

Technology and Biotechnology, 84: 1007-1014.

Foo K Y, Hameed B H (2010). Insights into the

modeling of adsorption isotherm systems,

Chemical Engineering Journal, 156: 2-10.

Gulipalli CS, Prasad B, Wasewar KL (2011). Batch

Study, Equilibrium and Kinetics of Adsorption

of Selenium using Rice Husk Ash (RHA),

Journal of Engineering Science and

Technology, 6(5): 586-605.

Hameed B H, Mahmoud D K, Ahmad AL (2008).

Equilibrium modeling and kinetic studies on the

adsorption of basic dye by a low-cost

adsorbent: Coconut (Cocos nucifera) bunch

waste, Journal of Hazardous Materials, 158: 65-

72.

Israel A, Ogali R, Akaranta O, Obot IB (2010).

Removal of Cu(II) from aqueous solution using

coconut (Cocos nucifera L.) coir dust, Der

Pharma Chemica, 2(5): 60-75.

Li W, Zhang L, Peng J, Li N, Zhang S, Guo S (2008).

Tobacco stems as a low cost adsorbent for the

removal of Pb(II) from wastewater: Equilibrium

and kinetic studies, Industrial Crops and

Products, 28: 294-302.

Malkoc E, Nuhoglu Y (2007). Potential of tea factory

waste for chromium(VI) removal from aqueous

solutions: Thermodynamic and kinetic studies,

Separation and Purification Technology 54:

291-298.

Martins AF, Wilde ML, da Silveira C (2006).

Photocatalytic Degradation of Brilliant Red

Dye and Textile Wastewater, Journal of

Environmental Science and Health, Part A, 41:

4, 675-685.

Mohanty K, Das D, Biswas MN (2006). Preparation

and characterization of activated carbons from

Sterculia alata nutshell by chemical activation

with zinc chloride to remove phenol from

wastewater, Adsorption, 12: 119-132.

Ozacar M, Sengil IA (2005). Adsorption of metal

complex dyes from aqueous solutions by pine

sawdust, Bioresource Technology, 96(7): 791-

795.

Patel H, Vashi RT (2011). Equilibrium, kinetic and

thermodynamic studies for adsorption of

Reactive dyes onto Activated Guava leaf

powder, Fresenius Environmental Bulletin,

20(2): 561-568.

Patel H, Vashi RT (2012). A comparison study of

removal of Methylene blue dye from aqueous

solution by adsorption on Neem Leaf Powder

(NLP) and activated NLP. Journal of

Environmental Engineering and Landscape

Management, 21(1): 36-41.

Patel H, Vashi RT (2010). Adsorption of Crystal

Violet dye onto Tamarind seed powder, E-

Journal of Chemistry, 7(3): 975-984.

Pathe PP, Biswas AK, Rao NN, Kaul SN (2005).

Physico-chemical Treatment of Wastewater

From Clusters of Small Scale Cotton Textile

Units, Environmental Technology, 26: 3, 313-

328.

O’Neill C, Hawkes FR, Hawkes DL, Lourenc ND,

Pinheiro HM, Delee W (1999). Colour in textile

effluents-sources, measurement, discharge

consents and simulation: a review, Journal of

Chemical Technology and Biotechnology, 74:

1009-1018

Quintelas C, Fernandes B, Castro J, Figueiredo H,

Tavares T (2008). Biosorption of Cr(VI) by a

Bacillus coagulans biofilm supported on

granular activated carbon (GAC), Chemical


Sulaymon A H, Ebrahim S E, Mohammed-Ridha MJ

(2013). Equilibrium, kinetic, and

thermodynamic biosorption of Pb(II), Cr(III),

and Cd(II) ions by dead anaerobic biomass

from synthetic wastewater, Environmental

Science and Pollution Research, 20(1): 175-

187.

Venkata Mohan S, Sailaja P, Srimurali M,

Karthikeyan J (1999). Color removal of

monoazo acid dye from aqueous solution by

adsorption and chemical coagulation,

Environmental Engineering and Policy, 1: 149-

154.



©2014 IJSRPUB

41

Fenton Oxidation for the Treatment of Liquid Waste with High COD and

Anionic/Non-ionic Surfactants

M.C. Collivignarelli1, S. Sorlini

2*, A. Abbà

1, M. Sordi

1

1Department of Civil Engineering and Architecture, University of Pavia, via Ferrata 1, 27100, Pavia, Italy

2Department of Civil Engineering, Architecture, Land, Environment and Mathematics, University of Brescia, via Branze 43,

25123, Brescia, Italy

*Corresponding Author: Email: [email protected]

Abstract. In the present work, liquid wastes from different kinds of industrial plants, characterized by high COD (up to

100,000 mg/L) and surfactants (up to 16,000 and 3,800 mg/L for non-ionic and anionic respectively) concentrations were

treated by means of Fenton oxidation. During the experimental research 28 tests at laboratory scale were performed. Each test

was characterized by specific dosage of reagents (Fe2+

e H2O2) and contact time. In particular, Fe2+

/H2O2 ratio varied from 0.25

to 0.50, H2O2/COD ratio from 0.72 to 1.5 while the contact time increased from 30 to 120 minutes. The main objective of this

work was to define the dosage of reagents and the reaction contact time able to optimize the process performance in terms of

removal yields of COD and surfactants. The results showed that the optimal treatment conditions could be obtained with a

Fe2+

/H2O2 ratio equal to 0.25, a H2O2/COD ratio of 1 and a reaction time of 30 minutes. An average removal yield of 70% for

AS (non-ionic surfactants) and COD and 95% for MBAS (anionic surfactants) respectively was obtained.

Keywords: Surfactants, Fenton Oxidation, High COD

1. INTRODUCTION

Surfactants are a group of compounds used daily in

huge amounts mainly in household applications and as

industrial cleaning agents (Gonzalez et al., 2007;

Hosseinnia et al., 2006). As is known, the surfactants

can be classified in different groups depending on the

electrostatic charge of its hydrophilic groups: anionic

(MBAS), non-ionic (TAS), cationic and amphoteric

surfactants (the last compounds behave as acids or

bases, depending on the solution acidity). The first

two groups are the most common and account for over

80% of the total usage in detergents. Cationic

surfactants are used mainly in fabric conditioners to

give a pleasant soft feel to the product while

amphoteric surfactants are used mainly for their skin

mildness properties (AISE, 2012). The new European

Detergent Regulation (Regulation 648/2004, that

entered in force on 2005) requires that the surfactants

used in household detergents must be biodegradable,

while derogation may be accepted for surfactants in

detergents used in special industrial or institutional

sectors. In Western Europe each year over 2.5 million

tonnes of surfactants are produced (45% anionic, 43%

non-ionic, 8% cationic, 4% amphoteric) (CESIO,

2012).

In the past, some types of surfactants (non-ionic)

resulted to be dangerous for the humans and the

environment also after biological treatment. In fact,

although they were easily removable in well operating

treatment plants with removal yields of about 90-95%,

the last oxidation step was much slower due to the

formation of by-products that are toxic and refractory

to biological treatment (Gonzalez et al., 2007). The

main methods for the treatment of sewage and liquid

wastes containing surfactants involve chemical and

physical processes such as coagulation, foaming,

advanced chemical oxidation, adsorption on different

types of active carbon and polyelectrolytes (Kowalska

et al., 2006). Liquid wastes produced in personal care

product manufacturing are primarily treated by

aerobic activated sludge systems coupled with

physico-chemical methods, although biological

methods are generally favoured because they produce

more tractable solids residuals (Ahammad et al.,

2013). Anaerobic biological processes typically are

not used for high color waster because of the presence

of compounds that inhibit anaerobic microorganisms

(Ahammad et al., 2013). Often liquid wastes

containing surfactants (such as those arising from

industries for the production of personal care

products) are characterized by high COD

concentrations and by poorly biodegradable

compounds that may be toxic (Dias de Melo et al.,

2013). Also liquid waste from tannery industries

belong to the same typology, so conventional

treatments produce effluents still do not meet the

required limits, at least for some parameters such as

COD, salinity, ammonia and surfactants (Kurt et al.,

2007). Hence chemical oxidation systems appear to be

more suitable in treating liquid wastes characterized

by poorly biodegradable and toxic compounds and

high concentrations contents in COD and surfactants.

In the present study Fenton oxidation was

experimented at laboratory scale in order to evaluate

the treatment of colored liquid wastes with high

concentrations of COD and surfactants.


Collivignarelli et al.

Fenton Oxidation for the Treatment of Liquid Waste with High COD and Anionic/Non-ionic Surfactants

42

2. MATERIALS AND METHODS

2.1. Equipment used for Fenton oxidation

Fenton experimental tests were carried out at lab scale

with a 2 liter-glass beaker placed over a magnetic

stirrer (Figure 1 a). A probe for the measurement of

pH was placed inside the beaker (Figure 1 b). Each experimental test was carried out according

to the following procedure:

(1) Addition of 1 L- volume of liquid waste into

the beaker;

(2) pH adjustment of the liquid waste through the

gradual addition of 1 N sulfuric acid until the desired

pH test;

(3) Dosage of reagents (hydrogen peroxide 40%

w/v solution and FeSO4∙7H2O ferrous salt in solid

form) and start of the reaction; both the two reagents

were dosed gradually inside the beaker in order to

allow a better distribution within the mixture;

(4) Monitoring of the reaction, at regular intervals,

maintaining the sample in stirring conditions: H2O2

and Fe2+

concentrations, temperature and pH were

measured;

(5) pH increasing up to 8.5-9 by gradual dosage of

lime milk (Ca(OH)2) in solution at 5%;

(6) settling of the mixture for 2 hours by means of

an Imhoff cone in order to observe the sedimentation

characteristics of the treated liquid waste (Figure 1c);

the level of the solid/liquid interface was recorded

every 30 minutes;

(7) Samples extraction by 0.45 µm (paper filter)

filtration of the supernatant obtained from step 6.

Fig. 1: Equipment used for Fenton oxidation tests 1 a - pH measurement 1 b - settling of the mixture 1 c

2.2. Characteristics of liquid wastes

Table 1 shows the qualitative characteristics of the

liquid wastes (LW) treated during the experimental

tests: raw LW II - III - IV and LW I, previously

mixed with an acid liquid waste (characterized by

COD concentration of 30000 mg/L and pH < 1.5) that

allowed to reach the optimal pH without the addition

of sulfuric acid (LW I - mix 1 and mix 2 contained

83% and 17% of LW I respectively).

Table 1: Qualitative characteristics of the liquid wastes

PARAMETER U.M. LW I LW II LW III LW IV

COD mg/L 17800 56680 50450 101900

TAS mg/L 24.7 1033 845 16000

MBAS mg/L 102 1490 13100 3870

BOD5 mg/L 8000 6000 - -

BOD20 mg/L - - - -

TN mg/L 307 < 0.5 - -

N-NH4+ mg/L 37.4 < 0.5 8.8 65

N-NO3- mg/L 17.3 < 0.5 - -

N-NO2- mg/L 0.2 < 0.1 - -

TP mg/L 393 < 0.5 - -

pH - 6.7 5.8 5.5 8.6

color/appearance - green/limpid white/mushy gray/mushy brown/limpid



43

2.3. Experimental tests

The tests were carried out with different dosages of

reagents (expressed as H2O2/COD and Fe2+

/H2O2),

reaction time and temperature in order to determine

the optimal process conditions. The detailed

experimental conditions are reported in Table 2.

Table 2: Experimental conditions applied during Fenton oxidation

Tests Liquid Waste (LW) Operative conditions

Contact time [minutes] (H2O2/COD)0 Fe2+/H2O2

F 1

I

(mix 1)

30 0.36 1/2

F 2 30 0.36 1/5

F 3 30 0.72 1/4

F 4 60 1 1/2

F 5 60 0.72 1/2

F 6 I

(mix 2)

60 0.72 1/2

F 7 60 1.17 1/2

F 8 60 2 1/4

F 11

II

60 0.72 1/4

F 12 60 1 1/4

F 13 60 0.72 1/2

F 14 30 0.72 1/4

F 15 60 1.5 1/4

F 16 120 0.72 1/4

F 17

III

60 0.72 1/4

F 18 60 1 1/4

F 19 60 1.5 1/4

F 20 30 0.72 1/4

F 21 60 1 1/2

F 22 120 1 1/4

F 23

IV

60 0.72 1/4

F 24 60 1 1/4

F 25 60 1.5 1/4

F 26 30 0.72 1/4

F 27 60 1 1/2

F 28 120 1 1/4

2.4. Analytical methods

The physical-chemical parameters such as COD, N-

NH4+, N-NO2

-, N-NO3

-, TN (total nitrogen), MBAS

(anionic surfactants), TAS (non-ionic surfactants), pH

were measured according to the standard methods for

water and wastewater (APHA et al., 2001). TP (total

phosphorus) was analyzed by ICP-MS. BOD5 was

determined at 20 °C by inoculation of active biomass

collected from an activated sludge wastewater

treatment plant (WWTP).


3.1. Effect of the Fe2+

dosage

In order to assess the optimal dosage of iron, each

liquid waste was tested with two ratios of Fe2+

/H2O2:

0.25 and 0.5 respectively. The removal yields of COD

and surfactants (MBAS and TAS) versus Fe2+

/H2O2

and with different contact time and H2O2/COD ratio

are reported in Figures 2 and 3 respectively. As

concerns COD the optimal dosage of iron seems to be

Fe2+

/H2O = 0.25, for which most of the processed LW

showed better removal yields for reaction times of 30

and 60 minutes and for H2O2/COD ratio equal to 0.36,

0.72 and 1 (LW I - mix1, LW II, LW III). The tests

performed on the LW II and LW IV show a decrease

of COD removal yields with increasing the iron

dosage. The best removal yields were achieved in the

treatment of liquid waste LW III with values slightly

below 90%.



44

Fig. 2: Effect of the iron dosage on COD removal

The removal yields of surfactants as a function of

the iron dosage are shown in Figure 3. In this case

doubling the iron dosage a reduction in the removal

yields of surfactants occurred (LW - mix 2, LW II,

LW IV), in particular for TAS that showed a higher

reduction than MBAS (LW I - mix 2, LW II).

Doubling Fe2+

/H2O2 ratio an improvement of

surfactant removal yields was observed only for LW I

- mix 1 and LW III. Moreover, the anionic surfactants

are the most easily removed, as shown by the average

yield obtained for all the tests equal to about 90%and

60% respectively for anionic and non-ionic

surfactants.

Fig. 3: Effect of the iron dosage on TAS and MBAS removal

3.2. Effect of the H2O2 dosage

After determining optimal dosages of iron, different

laboratory tests were carried out on the same liquid

wastes in order to evaluate the optimal dosage of

hydrogen peroxide. The H2O2/COD ratios were 0.72,

1, 1.5 and only for LWI - mix 2 the 1.17 ratio was

added. Figure 4 shows the removal yields of COD

obtained during these tests. All liquid wastes showed

an increase of COD removal yields with increasing

the dosage of hydrogen peroxide with the exception of

the LW III for which the removal yield was constantly

about 85%.



45

Fig. 4: Effect of the hydrogen peroxide dosage on COD removal

The anionic surfactants showed a similar behavior

to COD. Instead, for the non-ionic surfactants (TAS)

the yield of removal (Figure 5) significantly increased

with increasing the dosage of hydrogen peroxide with

respect to the initial COD from 0.72 to 1. A further

increase of the dosage of hydrogen peroxide (from 1

to 1.5 H2O2/COD ratio), does not lead to any

improvement in TAS removal yields (as concerns LW

II, LW III LW IV).

Fig. 5: Effect of the hydrogen peroxide dosage on TAS and MBAS removal

3.3. Effect of contact time

Finally, the effect of the contact time on the removal

of COD (Figure 6) and surfactants (Figure 7) was

evaluated. The results obtained showed that the

optimal contact time was equal to 30 minutes for each

liquid waste. Indeed, the low increase of removal

yields is observed for COD increasing the contact

time from 30 to 60 minutes may be partially caused by

an increase of the dosage of hydrogen peroxide (LW

III, LW IV) or iron (LW I - mix 2).



46

Fig. 6: Effect of reaction time on COD removal

The effect of contact time on surfactants removal is

shown in Figure 7. The optimal contact time for

anionic surfactants for all the tested liquid wastes is 30

minutes. Going into detail, for LW II a removal yield

of 90% was obtained after a 30 minutes contact time;

moreover, the slight increase of the removal efficiency

observed for the LW I - mix 2 at 60 minutes may be

due to the use of a higher dosage of iron. Regarding

TAS, as observed for COD, all tested liquid wastes

(LW - mix 2, LW III, LW IV) show the best removal

condition at 30 minute contact time, with the

exception of LW II for which the removal yield

increased from 40% to 60% and 80% with increasing

the test duration from 30 minutes to 60 and 120

minutes respectively.

Fig. 7: Effect of reaction time on TAS and MBAS removal

4. CONCLUSION

The results showed that the optimal treatment

conditions for surfactant removal from liquid wastes

can be obtained with a Fe2+

/H2O2 ratio equal to 0.25

and a H2O2/COD ratio of 1 and a reaction duration of

30 minutes. These process conditions allow obtaining

an average removal yield of 70% for TAS and COD

95% for MBAS. In fact, the experimental data

obtained show that higher dosages of Fe2 +

produce for

almost all tested liquid wastes a lower removal yield

of both the surfactants. Instead, regarding H2O2 the



47

used of dosages lower than the optimal one reduces

the removal yields of both TAS and MBAS, while

higher dosages do not significantly improve the

process efficiency). Finally, an increase of contact

time over 30 minutes does not generally improve the

removal of surfactants.

ACKNOWLEDGEMENT

The authors would like to give a special thanks to the

company ASMortara spa for the financial support to

the experimental research; moreover we want to thank

l'Eng. Chiara Clemente for the assistance in carrying

out the tests.

REFERENCES

Ahammad SZ, Zealand A, Dolfing J, Mota C,

Armstrong DV, Graham DW (2013). Low-

energy treatment of colourant wastes using

sponge biofilters for the personal care product

industry. Bioresource Technology, 129: 634–

638

AISE (2012). The new detergents regulation: fact

sheet on aerobic biodegradation of surfactants.

http://www.aise.eu/downloads/05_Fact%20shee

t%20biodegradability-

updated%2027082012.pdf

APHA, AWWA, WEF (2001). Standard methods for

the examination of water and wastewater, 21st

edition. American Public Health Association,

Washington DC, USA.

CESIO - European Committee of Organic Surfactants

and their Intermediates (2012). CESIO

surfactants statistics for Western Europe.

http://www.cefic.org/Documents/About-

Us/Industry%20sectors/CESIO/CESIO-

Statistics-2012.pdf

Dias de Melo E, Mounteer A H, Henrique de Souza

Leão L, Cibele Barros Bahia R, Ferreira

Campos IM (2013). Toxicity identification

evaluation of cosmetics industry wastewater.

Journal of Hazardous Materials, 244–245: 329–

334.

European Commission (2004). Regulation (EC) No

648/2004 of the European Parliament and of the

Council of 31 March 2004 on detergents.

Official Journal of the European Union.

http://eur-

lex.europa.eu/LexUriServ/LexUriServ.do?uri=

OJ:L:2004:104:0001:0035:en:PDF

Gonzalez S, Petrovic M, Barcelo D (2007). Removal

of a broad range of surfactants from municipal

wastewater – Comparison between membrane

bioreactor and conventional activated sludge

treatment. Chemosphere, 67: 335 – 343.

Hosseinnia A, Hashtroudi M S, Pazouki M,

Banifatemi M (2006). Removal of surfactants

from wastewater by Rice Husk. Iranian Journal

of Chemical Engineering, 3(3).

Kowalska I, Majewska-Nowak K, Kabsch-

Korbutowicz M (2006). Ultrafiltration

treatment of detergent solutions. Desalination

200: 274–276.

Kurt U, Apaydin O, Gonullu MT, (2007). Reduction

of COD in wastewater from an organized

tannery industrial region by electro-fenton

process. Journal of Hazardous Materials, 143:

33–40.

http://www.aise.eu/downloads/05_Fact%20sheet%20biodegradability-updated%2027082012.pdf



http://www.cefic.org/Documents/About-Us/Industry%20sectors/CESIO/CESIO-Statistics-2012.pdf



http://eur-lex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:2004:104:0001:0035:en:PDF





©2014 IJSRPUB

48

Ultrasound Irradiation on the Treatment of Aromatic Compounds in Wastewater

Wong Lai Peng, Guo Xinxin, Mohammed J. K. Bashir*

Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman,

31900 Kampar, Perak, Malaysia


Abstract. The presence of aromatic compounds such as, chlorinated aromatic compounds, phenolic compounds and dyes can

contaminate aquatic environment. Owing to the toxic, mutagenic and carcinogenic properties of some of aromatic pollutants,

their levels need to be controlled strictly in treated wastewater. Nevertheless, most of the conventional treatments methods are

not able to remove the toxicity of these pollutants completely. Ultrasonication treatment has been employed for the treatment

of hazardous materials in wastewater. Thus, the aim of this study is to investigate the characteristics, influencing factors and

applications of ultrasonic irradiation method. Also, review the current situation of applying cavitation process individually and

combined with other AOPs treatment methods for treatment of aromatic pollutants. In addition, the possible degradation

mechanism of aromatic compound is discussed and the configuration of ultrasonic equipment is suggested.

Keywords: Ultrasound Irradiation, Wastewater, Aromatic Compounds

1. INTRODUCTION

Recalcitrant aromatic compounds have been

extensively used in many industrial processes like

textile industry for dying leather, silk wool and paper.

Consequently, these industrial wastewater streams

always contain high concentration of aromatic

compounds which is harmful to human due to their

carcinogenicity and genotoxicity by accumulating in

fat tissue (Berberidou et al., 2007; Gogate and Pandit,

2004a; Gregory, 2009; Ju et al., 2008). Therefore it is

necessary and important to remove aromatic

compounds from wastewater before discharge and

keep the concentration of chemicals in the effluent

stream at a certain minimum level (Gogate and Pandit,

2004a; Zhou et al., 2012). Due to the stringent

environmental laws and regulations against hazardous

pollutants, the technologies for wastewater treatment

have been extensively developed in the last decades

(Pera-Titus et al, 2004). However, aromatic

compounds are resistant to conventional biological

wastewater treatment process due to high toxicity and

carcinogenicity of the pollutants. So, it is necessary to

develop novel and efficient methods degrade

biorefractory aromatic compounds into smaller

molecules, which can be complete minimized by

conventional biological or photochemical

technologies.

Researchers have tested various methods with

advanced oxidation processes (AOPs) for the

degradation of biorefractory aromatic compounds,

such as ozone oxidation, hydrogen peroxide oxidation,

Fenton with UV light or catalysts and cavitation

generated by ultrasonic irradiation (Ai et al., 2010;

Berberidou et al.,2007; Guinea et al., 2009; Meric et

al., 2005; Zhang and Zheng, 2009). Cavitation have

been found to be the most attractive and suitable

system in degradation or decolorization of recalcitrant

organic pollutants and also used as pretreatment to

convert pollutants into shorter chain compounds that

can be treated by conventional or biological methods

(Anjaneyulu et al., 2005; Gogate and Pandit, 2004a;

Gogate and Pandit, 2004b). This method is based on

the producing of highly reactive hydroxyl radicals

(·OH) with high oxidation potential compared to the

conventional oxidants like potassium permanganate

and hydrogen peroxide (Gogate and Pandit, 2004a).

The combination of different AOPs has been found to

be more efficient for the treatment of aromatic

compounds as compared to individual oxidation

process due to the high energy efficiency and

production of higher amount of free radicals

(Chakinala et al., 2008; Ioan et al., 2007; Namkung et

al., 2008; Papadaki et al., 2004; Sun et al., 2007).

The aim of this study is to consider the principles,

characteristics, influencing factors and applications of

ultrasonic irradiation method and takes an overview of

the current situation on applying cavitation process

individually and combined with other AOPs treatment

methods for treatment of aromatic pollutants including

chlorinated aromatic pollutants, phenolic pollutants,

dyes etc. In addition, the possible degradation

mechanism of aromatic compound is discussed and

the configuration of ultrasonic equipment is

suggested.

2. BASIC PRINCIPLES OF ULTRAONIC

CAVITATION AND SONOCHEMISTRY

Ultrasound is the term given to sound waves above 20

kHz which is the frequency above the range audible to

human being. Depending on the application and

frequency, ultrasonication is broadly divided into

three areas: i) low frequency or conventional power

ultrasound (20-100 kHz), (ii) medium frequency or

sonochemical effects ultrasound (300-1000 kHz) and


Peng et al.


49

iii) high frequency or diagnostic ultrasound (2-10

MHz) (Mason and Lorimer, 1988; Show and Wong,

2012). The effect of ultrasound in liquid medium was

first recognized by Alfred L. Loonis in 1927, however

the major developments in sonochemistry took place

in 1980s after the arrival of inexpensive high-intensity

ultrasound generators (Suslick, 1989). Recently, the

majority utilization of ultrasound in chemical reaction

included synthesis (organic, organometallic and

inorganic), polymer chemistry (degradation, initiation,

and copolymerization) and nanomaterial preparation

and some aspects of catalysis (Mason and Lorimer,

1988). In addition, ultrasound show great potential in

environmental engineering and protection, thus since

1990 there has been an increased interest in the use of

ultrasound to destroy organic contaminants present in

water and/or wastewater (Hao et al., 2003;

Chowdhury and Viraraghavan, 2009; Eden, 2012).

Environmental sonochemistry is a rapidly growing

area, and cavitation is an effective tool for degrading

different organic pollutants, including aromatic

compound (Hamdaoui et al., 2008; Minero et al.,

2008; Sonawane et al., 2009; Suslick, 1990).

Ultrasonic irradiation is generated by directly

introducing the source, i.e. a probe in the reactor or

immersing the reactor in an ultrasonic bath (Hong et

al., 1999; Vinodgopal and Peller; 2003). Ultrasonic

bath is the most widely used and cheapest source of

ultrasound however it does not produce better

sonochemical effects compare to the probe-type

ultrasonic irradiation (Goel et al., 2004).

2.1. Cavitation

The diffuse energy of sound is enhanced through

cavitation (Suslick, 1990). The process is principally

based on the phenomenon of acoustic cavitation,

which includes the formation, growth, and implosive

collapse of micro-bubbles in a liquid. When a liquid is

sonicated, positive and negative pressures are exerted

on a liquid by compression and expansion cycles

respectively of ultrasound waves (Suslick, 1989). The

dissolved gas molecules are entrapped by micro-

bubbles that grow and expand upon rarefaction of the

acoustic cycle; when a sufficiently large negative

pressure is applied to the liquid, the average distance

between the molecules would exceed the critical

molecular distance necessary to hold the liquid intact,

and the liquid will break down and voids or cavities

will be created; cavitation bubbles will then be formed

(Mason and Lorimer, 1988). These cavities, voids or

bubbles may grow in size until the maximum of the

negative pressure then extreme temperature release

upon adiabatic collapse (Mason and Lorimer, 1988;

Lin et al., 2008; Wang et al., 2007).

There are three possible reaction sites in

ultrasonically irradiated homogeneous liquids: i) the

gaseous interiors of collapsing cavities; ii) the

interfacial liquid region between cavitation bubbles

and the bulk solution, where high temperature and

high temperature gradients exist; and, iii) the bulk

solution at ambient temperature, where small amounts

of hydroxyl radicals (•OH) diffuse from the interface

(Ghodbane and Hamdaoui, 2009; Ozen et al., 2005).

The high temperatures (5000 K) and pressures (1000

atm) induced by cavitation in collapsing gas bubbles

in aqueous solution lead to the thermal dissociation of

water molecules into hydrogen radicals (•H) and

hydroxyl radicals (•OH) (Crum et al., 1999; Shimizu

et al., 2007).

The sonochemical effect takes place at the gas-

liquid interface due to the oxidation of organic

molecules to lesser extent by •OH in the bulk solution

or the pyrolytic decomposition inside the bubbles

(Goel et al., 2004; Li et al., 2008). Acoustic cavitation

generates hydro-mechanical forces and pyrolytic

reactions, in many cases, there are dominant factors in

the pollutant degradation.

However, it has been demonstrated that the

reaction mechanisms depending on the different

physico-chemical characteristics of the pollutant.

Hydrophobic pollutant degradation mainly happened

in the hydrophobic boundary where pyrolysis and

radical reactions contribute to the degradation. While

hydrophilic pollutants in the bulk liquid degradation

mainly contributed by free radicals or hydrogen

peroxide reactions in the cavitation bubbles

(Behnajady et al., 2008; Tezcanli-Guyer and Ince,

2003; Vajnhandl and Le Marechal, 2007). In brief the

cavitational collapse creates an unusual environment

for a chemical reaction in terms of enormous local

temperature and pressure (Suslick, 1990).

2.2. Basic Theories of Sonochemistry

The wide range of oxidations and reductions that

occurs with aqueous sonochemistry is often a

consequence of secondary reactions of the high

energy intermediates. The locally high pressure,

temperature and the formation of highly reactive

radicals could facilitate and accelerate the chemical

reactions inside the reactors. There are three popular

theories of reactivity have been proposed in applying

ultrasound on wastewater treatment: i) hot-spot

theory, ii) electrical theory, and iii) plasma discharge

theory. The localized short-lived (<10 μs) ―hot spots‖

in an irradiated liquid generated by the rapid collapse

of acoustic cavities implying the existence of

extremely high heating and cooling rates in the

vicinities of 1010 K/s (Ince et al., 2001). The reported

temperature inside the bubble is around 5000 K. The



50

collapse and the implosion of these cavitation bubbles

would result in light emission (sonoluminescence) of

more than 107 photons per flash (Flint and Suslick,

1991; Lohse, 2005).

According to Marguls (1992), during bubble

formation and collapse, large electrical field gradients

are produced. This would lead to sonochemical

reactions and sonoluminescence phenomena. Lepoint

and Mullie (1994) observed the analogies between

sonochemistry and coronachemistry, thus introduced

plasma theory to explain cavitation. They assumed it

as a fragmentation process due to an intense electrical

field rather than a true implosion. However, in the

environmental field, the hot-spot theory is widely

accepted in explaining sonochemical reactions

(Adewuyi, 2001).

2.3. Sonochemistry of Water

As reported by Neis (2000), the ultrasound induced

splitting of water molecules into extremely reactive

hydroxide radicals and hydrogen radicals. During the

quick cooling phase, these radicals recombined again

to form hydrogen peroxide (H2O2) and hydrogen

molecules (H2) as shows in equation in Eqs. 1-5 (Ai et

al., 2007; Ghodbane and Hamdaoui, 2009; Ince and

Tezcanli, 2004; Inoue et al., 2006; Minero et al., 2008;

Suslick, 1989; Wang et al., 2008)

H2O + ultrasound → •H +

•OH (1)

•OH +

•H→ H2O (2)

•H + O2 →

•HO2 (3)

2•OH → H2O2 (4)

2•HO2 → H2O2 + O2 (5)

Violent and fast collapse of the bubbles

compresses adiabatically gas and entrapped vapour,

which conducts to short and local hot spots (Crum,

1995). At the final step of the collapse, temperature

inside the residual bubble is thought to be above 2000

K. Under these conditions, entrapped molecules of

dissolved gases, vaporised water and solutes can be

brought to an excited state and dissociate. The local

concentration of •OH is at its maximum, and a large

majority of •OH are recombined at the gas-liquid

interface before being ejected into the bulk solution.

There is only a small fraction of •OH escapes from the

interfacial region and diffuse into the bulk solution

(Gultekin et al., 2009). In absence of any organic

compound, •OH radicals combine to produce H2O2 in

the bulk solution.

2.4. Cavitation on liquid-solid systems

The most important effects of ultrasound on liquid-

solid systems are mechanical and attributed to

asymmetric cavitation. When a bubble is collapsing in

a spherically asymmetric environment, the collapse

changes in a remarkable way: a flat solid surface

nearby caused the bubble to involutes from the surface

below the top and developed a micro-jet liquid with

speed up to 200 ms-1

. This resulting in newly

developed high reactive surface as well as corrosion

and erosion at the surface. In addition, the implosion

of cavitation bubbles also produced high energy

shockwaves within interfacial films surrounding

nearby solid particles that have the potential to create

microscopic turbulence (Hamdaoui et al., 2008).

These cavitation effects increase the rate of mass

transfer near the catalyst surface and enhance the

reaction rate (Song et al., 2009).

3. INFLUENCING FACTORS IN

SONOCHEMICAL REACTIONS

Under proper condition, there are at least three

successive stages in acoustic cavitation, i) nucleation,

ii) bubble growth and iii) implosive collapse (Suslick,

1990). The effect of cavitation is influenced by a

number of factors, namely ultrasonic frequency,

ultrasonic intensity, ultrasonic density, quid

temperature, type of pollutant, present of gas, external

pressure, viscosity and surface tension.

A few authors explained the cavitation theory with

the help of resonance frequency of bubble (Mason and

Lorimer, 1988). The resonant radius of a bubble is

inversely proportional to the ultrasonic frequency

(Hua and Hoffmann, 1997; Hung and Hoffmann,

1999). The important cavity effects were reported to

occur when the frequency of the wave was equal to

the resonating frequency of the bubbles (Mason and

Lorimer, 1988). Higher frequency was energetically

more favorable, It was also suggested that mass

transfer from liquid to vapor phase and •OH radical

formation were enhanced with high frequency

ultrasonication (Petrier et al., 1994; Hung and

Hoffmann, 1999).

Reaction temperature also influences sonochemical

reactions as it is directly related to solvent vapor

pressure. Higher temperature lowers the threshold

intensity required to produce cavitation. This could be

due to rising of vapour pressure or lowering of

viscosity or surface tension. The maximum

sonochemical benefit is achieved at as low

temperature as is feasible (Mason and Lorimer, 1988).

Cavitation is favored by the liquid of high vapor

pressure and low density. This was supported by the

results of Petrier et al. (1998) for sonochemical

degradation of chlorobenzene and 4-chlorophenol.

Chlorobenzene degradation was much faster than 4-

chlorophenol because of its higher vapor pressure.

Again solvents with higher viscosity, surface tension

Peng et al.


51

and density showed poor cavitation efficiency

(Adewuyi, 2001). The lower the vapour pressure for a

given external pressure, the larger the acoustic

pressure, thus the type of solvent in the systems plays

an important factor.

In general, it is assumed that the bubbles collapse

is an adiabatic process. A higher temperature and

pressure are expected with gases of higher heat

capacity ratio (γ) values. Lower thermal conductivity

and higher solubility provides better cavitation in the

system (Price, 1992).

Cavitation threshold is the minimum amount of

energy required to intitiate cavitation. Only the energy

applied above the threshold will contribute to the

formation of cavitation bubbles. The higher the

intensity level, the higher the acoustic amplitude and

the collapse pressure. Thus, the collapse is faster and

more violent. However, at frequencies greater than 1

MHz, the acoustic wave impacts on the liquid and

creates micro-currents together with stable and

oscillating gas bubbles which will not collapse but

occasionally rise to the surface of the water body

(Neis et al., 2000).

4.1. Sonochemical Degradation of Aromatic

Pollutants

A large variety of aromatic contaminants detected in

water bodies which included chlorinated aromatic

compounds, phenolic compounds and dyes. The

presence of these compounds even in low

concentration can contaminate aquatic environment.

Besides, due to the toxic, mutagenic and carcinogenic

properties of some of these pollutants, their

concentrations need to be controlled strictly in treated

wastewater. Some of these compounds are listed as

US-EPA and EU priority pollutants (Adewuyi, 2001,

2005). However, as consequences of the

characteristics of the aromatic compounds, most of the

conventional treatments methods are not able to

remove the toxicity of these pollutants completely.

Ultrasonication treatment have been proposed and

employed for the treatment of hazardous materials in

wastewater. Table 1 presented some studies conducted

of aromatic pollutant degradation by ultrasound

treatment.

4.2. Degradation of Chlorinated Aromatic

Pollutants

The application of high intensity ultrasound to a liquid

system may lead to physical and chemical reactions

that can significantly modify the structure and

materials present in the system. Ultrasonic

degradation of chlorobenzene (ClBz) takes place

predominantly within both the bubbles where it

undergoes pyrolysis and within the liquid gas

interface of bubble. Jiang et al. (2002) concluded that

during ClBz degradation by ultrasonication, more than

90% of the chlorine was recovered as chloride ions

along with carbon monoxide (CO), acetylene (C2H2),

methane (CH4) and carbon dioxide (CO2) as gaseous

products. Some hydroxylated intermediates, namely

4-chlorophenol, hydroquinone and 4-chlorocatechol

were also detected in a low yield (less than 2 μM),

however these compounds disappeared on extended

ultrasonic irradiation.

Petrier et al. (1998) examined preferential thermal

degradation of chlorobenzene inside the cavitation

bubble. In this study, the degradation rate increased at

a higher frequency which may be due to a lower

resonance radius of the cavitation bubble.

Peller et al. (2001) investigated high frequency

sonolysis of 2,4-dichlorophenoxyacetic acid (2,4-D)

in oxygen and argon saturated aqueous medium. The

study shown that the degradation rate was faster in

argon medium compared to oxygen medium due to

argon has a higher heat capacity ratio (γ) value than

oxygen.

4.3. Degradation of Phenolic Pollutants

Petrier et al. (1994) observed that ultrasonic

degradation of phenol was favored at a higher

frequency. At 20 kHz, only 2% of carbon was

recovered in gaseous phase after 300 min, however at

487 kHz, 15% of the carbon was recovered. Besides,

they found that the resonant radius of cavitation

bubble (6.6 μm) as well as the life of the bubble

(4.1×10−7

s) was lower in the case of 487 kHz

frequency. These factors enhanced the release of •OH

radical at a high frequency. Entezari et al. (2003)

reported only 20% degradation of phenol under

sonolysis (C0=60 mg/L; time=150 min). This study is

similar to the study conducted by Maleki et al. (2005)

who reported only 13% degradation of phenol under

sonolysis (C0 = 100 mg/L; time=300 min) and by

Mahamuni and Pandit (2005) who reported 17%

degradation (C0=85 mg/L; time=60 min). Phenol is

hydrophilic moderately soluble compound with a

relatively low vapor pressure. These characteristics

prevent the diffusion of phenol molecule into

cavitation bubble, so it remains in the bulk of the

solution during cavitation.



52

Table 1: Sonochemical degradation of aromatic compounds No. Aromatic compounds Sonochemical conditions Results Reference

1 Chlorobenzene (ClBz)

Ulrasonic power: 9.4 W

Ultrasonic frequency: 520 kHz

Undatim reactor

At concentration above 1000 µM,

pyrolysis was the dominant mechanisms,

at concentration 1-5µM, radical

mechanisms played a crucial role.

Dewulf et al., 2001

2

2,4-

dichlorophenoacetic

acid

Ultrasonic power: 50 W


Product: Oxalic acid,

Intermediate: 2,4-dichlorophenol,

hydroquinone, catechol were detected •OH radicals were primary reactive

species. With argon, 90% degradation

was achieved in less than 100 minutes

Peller et al., 2001

3 p-amino phenol (PAP)



Ultrasonic probe

The degradation increased with an

increased in the ultrasonic density

He et al., 2007

4 C.I. direct blue 168



Ultrasonic bath

88.3% decolorization

Song et al., 2009

5 Polycyclic aromatic

hydrocarbons (PAH)



Ultrasonic probe

•OH is the major process for complete

sonodegradation of less hydrophobic

PAHs while pyrolysis is the major

process for complete degradation of more

hydrophobic PAHs.

Sponza and

Oztekin, 2010

6 Bisphenol A (BPA) Ultrasonic frequency: 400 kHz

Piezoelectric ceramic transducer

Aromatic intermediates: 2-(4-

hydroxyphenyl)-2-(3,4-

dihydroxyphenyl)propane, commonly

known as 3-hydroxybisphenol A were

detected.

Bisphenol A (0.50mM) was completely

degraded after 10, 3 and 2 h of ultrasonic

irradiation at a frequency of 404 kHz, and

intensities of 3.5, 9.0 and 12.9kW/m2

respectively.

Inoue et al., 2008

In another study, Jiang et al. (2006) observed the

sonolysis of 4-ClPh in oxygen saturated aqueous

solution. Results showed that above 40 °C the

degradation rate and temperature were inversely

proportional whereas below 40°C a reverse relation

was observed. Nagata et al. (2000) compared the

degradation rate of 3-ClPh to 2-, 4-chlorophenol and

pentachlorophenol under air or argon atmosphere.

Faster degradation rate was observed in argon

medium. Argon medium was more favourable for

degradation because of high heat capacity ratio (γ).

Kotronarou et al. (1991) carried out the

degradation of 4-nitrophenol (4-NP) at low frequency

(20 kHz). They reported primary products after

degradation were nitrate, nitrite and hydrogen ions.

This study also found that the degradation rate was

greatly affected by the initial solute concentration.

Studies have shown AOPs can be applied for the

elimination of Bisphenol A (BPA) in industrial

wastewater (Chiang et al., 2004; Horikoshi, 2004).

Ultrasonic treatment shows some similarity to AOP

has been found efficient in BPA decomposition (Ioan

et al., 2007; Kitajima et al., 2006). Torres et al.

(2007) proved ultrasonic process could transforms

BPA in biodegradable aliphatic acids that could be

eliminated in a subsequent biological treatment.

Therefore, ultrasound systems represent a very

interesting AOP for the treatment of water

contaminated with phenolic pollutants.

4.4. Degradation of Dyes

In the past few years, synthetic dyes are used

extensively, especially in textile industries. It was

estimated that more than 100,000 different

commercial dyes and pigments and over 70,000 tons

of dyestuff were produced annually (Ghodbane and

Hamdaoui, 2009). Out of this figure, an estimated 10-

15% of dyestuff was lost in the effluent during the

dyeing process (Inoue et al., 2006). Synthetic dyes

containing one or more benzene rings that do not

easily decompose (Wang et al., 2005). In addition,

textile effluent is considered as high strength

wastewater which is highly variable in composition

with intense color, high concentration of chemical

oxygen demand, relatively low concentration of

biological oxygen demand, and high concentration of

suspended and dissolved salts that could hinder the

Peng et al.


53

penetration of oxygen into the water body (Entezari et

al., 2008; Song et al., 2009; Tezcanli Tezcanli-Guyer

and Ince, 2003). Conventional treatment processes for

textile wastewater usually involve coagulation,

flocculation, adsorption, and biological treatment.

However the treated water from these treatment most

of the time could not meet the standard discharged

regulations especially decolorization. Furthermore,

these processes not able to destroy or degrade dyes,

only remove the dye physically from the effluent and

disposal the sludge in landfill eventually (Destaillats

et al., 2000; Inoue et al., 2006; Ozen et al., 2005;

Tezcanli-Guyer and Ince, 2003; Velegraki et al., 2006;

Vinodgopal et al., 1998). In view of that, the

management of textile wastewater requires the

development of suitable treatments that could remove

or reduce the harmful pollutant from the effluents and

be able to recyclable the water for use. Recently,

several studies have promoted sonolysis as a feasible

method for the decolorization and mineralization of

dyes (Hong et al., 1999; Vinodgopal et al., 1998;

Voncina and Majcen-Le-Marechal, 2003).

Vajnhandl and LeMarechal (2007) used two

different types of ultrasonic devices to evaluate the

efficiency of degradation of reactive dye by

monitoring the oxidative species (•OH and H2O2)

during ultrasonic degradation. They found the radical

formation rate was 20-25 folds higher in plate type

system compare to probe type system at same acoustic

power. Tauber et al. (2005) reported that ultrasound

treatment successfully degraded six azobenzene dyes

(Acid Orange 5, Acid Orange 52, Direct Blue 71,

Reactive Black 5, Reactive Orange 16, Reactive

Orange 107) whereas enzyme treatment was unable to

degrade. In another study, Byun and Kwak (2005)

achieved the multi bubble sonoluminescence (MBSL)

condition by adjusting the ultrasound intensity, liquid

temperature and the distance between horn tip and the

bottom of the cell. At MBSL condition the

degradation efficiency of methylene blue was much

better than conventional photolytic degradation (TiO2

dispersion/UV). pH is interior optimization factor for

ultrasonic dye degradation. Acidic pH accelerates the

dye degradation by protonation of negatively charged

SO-3 sites, thus providing hydrophobic enrichment of

the molecules (Behnajady et al., 2008). Acidic

conditions enhance the probability of the dyes

approaching to the negatively charged cavity bubbles

where •OH are most abundant and undergo additional

oxidation/pyrolysis reactions at the gas-liquid

interface (Ozen et al., 2005; Wang et al., 2007).

In another study, Inoue et al. (2006) investigated

the effects of power input and frequency to the

degradation rate for Rhodmine B and Orange II. They

concluded that rate constant increased with increased

input power and at higher frequency. This degradation

mechanism can be explained with the production of •OH radicals. At lower frequency, the degradation rate

was lower due to the •OH radical production rate was

slow.

5. REACTIONS AND MECHANISMS OF

AROMATIC COMPOUND DURING

SONOCHEMICAL DEGRADATION

The reaction occurs in the solution bulk for highly

water-soluble dyes are largely hydrophilic. The •OH

radicals ejected into the solution bulk cause oxidative

dye destruction, whereas hydrophobic and volatile

species degrade thermally to the gas phase or gas-

liquid interface (Ince and Tezcanli- Güyer, 2004;

Vajnhandl and Le Marechal, 2007). However, for

hydrophilic compounds with high concentration, an

additional degradation mechanism occurs quickly via

reactions with •OH radicals or pyrolysis in the

interface region of the collapsing bubbles (Ince and

Tezcanli-Güyer, 2001; Okitsu et al., 2005). For dye

degradation and mineralization, hydrodynamic

cavitation had been reported to be more energy

effective compared to acoustic cavitation (Chakinala

et al., 2008; Wang et al., 2009). Decolorization with

hydrodynamic cavitation increased by increasing the

inlet pressure and it is substantially enhanced with the

addition of H2O2.

Petrier et al. (1998) have proposed a preferential

thermal degradation in terms of pseudo first order

kinetic law of ClBz by conducting a comparative

study with 4-ClPh and ClBz as models for hydrophilic

and hydrophobic substrates respectively. They found

degradation of 4-ClPh was governed by •OH

formation whereas ClBz degradation was basically a

thermal degradation inside the cavitation bubble. In

the similar study by Stavarache et al. (2002) showed

the decomposition of ClBz took place inside

cavitation bubbles forming phenyl (C6H5•) and

chlorine (Cl•) radicals. According to the authors, at

micromolar concentrations the degradation induced by

the •HO radical is significant compare to the pyrolytic

process.

As reported by Jiang et al. (2002) the sonolysis of

ClBz solution was following the pseudo-first-order

rate constant. Similar trends were observed for 1,4-

dichlorobenzene (1,4-DClBz) and 1-

chloronaphthalene (ClNt) degradation where the rate

constants of ultrasonic degradation for 1,4-DClBz and

ClNt solutions also increased with ultrasonic intensity.

The authors proposed two possible schemes of ClBz

degradation as, i) the degradation followed by

chlorine removal by high temperature combustion and

ii) degradation with reactive •OH radicals. However,

the second reaction scheme was recognized as a minor



54

pathway as the hydroxylated intermediate production

was too low.

The degradation of phenol compound can be

proved by the formation of hydroxylated

intermediates (hydroquinone, catechol and resorcinol).

According to Petrier et al. (1994), phenol degradation

in terms of free radical reaction was following two

possible reactions as presented in Eqs 6-7:

•OH +

•OH→ H2O2 (6)

C6H5-OH + •OH → C6H5(OH)2 (7)

However, the degradation of phenol by ultrasound

was comparable low due to the hydrophilic nature of

phenol which made it unavailable for free radical

reaction with •OH radicals (Maleki et al., 2005).

Zheng et al. (2005) conducted a study to enhance the

sonochemical degradation of phenol by applying the

hydrogen atom scavenger. The additional of CCl4

traps the hydrogen atoms and prevents their

recombination with •OH radicals. This increased

availability of •OH radical that ultimately accelerated

the sonochemical degradation of phenol.

To understand the degradation mechanisms of 4-

ClPh during sonolysis, Hao et al. (2003) have

performed an oxidation experiment between 4-ClPh

and H2O2. The experiment result showed that high

temperature pyrolysis was the dominant degradation

for 4-ClPh. On the other hand, Jiang et al. (2006)

reported that the 4-ClPh degradation mechanism was

strongly dependent on solute concentration. Teo et al.

(2001) used ion selective electrode to measure the

produced Cl− ions during 4-ClPh degradation. The

results precisely showed the presence of Cl−

from the

cleavage of C–Cl bond of the 4ClPh due to ultrasonic

irradiation. The Cl− balance in the system suggested

the formation of chlorinated intermediates.

Kotronarou et al. (1991) expressed the degradation

of p-nitrophenol (PNP) as a first order reaction. This

was because PNP did not diffuse into the imploding

cavities at a low vapor pressure, but collected at the

liquid/gas interface where it was pyrolyzed and/or

attacked by the solvent radicals. The decomposition

mechanism of PNP was based on a combination of

both pyrolysis and radical reaction (Colarusso and

Serpone, 1996). In another study, Tauber et al. (2000)

found that the reaction mechanism was pH dependent.

At acidic pH (4.0) oxidative pyrolytic decomposition

was predominant whereas at alkaline pH (10.0) free

radical mechanism mainly occurred.

The ultrasonic degradation of hydrophobic

organics such as PAHs can occur when they penetrate

to the surrounding of the hot heart of the cavitation

bubble being pyrolyzed, burnt and/or ionized in the

plasma core (Flannigan and Suslick, 2005). According

to Sponza and Oztekin (2011), sonication alone had a

potential for use in the decomposition of PAHs from

petrochemical industry wastewater. Radical attack

was an important degradation mechanism of less

hydrophobic PAHs while pyrolysis was an important

destruction pathway for more hydrophobic PAHs

In the homogenous sonochemical reactor for

aromatic pollutants degradation, most of the

hydrophobic compounds reacted inside the cavitation

bubble whereas hydrophilic substances reacted at bulk

phase (Adewuyi, 2001; Liang et al., 2007). The

heterogeneous systems also follow the same physical

mechanism but differ in terms of cavitational

threshold, high speed liquid jet and other related

physical effects from acoustic cavitation (Liang et al.,

2007).

6. COMBINATION OF ULTRASONICATION

AND OTHER DEGRADATION

TECHNOLOGIES

Although sonochemical reactions were quite efficient

for degradation of organic compounds, however,

complete mineralization was not achieved sometimes.

This might be due to higher polarity of the organic

compound, low availability of •OH radical or lack of

dissipated power. Therefore, to overcome these

disadvantages in the sonochemical process, ultrasound

has been widely used as an auxiliary process with

other treatment techniques mostly biochemical,

electrochemical, ozonation, photolysis, photocatalysis

and Fenton processes.

Studies of the combined process involving

ultrasound and ozone had shown faster degradation

rates for a range of chemical contaminants than either

method applied alone (Mason and Petrier, 2004). As

the mass transfer rate of ozone in the solution was the

limiting parameter, thus the combined operation of

sonolysis and ozonation renders synergistic effected in

degradation and accelerateed the mass transfer of

ozone in the solution due to ultrasonic effects. When a

liquid was sonicated in the presence of ozone (O3), the

thermal decomposition of ozone in the cavitation

bubbles enhanced •OH and H2O2 yields. Xu et al.

(2005) studied on synergetic effect of ozonation with

ultrasonic treatment for p-nitrophenol degradation. In

O3/ultrasonic system produced more •OH and H2O2

than the sonolysis system alone. This combined

sonolysis/ozonolysis process showed 116%

enhancement in removal rate for p-nitrophenol.

Gultekin and Ince (2006) investigated the

degradation of aryl-azo-naphthol dyes with

ultrasound, ozone and combination of the two

processes at optimized condition. The US/O3

combination was the best option for the degradation of

these dyes. The synergetic effect incorporated three

main factors, i) increased mass transfer of ozone, ii)

Peng et al.


55

excess hydroxyl radical generation, and iii)formation

of secondary oxidation species (•O2

- and

•O2H). The

major part of the dissolved ozone was therefore

efficiently decomposed by ultrasonic irradiation

because of the additional pathways involving the

production of these secondary oxidation species. The

combination of these two processes provided three

sources of •OH which are from (i) sonochemical

decomposition of water,ii) the normal chemical

degradation of ozone, and iii) the thermolytic

decomposition of ozone in the acoustic cavitation

bubble. (Mason and Petrier, 2004).

The degradation and mineralization study showed

that dye degradation rate under high frequency

ultrasound (300 kHz) and ozone combined reactor is

higher than the sum of the individual processes

(Gultekin and Ince, 2006). Ozone concentration was

an important parameter in the sonolytic ozonation

process. An increase in ozone dose could improve the

mass transfer of ozone due to an increase in ozone

concentration in the liquid phase (Song et al., 2007).

The pH value in the dye solution was an important

factor in determining the rate constants (Ince and

Tezcanli, 2001). The dye degradation rate increased

when the pH value increased to 8-10. However,

further pH increase would result in reducing dye

degradation due to the existence of the radical

scavengers such as CO32-

, SO42-

and PO43-

(He et

al.,2007; Song et al., 2007; Zhang et al., 2006; Zhang

et al., 2008a,b;Zhang et al., 2009).

Tauber et al. (2005) incorporated ultrasound

treatment followed by enzyme treatment for azo dyes

(Acid Orange 5, Acid Orange 52, Direct Blue 71,

Reactive Black 5, Reactive Orange 16 and Reactive

Orange 107). It was predicted that the intermediates

created during the ultrasonication had acted as the

internal mediator for the subsequent enzyme (laccase)

treatment. Thus, the degradation rate was higher and

showed better efficiency in these dyes removal.

Another study of ultrasonic as pre-treatment for

Congo red solution followed by a biological treatment

presented that sonolysis reduced the biological

treatment time from 23-29 h to 6-8 h, thus reduced the

power consumption and cost (Basto et al., 2007). The

effects of ultrasonic irradiation on biological treatment

can be evaluated by following justification: (i) the

facilitation of substrate diffusion and (ii) the

enhancement of cell enzyme secretion. In addition, the

formation of ultrasonic cavitation could promote

particle movement in reaction solution thus

accelerated mass transfer in the reactor and enhanced

the permeability of cell membrane and wall around

cavitation bubbles (Liu et al., 2005).

The first-order rate constant of the sonophotolytic

process in the presence of hydrogen peroxide is 17.87

times greater than that of the sonophotolytic process

alone for Malachite oxalate green degradation

(Behnajady et al., 2008). In general, the utilization

efficiency of photocatalysis was low because of the

screening effect of catalysts; this disadvantage could

be overcome with the combined use of ultrasonic

irradiation. Ultrasonic irradiation might increase the

photocatalytic reaction rate by increasing the catalytic

activity and reducing the size of photocatalyst

particles following particle disaggregation, which

increased the surface area (Stock et al., 2000). Peller

et al. (2003) study showed the complete

mineralization of chlorinated organic compounds was

achieved with no toxic intermediate build up in low

calatyst loading from the combination of

ultrasound/photocatalysis treatment system. The

beneficial effect of photocatalysis coupled with

ultrasonic irradiation could also be attributed to: (i)

the increased production of •OH by ultrasound, (ii) the

enhanced mass transfer of dye between the liquid

phase and the catalyst surface (Stock et al.,2000; Yuan

et al., 2009), (iii) the excitation of catalyst by

ultrasonically induced wavelength (<375 nm) i.e.

sonoluminescence and (iv) the constantly refreshed

catalyst surface due to acoustic micro-streaming

cleaning and sweeping, which allowd more active

catalyst sites to be available for dye degradation

(Kritikos et al.,2007).

As showed in Eqs (4) and (6), when an aqueous

solution was irradiated with ultrasound, hydrogen

peroxide would be produced. If ferrous ions were

added to this solution, they react with ultrasonically

generated H2O2, thus enhanced the production of •OH

which could be readily used for dye degradation. The

pseudo first-order rate constant of dye by sono-Fenton

oxidation was three fold greater than that of its

sonolysis alone in the presence of Fe (II) in the

concentration range of 0.1 mM to 0.5 mM. The

increment was due to the formation of more •OH

radical via Fenton reaction (Joseph et al. (2000). In

ultrasonic assisted Fenton process, the highest

decolorization efficiency of dye is achieved at pH 3

(Song et al., 2009). This process was pH dependent

when pH was below 3, hydrogen radicals could be

consumed by the scavenging effects of H+ and

hydrogen peroxide would formed ozonium ion which

could enhanced the stability and reduced the reactivity

between hydrogen peroxide and ferrous ions (Sun et

al., 2007; Wang et al., 2008a). Another similar study

by Zhang et al. (2009) evaluated the decolorization of

C.I. Acid Orange 7 with the combination of

ultrasound and advanced Fenton process. The

decolorization fit the modified pseudo-first order

kinetic model where the decolorization rate increased

with hydrogen peroxide concentration and power

density but decreased with the increased of initial pH

value.



56

From the studies can concluded that the

combination of ultrasound with relevant treatment

were more economically attractive than the use of

ultrasound alone for aromatic pollutants treatment.

7. CONFIGURATION OF ULTRASONIC

EQUIPMENT

In sonochemical treatment the electrical energy was

basically converted to vibration via a transducer

(Price, 1992). Depending upon the application or the

reaction to be achieved, one can choose different

generators and design appropriate transducers as

required (Ashokkumar and Grieser, 1999).

In general there were four types of laboratory

ultrasonic equipments that were widely used namely,

i) whistle, ii) bath, iii) probe (horn) and iv) cup-horn

system (Mason and Lorimer, 1988). The whistle

reactors were primarily applicable for emulsification,

polymerization and phase transfer reactions.

Ultrasonic bath system was the low intensity load

system with a power density of 1–2 W/cm2 (Mason

and Lorimer, 1988). The main advantages of bath

system were simplicity and economy. While the

disadvantages were the limitation in temperature

control and ultrasonic intensity control. The ultrasonic

probe system was a high intensity system with better

temperature control. However, horn configuration

became a major limiting factor when dealing with

high amplitudes. Sometimes there were some

chemical interference issues due to the cavitational

erosion of the horn tip metal. The horn design could

essentially limit its ability to achieve greater

cavitation levels and power output. The cup-horn

system was a combination of bath and horn (probe)

system with a better temperature control device. It

allows much higher intensity without any

contamination by the horn tip material (Price, 1992).

Most of the studies describing the degradation of

organic dyes (Rehorek et al., 2004; Vajnhandl and

LeMarechal, 2007 , phenolic compounds (Petrier et

al., 1994; Nagata et al., 2000; Maleki et al., 2005; Xu

et al., 2005) and chlorinated organic compounds

(Drijvers et al., 1998; Gaddam and Cheung, 2001)

used probe (horn) type systems because of its high

power output and optimum performance at different

amplitudes. In the case of a low intensity irradiation a

batch reactor with a submersible transducer was

shown to be the best option (Mason and Tiehm,

2001). Gogate and co-researches (2004) suggested a

few important points on large scale sonochemical

reactors:

(i) multiple frequencies gave higher intensities of

cavitation compared to a single frequency operation.

(ii) Hexagonal geometry was more efficient in

terms of better distribution of cavitational activity.

Bhirud et al. (2004) used ultrasonic bath having

longitudinally vibrating transducer for formic acid

degradation. The longitudinally vibrating reactor

provided 4–5 times more cavitational yield than

multiple frequency flow cells. The design of the

reactor and the reactor performance also greatly

affected by the ultrasonic poser and free volumetric

flow rate (Gondrexon et al. (1999). According to

Asakura et al. (2008) the sonochemical reactor

volume for pilot scale and industrial scale should be

more than 100 L and 1000 L, respectively. The

observed sonochemical efficiency of the large reactor

was comparable with the laboratory scale reactors. It

was found that the sonochemical efficiency was

dependent on frequency and liquid height in the

frequency range of 45–490 kHz.

8. CONCLUSION

Ultrasound technologies have been widely used for

the degradation of aromatic compounds at laboratory

scale and the combination of ultrasound with other

relevant treatments can prove to achieve complete

mineralization and be more economically attractive

than the use of individual techniques due to the

generation of higher quantum of free radicals and high

degree of energy efficiency. It has been observed that

the efficacy of sonochemical reactions in cavitation

process is influenced by many important operating

parameters such as resonance frequency of bubbles,

reaction temperature and energy intensity. An

optimised reactor configuration should be able to

maximize the expected output. Thus, the

configurations of the sono-reactor may significantly

affect the efficacy of ultrasonic system for

sonochemical degradation.

REFERENCES

Adewuyi YG (2001). Sonohemistry: environmental

science and engineering applications. Industrial

& Engineering Chemistry Research, 40: 4681–

715.

Adewuyi YG (2005). Sonochemistry in environmental

remediation.1. Combinative and hybrid

sonophotochemical oxidation processes for the

treatment of pollutants. Environmental Science

& Technology, 39: 3409–3420.

Ai ZH, Li JP, Zhang LZ, Lee SC (2010). Rapid

decolorization of azo dyes in aqueous solution

by an ultrasound-assisted electrocatalytic

oxidation process. Ultrasonics Sonochemistry,

17 (2): 370–375.

Ai ZH, Lu LR, Li JP, Zhang LZ, Qiu JR, Wu MH

(2007). Fe@Fe2O3 core-shell nanowires as iron

reagent. 1. Efficient degradation of Rhodamine

Peng et al.


57

B by a novel sono-Fenton process. Journal of

Physical Chemistry C, 111 (11): 4087-4093.

Anjaneyulu Y, Chary NS, Raj DSS (2005).

Decolourization of industrial effluents- available

methods and emerging technologies-a review.

Reviews in Environmental Science and

Biotechnology, 4: 245–273.

Asakura Y, Yasuda K, Kato D, Kojima Y, Koda S

(2008). Development of a large sonochemical

reactor at a high frequency. Journal of Chemical

Engineering, 139: 339–43.

Ashokkumar M, Grieser F (1999). Ultrasound assisted

chemical processes. Reviews in Chemical

Engineering, 15(1): 41–83.

Basto C, Silva CJ, Gubitz G, Cavaco-Paulo A (2007).

Stability and decolourization ability of Trametes

villosa laccase in liquid ultrasonic fields.

Ultrasonics Sonochemistry, 14(3): 355-362.

Behnajady MA, Modirshahla N, Shokri M, Vahid B

(2008). Investigation of the effect of ultrasonic

waves on the enhancement of efficiency of

direct photolysis and photooxidation processes

on the removal of a model contaminant from

textile industry. Global Nest Journal, 10(1): 8-

15.

Berberidou C, Poulios I, Xekoukoulotakis NP,

Mantzavinos D (2007). Sonolytic,

photocatalytic and sonophotocatalytic

degradation of malachite green in aqueous

solutions. Applied Catalysis B- Environment,

74(1–2): 63–72.

Bhirud US, Gogate PR, Wilhelm AM, Pandit AB

(2004). Ultrasonic bath with longitudinal

vibrations: a novel configuration for efficient

wastewater treatment. Ultrasonics

Sonochemistry, 11(1): 43–7.

Byun KT, Kwak HY (2005). Degradation of

methylene blue under multibubble

sonoluminescence condition. Journal of

Photochemistry and Photobiology A, 175: 45–

50.

Chakinala AG, Bremner DH, Gogate PR, Namkung

KC, Burgess AE (2008). Multivariate analysis

of phenol mineralisation by combined

hydrodynamic cavitation and heterogenous

advanced Fenton processing. Applied Catalysis

B-Environment, 78: 11–18.

Chiang K, Lim TM, Tsen L, Lee CC (2004).

Photocatalytic degradation and mineralization of

bisphenol A byTiO2 and platinized TiO2,

Applied Catalysis A-Environment, 261: 225–

237.

Chowdhury P, Viraraghavan T (2009). Sonochemical

degradation of chlorinated organic compounds,

phenolic compounds and organic dyes – A

review. Science of The Total Environment, 407:

2474-2792

Colarusso P, Serpone N (1996). Sonochemistry II.—

effects of ultrasounds on homogeneous chemical

reactions and in environmental detoxification.

Research on Chemical Intermediates, 22(1):61–

89.

Crum LA (1995). Comments on the evolving field of

sonochemistry by a cavitation physicist,

Ultrasonics Sonochemistry, 2: S147–S152.

Crum LA, Mason TJ, Reisse JL, Suslick KS (Eds.)

(1999). Sonochemistry and Sonoluminescence.

Kluwer Academic Publishers, Dordrecht.

Destaillats H, Colussi AJ, Joseph JM, Hoffmann MR

(2000). Synergistic effects of sonolysis

combined with ozonolysis for the oxidation of

azobenzene and methyl orange. Journal of

Physical Chemistry A, 104(39): 8930-8935.

Dewulf J, Van Langenhove H, De Visscher A, Sabbe

S (2001). Ultrasonic degradation of

trichloroethylene and chlorobenzene at

micromolar concentrations: kinetics and

modeling. Ultrasonics Sonochemistry, 8: 143.

Drijvers D, Langenhove HV, Vervaet K (1998).

Sonolysis of chlorobenzene in aqueous solution:

organic intermediates. Ultrasonics

Sonochemistry, 5:13–9.

Entezari MH, Petrier C, Devidal P (2003).

Sonochemical degradation of phenol in water: a

comparison of classical equipment with a new

cylindrical reactor. Ultrasonics Sonochemistry,

10:103–108.

Entezari MH, Al-Hoseini ZS, Ashraf N (2008). Fast

and efficient removal of Reactive Black 5 from

aqueous solution by a combined method of

ultrasound and sorption process. Ultrasonics

Sonochemistry, 15(4): 433-437.

Eren Z (2012). Ultrasound as a basic and auxiliary

process for dye remediation: A review. Journal

of Environmental Management, 104: 127-141

Flannigan DJ, Suslick KS (2005). Plasma formation

and temperature measurement during single-

bubble cavitation. Nature, 434: 52–55.

Flint EB, Suslick KS (1991). The temperature of

cavitation. Science, 253:1397–1399.

Gaddam K, Cheung HM (2001). Effect of pressure

temperature, and pH on the sonochemical

destruction of 1,1,1-trichloroethane in dilute

aqueous solution. Ultrasonics Sonochemistry,

8:103–109.

Ghodbane H, Hamdaoui O (2009). Intensification of

sonochemical decolorization of anthraquinonic

dye Acid Blue 25 using carbon tetrachloride.


Goel M, Hu HQ, Mujumdar AS, Ray MB (2004).

Sonochemical decomposition of volatile and



58

non-volatile organic compounds e a comparative

study. Water Research, 38 (19): 4247-4261.

Gogate PR, Pandit AB (2004a). A review of

imperative technologies for wastewater

treatment I: oxidation technologies at ambient

conditions. Advances in Environment Research,

8: 501-551.

Gogate PR, Pandit AB (2004b). A review of

imperative technologies for wastewater

treatment II: hybrid methods. Advances in

Environment Research, 8: 553-597.

Gogate PR (2008). Treatment of wastewater streams

containing phenolic compounds using hybrid

techniques based on cavitation: a review of the

current status and the way forward. Ultrasonics

Sonochemistry, 15: 1–15.

Gogate PR, Pandit AB (2001). Hydrodynamic

cavitation reactors: a state of the art review.

Reviews in Chemical Engineering, 17(1): 1.

Gogate PR, Majumdar S, Thampi J, Wilhelm AM,

Pandit AB (2004). Destruction of phenol using

sonochemical reactors: scale up aspects and

comparison of novel configuration with

conventional reactors. Separation and

Purification Technology, 34:25–34.

Gondrexon N, Renaudin V, Petrier C, Boldo P, Bernis

A, Gonthier Y (1999). Degradation of

pentachlorophenol aqueous solutions using a

continuous flow ultrasonic reactor: experimental

performance and modeling. Ultrasonics


Gregory P (2009). Dyes and dye intermediates. in:

Kirk-Othmer Encyclopedia of Chemical


Guinea E, Brillas E, Centellas F, Canizares P, Rodrigo

MA, Saez C (2009). Oxidation of enrofloxacin

with conductive-diamond electrochemical

oxidation, ozonation and Fenton oxidation. A

comparison. Water Research, 43 (8): 2131–

2138.

Gultekin I, Ince NH (2006). Degradation of aryl-azo-

naphthol dyes by ultrasound, ozone and their

combination: effect of α-substituents.

Ultrasonics Sonochemistry, 13:208–214.

Gultekin I, Tezcanli-Guyer G, Ince NH (2009).

Sonochemical decay of CI Acid Orange 8:

effects of CCl4 and t-butyl alcohol. Ultrasonics

Sonochemistry, 16 (4): 577-581.

Hamdaoui O, Chiha M, Naffrechoux E (2008).

Ultrasound-assisted removal of malachite green

from aqueous solution by dead pine needles.


Hao H, Wu M, Chen Y, Yin Y, Lu Z (2003).

Cavitation-induced pyrolysis of toxic

chlorophenol by high frequency ultrasonic

irradiation. Wiley Periodicals, Inc. Environ

Toxicol. 18: 413–417.

He ZQ, Song S, Xia M, Qiu JP, Ying .P, Lu BS, et al.

(2007). Mineralization of CI Reactive Yellow

84 in aqueous solution by sonolytic ozonation.

Chemosphere, 69: 191-199.

Hong Q, Hardcastle JL, McKeown RAJ, Marken F,

Compton RG (1999). The 20 kHz sonochemical

degradation of trace cyanide and dye stuffs in

aqueous media. New Journal of Chemistry, 23

(8): 845-849.

Horikoshi S, Tokunaga A, Hidaka H, Serpone N

(2004). Environmental remediation by an

integrated microwave/UV illumination method.

VII. Thermal/non-thermal effects in the

microwave assisted photocatalyzed

mineralization of bisphenol-A. Journal of

Photochemistry and Photobiology A, 162: 33–

40.

Hua I, Hoffmann MR (1997). Optimization of

ultrasonic irradiation as an advanced oxidation

technology. Environmental Science &

Technology, 31:2237–2243.

Hung HM, Hoffmann MR (1999). Kinetics and

mechanism of the sonolytic degradation of

chlorinated hydrocarbons: frequency effect.

Journal of Physical Chemistry A, 103: 2734–

2739.

Ince NH, Tezcanli G, Belen RK, Apikyan IG (2001).

Ultrasound as a catalyzer of aqueous reaction

systems: the state of the art and environmental

applications. Applied Catalysis B-

Environmental, 29:167–76.

Ince NH, Tezcanli G (2001). Reactive dyestuff

degradation by combined sonolysis and

ozonation. Dyes and Pigments, 49 (3): 145-153.

Inoue M, Masuda Y, Okada F, Sakurai A (2008).

Degradation of bisphenol A using sonochemical

reactions. Water Research, 42: 1379 – 1386.

Inoue M, Okada F, Sakurai A, Sakakibara M (2006).

A new development of dyestuffs degradation

system using ultrasound. Ultrasonics

Sonochemistry, 13(4): 313-320.

Ioan I, Wilson S, Lundanes E, Neculai A (2007).

Comparision of Fenton and sono-Fenton

bisphenol A degradation. Journal of Hazardous

Materials, 142: 559–563.

Jiang Y, Petrier C, Waite TD (2002). Kinetics and

mechanism of ultrasonic degradation of volatile

chlorinated aromatics in aqueous solutions.

Ultrasonics Sonochemistry, 9:317–323.

Jiang Y, Petrier C, Waite TD (2006). Sonolysis of 4-

chlorophenol in aqueous solution: effect of

substrate concentration, aqueous temperature

and ultrasonic frequency. Ultrasonics


Peng et al.


59

Joseph JM, Destaillats H, Hung HM, Hoffmann MR

(2000). The sonochemical degradation of

azobenzene and related azo dyes: rate

enhancements via Fenton's reactions. Journal of

Physical Chemistry A, 104: 301–307.

Ju YM, Yang SG, Ding YC, Sun C, Zhang AQ, Wang

LH (2008). Microwave assisted rapid

photocatalytic degradation of malachite green in

TiO2 suspensions: mechanism and pathways.

Journal of Physical Chemistry A, 112(44):

11172–11177.

Kitajima M, Hatanaka S, Hayashi S (2006).

Mechanism of O2-mediated sonolysis of

bisphenol A. Ultrasonics, 44: 371–373.

Kotronarou A, Mills G, Hoffmann MR (1991).

Ultrasonic irradiation of p-nitrophenol in

aqueous solution. Journal of Physical Chemistry

A, 95:3630–3638.

Kritikos DE, Xekoukoulotakis NP, Psillakis E,

Mantzavinos D (2007). Photocatalytic

degradation of reactive black 5 in aqueous

solutions: effect of operating conditions and

coupling with ultrasound irradiation. Water

Research, 41(10): 2236-2246.

Lepoint T, Mullie F (1994). What exactly is cavitation

chemistry? Ultrasonics Sonochemistry, 1(1):

S13–S22.

Li M, Li JT, Sun HW (2008). Decolorizing of azo dye

Reactive red 24 aqueous solution using

exfoliated graphite and H2O2 under ultrasound

irradiation. Ultrasonics Sonochemistry, 15(5):

717-723.

Liang J, Komarov S, Hayashi N, Kasai E (2007).

Recent trends in the decomposition of

chlorinated aromatic hydrocarbons by

ultrasound irradiation and Fenton's reagent.

Journal of Material Cycles and Waste

Management, 9:47–55.

Lin JJ, Zhao XS, Liu D, Yu ZG, Zhang Y, Xu H

(2008). The decoloration and mineralization of

azo dye CI Acid Red 14 by sonochemical

process: rate improvement via Fenton’s

reactions. Journal of Hazardous Materials, 157

(2-3): 541-546.

Liu H, He YH, Quan XC, Yan YX, Kong XH, Lia AJ

(2005). Enhancement of organic pollutant

biodegradation by ultrasound irradiation in a

biological activated carbon membrane reactor.

Process Biochemistry, 40(9): 3002-3007.

Lohse D (2005). Cavitation hots up. Nature, 434: 33–

34.

Mahamuni NN, Pandit AB (2005). Effect of additives

on ultrasonic degradation of phenol. Ultrasonics

Sonochemistry. 13(2): 165–174.

Maleki A, Mahvi AH, Vaezi F, Nabizadeh R (2005).

Ultrasonic degradation of phenol and

determination of the oxidation by-product

toxicity Iran. Journal of Environmental Health

Science and Engineering, 2(3):201–206.

Marguls MA (1992). Fundamental aspects of

sonochemistry. Ultrasonics, 30(3):152–155.

Mason TJ, Lorimer JP (1988). Sonochemistry:

Theory, applications and uses of ultrasound in

chemistry. Ellis Horwood Ltd New York:.

Mason TJ, Petrier C (2004). Ultrasound processes. In:

Parsons S, editor. Advanced oxidation processes

for water and wastewater treatment. IWA

publishing. 185–208.

Mason TJ, Tiehm A (2001). The design of Ultrasonic

reactors for environmental remediation.

Advances in Sonochemistry. Jai Press Inc.

Stamford, Connecticut, USA.

Meric S, Selcuk H, Belgiorno V (2005). Acute

toxicity removal in textile finishing wastewater

by Fenton’s oxidation, ozone and coagulation–

flocculation processes. Water Research, 39 (6):

1147–1153.

Minero C, Pellizzari P, Maurino V, Pelizzetti E, Vione

D (2008). Enhancement of dye sonochemical

degradation by some inorganic anions present in

natural waters. Applied Catalysis B-

Environmental, 77(3-4): 308-316.

Nagata Y, Nakagawa M, Okuno H, Miznkoshi Y, Yim

B, Maeda Y (2000). Sonochemical degradation

of chlorophenols in water. Ultrasonics


Namkung KC, Burgess AE, Bremner DH, Staines H

(2008), Advanced Fenton processing of aqueous

phenol solutions: a continuous system study

including sonication effects, Ultrasonics


Neis U (2000). Ultrasound in water, wastewater and

sludge treatment. Sewage treatment, 21: 36-39.

Neis U, Nickel K, Tiehm A (2000). Enhancement of

anaerobic sludge digestion by ultrasonic

disintegration. Water Science Technology, 42:

73-80.

Okitsu K, Iwasaki K, Yobiko Y, Bandow H,

Nishimura R, Maeda Y (2005). Sonochemical

degradation of azo dyes in aqueous solution: a

new heterogeneous kinetics model taking into

account the local concentration of OH radicals

and azo dyes. Ultrasonics Sonochemistry, 12

(4): 255-262.

Ozen AS, Aviyente V, Tezcanli-Guyer G, Ince NH

(2005). Experimental and modeling approach to

decolorization of azo dyes by ultrasound:

degradation of the hydrazone tautomer. Journal

of Physical Chemistry A, 109(15): 3506-3516.

Papadaki M, Emery RJ, Hassan MAA, Bustos AD,

Metcalfe IS, Mantzavinos D (2004).

Sonocatalytic oxidation processes for the



60

removal of contaminants containing aromatic

rings from aqueous effluents. Separation and

Purification Technology, 34: 35–42.

Peller J, Wiest O, Kamat PV (2001). Sonolysis of 2,4-

dichlorophenoxyaceti acid in aqueous solution:

evidence for OH-radical-mediated degradation.

Journal of Physical Chemistry A, 105:3176–

3181.

Peller J, Wiest O, Kamat PV (2003). Synergy of

combining sonolysis and photocatalysis in the

degradation and mineralization of chlorinated

aromatic compounds. Environmental Science &

Technology, 37:1926–1932.

Pera-Titus M, Garcıa-Molina V, Banos MA, Gimenez

J, Esplugas S (2004). Degradation of

chlorophenols by means of advanced oxidation

processes: a general review. Applied Catalysis

B-Environmental, 47: 219–256.

Petrier C, Jiang Y, Lamy MF (1998). Ultrasound and

environment: sonochemical degradation of

chloroaromatic derivatives. Environmental

Science & Technology, 32:1316–1318.

Petrier C, Lamy MF, Francony A, Benahcene A,

David B (1994). Sonochemical degradation of

phenol in dilute aqueous solutions: comparison

of the reaction rates at 20 and 487 kHz. Journal

of Physical Chemistry A, 98:10514–10520.

Price GJ (1992). Introduction to Sonochemistry. In:

Price GJ, editor. Current Trends in

Sonochemistry. Cambridge: The Royal Society

of Chemistry, 1–7.

Rehorek A, Tauber M, Gubitz G (2004). Application

of power ultrasound for azo dye degradation.

Ultrasonics Sonochemistry, 11: 177–182.

Shimizu N, Ogino C, Dadjour MF, Murata T (2007).

Sonocatalytic degradation of methylene blue

with TiO2 pellets in water. Ultrasonics

Sonochemistry, 14(2):184-190.

Show KY, Wong LP (2012). Application of

Ultrasound Pretreatment for Sludge Digestion.

In: Anaerobic Digestion, Editor: Ackmez

Mudhoo, Scrivener Publishing LLC. 91-132.

Sonawane S, Chaudhari P, Ghodke S, Phadtare S,

Meshram S (2009). Ultrasound assisted

adsorption of basic dye onto organically

modified bentonite (nanoclay). Journal of

Scientific & Industrial Research, 68(2): 162-

167.

Song S, Ying HP, He ZQ, Chen JM (2007).

Mechanism of decolorization and degradation of

CI Direct Red 23 by ozonation combined with

sonolysis. Chemosphere, 66(9): 1782-1788.

Song Y, Li J, Chen H (2009). Degradation of C.I.

Acid Red 88 aqueous solution by combination

of Fenton’s reagent and ultrasound irradiation.

Journal of Chemical Technology and

Biotechnology, 84 (4): 578-583.

Sponza, DT, Oztekin R (2010). Destruction of some

more and less hydrophobic PAHs and their

toxicities in a petrochemical industry

wastewater with sonication in Turkey.

Bioresource Technology, 101: 8639–8648

Sponza DT, Oztekin R (2011). Removals of some

hydrophobic poly aromatic hydrocarbons

(PAHs) and Daphnia magna acute toxicity in a

petrochemical industry wastewater with

ultrasound in Izmir-Turkey. Separation and

Purification Technology, 77: 301–311

Stavarache C, Yim B, Vinatoru M, Maeda Y. (2002).

Sonolysis of chlorobenzene in Fenton type

aqueous systems. Ultrasonics Sonochemistry. 9:

291–296.

Stock NL, Peller J, Vinodgopal K, Kamat PV (2000).

Combinative sonolysis and photocatalysis for

textile dye degradation. Environmental Science

& Technology, 34(9): 1747-1750.

Sun JH, Sun SP, Sun JY, Sun RX, Qiao LP, Guo HQ,

Fan MH (2007). Degradation of azo dye Acid

black 1 using low concentration iron of Fenton

process facilitated by ultrasonic irradiation.


Suslick KS (1989). The chemical effect of ultrasound.

Sci. Am, 260:80–86.

Suslick KS (1990). Sonochemistry. Science,

247(4949): 1439-1445.

Tauber A, S chuchmann HP, Sonntag CV (2000).

Sonolysis of aqueous 4-nitrophenol at low and

high pH. Ultrasonics Sonochemistry, 7: 45–52.

Tauber MM, Guebitz GM, Rehorek A (2005).

Degradation of azo dyes by laccase and

ultrasound treatment. Applied and

Environmental Microbiology, 71(5):2600–2606.

Teo KC, Xu Y, Yang C (2001). Sonochemical

degradation for toxic halogenated organic

compounds. Ultrasonics Sonochemistry, 8:241–

246.

Tezcanli-Guyer, G, Ince NH (2003). Degradation and

toxicity reduction of textile dyestuff by

ultrasound. Ultrasonics Sonochemistry, 10(4-5):

235-240.

Torres RA, Abdelmalek F, Combet E, Petrier C,

Pulgarin C (2007). A comparative study of

ultrasonic cavitation and Fenton’s reagent for

bisphenol A degradation in deionised and

natural waters. Journal of Hazardous Materials,

146: 546–551.

Vajnhandl S, LeMarechal AM (2007). Case study of

the sonochemical decolouration of textile azo

dye Reactive Black 5. Journal of Hazardous

Materials, 141(1): 329-335.

Peng et al.


61

Velegraki T, Poulios I, Charalabaki M, Kalogerakis

N, Samaras P, Mantzavinos D (2006).

Photocatalytic and sonolytic oxidation of acid

orange 7 in aqueous solution. Applied Catalysis

B-Environmental, 62 (1-2): 159-168.

Vinodgopal K, Peller J (2003). Hydroxyl radical-

mediated advanced oxidation processes for

textile dyes: a comparison of the radiolytic and

sonolytic degradation of the monoazo dye Acid

Orange 7. Research on Chemical Intermediates,

29(3): 307-316.

Vinodgopal K., Peller J, Makogon O, Kamat PV

(1998). Ultrasonic mineralization of a reactive

textile azo dye, Remazol black B. Water

Research, 32 (12): 3646-3650.

Voncina DB, Majcen-Le-Marechal A (2003).

Reactive dye decolorization using combined

ultrasound/H2O2. Dyes and Pigments, 59 (2):

173-179.

Wang J, Guo B, Zhang X, Zhang Z, Han J,Wu J

(2005). Sonocatalytic degradationof methyl

orange in the presence of TiO2 catalysts and

catalytic activity comparison of rutile and

anatase. Ultrasonics Sonochemistry, 12 (5): 331-

337.

Wang J, Jiang YF, Zhang Z, Zhang XD, Ma T, Zhang

G, et al. (2007). Investigation on the

sonocatalytic degradation of acid red B in the

presence of nanometer TiO2 catalysts and

comparison of catalytic activities of anatase and

rutile TiO2 powders. Ultrasonics


Wang XK, Yao ZY, Wang JA, Guo WL, Li GL

(2008). Degradation of reactive brilliant red in

aqueous solution by ultrasonic cavitation.

Ultrasonics Sonochemistry, 15: 43-48.

Wang XK, Zhang SY, Li SP (2009). Decolorization of

reactive brilliant red K-2BP in aqueous solution

by using hydrodynamic cavitation.

Environmental Engineering Science, 26 (1): 53-

59.

Xu XW, Shi HX, Wang DH (2005). Ozonation with

ultrasound enhancement of p-nitrophenol

wastewater. J Zhejiang Univ Sci. 6B(5): 319–

323.

Yuan SA, Yu L, Shi LY, Wu J, Fang JH, Zhao Y

(2009). Highly ordered TiO2 nanotube array as

recyclable catalyst for the sonophotocatalytic

degradation of methylene blue. Catalysis

Communications, 10(8): 1188-1191.

Zhang H, Duan LJ, Zhang DB (2006). Decolorization

of methyl orange by ozonation in combination

with ultrasonic irradiation. Journal of Hazardous

Materials, 138(1): 53-59.

Zhang H, LV, YJ, Liu F, Zhang DB (2008a).

Degradation of CI Acid Orange 7 by ultrasound

enhanced ozonation in a rectangular air-lift

reactor. Chemical Engineering Journal, 138(1-

3): 231-238.

Zhang H, Zhang J, Zhang C, Liu F, Zhang D (2009).

Degradation of C.I. Acid Orange 7 by the

advanced Fenton process in combination with

ultrasonic irradiation. Ultrasonics

Sonochemistry, 16: 325–330

Zhang Z, Zheng H (2009). Optimization for

decolorization of azo dye acid green 20 by

ultrasound and H2O2 using response surface

methodology. Journal of Hazardous Materials,

172(2–3): 1388–1393.

Zhang ZH, Yuan Y, Liang LH, Fang YJ, Cheng YX,

Ding HC, et al. (2008b).

Sonophotoelectrocatalytic degradation of azo

dye on TiO2 nanotube electrode. Ultrasonics


Zheng W, Maurin M, Tarr MA (2005). Enhancement

of sonochemical degradation of phenol using

hydrogen atom scavengers. Ultrasonics


Zhou XJ, Guo WQ, Yang SS, Ren NQ (2012). A rapid

and low energy consumption method to

decolorize the high concentration

triphenylmethane dye wastewater: operational

parameters optimization for the

ultrasonicassisted ozone oxidation process.

Bioresource Technology, 105: 40–47.



©2014 IJSRPUB

62


Treatment Systems

Chapter 3: Wastewater Treatment by Biological

Methods



©2014 IJSRPUB

63

Wastewater Treatment by Biological Methods

Dadrasnia A.1,*

, Shahsavari N.2, Emenike C.U.

1

1Institute of Biological Sciences, Faculty of Science, University of Malaya, 50603 Kuala Lumpur, Malaysia

2Science and Research Branch, Islamic Azad University, Sirjan, Iran


Abstract. Wastewater generation is just inevitable but discharging it into the environment in a sustainable way is the subject of

concern. Hence, treatments adopted have the common goal of conforming to water quality standards to ensure environmental

protection. Different sources that range from industrial productions to domestic use of water has continued to generate

wastewater and various treatment methods are employed, which includes biological methods. For effective treatment,

considerations are paid to wastewater characteristics, requirement of treated wastewater quality, treatment alternatives and

associated sampling techniques with analysis. Most biological treatments often take care of the removal of biochemical oxygen

demand, chemical oxygen demand, suspended solids, ammonia and even other inorganic compounds that includes heavy

metals. This chapter evaluated the use of oxidation, anaerobic, aerobic ponds, and activated sludge process as biological

treatment methods for wastewater. Common to the options is the utilization of microorganisms to degrade pollutants of

interest, but varied on certain conditions that border on design and oxygen utilization and limitation. Biological treatments can

be achieved under reduced cost when compared to physical and chemical methods, yet time consumption appear to be its

drawback.

Keywords: Wastewater, Treatment, Biological Methods

1. INTRODUCTION

Wastewater emanates from many sources which

includes industrial production and domestic use of

water. Hence, wastewater characteristics are highly

dependent on the source, and when it is not properly

treated before discharge, it can significantly pollute

watercourses. Regardless of the treatment option

adopted for any form of wastewater, the end

expectation is to relate to discharge/effluent standards

and quality. With high level of varied forms of

production and changing lifestyle, chemical

composition and loading of most wastewater are of

significant interest world over, excess organic

Chemical pollution from heavy metals, dyes,

solvents and pesticides is one of the major threats to

water quality. Chemicals enter the aquatic

environment through different ways, among which are

via wastewater treatment plants (WWTP) that do not

fulfill their obligations (Oller et al., 2011), and direct

discharge of industrial effluents. Similarly, indirect

chemical pollution of water arises from the use of

plant health products, such as biocides and, fertilizers

in agriculture (Oller et al., 2011). Considering the fact

that water soluble substances can be distributed and

transported more easily in the water cycle, treatment

of wastewater by discharge becomes paramount.

Wastewater treatment refers to the process of

removing pollutants from water, especially those

associated with agricultural, industrial and municipal

activities. This process is designed to achieve the

expectation and/or standard level in the quality of the

wastewater. Current methods used for removing

pollutants present in the wastewater can be divided

into chemical, physical, biological and energetic

methods.

The aim of applying various treatment processes

are to reduce:

(a) Biodegradable organic parts [chemical oxygen

demand (COD) and Biochemical oxygen demand

(BOD)] serve as ―food‖ for bacteria. Microorganisms

combine oxygen with organic part of the wastewater

to yield the energy they need to multiply and thrive; in

addition, fish and other organisms in the river need

this oxygen as well. Heavy organic pollution of the

environment may result in ―dead zones‖ and simply

imply that no fish is found; sudden releases of heavy

organic may result in dramatic ―Fishkill‖.

(b) Suspended solid part (including those particles

that can block channels or rivers)

(c) Pathogenic bacteria/disease organisms. These

are most appropriate where the water is necessary for

drinking, or where people would otherwise have close

contact by using it; and

(d) Nutrients, which includes nitrates as well as

phosphates. These nutrients can bring about excessive

levels of algal distribution, which can significantly

contribute to higher loading of biodegradable organic

matter. Treatment techniques could also remove or

neutralize industrial wastes and dangerous chemicals.

Such remedy goal should be practiced in-situ

industrial plant, before final effluent discharge,

especially to watercourses.

1.1. Levels of wastewater treatment

In order to be able to bring about effective treatment

of the effluent, it is necessary to know more about:

Dadrasnia et al.


64

(a) Characteristics of wastewater to be treated

(b) Requirement of treated wastewater quality

(c) Types of treatment alternatives available

(d) Techniques of wastewater sampling and analysis

These are the additions to the preventive and

corrective maintenance of the treatment machinery,

knowledge of repairs to and replacement of various

parts of equipment, record keeping, report

preparation, and aspects of safety in treatment plants.

The characterization involves determination of

biological, physical and chemical characteristics of

the samples of wastewater using laboratory techniques

such as gravimetry, colorimetry and titrimetry.

Knowledge of the characteristics helps the plant

operators to provide the information on:

(a) The strength of the raw and treated wastewater

(b) The efficiency of the plant operation as a whole

and each of the treatment processes

(c) The nature of treatment required in the case of the

given wastewater to meet the quality standard.

1.1.1. Primary treatment (mechanical)

Primary treatment (mechanical) is a preliminary

treatment used to protect and facilitate other

equipment in the treatment processes. It is designed to

cut up or remove the large suspended, gross and

floating solids, and heavy organic solids or amount of

oil from raw sewage. However, this stage is called

―mechanical treatment‖, while chemical substances

can be used to increase the actual sedimentation

procedure. It includes screening or sieving to trap and

remove sediments and solid parts based on principle

of gravity. By this way, the BOD of the incoming

wastewater can be reduced by 20-30% alongside 50-

60% of total suspended solids. Primary treatment is

usually done at the initial period of wastewater

treatment. Many devices are employed for preliminary

treatment, including screens bar, rack, pre-aeration

tanks, as well as grit chambers. These devices require

careful operation and design. Several advanced

wastewater treatment facilities in industrialized places

started with primary treatment, and have after that

included other treatment phases as wastewater load

has grown, and there is requirement for

comprehensive treatment.

1.1.2. Secondary treatment

Secondary treatment method eliminates the dissolved

organic matter which escapes preceding treatment

method. This is realized by microbes consuming the

organic matter as meal, and converting it to CO2,

water, and energy for their own reproduction and

growth. The biological method is then followed by

inclusion more sedimentation tanks to adequately

remove remaining suspended solids. In relation to

removal of 85% of the suspended solids, BOD may be

taken off by a well designed plant with secondary

treatment part. Secondary treatment method

technologies are classified as the necessary stimulated

sludge procedure, the variants of pond and constructed

wetland systems, trickling filtration and other forms of

treatment method designed to use biological activity

in order to break down organic matter.

1.1.3. Tertiary treatment

Tertiary treatment is simply an added treatment option

that is more advanced than the secondary treatment.

Tertiary treatment can certainly take away

approximately 99% of all the impurities from sewage,

producing great effluent associated with practically

drinking-water top quality. This associated technology

are often very expensive, requiring a high level of

technical know-how as well as expert plant operators,

a steady energy supply, chemicals and specific

products which in turn may not be easily obtainable.

An example of a typical tertiary treatment action will

be the change of a secondary treatment method to take

out more phosphorus as well as nitrogen.

Disinfection, usually together with chlorine, could

possibly be the last move ahead before discharge of

the effluent. Even so, a few environmental authorities

are concerned which chlorine residuals in the effluent

can be a problem. Disinfection is often that are part of

treatment plant design, although not properly used, as

a result of the higher price of chlorine, as well as

decreased efficiency of ultraviolet radiation where the

normal water is not completely clear, and without any

contaminants.

2. BASICS OF BIOLOGICAL TREATMENTS

2.1. BOD, COD and suspended solids

Biochemical oxygen demand (BOD) is similar in

function to chemical oxygen demand (COD), Organic

waste and natural organic detritus from wastewater

treatment plants, failing septic methods, and farm and

downtown runoff, functions as a food source for

water-borne microorganisms. Bacteria decompose

these natural components employing dissolved

oxygen, thus reducing the DO present for fish or even

marine organisms. Biochemical oxygen demand

(BOD) is usually a way of measuring how much

oxygen bacteria utilize while decomposing organic

matter under aerobic conditions. Biochemical oxygen

need is dependent upon incubating a sealed sample of

water for several times and also measuring the loss of

oxygen from the beginning to the end of the test.

Samples must be diluted prior to incubation; else the

http://en.wikipedia.org/wiki/Chemical_oxygen_demand


Chapter 3: Wastewater Treatment by Biological Methods

65

bacteria will drastically reduce the available oxygen

inside the container prior to end of the test. The main

focus of all wastewater treatment facilities is to reduce

the BOD in the effluent discharged to natural waters.

Wastewater treatment plants are created to function as

bacteria farms, where bacteria are usually given

oxygen and organic waste. The excess bacteria grown

in the system are usually taken away as sludge, and

disposed of on land as ―solid‖ waste.

Chemical oxygen demand (COD) does not

basically differentiate between organic matter that is

biologically available and the inert one. However, it is

a measure of the total amount of oxygen required to

oxidization of all organic materials to carbon dioxide

and water. When values are compared, BOD is often

less than COD, though COD measurements can be

produced in some hours though BOD measurements

take five days (BOD5).

3. BIOLOGICAL WASTEWATER TREATMENT

3.1. Oxidation Ponds

Before the discharge of wastewater into an aquatic

ecosystem, oxidation ponds are often used for

effective reduction of BOD considering its relatively

low cost and simple technology. It is involves ring

channel that is equipped with mechanical ventilation

apparatuses. Monitored wastewater penetrating the

pool is ventilated by mechanical apparatuses which

circulate at about 0.25–0.35 m/s. Oxidation pools

normally operate in an expanded aeration mode with

long confinement and solids retention times (Sperling,

et al., 2005). Kotsou et al (2004) had studied an

aerobic biological method by utilizing an Aspergillus

niger strain in a bubble column bio-reactor that was

combined with chemical oxidation treatment of

process wastewater from table olive. Authors reported

that a relatively significant decline of chemical

oxygen require (COD) was found 2 days after

biological treatment, just as the simple and total

phenolic complexes were also reduced by 41 and

85%, respectively. The stage of chemical oxidation

principally influenced the clearance of constant

phenolic complexes during the biological treatment of

total phenolic compounds. Furthermore, coagulation

support from CaO considerably enhanced the

treatment efficiency.

3.2. Anaerobic Ponds

Anaerobic ponds (Figure 1) are planned in a manner

that such will enable effective pretreatment of high

resistance wastewaters. This is applied to the aerobic

treatment and is regularly very efficient and cost-

effective for eliminating BOD and COD high

concentrations (Dewil et al., 2006). These ponds have

more organic loading in comparison to the oxygen

content entering the pond for sustaining the anaerobic

state of the pond surface. Anaerobic bacteria which

occur naturally degrade the organic components in the

wastewater, releasing carbon dioxide and methane.

The ensuing sludge will settle at the bottom while a

crust may aggregate at the surface (Doorn et al.,

2006). Fu et al. (2011) demonstrated textiles

wastewater treatment with the application anoxic filter

bed (AFB) and (BWB-BAF). The study showed that

the AFB effluent COD content reduction upon

addition of new carriers and the average efficiency for

the COD removal was 20.2%.

Leal et al. (2010) studied the occurrence of

xenobiotics in gray water and the associated removal

via three different biological treatment systems. In this

study, 18-selected xenobiotics related to chemicals

from personal care and household products were

estimated in gray water from different houses (32) and

in effluents of three different biological treatment

systems (aerobic, anaerobic, and anaerobic-aerobic).

The authors pointed out that the degree of removal

was optimal was highest at aerobic condition than

when anaerobic situation was used. 3.3. Aerobic Ponds

Aerobic ponds contain suspensions, typically bacteria

and algae that sustain aerobic conditions throughout

their depth. Two types of aerobic wetlands namely,

aerated and shallow ponds are often used

(Vijayaraghavan, 2007). Souza et al., (2011)

investigated the treatment of refinery wastewater by

AOPs along with biological activated carbon (BAC)

with the objective to produce water intended for reuse.

BAC filtration were being observed to be effective,

attaining average efficiencies up to 65% in a

sufficiently long period (84 days), while granular

activated carbon filters were saturated after twenty-

eight days. Cao et al. (2011) suggested internal

electrolysis biological contact oxidation process for

the treatment wastewater containing linear

alkylbenzene sulfonate.

Dadrasnia et al.


66

Figure 1- Anaerobic pond

4. MODELING THE BIOLOGICAL

PROCEEDING

Biology aims to understand the mechanisms by which

organisms survive, grow, and reproduce. It collects

remarks, identifies returning phenomena, and using

existing knowledge attempts to explain these

phenomena, Like other scientific fields. However, this

attempt is a compound one, and biologists would

advantage from computational tools to support them

in construct and evaluate their models.

The achievement of apparatus learning and data

extraction in business areas has led to enhanced

interest in using comparable methods to find out

information in biology and other scientific

arrangements (Fayyad et al., 1996). However, in the

absence of background knowledge, the best-developed

techniques are designed to operate on large data sets.

Regardless the contrary of rhetoric, biology stay puts

a data-sparse field, but it has significant knowledge

available to limit the study for models.

The construction descriptive models are another

disadvantage of standard induction methods. These

can construct precise forecasts on new assessment

cases, which may be adequate for commercial uses,

but biologists naturally desire descriptive models of

behavior. A reason of some occurrence is cast in terms

of other knowledge, such as constructions or courses

that are well-known to field specialists.

Finally, customary instruction techniques create

models that are articulated in documents designed by

computer scientists, few of which biologists discover

intelligible. Even effort on inducing causal models,

which often have an instructive taste, spotlights on

summary formalisms that compose little contact with

implications from biomedical science. Information

that supports the integration of area impressions more

directly would most probably be easier to recognize

and make available additional controls on model

mechanism (Langley et al., 2006).

Biologists require computational methods for data

analysis and organization, with the increasing volume

of genomic data available. We first require a

knowledge model that can indicate biological systems,

to develop computer applications that aid in this task.

Such a model should demonstrate the advanced

physiological processes and attach them to molecular-

level applications. The appearance of structural

reforms, such as the Gene Ontology (The Gene

Ontology Consortium, 2000) is an important first step

in making the communications requisite for a

biological knowledge model (Peleg et al., 2001).

A model of biological process can (1) constantly

indicate the dynamic knowledge about advanced

biological processes, in the framework of their part

molecular-level sub-processes, and (2) be agreeable to

deduction, validation of dynamic (control-flow)

characteristics, and qualitative simulation.

There are a set of desirable requirements for a

model of biological process. First, the model has to

show the following three cases of a biological system.

(1) The Static-structural outlook of bio-molecular

compounds, chemical, and biopolymers that

contribute in the system, their characteristics, and the

relationships between them;

(2) The Dynamic view that indicates how methods

are arranged over time (control flow) and how a

course is recursively decomposed to component

processes and reactions (atomic processes). The

dynamic model have to carry parallel, consecutive,

restrictive and iterative processes; and

(3) The practical view that demonstrates the

performers (e.g., enzyme) that carries out each

purpose, the substrates (input) of each function, and

the results of the function (outputs). The cellular

position of the substrates and products has to be

particular (Peleg et al., 2001).



67

Second, the model has to comprise a biological

ontology that will characterize biological thoughts and

organize them in classification hierarchies. Ontologies

make available constant definitions and explanations

of biological concepts, and make possible software

applications to share and reprocess the knowledge

constantly (Gruber, 1995). Ontologies can be used to

make logical conclusion over the set of impressions to

provide for simplification and clarification facilities

(Schulze- Kremer, 1998).

Third, the demonstration should be sensitive.

Biologists have to find it easy to make and realize a

system that is demonstrated using the model.

Fourth, the model has to be hierarchical to handle the

density of the represented system.

Fifth, the model has to be mathematically based to

permit verification of characteristics that are

advantageous in biological systems, and simulation of

system activities. For example, we were able to

validate that there is no toxic aggregation of

metabolites in pathways (boundedness property).

For a high-level biological process consisting of

lower-level processes, could validate that all the

component processes can contribute (liveness

property). We are able to also investigate to processes

that are constantly in use (violate the fairness

property) and may consequently be good objectives

for affecting system manners. A formal model have to

also make possible us to ensure whether we are able to

move from one system state (e.g., parasite inside host

liver cell) to another status (e.g., parasite cleared).

And sixth, the model should permit important

deduction abilities. For example, proposing the

consequences of knockout experiments – substrates

that may accumulate, other reactions that might

procedure these substrates, etc. Other examples

contain recognizing processes that are operated in

response to environmental effects (e.g., heat shock),

or discovery procedures that are active during a

determined developmental stage. The argument

mechanism should apply the biological concept model

to structure abstract concepts (e.g., tight-junction

formation is a kind of adhesion process) (Peleg et al.,

2001).

4.1. TAMBIS model

Clear Access to numerous Biological Information

resources (TAMBIS) (Baker et al., 1999) is ontology

for a description of data to be fined from

bioinformatics resources. TAMBIS goes beyond

taxonomic models. It demonstrates biological

concepts and appearances a semantic template of

concept relations that can be used to create deductions

from biological data. TAMBIS is defined as a logic-

based knowledge representation formalism using

explanation Logic (Bordiga, 1995) that defines

concepts using argument to sort the concepts based on

those descriptions in terms of their properties.

4.2. The Workflow model of the Workflow

Management Coalition

The Workflow Management Coalition is defined as a

Workflow model for business procedures (Workgroup

Management Coalition, 1999). Definition of the

Workflow model is a representation of a business

(high-level biological) process in an automated

manipulation form. The process definition presents a

functional and dynamic model that includes of a

network of activities (logical steps in the process) and

their relationships, criteria to demonstrate the start and

ending of the process, and information about the

participants of individual activities.

4.3. Petri Nets model

A Petri Net (Peterson, 1981) which is used to model

concurrent systems defined a formal model. A Petri

Net is represented by a directed, bipartite graph in

which nodes are either locations or transmissions,

where locations indicate conditions (e.g., parasite in

blood stream) and transitions demonstrate activities

(e.g., invasion of host erythrocytes).

5. ACTIVATED SLUDGE PROCESS

The activated-sludge process (ASP) is one of the

forms of wastewater treatment via biological method.

It is principally based on the utilization of different

but mixed community of microbes within an aerobic

aquatic environment. The carbonaceous organic

matter in the aerated wastewater serve as the source of

energy for the microbes which they use for generation

of new cells via biosynthesis, while undergoing

respiration simultaneously by giving out energy when

the organic matter is converted into compounds of

lower energy, especially carbon dioxide and water.

Similarly, some of the microbes in this system

generate energy through nitrification process;

ammonia nitrogen is converted to nitrate nitrogen.

Therefore, activated sludge refers to the consortium of

microbes with the associated biological component of

the process.

5.1. Principles of Activated Sludge Process (ASP)

Dadrasnia et al.


68

Storage of the wastewater in aeration tank

• Bacteria growth is enhanced by supplying

oxygen, food (BOD), time and optimal temperature.

The consortium of BOD substrate enhances bacteria

growth and increased population.

• Secondary clarifiers receive the treated

wastewater, and the bacteria cells that settle are

collected from the clarifiers as sludge.

• A portion of the sludge is returned back to the

activated sludge tank (partial recycling) to ensure that

certain bacteria population is maintained in the

system.

• Remaining sludge is totally removed as waste.

-ASP 1

The ASP comprised of interrelated equipment (5) and

components. Aeration tanks(s) are the first

components which serve as an aerobic environment

when air or oxygen is injected/supplied to the system.

Such environment basically satisfies the requirement

of the biological community and thoroughly ensures

proper mixing of the activated sludge (Figure 2).

Fig. 2: Schematic presentation of aeration reactors in Activated Sludge Process

-ASP 2

An aeration source is the second vital component

which supplies sufficient oxygen to the tanks and

ensures adequate mixing within the system. The

oxygen supply can be in form of mechanical aeration

or introduction of pure compressed oxygen.

Equipment for the supply of oxygen to this system can

vary in design and operation due to the number and

modifications in shape of aeration tanks used.

-ASP 3

Third in the line of ASP equipment are the secondary

clarifiers. Here, separation of activated sludge solids

from water occurs. This is based on the principle of

flocculation (aggregation of large particles or flocs

due to adherence of floc-forming organisms to

filamentous organisms) and gravity sedimentation,

where flocs settle at the base of the clarifier sluggish

environment. Subsequently, a secondary effluent

(wastewater characterized of low-level activated-

sludge solids that are in suspension) is formed from

the separation at the upper segment of the clarifier.

However, at the base of the clarifier is a thickened

sludge characterized of flocs and referred to as ―return

activated sludge‖ (RAS).

-ASP 4

The RAS generated must be removed from the

clarifiers and reintroduced into the aeration tank(s)

before depletion of the dissolved oxygen present. By

so doing, there is replenishment of the biological



69

community required to metabolize influent organic or

inorganic matter in the wastewater stream.

-ASP 5

Here, the activated sludge which is characterized of

high microbial load must be removed or discarded

(referred to as waste activated sludge or WAS) from

the treatment process. It is done using pumps and

carried out in parts so as to regulate the food-to-

microorganism ratio (F:M) inside the aeration tank.

5.2. ASP Process Requirements

A standard activated sludge process will require the

following for effective operation;

(a) Formation of flocs as it ensures effective

separation of the microbial biomass.

(b) The mean cell residence time (MCRT) should

be increased alongside adequate aeration in order to

enhance settling ability of the microbial biomass.

(c) Oxygen level needs to be properly regulated.

(d) Microorganisms constitute the biological

component of activated sludge system. The microbes

can contain about 70 – 90% organic matter. In fact,

the wastewater composition and the specific

characteristics of microbes in the biological

community determine the cell makeup.

(e) Most commonly found organisms within the

biological component of activated sludge are bacteria,

fungi, protozoa and rotifiers. In some cases, nematode

worms which are from metazoan can be present.

However, the growth of higher microbes can be

impaired by the constant agitation in the aeration

tanks in addition to recirculation connected with

sludge.

In 1990s’ when ASP was developed, it was

characterized of fill and draw reactors (batch process).

The continuous flow reactors were developed in order

to tackle the problem of regulating a number of batch

reactors that always have varied influent flow rate.

Interestingly, over the years, sequential batch reactor

(SBR) has replaced the old system and aeration

process is the significant energy consumption

operation of the system, whereas operations on RAS

and clarifiers are insignificant. Yet, the major

expenses can arise from the sludge processing and

disposal.

6. CONCLUSION

Previously, water was regarded as the most abundant

natural resource based on its global distribution, and

as such less caution was taken to on securing its

natural quality. Unfortunately, it has turned out to be

the scarcest resource now and no thanks to some

socio-economic developments that do not take into

account environmental impact assessment. Therefore,

wastewater treatment has been adopted as a mitigation

agent that will prevent aquatic pollution. Hence many

treatment options abound; physical, chemical and

biological. Adoption of biological method ahead of

others is being canvassed for due to its greener nature

and relatively reduced cost. Yet, the time consumed

when using biological methods of wastewater

treatment is a recurring factor that is a drawback, and

the broad spectrum of pathogens in wastewater often

rise concerns on the use of microbes in treatments.

Acknowledgement

We wish to express our deepest gratitude to all the

researchers whose valuable data as reported in their

respective publications and cited in this chapter have

been of considerable significance in adding substance.

REFERENCES

Baker PG, Goble CA, Bechhofer S, Paton NW,

Stevens R, Brass A (1999). An ontology for

bioinformatics applications. Bioinformatics, 15:

510-520.

Bordiga A (1995). Description Logics in Data

Management. IEEE Transactions on

Knowledge and Data Engineering 7, 671-682.

Cao XZ, Li YM (2011). Treatment of linear

alkylbenzene sulfonate (LAS) wastewater by

internal electrolysis-biological contact

oxidation process. Water Sci. Technol., 64(1):

147–154.

Dewil R, Baeyens J, Neyens E (2006). Reducing the

heavy metal content of sewage sludge by

advance sludge treatment methods. Environ.

Eng. Sci., 23: 994–999.

Doorn MRJ, Towprayoon S, Maria S, Vieira M,

Irving W, Palmer C, Pipatti R, Wang C (2006).

Wastewater treatment and discharge, in: 2006

IPCC Guidelines for National Greenhouse Gas

Inventories. vol. 5, (Waste). WMO, UNEP, pp.

6, 1–6, 28.

Fayyad U, Haussler D, Stolorz P (1996). KDD for

science data analysis: Issues and examples.

Proceedings of the Second International

Conference of Knowledge Discovery and Data

Mining (pp. 50{56). Portland, OR: AAAI Press.

Fu Z, Zhang Y, Wang X (2011). Textiles wastewater

treatment using anoxic filter bed and biological

wriggle bed-ozone biological aerated filter.

Bioresource Technol., 102(4): 3748–3753.

Gruber TR (1995). Toward Principles for the Design

of Ontologies Used for Knowledge Sharing. Int.

Journal of Human-Computer Studies 43.

Dadrasnia et al.


70

Leal LH, Vieno N, Temmink H, Zeeman G, Buisman

CJ (2010). Occurrence of xenobiotics in gray

water and removal in three biological treatment

systems, Environ. Sci. Technol., 44(17): 6835–

6842.

Langley P, Shiran O, Shrager J, Todorovski

L, Pohorille A (2006). Constructing explanatory

process models from biological data and

knowledge. Artificial intelligence in medicine,

37(3): 191-201.

Oller I, Malato S, Sánchez-Pérez J A (2011).

Combination of Advanced Oxidation Processes

and biological treatments for wastewater

decontamination—A review. Science of the

Total Environment, 409(20), 4141-4166.

Peleg M, Yeh I, Altman RB (2001). Modeling

biological processes using Workflow and Petri

Net models. Stanford Medical Informatics.

Stanford University, Stanford, CA, 94305,

USA.

Peterson JL (1981). Petri Net Theory and the

Modeling of Systems (Englewood Clifs, NJ:

Prentice-Hall). Schulze-Kremer S (1998). On

tologies for Molecular Biology. In Proceedings

of the Third Pacific Symposium on

Biocomputing, pp. 693-704.

Souza B.M, Cerqueira A.C, Sant Anna G.L, Dezotti

M (2011). Oil-refinery wastewater treatment

aiming reuse by advanced oxidation processes

(AOPs) combined with biological activated

carbon, Ozone. Sci. Eng., 33(5): 403–409.

The Gene Ontology Consortium (2000). Gene

Ontology: tool for the unification of biology.

Nature Genetics, 25: 25-29.

Vijayaraghavan K, Ahmada D, Mohd A.Z (2007).

Aerobic treatment of palm oil mill effluent, J.

Environ. Manage., 82; 24–31.

Workgroup Management Coalition (1999). Interface

1: Process Definition Interchange.

http://www.wfmc.org/standards/docs/

Sperling M, Lemos Chernicharo CA (2005).

Biological Wastewater Treatment in Warm

Climate Regions, IWA Publishing, London.

Kotsou M, Kyriacou AK, Lasaridi G (2004).

Integrated aerobic biological treatment and

chemical oxidation with Fenton’s reagent for

the processing of green table olive wastewater.

Process Biochem, 39: 1653–1660.

http://www.ncbi.nlm.nih.gov/pubmed?term=Langley%20P%5BAuthor%5D&cauthor=true&cauthor_uid=16781850

http://www.ncbi.nlm.nih.gov/pubmed?term=Shiran%20O%5BAuthor%5D&cauthor=true&cauthor_uid=16781850

http://www.ncbi.nlm.nih.gov/pubmed?term=Shrager%20J%5BAuthor%5D&cauthor=true&cauthor_uid=16781850

http://www.ncbi.nlm.nih.gov/pubmed?term=Todorovski%20L%5BAuthor%5D&cauthor=true&cauthor_uid=16781850

http://www.ncbi.nlm.nih.gov/pubmed?term=Todorovski%20L%5BAuthor%5D&cauthor=true&cauthor_uid=16781850

http://www.ncbi.nlm.nih.gov/pubmed?term=Pohorille%20A%5BAuthor%5D&cauthor=true&cauthor_uid=16781850



©2014 IJSRPUB

72

Biological Treatment of Recycled Paper Mill Wastewater Using Modified Anaerobic

Inclining-Baffled Bioreactor (MAIB-R)

Haider M. Zwain1, Irvan Dahlan

2*

1School of Civil Engineering, Universiti Sains Malaysia, Engineering Campus, Seri Ampangan, 14300 Nibong Tebal,

Pulau Pinang, Malaysia. 2School of Chemical Engineering, Universiti Sains Malaysia, Engineering Campus, Seri Ampangan, 14300 Nibong Tebal,

Pulau Pinang, Malaysia.


Abstract. Industrial wastewater treatment has gained importance due to the insistence of the environmental groups,

environmental regulation and public awareness. Wastewater from pulp and paper industry is among the most harmful

wastewater that needs to be treated since it contains high organic substances. Although the raw materials of this kind of

industry has been replaced by recycled materials in order to reduce wastes, however effluent from this recycled paper mill is

still highly polluted that need to be treated. In this study, the start-up performance of treating recycled paper mill wastewater

using a modified anaerobic inclining-baffled bioreactor (MAIB-R) has been studied. The MAIB-R was started with a hydraulic

retention time (HRT) of 5 days at 37˚C. The start-up process took 14 days before it reached steady state on day 15. The MAIB-

R, after successful start-up, can achieve relatively high average COD removal of 92% with pH level of effluent relatively

stable at 6.63. In addition, the biogas production reached 0.77L/d with methane content up to 57% on day 15. The result shows

that during the start-up period, MAIB-R was successfully operated in treating recycled paper mill wastewater.

Keywords: Anaerobic Inclining-Baffled, Biological Treatment, Paper Mill Wastewater

1. INTRODUCTION

1.1. Recycled Paper Mills

Paper and pulp industry together with recycled paper

mill ranked among the largest industries in the world.

There are mills found in more than 100 countries in

every region of the world. The major pulp and paper

producing nations include the United States, China,

Japan, Canada, Germany, Brazil, Sweden, Finland and

France. In Malaysia, recycled paper and boards are

important sources for pulp and paper industry. Of 20

paper mills in the country, only Sabah Forest

Industries Sdn. Bhd. is an integrated pulp and paper

mill. All of the 19 others utilize 95% waste papers as

their main raw material except for Kimberly Clark

(M) Sdn. Bhd. which uses 80% imported virgin pulp.

Table 1 shows the existing pulp and paper companies

in Malaysia (Roda and Rathi, 2006).

Table 1: Pulp and Paper Industry in Malaysia.

Company Total Capacity Per Annum (mt)

Cita Peuchoon 30,000

Johmewah 35,000

Genting Sanyen 300,000 Muda Paper (Kajang) 170,000

Muda Paper (S. Prai) 130,000 Malaysia Newsprint 250,000

Nibong Tebal 60,000

Pascorp Paper 140,000 Pembuatan Kertas (Perak) 3,000

Sabah Forest 165,000

See Hua Paper 12,000 Taiping Paper 2,400

Theen Seng Paper 15,000

Trio Paper 30,000 Union Paper 12,400

United Paper Board 80,000

Yeong Chaur S 3,600

1.2. Recycled Paper Mill Wastewater and Its

Environmental Hazards

Recycled paper mills (RPM) are using recovered

paper as their feedstock, unlike paper and pulp mills

which are using wood pulp, mixture of cellulose fibers

and water as the basis of all paper products. The

recovered paper is used to produce new paper

products. This is an effort emphasize on the present

green awareness which can help to reduce the cutting

of trees. However, RPM is one of the industries that

have caused concerns about hazardous water

pollutants being continuously discharged into streams

and other water bodies without clean-up treatment.

Zwain and Dahlan

Biological Treatment of Recycled Paper Mill Wastewater Using Modified Anaerobic Inclining-Baffled (MAIB)

Bioreactor

73

The RPM wastewater characteristics is shown in

Table 2 (Žarković et al., 2011). Furthermore, the RPM

wastewater contain heavy metal is given in Table 3

(Hassan et al., 2014).

Table 2: Typical Physico-chemical characterization of RPM wastewater

Parameter (in mg/L, except pH value) Range Mean ± S.D.

pH 7.0-7.7 7.3±0.22

Chemical oxygen demand, COD 3770-9330 5539±412

Biochemical oxygen demand, BOD 816-2495 1372±108

Ammonia, as NH4 1.4-3.8 2.36±0.8

Total solids, TS 2051-11161 4654±276

Total dissolved solids, TDS 200-892 595±50

Total suspended solids, TSS 603-8495 3993±216

Settleable solids, SS 40-850 480±58

Table 3: Heavy metals present in RPM wastewater

Heavy metals Concentration (mg/L) Standard B*(DOE, 2009)

As 1.06 0.1

Be 0.01 –

Ca 398.78 –

Cd 0.39 0.02

Co 0.64 –

Cr 0.51 0.05

Cu 1.10 1.00

Fe 2.39 1.00

Li 0.65 –

Mg 12.19 –

Mn 0.85 0.20

Mo 0.92 –

Ni 0.72 0.20

Pb 1.11 0.50

Sb 1.47 –

Se 1.63 –

Ti 0.99 –

Tl 0.82 –

V 0.86 –

Zn 1.39 1.00

*Acceptable conditions for discharge of industrial effluent or mixed effluent of standards B.

Untreated recycled paper mill effluent (RPME) can

cause disastrous environmental consequences,

including the destruction of fisheries and the

contamination of drinking and irrigation water.

Studies have showed that pulp and paper industry is

considered as the third largest polluter in the United

States (Sinclair, 1991). The effluent generated by pulp

and paper industry contains a considerable amount of

pollutants characterized by biochemical oxygen

demand (BOD), chemical oxygen demand (COD),

suspended solids (SS), toxicity, and colour when

untreated or poorly treated effluents are discharged to

receiving waters (Pokhrel and Viraraghavan, 2004).

Black liquor, an aqueous solution of lignin

residues, hemicelluloses and inorganic chemicals used

in the paper manufacturing process is by-product of

many paper pulp mill and recycled paper industries. It

comprises 15% solids by weight of which 10% are

inorganic and 5% are organic. Black liquor released

by paper mills constitutes only 10-15% of total

wastewater, but contributes approximately 95% of the

total pollution load of pulp and paper mill effluents

(Grover et al., 1999). The black liquor has

characteristically contained high biochemical oxygen

demand (BOD), chemical oxygen demand (COD), and

total solids (TS) along with slowly degradable lignin

compounds, which make it significantly toxic to the

environment. Thus pulp and paper mill effluent

treatment has to be taken seriously as it can cause

serious environmental pollution and threat to human

health.

1.3. Biological Treatment

Many new wastewater treatment methods are being

developed and designed from recent research findings

to create more efficient and cost effective wastewater

treatment systems. Wastewater is treated by removing

or reducing certain harmful constituents found in

wastewater. These constituents are removed by

physical, chemical and biological methods. Table 4

listed the processes involve for each method (United

Nations, 2003).

Almost all wastewater contains biodegradable

constituents, therefore it can be treated biologically

with proper analysis and understanding of biological

treatment processes. Biological unit operations shown

in Table 1.4 are used to transform (i.e. oxidize)



74

dissolved and particulate biodegradable constituents

into desired end products (i.e. biogas and cell tissues).

Besides that, these unit operations are used to remove

or reduce the concentration of organic content which

measured as BOD, TOC or COD and nutrient content

(mainly nitrogen and phosphorus) that are relatively

high and above allowable limit in regulations.

Physical and chemical methods, although quite

effective in reducing the constituents of wastewater

which are harmful, are unattractive for industrial

applications because of the high costs involved.

However, biological methods of effluent treatment

have the advantage of being cost-effective and they

can also reduce both the biological oxygen demand

(BOD) and chemical oxygen demand (COD) of the

wastewater. Biological treatment can be divided to

aerobic process, anaerobic process, and anoxic

process.

Table 4: Wastewater treatment unit operations and processes Methods Unit operations

Physical unit operation Screening

Comminution

Flow equalization

Sedimentation

Floatation

Granular-medium filtration

Chemical unit operation Chemical precipitation

Adsorption

Disinfection

Dechlorination

Other chemical applications

Biological unit operation Activated sludge process

Aerated lagoon

Trickling filters

Rotating biological contractors

Pond stabilization

Anaerobic digestion

Biological nutrient removal

1.4. Anaerobic Process

Since the early 1980’s, anaerobic process for

wastewater treatment has attracted a lot of attention.

This process has advantages as design simplicity, use

of non-sophisticated equipment, high treatment

efficiency, low excess sludge production and low

operating and capital cost (Saktaywin et al., 2005).

There are three basic steps are involved in the overall

anaerobic oxidation of a waste: (1) hydrolysis, (2)

fermentation (also referred to as acidogenesis) and (3)

methanogenesis. The first step for most fermentation

processes, in which particulate material is converted

to soluble compounds that can then be hydrolyzed

further to simple monomers that are used by bacteria

that perform fermentation, is termed hydrolysis.

The second step is fermentation (also referred to as

acidogenesis). In the fermentation process, amino

acids, sugars, and some fatty acids are degraded

further. Organic substrates serve as both the electron

donors and acceptors. The principal products of

fermentation are acetate, hydrogen, CO2, and

propionate and butyrate. The propionate and butyrate

are fermented further to produce hydrogen, CO2, and

acetate. Thus, the final products of fermentation

(acetate, hydrogen, and CO2) are the precursors of

methane formation (methanogenesis) (Wang, 2012).

The third step, methanogenesis, is carried out by a

group of organisms known collectively as

methanogens. Two groups of methanogenic organisms

are involved in methane production. One group,

termed aceticlastic methanogens, split acetate into

methane and carbon dioxide. The second group,

termed hydrogen-utilizing methanogens, uses

hydrogen as the electron donor and CO2 as the

electron acceptor to produce methane. Bacteria within

anaerobic processes, termed acetogens, are also able

to use CO2 to oxidize hydrogen and form acetic acid.

However, the acetic acid will be converted to

methane, so the impact of this reaction is minor

(Wang, 2012).

1.5. Methane Gas Production

The microorganisms responsible for methane

production, classified as archaea, are strict obligate

anaerobes. Many of the methanogenic organisms

identified in anaerobic digesters are similar to those

found in the stomachs of ruminant animals and in

organic sediment taken from lakes and rivers. In

methanogenesis, methanogens consume the acids

produced by the acidogens, generating biogas

(methane gas) as byproduct. Instead of oxygen

accounting for the change in COD, the COD loss in

Zwain and Dahlan


Bioreactor

75

the anaerobic reactor is accounted for by the methane

production (Nègre and Jonsson, 2010). By

stoichiometry the COD equivalent of methane can be

determined as shown in equation 1.1.

(1.1)

1.6. Anaerobic Baffled Reactor

Nowadays, many researchers have focused on

anaerobic reactors for the treatment of wastewater. As

one of the high-rate anaerobic reactors, ABR was

extensively used in treating wastewater (Liu et al.,

2010). The reactor design has been developed since

the early 1980s. ABR comprises a series of vertical

baffles to force the wastewater to flow under and over

them and therefore, the wastewater comes into contact

with a large active biological mass (Grover et al.,

1999). Bacteria within the reactor gently rise and

settle due to flow characteristics and gas production,

but move down the reactor at a slow rate. Figure 1

shows a schematic diagram of anaerobic baffled

reactor (Zwain et al., 2013).

Fig. 1: Schematic diagram of anaerobic baffled reactor.

Each gas chamber was separated permitting the

measurement of gas composition and production from

each compartment. The separation of the gas can also

enhance the reactor stability by shielding syntrophic

bacteria from the elevated levels of hydrogen which

are found in the front compartments of the baffled

reactor (Barber and Stuckey, 1999).

The performance of the ABR in treating a variety

of wastewaters, in particular, low and high strength

wastewater and other refractory wastewaters has been

well reviewed in the literature (Liu et al., 2010).

Table 5 shows low strength wastewater treated by

ABR. Dilute wastewaters inherently provide a low

mass transfer driving force between biomass and

substrate, and subsequently biomass activities will be

greatly reduced according to Monod kinetic. Thus,

ABR is effective in treating low strength wastewater

with no substantial change occurred in the population

of acid producing bacteria down the length of the

reactor. Lower gas production rates can help to

overcome the problem of sludge washout with low

HRTs (6-2 days) in which can increase hydraulic

turbulence, which can lower apparent KS values, thus

enhance treatment efficiency. When treating dilute

wastewater, baffled reactors should be started-up with

higher biomass concentrations in order to obtain a

sufficiently high sludge blanket (and better mixing) in

as short time as possible to counteract the problem of

low sludge blankets caused after long periods of

biomass settling (Kato et al., 1997).

In addition, ABR is also applied for the treatment

of high strength wastewaters. A brief summary of the

literatures on high strength treatment is shown in

Table 6. In high strength wastewater treatment, longer

retention times are necessary due to the high gas

mixing caused by improved mass transfer between the

biomass and substrate. Longer retention time

enhances biomass settling ability. According to

kinetic consideration, high substrate concentrations

will encourage to increase the rate of bacteria growth

and gas production.



76

Table 5: Selected low strength performance data (Liu et al., 2010). Wastewater HRT

(h)

Influent COD

(mg/L)

COD Removal

(%)

OLR

(kg/m3d)

Gas Produced

(LCH4/g COD d)

Reference

Synthetic

wastewater

10 501 90.7 1.2 0.36 (Gopala Krishna and

Kumar, 2007)

Domestic

wastewater1

48 305.18±36.22 74±5 - - (Feng et al., 2008)

Municipal

wastewater

6 350 86 2.62 0.34 (Bodkhe, 2009)

Complex

wastewater

8 500 88 2 0.31 (Gopala Krishna et al.,

2008)

Diluted wastewater 10 500 95 0.13 0.3 (Langenhoff and

Stuckey, 2000)

Domestic

wastewater2

22 716±54.4 72±3 - - (Foxon et al., 2004)

Low strength

wastewater

12 550 89 1.69 - (Shen Y.L., 2004)

Low strength

wastewater

3 850 90 6.9 - (Shen Y.L., 2004)

1= at 28±1˚C; 2= on pilot scale; others carried out on laboratory scale.

Table 6: Selected high strength performance data (Liu et al., 2010)

Wastewater HRT

(h)

Influent COD

(mg/L)

COD Removal

(%)

OLR

(kg/m3d)

Gas Produced (LCH4/g

COD d)

Reference

Palm oil mill wastewater

72 16000 77.3 5.33 0.33 (Faisal and Unno,

2001)

Whisky distillery

wastewater

96 9500 96.1 2.38 - (Akunna and Clark,

2000)

Brewery wastewater

19.23 10720 93 13.38 - (Baloch et al.,

2007)

Soybean protein processing

wastewater

39.5 1000 97 6.0 - (Zhu et al., 2008)

ABR shows a promise for industrial wastewater

treatment due to its ability to withstand severe

hydraulic and organic shock loads, intermittent

feeding, temperature changes, and tolerant certain

toxic materials. The successful application of

anaerobic technology in the treatment of industrial

wastewater critically dependent on the development

and use of high rate anaerobic bioreactors (Barber and

Stuckey, 1999). Table 7 shows the recommendations

based on literature findings on anaerobic baffled

reactor.

1.7. Advantages of Anaerobic Baffled Bioreactor

The most significant advantage of the ABR is its

ability to separate acidogenesis and methanogenesis

longitudinally down the reactor, allowing the reactor

to behave as a two-phase system without the

associated control problems and high costs (Hassan et

al., 2013). Barber and Stuckey (1999) study showed

that ABR is capable of treating a variety of

wastewaters of varying strength, over a large range of

loading rates, and with high solids concentrations.

Table 8 shows the advantages associated with

anaerobic baffled reactor. 2. CASE STUDY

The development of effective and simple methods for

treatment recycled paper mills (RPM) wastewater is a

challenging task to environmental engineers and

scientists. Therefore, in this study, a novel modified

anaerobic inclining-baffled bioreactor (MAIB-R) was

developed and tested for RPM wastewater treatment.

The aims of this study are to characterize the RPM

wastewater and to perform the start-up of MAIB-R.

Zwain and Dahlan


Bioreactor

77

Table 7: Recommendations based on literature findings (Barber and Stuckey, 1999) Recommendations

Start-up

Low initial loading rates will encourage granule/ floc growth. Pulses of methane precursors (e.g. acetate) have been successfully used to encourage

methanogenic growth and dampen the effects of increases in loading rate.

Start-up with long retention times reduces solids loss due to low liquid upflow velocities and promotes higher methanogenic populations in every compartment.

Recycle

Recycle is beneficial with respect to diluting toxicants in feed stream, increasing front pH and reducing production of foam and SMPs.

Low strength wastewater Low retention time enables better mass transport due to improved hydraulic mixing and reduces biomass starvation in latter Compartments.

methane production will originate from scavenging bacteria (Methanosaeta).

High strength wastewater

Long retention times reduce solids washout caused by high gas production, otherwise

the reactor may be modified (by adding packing) to decrease biomass loss. Methane production will be mainly due to Methanosarcina, and hydrogen scavenging

methanogens.

High solids wastewater

A larger front compartment has proved to be effective in treating wastewater with high

solids content.

Temperature Reducing temperature to 25˚C from 35˚C has no effect on easily degradable waste,

further decreases in temperature are detrimental on reactor performance, this may be

due to potential toxicity, nutrient bioavailability and slower kinetic rates. Reactors started-up and kept at lower temperatures perform consistently well

2.1. Reactor Design and Fabrication

MAIB-R used in this study was rectangular, consisted

of five compartments and constructed by using plexi-

glass with a dimension of 80cm long, 15cm wide and

30cm in height, having a total effective volume of 35

L as shown in Figure 2. Each compartment composed

of upflow and downflow sections by using modified

inclining baffles in series in which this bioreactor is

called as modified anaerobic inclining-baffled

bioreactor (MAIB-R). It was a modification from

ABR. Feed tank was equipped with a mixer to mix the

RPM wastewater well before it was pumped into

MAIB-R. Peristaltic pump was used to adjust the flow

rate of incoming RPM wastewater into the MAIB-R.

The biogas generated from each compartment was

collected by using biogas bag. The volume of the

biogas was measured according to the volume of

biogas bag. Each compartment was equipped with

liquid sampling ports for the convenient of grabbing

liquid sample to be analyzed for its characteristic. The

MAIB-R was maintained at 37 ºC in a water bath. The

treated RPM wastewater was collected in effluent

collection tank to be analyzed for its characteristic. 2.2. Seeding and Acclimatization

The inoculum for seeding the MAIB-R was

anaerobically digested sludge taken from Palm Oil

Mill Effluent (POME) (Malpom Palm Industries Bhd,

Penang, Malaysia). While the RPM wastewater

samples were taken from Muda Paper Mill Penang,

Malaysia. The sludge and RPM wastewater samples

were kept in refrigerator at 4°C before they were used

for seeding process to avoid biological contamination

at different temperature.

Initially, the MAIB-R was inoculated with mixture

of POME sludge (about 10% of total volume of

MAIB-R) and fresh RPM wastewater taken from

Muda Paper Mills Penang (about 10% of total sludge

mixture). Then, the MAIB-R was sealed tightly so that

no air could enter the MAIB-R. Besides that, the

MAIB-R was flushed with nitrogen gas to displace

any air inside the bioreactor and purge out to remove

oxygen content before feeding with RPM wastewater.

Then, the MAIB-R was daily batch-fed with fresh

RPM wastewater which was diluted to 1000 mg

COD/L as much as 10% of total volume of bioreactor

(3.5 liter) every day until the MAIB-R reach its

capacity. Then, continuous phase was started. The

acclimatization of sludge with RPM wastewater

during start-up period was monitored by daily

measurement of physico-chemical parameters (pH,

temperature, COD and biogas production) until steady

state is achieved.



78

Table 8: Advantage associated with the anaerobic baffled reactor (Barber and Stuckey, 1999)

Advantage

Construction

1. Simple design

2. No moving parts

3. No mechanical mixing

4. Inexpensive to construct

5. High void volume

6. Reduced clogging

7. Reduced sludge bed expansion

8. Low capital and operating costs

Biomass

1. No requirement for biomass with unusual settling properties

2. Low sludge generation

3. High solids retention times

4. Retention of biomass without fixed media or a solid-settling chamber

5. No special gas or sludge separation required

Operation

1. Low HRT

2. Intermittent operation possible

3. Extremely stable to hydraulic shock loads

4. Protection from toxic materials in influent

5. Long operation times without sludge wasting

6. High stability to organic shocks

Fig. 2: Schematic diagram of the MAIB bioreactor experimental set-up.

2.3. Experimental Analysis procedure

The RPM wastewater collected from Muda Paper Mill

Penang were subjected to the analysis of the following

parameters, i.e. pH, chemical oxygen demand (COD),

biochemical oxygen demand (BOD), total dissolved

solids (TDS), total suspended solids (TSS), total

solids (TS), volatile suspended solids (VSS) and

biogas methane concentration. These parameters

analyses were performed according to American

Public Health Association (APHA, 1995). They were

analyzed using specialized equipments, such as

Spectrophotometer DR-2800, Shimadzu Gas

Chromatography-FID (GC) with FID column and

COD reactor digest.

The MAIB-R was operated continuously until

steady-state condition achieved. The MAIB-R

samples were monitored each two days for pH,

temperature, COD and biogas. During the steady state,

the influent, each compartment and effluent samples

were collected. Samples were collected starting from

the last compartment towards the first compartment to

Zwain and Dahlan


Bioreactor

79

prevent air intrusion and maintain anaerobic

condition. All the parameters were determined

according to APHA Standard Method (APHA, 1995)

Biogas generated (methane gas) in each

compartment of MAIB bioreactor was collected from

the upper part of each compartment using biogas bag.

Volume of biogas was measured daily by using biogas

bag with a volume of 1 L per bag and pipette bulb

with 40 ml per pump. Shimadzu Gas

Chromatography-FID (GC) with propack N column

equipped with Flame ionization detector (FID) and

data acquisition system with computer software was

used to determine methane concentration in biogas

collected. Nitrogen gas was used as carrier gas. The

temperature of detector was set to 100˚C and injector

temperature at 90˚C. Methane concentration was

analyzed by injecting 1.0 ml of biogas sample in

column of GC using airtight glass syringe. Then, the

methane concentration was taken.

3. RESULTS AND DISCUSSION

3.1 Characteristic of Recycled Paper Mill

Wastewater

The characteristic of the RPM wastewater generated

depends on the type of processes, process technology

applied, management practices internal recirculation

of the effluent for recovery, and the amount of water

to be used in the particular process (Pokhrel and

Viraraghavan, 2004). Table 9 shows the characteristic

of RPM wastewater generated by Muda Paper Mill

Penang, Malaysia. The RPM wastewater generated

contained a considerable high amount of pollutants

characterized by various physico-chemical

constituents. If the untreated RPM wastewater

discharges into streams and other water bodies, it can

cause disastrous environmental consequences,

including the destruction of fisheries and the

contamination of drinking and irrigation water.

Table 9: Characterization of RPM wastewater

pH - 6.2-7.8

Temperature ˚C 35-45

Biological oxygen demand (BOD) mg/L 1650-2565

Chemical oxygen demand (COD) mg/L 3380-4930

BOD5/ COD - 0.488-0.52

Total Solids (TS) mg/L 3530-6163

Total dissolved solids (TDS) mg/L 1630-3025

Total suspended solids (TSS) mg/L 1900-3138

Total volatile solids (VSS) mg/L 840-2920

These physico-chemical constituents were

relatively high compared to Malaysian Standard of

parameter limits of effluent discharge in Fifth and

Seventh Schedules of Environmental Quality

(Industrial Effluent) Regulations as shown in Table 10

and Table 11, respectively. There are two standards

stipulated in EQA 1974, namely Standard A and

Standard B. Standard A is for effluent discharges

upstream of a raw water intake and is more stringent.

On the other hand, Standard B is for effluent

discharges downstream of raw water intake. These

standards are interpreted as absolute standards, which

sewage and industrial effluent must comply before

being discharged; in this study is RPM wastewater.

This wastewater standard is established for the main

purpose of regulating the disposal pollutants into a

receiving water course to protect aquatic ecosystems,

public health and welfare.

Table 10: Environmental Quality Act 1974 and Regulations. Environmental Quality (Industrial Effluent) Regulations (PU (A)

434), Fifth Schedule

Parameters Units Standard

A B

Temperature ˚C 40 40

pH value - 6.0-9.0 5.5-9.0

Suspended solids mg/l 50 100

Table 11: Acceptable conditions for discharge of industrial effluent containing chemical oxygen demand (COD) for specific

trade or industry sector Trade/Industry Unit Standard A Standard B

(a) Pulp and paper industry

(i) pulp mill mg/L 80 350

(ii) paper mill (recycled) mg/L 80 250

(iii) pulp and paper mill mg/L 80 300



80

By comparing Table 9 with Table 10 and 11, the

characterization of RPM wastewater was far from the

allowable limit set by EQA 1974 for discharge

industrial effluent, therefore a proper and systematic

wastewater treatment is required to treat the RPM

wastewater to make sure the parameters are in

allowable range before it is discharged.

From Table 9, the RPM wastewater generated had

relatively neutral pH with the range of pH 6.2 to 7.8.

Thus, the wastewater does not require alkaline

adjustment because the pH is within the allowable

discharge range by EQA 1974. The BOD5/COD

relationship of 0.5 showed that the wastewater is of

high strength organic type and has a potential to

increase in strength with time, hence it’s suitable to be

treated anaerobically (Nyanchaga and Elkanzi, 2002).

The BOD5/COD was found comparable to the

average BOD5/COD ratio of 0.59 reported by

Nyanchaga and Elkanzi (2002).

The ratio of BOD5/COD of a wastewater is a good

indication of amount of the total organic load (or

oxygen demand) is available for degradation.

Indirectly, BOD5/COD ratio is a measure of food

value. If the ratio is higher, it has higher food and less

toxicity. However, researchers used to describe

BOD5/COD ratio as biodegradability level of

materials by which organic matter containing

wastewater is readily broken down in the

environment. BOD5/COD ratio can be categorized

into toxic, biodegradable and acceptable or stable

zones (Samudro and Mangkoedihardjo, 2010).

3.2 Start-up at Continuous Feeding

The start-up was the most crucial and important part

of overall MAIB-R processes, it requires a long period

of time to reach steady state due to slow growth rates

of the anaerobic microorganisms inside the MAIB-R,

especially methanogens. The adaptation during this

period is very important for the bacterial population to

develop the microbial culture for treating wastewater

in an anaerobic environment in the MAIB-R.

The anaerobic digestion process was begun in the

absence of oxygen. During this stage, MAIB-R was

under close monitoring to ensure that the anaerobic

active sludge is well grown by monitoring the

temperature of the bioreactor, pH, COD. The

continuous operation of MAIB-R was started at

hydraulic retention time (HRT) of 5 days. To obtain a

good start-up, low initial organic loading rate was

being carried out to provide good condition for the

growth of anaerobic active sludge as well as to

prevent slow growing microorganisms from

overloaded in MAIB-R.

Later, COD removal efficiency and the varying of

pH curve of RPM wastewater were closely monitored

and studied to determine the performance efficiency

of MAIB-R during start-up. These parameters were

analyzed using the standard methods by American

Public Health Association (APHA, 1995).

3.2.1. pH Level at Continuous Feeding

The obtained result for pH of MAIB-R in each

compartment during continuous feeding period for 15

days is shown in Figure 3. The pH level of RPM

wastewater in influent was relatively stable at about

7.90. However, in first compartment, the pH level was

decreased drastically with the range from 6.70 to 6.21.

On day 2 to day 6, the pH in compartment 2 to 4 were

increased, but the pH level was decreased in

compartment 5 and slightly increased again to 7.20,

7.10 and 7.03 for the effluent pH on day 2, day 4 and

6, respectively.

The decreasing of the pH in first compartment

(acidification zone) is due to the presence of large

population of acidogenic bacteria. The optimum pH of

acidogenic bacteria was recorded in the range

between 5.8 and 6.2 (Zoetemeyer et al., 1982) which

has shown on day 15 in first compartment. The

substrate level was high thus contributing to the fast

growing bacteria. According to Hu and Li (2008), the

pH level of each compartment increased towards the

end of the reactor from compartment 2 to 5 which has

been proven through this study.

On day 11 to day 15, pH in compartment 2 to 4 is

relatively stable and not much of fluctuation. On day

11, pH of MAIB-R from compartment 1 to 5 showed

a smooth increase which is the same as predicted by

Hu and Li (2008) in their study. If the first

compartment has higher pH compared to other

compartment, it could be due to the substrate level

was insufficient to feed the fast growing bacteria and

some of them starved thus died. This shows that pH

has an important effect on the selection, survival, and

growth of microorganism. On day 15, the pH in every

compartment showed a smooth increase and had

almost negligible fluctuation, hence steady state was

reached.

Zwain and Dahlan


Bioreactor

81

Fig. 3: pH profiles for each compartment of MAIB bioreactor during continuous feeding period

3.2.2. COD Removal at Continuous Feeding

As shown in Figure 4, the influent COD concentration

was ranging from 717-752mg/L with an average of

737mg/L. It was greatly reduced with the COD

concentration in effluent was varied in the range of

264 to 60 mg/L. On day 2, the COD removal

efficiency was 63% with influent and effluent COD at

720mg/L and 264mg/L, respectively. COD removal

efficiency was slightly increased to 74% on day 4.

Then, the COD removal rate was further increased

steadily with the influent COD relatively constant at

737mg/L. It increased gradually from 63% to 92%

from day 2 to 15. On day 11 to 15, the COD removal

efficiency of MAIB-R showed relatively stable with

the range from 91% to 92%.

On day 15, the influent COD concentration

reduced from 750mg/L to 60mg/L with COD removal

efficiency at 92%. This could be the beginning day of

steady state in MAIB-R. Steady state of COD removal

of more than 80% is considered acceptable for

anaerobic bioreactor start-up and acclimatization

(Buitron G., 2003; Enright A.M., 2005). As for other

type of wastewater, ABR can remove up to 77% of

COD for treatment of textile dye effluent with HRT

of 4 days (Goel, 2010). The average COD removal

efficiency of ABR was constantly maintained at

around 65% and the total COD removal rate of the

system in treating chemical synthesis-based

pharmaceutical wastewater was as high as 95% to

98% (Lili et al., 2009).

Fig. 4: COD concentration and removal efficiencies in the MAIB bioreactor during continuous feeding



82

3.2.3. Methane Gas Production at Continuous

Feeding

Table 12 shows the result of methane concentration in

biogas analyzed by GC. The methane gas production

rate during continuous feeding and at steady state was

monitored and analyzed from day 2 to 15. Methane

gas was produced along with degradation of organic

matter in MAIB-R. Methane concentration in biogas

was varied in the range of 46 to 70%. The average

methane concentration from day 2 to 15 was 67% to

57%. On day 2 to 6, the methane content was in the

range of 59 to 67% with 0.65L/d, 0.44L/day and

0.57L/day of biogas for day 2, 4 and 6, respectively.

The biogas yield was obtained by measuring the

volume of biogas produced within 24hours. Then, on

day 11 to 15, the methane content in biogas was

slightly higher than previous days in the range of 47 to

57% with varied amount of biogas by 0.67L/day,

0.77L/day and 0.77L/day for day 11, 13 and 15,

respectively.

Organic matter which is broken down by bacteria

without oxygen will produce significant quantities of

methane gas (CH4) with sufficient supply of nutrients

like nitrogen and phosphorus. The lack of nutrients or

substrate will limit biogas production. Having more

organic content in sludge will generate more methane

gas. Biogas mainly composed of methane and carbon

dioxide is a byproduct of anaerobic decomposition of

organic waste. Increase in biogas production and

methane content may be due to greater activity of

methanogenic bacteria in MAIB-R (Goel, 2010).

However, the quality of biogas and methane content

generated by organic waste materials does not remain

constant but varies with the period of digestion in

MAIB-R (Mahdi, 1986).

As for textile dye effluent, ABR generated up to

83% of methane content in biogas (Goel, 2010).

Meanwhile, methane content in biogas generated by

ABR in treating palm oil mill effluent was up to

67.4% (Setiadi et al., 1996).

Table 12: Methanogenesis of RPM wastewater using MAIB bioreactor at different time

Day Methane content (% v/v) Biogas yield

(L/d) 1 2 3 Average

2 70 65 66 67 0.65

4 69 64 62 65 0.44

6 52 60 65 59 0.57

11 46 50 46 47 0.67

13 58 53 59 57 0.77

15 55 57 59 57 0.77

3.3. Steady State Performance

3.3.1. pH Level

Steady state was achieved on day 15 of continuous

feeding. As shown in Figure 5, on day 15, influent pH

of RPM wastewater was 7.94 while effluent pH of

RPM wastewater dropped to 6.63 after anaerobic

digestion in MAIB-R. The pH level decreased

drastically to 6.21 in first compartment. However, the

pH level from compartment 2 to 5 increased smoothly

towards the end of the MAIB-R with pH of 6.39, 6.47,

6.59 and 6.60, respectively.

In the anaerobic digestion process, there are three

types of bacteria exist in anaerobic digestion process,

namely acid producing fermentation bacteria,

hydrogen-producing acetogenic bacteria and

methanogens. However, the pH demanded by the

three bacterial communities was different. The best

pH for the growth of methanogen ranged from 6.6 to

7.8. If the pH falls below 6, methanogenic bacteria

cannot survive. Hydrogen-producing acetogenic

bacteria and methanogen generate together. The pH

for hydrogen-producing acetogenic bacteria was

similar to that of methanogen, whose pH ranged from

6.0 to 8.0 (Guochen et al., 2009).

The pH level in the first compartment of MAIB-R

(acidification zone) has lower pH of 6.21. In the

acidogenesis process, acidogenic bacteria converts the

product of hydrolysis (amino acids, fatty acids, simple

sugars and glucose) into simple organic compounds,

mostly short chain acids, volatile acids (Metcalf &

Eddy, 2004). The transition of the high substrate

(carbon and energy source) level from organic

material to organic acids in the acid forming stages

causes the pH of the first compartment to drop

drastically and increase the growth rate of the

bacteria. This is beneficial for the acidogenic and

acetagenic bacteria that prefer a slightly acidic

environment and are less sensitive to changes in the

incoming feed stream.

The growth rate of scavenging bacteria becomes

slower towards the end of the reactor (methanogenic

zone) at higher pH (Hu and Li, 2008). In

methanogenesis, methane-producing bacteria,

methanogens are very sensitive towards changes and

prefer a neutral to slightly alkaline environment (Gas

Technology Inc, 2003). If the pH falls below 6,

Zwain and Dahlan


Bioreactor

83

methanogenic bacteria cannot survive. Thus, the

change in pH can be both an indicator to determine

the performance efficiency and the cause of process

imbalance in MAIB-R. At steady state, effluent pH of

RPM wastewater produced using MAIB-R was 6.63.

If compared with the regulations in EQA 1974 shown

in Table 10 and it was found to be complying and

within the limit of both Standard A (pH6.0-9.0) and

Standard B (pH5.5-9.0).

Fig. 5: pH profiles for each compartment of the MAIB bioreactor at steady state on day 15

3.3.2. COD Removal

Figure 6 shows the COD concentration in each

compartment at the steady state performance. The

removal efficiency of COD on day 15 was 92% when

the pH was recorded above 6. These results proved

that MAIB-R reactor has showed good bioreactor

performance. The COD concentration was further

decreased towards the end of the MAIB-R from

compartment 1 to 5. In compartment 1 of MAIB

bioreactor, the COD concentration was dropped

drastically to 328mg/L from initial 988mg/L of

influent COD concentration. As for compartment 2,

the COD concentration measured was 322 mg/L while

in compartment 3 was 316 mg/L. The COD

concentration for compartment 4 and 5 of MAIB-R

was 298 mg/L and 294 mg/L, respectively.

From this result clearly indicates that MAIB-R led to

a higher biological degradation of the organic matter

in RPM wastewater, and a better adaptation of the

biomass for the degradation of the substrates. Similar

observation was reported by Guochen et al (2009)

using ABR bioreactor, where the COD removal

efficiency up to 89%. Guochen et al. (2009) reported

that during stable stage, average COD removal rate of

ABR treating high-concentration sugar producing

wastewater was 85.5%. COD removal of ABR in

treating wheat flour starch industry wastewater was up

to 67% (Movahedyan et al., 2007).

At steady state, effluent COD concentration of MAIB-

R was 60mg/L. From Table 11, it showed that it

complied to EQA 1974 Standard B with COD limit of

100mg/L. However it failed to comply with Standard

A with COD limit of 50mg/L with a slight difference

by 10mg/L. Further studies need to be done for better

COD removal efficiency so that the effluent (RPM

wastewater) can be reached below 50mg/L.



84

Fig. 6: COD concentration in the MAIB bioreactor at steady state on day 15

4. CONCLUSIONS AND RECOMMENDATIONS

Recycled paper mill wastewater contained a

considerable high amount of COD with the range of

3380mg/L to 4930mg/L; BOD with the range of

1650mg/L to 2565mg/L. The BOD5/COD relationship

(0.488-0.52) showed that RPM wastewater was of

high strength wastewater. In addition, RPM

wastewater also contained relatively high total solids,

total suspended solids and total dissolved solids. The

start-up process of MAIB-R in this study has been

shortly achieved and the steady state was obtained on

day 15. The MAIB-R after successful start-up can

achieve relatively high average COD removal of 92%.

The pH showed a smooth increased from

compartment 1 towards the end of the MAIB-R,

compartment 5. Influent pH of RPM wastewater was

7.94 and dropped to 6.63 (effluent pH). On the other

hand, the COD concentration in each compartment

was slightly decreased from compartment 1 to

compartment 5. The removal efficiency of COD on

day 15 was 92%. Methane content in biogas generated

by MAIB-R was up to 57% with biogas yield of

0.77L/d on day 15. Although, this MAIB-R start-up

was successfully carried out, however further study is

needed to enhance the performance of the MAIB-R at

wide range of OLR, HRT and temperature. It is also

suggested that effort should be made to investigate the

microbial growth kinetic of the MAIB-R.

ACKNOWLEDGMENT

The authors wish to acknowledge the financial

support from the Universiti Sains Malaysia (RU-I

Grant A/C. 1001/PJKIMIA/814148).

REFERENCES

Akunna JC, Clark M (2000). Performance of a

granular-bed anaerobic baffled reactor

(GRABBR) treating whisky distillery

wastewater. Bioresource Technology, 74: 257-

261.

APHA (1995). Standard Methods for the Examination

of Water and Wastewater. American Public

Health Association, Washington, DC.

Baloch MI, Akunna JC, Collier PJ (2007). The

performance of a phase separated granular bed

bioreactor treating brewery wastewater.

Bioresource Technology, 98: 1849-1855.

Barber WP, Stuckey DC (1999). The use of the

anaerobic baffled reactor (ABR) for wastewater

treatment: a review. Water Research 33: 1559-

1578.

Bodkhe SY (2009). A modified anaerobic baffled

reactor for municipal wastewater treatment.

Journal of Environmental Management, 90:

2488-2493.

Buitrón G, Melgoza RM, Jiménez L (2003).

Pharmaceutical wastewater treatment using an

anaerobic-aerobic sequencing batch biofilm.

Journal of Environment Science Health 38:

2077-2088.

DOE (2009). Environmental Quality Act 1974 and

Regulations. Environmental Quality (Industrial

Effluent) Regulations (PU (A) 434), Fifth

Schedule, in: Environment, D.o. (Ed.),

Malaysia.

Enright AM, Collinc G, O'Flaherty V (2005). Low

temperature anaerobic biological treatment of

solvent containing pharmaceutical wastewater.

Water Research, 39: 4587-4596.

Zwain and Dahlan


Bioreactor

85

Faisal M, Unno H (2001). Kinetic analysis of palm oil

mill wastewater treatment by a modified

anaerobic baffled reactor. Biochemical


Feng H, Hu L, Mahmood Q, Qiu C, Fang C, Shen D

(2008). Anaerobic domestic wastewater

treatment with bamboo carrier anaerobic

baffled reactor. International Biodeterioration &

Biodegradation, 62: 232-238.

Foxon KM, Pillay S, Lalbahadur T, Rodda N, Holder

F, Buckley CA (2004). The anaerobic baffled

reactor (ABR): An appropriate technology for

on-site sanitation. Water SA, 30: 592-598.

Gas Technology Inc (2003). HIMET- A two stage

anaerobic digestion process for converting

waste to energy. Institute of Gas Technology,

Chicago IL.

Goel S (2010). Anaerobic baffled reactor for treatment

of textile dye effluent. Journal of Scientific and

Industrial Research, 69: 305-307.

Gopala Krishna GVT, Kumar P (2007). Complex

wastewater treatment using an anaerobic

baffled reactor. Environmental Progress, 26:

391-398.

Gopala Krishna GVT, Kumar P, Kumar P (2008).

Treatment of low strength complex wastewater

using an anaerobic baffled reactor (ABR).


Grover R, Marwaha SS, Kennedy JF (1999). Studies

on the use of an anaerobic baffled reactor for

the continuous anaerobic digestion of pulp and

paper mill black liquors. Process Biochemistry,

34: 653-657.

Guochen Z, Jianzheng L, Junguo H, Weiguang L,

Chao M (2009). The start-up and operational

performance of anaerobic baffled reactor

(ABR)treating high-concentration sugar-

producing wastewater. 3rd International

Conference on Bioinformatics and Biomedical

Engineering, Beijing.

Hassan SR, Zwain HM, Zaman NQ, Dahlan I (2014).

Recycled paper mill effluent treatment in a

modified anaerobic baffled reactor: start-up and

steady-state performance. Environmental


Hu XQ, Li ZH (2008). Operational characteristics of

an anaerobic baffled reactor treating low

strength wastewater. 3135-3139.

Kato MT, Field JA, Lettinga G (1997). The anaerobic

treatment of low strength wastewaters in UASB

and EGSB reactors. Water Science and

Technology, 36, 375-382.

Langenhoff AAM, Stuckey DC (2000). Treatment of

dilute wastewater using an anaerobic baffled

reactor: effect of low temperature. Water

Research, 34: 3867-3875.

Lili Z, Min J, Guozhong Z, Cuixia Y (2009).

Treatment of chemical synthesis-based

pharmaceutical wastewater with combined

ABR and MBR system.

Liu R, Tian Q, Chen J (2010). The developments of

anaerobic baffled reactor for wastewater

treatment: A review. African Journal of

Biotechnology, 9: 1535-1542.

Mahdi KCKSS (1986). Biogas technology: practical

handbook. Tata McGraw Hill, New Delhi.

Metcalf & Eddy I (2004). Wastewater Engineering

Treatment and Reuse, Fourth ed. McGraw Hill,

New York, p. 11.

Movahedyan H, Assadi A, Parvaresh A (2007).

Performance evaluation of an anaerobic baffled

reactor treating wheat flour starch industry

wastewater. Iranian Journal of Environmental

Health Science and Engineering, 4: 77-84.

Nègre J, Jonsson L (2010). Sludge treatment in an

Anaerobic BioReactor with external

Membranes. Stockholm Vatten AB, Stockholm,

Sweden.

Nyanchaga EN, Elkanzi MA (2002). Characteristic

Strength and Treatability of a Recycled Paper

Mill Wastewater in a UASB Reactor. Journal of

Civil Engineering, JKUAT 8: 61-77.

Pokhrel D, Viraraghavan T (2004). Treatment of pulp

and paper mill wastewater—a review. Science

of The Total Environment, 333: 37-58.

Roda JM, Rathi S (2006). Feeding China’s Expanding

Demand for Wood Pulp: A Diagnostic

Assessment of Plantation Development, Fiber

Supply, and Impacts on Natural Forests in

China and in the South East Asia Region,

Malaysia Report. Center for International

Forestry Research (CIFOR), Indonesia.

Saktaywin W, Tsuno H, Nagare H, Soyama T,

Weerapakkaroon J (2005). Advanced sewage

treatment process with excess sludge reduction

and phosphorus recovery. Water Research, 39:

902-910.

Samudro G, Mangkoedihardjo S (2010). REVIEW

ON BOD, COD AND BOD/COD RATIO: A

TRIANGLE ZONE FOR TOXIC,

BIODEGRADABLE AND STABLE LEVELS.

International Journal of Academic Research 2:

235-239.

Setiadi T, Husaini, Djajadiningrat A (1996). Palm oil

mill effluent treatment by anaerobic baffled

reactors: Recycle effects and biokinetic

parameters, pp. 59-66.

Shen YL, Guan W, Ye F, 2004. Treatment of low

strength wastewater by anaerobic baffled

reactor (ABR). Jiangsu Environ. Sci. Technol.,

38.



86

Sinclair WF (1991). Controlling effluent discharges

from Canadian pulp and paper manufacturers.

Canadian Public Policy/Analyse de Politiques

17, 86-105.

United Nations (2003). Economic and Social

Commission for Western Asia Wastewater

Treatment Technologies: A General Review,

United Nations, New York.

Wang J (2012). Fundamentals of Biological Processes

for Wastewater Treatment, Biological Sludge

Minimization and Biomaterials/Bioenergy

Recovery Technologies. John Wiley & Sons,

Inc., pp. 1-80.

Žarković DB, Todorović ŽN, Rajaković LV (2011).

Simple and cost-effective measures for the

improvement of paper mill effluent treatment –

A case study. Journal of Cleaner Production,

19: 764-774.

Zhu GF, Li JZ, Wu P, Jin HZ, Wang Z (2008). The

performance and phase separated characteristics

of an anaerobic baffled reactor treating soybean

protein processing wastewater. Bioresource

Technology, 99: 8027-8033.

Zoetemeyer RJ, Matthijsen AJCM, Cohen A,

Boelhouwer C (1982). Product inhibition in the

acid forming stage of the anaerobic digestion

process. Water Research, 16: 633-639.

Zwain HM, Hassan SR, Zaman NQ, Aziz HA, Dahlan

I (2013). The start-up performance of modified

anaerobic baffled reactor (MABR) for the

treatment of recycled paper mill wastewater.

Journal of Environmental Chemical

Engineering, 1: 61-64.



©2014 IJSRPUB

87

Augmentation of Biological Nitrogen Removal via Optimization of Support Media

Size and Aeration Strategy in Moving Bed Sequencing Batch Reactor

Jun-Wei Lim1, Mohammed J.K. Bashir

1,*, Si-Ling Ng

2, Sumathi Sethupathi

1, Lai-Peng Wong

1

1Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul

Rahman, 31900 Kampar, Perak, Malaysia 2School of Science, Monash University Sunway Campus, 46150 Petaling Jaya, Selangor, Malaysia


Abstract. As the discharge of wastewater containing nitrogen is in soaring trend, the objectives of this research aim to

optimize the size of support media and aeration strategy in moving bed sequencing batch reactor (MBSBR) for the biological

nitrogen removal enhancement. In accordance with these objectives, MBSBRs packed with polyurethane (PU) foam cubes of

sizes 8 to 125-mL as the support media and operated with continuous aeration (CA) strategy was initially experimented to

determine the optimum size of support media, reflected by the highest total nitrogen (TN) removal. Subsequently, the

operation of MBSBR packed with optimum size of support media was converted to intermittent aeration (IA) strategy and its

performance in terms of TN removal was investigated again. The results of these two steps experiments vindicated that the

MBSBR packed with 8-mL PU foam cube and operated with IA strategy removed the highest TN removal, signifying the

optimum condition. This optimum condition had eventually resulted the MBSBR procured the highest total weight of attached-

growth biomass that was associated to the maximum amount of internally stored carbon source in which favorable the

simultaneous nitrification and denitrification (SND) process.

Keywords: Nitrogen Removal, Sequencing Batch Reactor, Biological Method

1. INTRODUCTION

The increasingly discharge of nitrogen wastes from

industrial, livestock and agricultural wastewaters

without prior channeling to a proper treatment can

adversely contaminate the quality of receiving waters.

Serious concern related to the adverse impacts of

nitrogen wastes particularly ammonium-nitrogen

(NH4+-N) includes dissolved oxygen (DO) depletion,

toxicity and eutrophication (Gerardi, 2002; Magri et

al., 2013). Therefore, the necessity to remove NH4+-N

in wastewaters before the disposal is utmost crucial to

comply with the stringent discharged limits for

wastewaters containing nitrogen.

Biological nitrogen removal has been widely

recognized to offer a more economical and

environmentally friendly approach in removing NH4+-

N via nitrification and denitrification processes

(Gerardi, 2002; Jokela et al., 2002; Leta, 2004;

Dempsey et al., 2005; Aslan et al., 2009; Yao et al.,

2013). Over the decades, various operational

parameters had been studied with intention to

optimize the biological nitrogen removal (Deguchi

and Kawashiwaya, 1994; Pochana and Keller, 1999;

Katsogiannis et al., 2003; Daniel et al., 2009; Chu and

Wang, 2011).

1.1. Nitrification Process

The nitrification process is a two steps aerobic

process. The first step involves the oxidation of NH4+-

N to nitrite-nitrogen (NO2--N) which is subsequently

oxidized to nitrate-nitrogen (NO3--N) in the second

step mediated by Nitrosomonas sp. and Nitrobacter

sp., respectively (Kotlar et al., 1996; Qiao et al.,

2008). The oxidation of NH4+-N to NO2

--N is

conventionally known as nitritation process, whereas

the oxidation of NO2--N to NO3

--N is typically

identified as nitratation process (Zeng et al., 2009;

Fukumoto et al., 2011).

Nitritation process performed by Nitrosomonas sp.:

NH4+ + 1.5O2 → NO2

- + 2H

+ + H2O (1)

Nitratation process performed by Nitrobacter sp.:

NO2- + 0.5O2 → NO3

- (2)

Overall nitrification process:

NH4+ + 2O2 → NO3

- + H2O + 2H

+ (3)

The nitrifying bacteria of Nitrosomonas spp. and

Nitrobacter spp. are categorized as

chemolithoautotrophs, indicating that the nitrifying

bacteria obtain carbon from the inorganic carbon, i.e.,

carbon dioxide, and energy from the chemical

reactions (Eqs. 1 and 2, respectively) for growth and

multiplication. Carbon dioxide is accessible to

nitrifying bacteria in the form of bicarbonate in the

wastewaters (Gerardi, 2002; Liao et al., 2008). As

more energy can be reaped from the oxidation of

NH4+-N to NO2

--N than NO2

--N to NO3

--N, the

population size of Nitrosomonas spp. is always greater

than Nitrobacter spp. Besides, Nitrosomonas spp. also

has a shorter generation time which enables its

numbers to faster than Nitrobacter spp. (Gerardi,

2002). Therefore, in the activated sludge system, the


Lim et al.

Augmentation of Biological Nitrogen Removal via Optimization of Support Media Size and Aeration Strategy in

Moving Bed Sequencing Batch Reactor

88

ability of being oxidized for NH4+-N is always higher

than for NO2--N, causing a transient accumulation of

NO2--N which is removed when NH4

+-N is completely

oxidized. However, during unfavourable operational

conditions such as low DO concentration, excessive

sludge washout, extreme change in temperature, etc.,

significant accumulation of NO2--N concentration can

be detected in the mixed liquor of the bioreactors

(Ruiz et al., 2006; Jubany et al., 2009; Lim et al.,

2010; Gu et al., 2012).

Although the activated sludge system is commonly

used to treat wastewaters containing NH4+-N in many

treatment plants, it is actually not an ideal system for

the nitrification process. As demonstrated by Gerardi

(2002), the typical population size of nitrifying

bacteria in the activated sludge system is merely in the

range of 3% to 10% with the remaining consisting of

organotrophs. This is due to the nitrifying bacteria has

a very low reproductive rate as large amounts of

NH4+-N and NO2

--N are required during the

assimilation process.

1.2. Denitrification Process

The sole nitrification process is only partially

fulfilling the objective of biological nitrogen removal

as the NH4+-N is merely bio-transformed to the

oxidized nitrogen, viz., NO3--N and NO2

--N, by the

nitrifying bacteria. These oxidized nitrogen species

must be later reduced to nitrogen gas (N2) and

released to the atmosphere in order to remove nitrogen

from wastewaters. The reduction process is known as

denitrification process which reduces the oxidized

nitrogen to N2 with the sequence shown hereafter

(Gerardi, 2002):

Nitrate ion (NO3-) → Nitrite ion (NO2

-) → Nitric

oxide (NO) → Nitrous oxide (N2O) → Ntrogen gas

(N2)

The process of denitrification can be achieved by a

broad range of facultative anaerobes which make up

an approximately 80% of the bacteria in the activated

sludge system. The largest numbers of facultative

anaerobes that can perform the denitrification process

are from genera of Alcaligenes, Bacillus and

Pseudomonas (Gerardi, 2002). More so, the

facultative anaerobes that denitrify are also termed by

several names including denitrifying bacteria,

organotrophs, denitrifiers and heterotrophs.

In assuring the feasibility of denitrification

process, the carbon source must be made available to

the denitrifying bacteria under an anoxic environment

with DO concentrations of lower than 1.0 mg/L or 2%

saturation in the mixed liquor of the bioreactors (Goh,

2007; Lim et al., 2012). As the denitrifying bacteria

can reap more energy through the aerobic respiration,

the denitrification process has to be strictly

commenced in the anoxic environment to enable the

added carbon source is primarily utilized to reduce the

oxidized nitrogen instead of being aerobically

degraded (Gerardi, 2002). According to Gerardi

(2002), the reduction of NO3- ion in the presence of

carbon source under the anoxic environment can be

expressed in two simplified biochemical reactions

shown hereafter:

NO3- + Carbon Source → NO2

- + CO2 + H2O (4)

NO2- + Carbon Source → N2 + CO2 + OH

- (5)

1.3. Moving Bed Sequencing Batch Reactor -

Support Media Size and Aeration Strategy

Sequencing Batch Reactor (SBR) system has been

extensively manipulated for the treatment of domestic,

municipal and industrial wastewaters and also offers

an attractive option in the biological wastewater

treatment systems with the advantages as below

(Irvine and Ketchum, 1989; Wobus et al., 1995;

Louzeiro et al., 2002; Goh, 2007):

(a) Resistant to influent loading fluctuation;

(b) Simplicity and cost effective;

(c) Flexibility in terms of sequence and cycle time;

(d) Ideal quiescent settling condition;

(e) Ability of combining aerobic and anoxic phases in

a single reactor.

Incorporated with an activated sludge system, SBR

is primarily running on the fill and draw principle.

The operation of the SBR is usually achieved by five

sequential periods, namely FILL, REACT, SETTLE,

DRAW and IDLE in a single tank instead of

simultaneously in the separate tanks as typically

observed in the conventional activated sludge system

(Arora et al., 1985). Thus, permits the construction of

smaller treatment plant which is suitable in the area

with limited land availability. Nevertheless, a

conventional SBR system is generally incapable of

executing an effective biological nitrogen removal

(Goh, 2007; Li et al., 2011). As a result, an

improvement of SBR system is deemed essential

when the bio-treatment of wastewater containing

nitrogen is to be addressed.

For years, numerous attempts have been

committed to modify the SBR system in order to

enhance the performance. Among others, the moving

bed sequencing batch reactor (MBSBR) which is

developed by retrofitting support media into the SBR

has attracted much interest among the researchers in

the field of biological wastewater treatment (Garzon-

Zuniga and Gonzalez-Martinez, 1996;

Sirianuntapiboon and Yommee, 2006; Goh et al.,

2009; Hosseini Koupaie et al., 2012). In general, three



89

categories of support media, namely Kaldnes

(polyethylene media), Liapor (ceramic media) and

Linpor (plastic media with high porosity) are used to

pack the MBSBR (Valdivia et al., 2007). The mobility

of the support media in MBSBR is usually conserved

by the aeration or mechanical mixing systems

(Odegaard et al., 1994). Once the attached-growth

biomass (biofilm) has grown on the surfaces of

support media, the MBSBR which contains both

suspended-growth and attached-growth biomasses are

anticipated to outperform the SBR which consists of

only suspended-growth biomass.

Among of these three types of support media used,

Valdivia et al. (2007) reported that the MBSBR

packed with Linpor performed better than either

Kaldnes or Liapor in removing chemical oxygen

demand (COD) and NH4+-N at the organic loading

rate higher than 3.0 g COD/m3 day. The polyurethane

(PU) foam which belongs to the Linpor type of

support media has a high porosity and is an ideal

medium for biomass immobilization through

attachment. In addition, PU foam also has a good

mechanical strength and is relatively low cost (Golla

et al., 1994; Chu and Wang, 2011).

Nitrogen removal via simultaneous nitrification

and denitrification (SND) process had been

investigated in previous studies (Kotlar et al., 1996;

Daniel et al., 2009; Chu and Wang, 2011). The SND

process is potentially reducing the operational period

which ultimately results in the reduction of operation

cost. The oxygen concentration and the availability of

the carbon source for denitrification were found to be

the important parameters for the commencement of

SND process (Pochana and Keller, 1999). By using

the PU foam cubes as a support media for biofilm

formation, studies had demonstrated the occurrence of

decreasing dissolved oxygen (DO) gradient within the

inner layer of biofilm and also deep inside the PU

foam which leads to the creation of an anoxic zone

(Morper, 1994; Guo et al., 2010). Furthermore, as the

carbon source is also an important constituent to

ensure the viability of the denitrification process, the

SND process is frequently inhibited in treating

wastewaters containing low COD/N ratio.

Nevertheless, in the study of Guo et al. (2010), the

SND process was still detected in the anoxic zone of

the PU foam. The occurrence of the SND process

without the addition of the external carbon source in

their study could be explained by the possibility of

carbon storage in the deeper biofilm layers which was

also reported by several researchers (Morgenroth and

Wilderer, 1999; Pastorelli et al., 1999; Gieseke et al.,

2002). The presence of carbon storage together with

anoxic zone within the biofilm and deep inside the PU

foam could be used to kindle the denitrification

process which eventually stimulated the SND process.

Based on the results of previous studies, PU foam

was found to be a good support media in the MBSBR

system. Its porous structure will allow the formation

of attached-growth biomass, the establishment of

anoxic zone and the storage of carbon which enhance

the nitrogen removal via SND process. However, to

date, relatively little is documented on the effect of

different sizes of the PU foam cubes on the nitrogen

removal in the MBSBR system. In addition, the

biomass growth onto the surfaces and into the interior

porous structure of the various sizes of PU foam cubes

along with the role of these various sizes of PU foam

cubes in the SND process in treating low COD/N ratio

wastewaters have not been thoroughly investigated.

The main advantage of intermittent aeration (IA)

strategy application in treating wastewater containing

nitrogen includes the reduction of the operational cost

due to a decrease of continuous aeration (CA) supply

and the amount of external carbon source required for

the subsequent denitrification phase (Li et al., 2008).

In the IA system, the high DO concentration during

the aeration period enables the aerobic nitrifiers to

oxidize NH4+-N to NO2

--N and then to NO3

--N.

During the subsequent non-aeration period, the DO

concentration will decrease to such a low level that

the NO2--N and NO3

--N species are reduced to N2

mediated by the denitrifiers (Li et al., 2008).

Nonetheless, the feasibility of IA strategy

application in removing nitrogen from low COD/N

ratio wastewaters is still questionable due to

inadequate supply of carbon source which results in

an incompetency of denitrification process. In view of

the carbon storage in the deeper layers of biofilm

within the support media which can serve as the

carbon source, the integration of IA strategy in

MBSBR (IA-MBSBR) packed with PU foam cubes is

not expected to face the aforesaid problem. The

carbon storage is not only found in the biofilm on the

external surfaces which is typically exhibited by most

of the support media but is also located within the

interior porous structure of the PU foam cubes which

is being occupied by the attached-growth biomass.

However, to date, the available information to sustain

the viability of IA-MBSBR in the enhancement of

nitrogen removal is still lacking. The availability of

the carbon storage located onto and into the

acclimated PU foam cubes with attached-growth

biomass also needed to be further substantiated.

In light of relatively little information on the

potential of MBSBR in the removal of nitrogen, the

objectives of this study are to optimize the size of

support media, i.e., PU foam cubes, and aeration

strategy in MBSBR system for biological nitrogen

removal via SND process.

Lim et al.



90

2. METHODOLOGY

To achieve the aforementioned objectives, two major

experiments were carried out with the first experiment

aimed to determine the optimum size of support media

used to pack the MBSBR and the subsequent

experiment targeted to investigate the optimum

aeration strategy; both emphasized on the

enhancement of biological nitrogen removal via SND

process.

2.1. Determination of Optimum Support Media

Size Packed in MBSBR

Five plexiglass reactors were inoculated with the

activated sludge collected from municipal sewage

treatment plant and fed with synthetic wastewater

simulating the municipal wastewater composition

with NH4+-N and COD concentrations of

approximately 48 and 200 mg/L, respectively. The

reactors were operated with a cycle time of 24 h in the

following sequencing periods: instantaneous FILL, 0

h; aerobic REACT, 10 h; anoxic REACT, 2 h;

SETTLE, 1.5 h; DRAW, 1 h and IDLE, 9.5 h.

Adequate amount of ethanol solution was

instantaneously added into the reactor as a carbon

source at the beginning of the anoxic REACT period

to reduce the oxidized nitrogen to N2. At the end of

the REACT period, mixed liquor was wasted to

maintain the sludge age of suspended-growth biomass

at 40 days. During the DRAW period, the supernatant

or treated effluent was drawn out with an exchange

volume of the reactor being retained at 70.3%. The

residual settled solids in the reactor were left to rest

throughout the IDLE period in preparation for the next

cycle.

After achieving the quasi-steady state, four rectors

were transformed to the MBSBR by introducing PU

foam cubes into its mixed liquor for the development

of attached-growth biomass. The packing volume of

the PU foam cubes in each MBSBR was fixed at 8%

(v/v) of the working volume. The performance of

MBSBRs were investigated using four different sizes

of PU foam cubes, namely 8-, 27-, 64- and 125-mL

cubes which were labeled as R2 - R5, respectively.

One of the reactors was continuously operated without

the addition of support media to serve as the SBR

system (R1) for comparison. The schematic diagram

of the MBSBR is shown in Figure 1. The

performances of all reactors were monitored in terms

of nitrogen removal consisting of NH4+-N, NO2

--N

and NO3--N species.

Fig. 1: Schematic diagram of the MBSBR: (a) plexiglass reactor; (b) aquarium air pumps; (c) addition of feed solution by the

peristaltic pump; (d) addition of ethanol solution by the peristaltic pump; (e) draw out tubing to remove treated effluent

containing unsettled suspended solids by the peristaltic pump; (f) support media of PU foam cubes; (g) ejector; (h) air

diffusers; (i) aeration tubing; (j) working volume of 12.8 L; (k) residual volume of 3.8 L.



91

2.2. Determination of Optimum Aeration Strategy

in MBSBR

The selected MBSBR with optimum size of PU foam

cubes from the experiment as outlined in Section 2.1

was then converted to IA-MBSBR in which the

REACT period was converted to begin with 1 h of

aeration period followed by 1 h of non-aeration period

and the pattern was repeated throughout the 12 h

REACT period. The instantaneous addition of

adequate amount of ethanol solution to serve as a

carbon source was only implemented at the beginning

of the last non-aeration period in every cycle to reduce

the oxidized nitrogen to N2. The performance of IA-

MBSBR in terms of nitrogen removal was compared

with its former stage of aeration strategy, i.e., CA

strategy.


The comparative study of MBSBRs packed with

various sizes of PU foam cubes each and with SBR

was initially discussed in this section. Optimum size

of PU foam cubes used to pack the MBSBR was

culled based on the highest nitrogen removal

efficiency via SND process. Subsequently, the

MBSBR packed with optimum size of PU foam cubes

was operated with IA strategy during the REACT

period and the results obtained were compared with its

former stage, i.e., CA strategy, to determine the

optimum aeration strategy for the enhancement of

biological nitrogen removal.

The image of Scanning Electron Microscope of

support media, i.e., PU foam cube, added into the

MBSBR is presented in Figure 2. There are many

pores on and into the surfaces of PU foam cubes; thus,

permitting the growth and attachment of biomass.

Figure 2: SEM image of PU foam cube with 40X magnification and captured at WD of 7 mm using EHT and

signal of 5.00 kV and SE2, respectively.

3.1. Optimization of Support Media Size Packed in

MBSBR

The concentration profiles of nitrogen species during

the REACT period for SBR and MBSBRs are shown

in Fig. 3. For SBR, a small NO2--N concentration peak

was observed during the initial stage of the aerobic

REACT period and a constant NO3--N concentration

was attained when the nitrification process was

completed. In contrast, for MBSBRs, the formation of

NO2--N was barely observed in the aerobic REACT

period. Figure 3 shows that the NO3--N concentrations

in R2 to R5 decreased gradually after the completion

of nitrification, with the most obvious reduction seen

in R2. The percentages of total nitrogen (TN) removal

during the aerobic REACT period for R1 to R5 were

15%, 37%, 31%, 24% and 19%, respectively. If only

nitrification process was involved, the TN should

remain constant during the aerobic REACT period.

Instead, an increasing percentage of TN removal was

observed in the order of R1 < R5 < R4 < R3 < R2.

The removal of nitrogen can be plausibly

explained by the assimilation and SND processes

(Valdivia et al., 2007). Several studies have reported

the aeration condition and carbon source as the

parameters affecting the yield of the SND process

(Pochana and Keller, 1999; Third et al., 2003; Walters

et al., 2009; Khor et al., 2011). The percentage of TN

removal was found to be the lowest in R1. The

occurrence of SND process was not favored in R1 due

to the high DO and low COD concentrations in the

mixed liquor (Pochana and Keller, 1999; Chu and

Wang, 2011; Seifi and Fazaelipoor, 2012). Therefore,

the TN removal in R1 was deemed to be caused by

only assimilation process in which the lost TN was

used to build cell mass. In MBSBRs, aerobic

condition was kept on the surface of the acclimated

PU foam cubes whereas, distinctive DO gradient

Lim et al.



92

occurred along the acclimated PU foam cube’s inward

depth (Uemura et al., 2002; Guo et al., 2010).

Moreover, the carbon source required for

denitrification process could be obtained from the

storage inside the deep layer of acclimated PU foam

cubes (Pastorelli et al., 1999; Gieseke et al., 2002).

According to Gray (2004), the denitrification process

is dependent on the size of microorganism population.

The total weight of attached-growth biomass of all

MBSBRs were measured and found to be 49, 35, 33

and 30 g for R2 to R5, respectively, which was in

accordance with the decreasing trend of TN removal

from R2 to R5 as abovementioned. The lower weight

of attached growth biomass measured in larger PU

foam cubes was due to the failure of biomass to

diffuse into the internal surface of cubes to form

attachment viz-à-viz the smaller cubes as shown by

the SEM images in Figure 4. Hence, among the

MBSBRs, R2 packed with 8-mL acclimated PU foam

cubes was the most optimum reactor due to the

highest achievable of TN removal.

Fig. 3: Profiles of nitrogen species concentrations during the REACT period for (a) R1, (b) R2, (c) R3, (d) R4 and (e) R5.

Fig. 4: SEM images of (a) internal surface of acclimated 8-mL PU foam cube, (b) internal surface of acclimated 27-mL PU

foam cube, (c) internal surface of acclimated 64-mL PU foam cube and (d) internal surface of acclimated 125-mL PU foam

cube all with 40X magnification and captured at WD of 7 mm using EHT and signal of 5.00 kV and SE2, respectively. The

internal surface of each size of the PU foam cube is referring to the surface of the cube cross sectional area.

3.2. Optimization of Aeration Strategy in MBSBR

The study in Section 3.1 had underscored the

advantages of using 8-mL PU foam cubes as the

support media in MBSBR for the removal of nitrogen

from low COD/N ratio wastewater as opposed to

larger PU foam cubes. The highest growth of

attached-growth biomass was achieved in the MBSBR

packed with this size of PU foam cubes which led to

the highest TN removal mainly via SND process.

However, the application of continuous aeration (CA)

strategy in MBSBR (CA-MBSBR) during the 10 h of

(a) (b) (c) (d)



93

aerobic REACT period increases the operation cost

especially when the nitrification process has

completed. Besides, the CA strategy would also

deplete the carbon storage in the acclimated PU foam

cubes which plausibly led to the slowing down of

SND process. Thus, the potential of integrating IA

strategy in MBSBR (IA-MBSBR) packed with 8-mL

PU foam cubes as the support media for nitrogen

removal enhancement was investigated.

Figure 5 shows that the NH4+-N was completely

removed only at the third aeration period. During the

first and second non-aeration periods, the oxidation of

NH4+-N was impeded due to the exhausted supply of

DO in the mixed liquor, resulting in a relatively flat

region of the NH4+-N concentration profile. The NO3

--

N concentration, however, decreased gradually during

the non-aeration period due to the denitrification

process. Figure 5 also shows that the NH4+-N was

swiftly oxidized to NO3--N during the nitrification

process with negligible accumulation of NO2--N.

Based on Figure 5, the calculated percentage of TN

removal in IA-MBSBR during the REACT period

prior to the addition of ethanol solution was 57%. The

TN removal could be explained by the occurrence of

the SND process within the anoxic zone in the

acclimated PU foam cubes. The percentage of TN

removal in the IA-MBSBR was also found to be

higher than the CA-MBSRB of R2, which was

reported to be only 37%. The better performance of

IA-MBSBR was plausibly due to the presence of

higher total weight of attached-growth biomass in the

IA-MBSBR (57 g) than in the CA-MBSRB of R2 (49

g) as the quantity of stored carbon source which is

used to facilitate the denitrification process, increases

with increasing amount of attached-growth biomass.

Therefore, the integration of IA strategy in MBSBR

was considered optimum since higher percentage of

TN removal could be attained as compare with the

CA-MBSBR. Additionally, IA strategy was also more

economical since the overall aeration period of IA-

MBSBR had reduced to almost half of the CA-

MBSBR.

4. CONCLUSIONS

The highest TN removal could be noticed when the

CA-MBSBR was packed with 8-mL PU foam cubes

as the support media as compare with other sizes of

cubes. The percentage of TN removal was later

improved when the IA strategy was retrofitted into the

REACT period of MBSBR packed with similar size of

support media, signifying the optimum condition. The

dominance of this MBSBR in removing nitrogen was

due to the presence of the highest amount of attached-

growth biomass that could store carbon source, a

crucial substance to promote the SND process.

Fig. 5: Profiles of nitrogen species concentrations during the REACT period for IA-MBSBR.

REFERENCES

Arora ML, Barth EF, Umphres MB (1985).

Technology Evaluation of Sequencing Batch

Reactors. Journal (Water Pollution Control

Federation), 57: 867-875.

Aslan S, Miller L, Dahab M (2009). Ammonium

Oxidation via Nitrite Accumulation under

Limited Oxygen Concentration in Sequencing

Batch Reactors. Bioresource Technology, 100:

659-664.

Chu L, Wang J (2011). Comparison of Polyurethane

Foam and Biodegradable Polymer as Carriers in

Moving Bed Biofilm Reactor for Treating

Wastewater with a Low C/N Ratio.

Chemosphere, 83: 63-68.

Daniel LMC, Pozzi E, Foresti E, Chinalia FA (2009).

Removal of Ammonium via Simultaneous

Nitrification-denitrification Nitrite-shortcut in a

Single Packed-bed Batch Reactor. Bioresource

Technology, 100: 1100-1107.

Lim et al.



94

Deguchi H, Kawashiwaya M (1994). Study on

Nitrified Liquor Recycling Process Operations

Using Polyurethane Foam Sponge Cubes as a

Biomass Support Medium. Water Science and


Dempsey MJ, Lannigan KC, Minall RJ (2005).

Particulate-biofilm, Expanded-bed Technology

for High-rate, Low-cost Wastewater Treatment:

Nitrification. Water Research, 39: 965-974.

Fukumoto Y, Suzuki K, Kuroda K, Waki M, Yasuda

T (2011). Effects of Struvite Formation and

Nitratation Promotion on Nitrogenous

Emissions such as NH3, N2O and NO during

Swine Manure Composting. Bioresource

Technology, 102: 1468-1474.

Garzon-Zuniga MA, Gonzalez-Martinez S (1996).

Biological Phosphate and Nitrogen Removal in

a Biofilm Sequencing Batch Reactor. Water

Science and Technology, 34: 293-301.

Gerardi MH (2002). Nitrification and Denitrification

in the Activated Sludge Process, John Wiley

and Sons, Inc., New York.

Gieseke A, Arnz P, Amann R, Schramm A (2002).

Simultaneous P and N Removal in a

Sequencing Batch Biofilm Reactor: Insights

from Reactor- and Microscale Investigations.


Goh CP (2007). Comparative Study of Sequencing

Batch Biofilm Reactor and Sequencing Batch

Reactor in p-Nitrophenol Removal. Master

Dissertation from the School of Chemical

Sciences, Universiti Sains Malaysia, Malaysia.

Goh CP, Seng CE, Ahmad Sujari AN, Lim PE (2009).

Performance of Sequencing Batch Biofilm and

Sequencing Batch Reactors in Simultaneous p-

Nitrophenol and Nitrogen Removal.

Environmental Technology, 30: 725-736.

Golla PS, Reddy MP, Simms MK, Laken TJ (1994).

Three Years of Full-scale CAPTOR Process

Operation at Moundsville WWTP. Water


Gray NF (2004). Biology of Wastewater Treatment.

2nd ed., Imperial College Press, London.

Gu S, Wang S, Yang Q, Yang P, Peng Y (2012). Start

Up Partial Nitrification at Low Temperature

with a Real-time Control Strategy Based On

Blower Frequency and pH. Bioresource


Guo W, Ngo HH, Dharmawan F, Palmer CG (2010).

Roles of Polyurethane Foam in Aerobic

Moving and Fixed Bed Bioreactors.


Hosseini Koupaie E, Alavi Moghaddam MR, Hashemi

SH (2012). Investigation of Decolorization

Kinetics and Biodegradation of Azo Dye Acid

Red 18 Using Sequential Process of Anaerobic

Sequencing Batch Reactor/Moving Bed

Sequencing Batch Biofilm Reactor.

International Biodeterioration and

Biodegradation, 71: 43-49.

Irvine RL, Ketchum LH Jr (1989). Sequencing Batch

Reactors for Biological Wastewater Treatment.

CRC Critical Reviews in Environmental

Control, 18: 255-294.

Jokela JPY, Kettunen RH, Sormunen KM, Rintala JA

(2002). Biological Nitrogen Removal from

Municipal Landfill Leachate: Low-cost

Nitrification in Biofilters and Laboratory Scale

In-situ Denitrification. Water Research, 36:

4079-4087.

Jubany I, Lafuente J, Baeza JA, Carrera J (2009).

Total and Stable Washout of Nitrite Oxidizing

Bacteria from a Nitrifying Continuous

Activated Sludge System Using Automatic

Control Based On Oxygen Uptake Rate

Measurements. Water Research, 43: 2761-2772.

Katsogiannis AN, Kornaros M, Lyberatos G (2003).

Enhanced Nitrogen Removal in SBRs

Bypassing nitrate Generation Accomplished by

Multiple Aerobic/Anoxic Phase Pairs. Water


Khor SM, Ng SL, Lim PE, Seng CE (2011). The

Effects of Nickel(II) and Chromium(VI) on

Oxygen Demand, Nitrogen and Metal Removal

in a Sequencing Batch Reactor. Environmental

Technology, 32: 1903-1914.

Kotlar E, Tartakovsky B, Argaman Y, Sheintuch M

(1996). The Nature of Interaction between

Immobilized Nitrification and Denitrification

Bacteria. Journal of Biotechnology, 51: 251-

258.

Leta S (2004). Developing and Optimizing Processes

for Biological Nitrogen Removal from Tannery

Wastewaters in Ethiopia. Doctoral Dissertation

from the Department of Biotechnology, Royal

Institute of Technology, Stockholm, Sweden.

Li J, Elliott D, Nielsen M, Healy MG, Zhan X (2011).

Long-term Partial Nitrification in an

Intermittently Aerated Sequencing Batch

Reactor (SBR) Treating Ammonium-rich

Wastewater under Controlled Oxygen-limited

Conditions. Biochemical Engineering Journal,

55: 215-222.

Li JP, Healy MG, Zhan XM, Rodgers M (2008).

Nutrient Removal from Slaughterhouse

Wastewater in an Intermittently Aerated

Sequencing Batch Reactor. Bioresource

Technology, 99: 7644-7650.

Liao D, Li X, Yang Q, Zeng G, Guo L, Yue X (2008).

Effect of Inorganic Carbon on Anaerobic

Ammonium Oxidation Enriched in Sequencing



95

Batch Reactor. Journal of Environmental

Science, 20: 940-944.

Lim JW, Ng SL, Khor SM, Seng CE (2012).

Inhibitory Effect of 2,4-Dichlorophenol on

Nitrogen Removal in a Sequencing Batch

Reactor. Korean Journal of Chemical


Lim PE, Lim SP, Seng CE, Md Noor A (2010).

Treatment of Landfill Leachate in Sequencing

Batch Reactor Supplemented with Activated

Rice Husk as Adsorbent. Chemical Engineering

Journal, 159: 123-128.

Louzeiro NR, Mavinic DS, Oldham WK, Meisen A,

Gardner IS (2002). Methanol-induced

Biological Nutrient Removal Kinetics in a Full-

scale Sequencing Batch Reactor. Water

Research, 36: 2721-2732.

Magri A, Beline F, Dabert P (2013). Feasibility and

Interest of the Anammox Process as Treatment

Alternative for Anaerobic Digester

Supernatants in Manure Processing - An

Overview. Journal of Environmental

Management, 131: 170-184

Morgenroth E, Wilderer PA (1999). Controlled

Biomass Removal - The Key Parameter to

Achieve Enhanced Biological Phosphorus

Removal in Biofilm Systems. Water Science

and Technology, 39: 33-40.

Morper MR (1994). Upgrading of Activated Sludge

System for Nitrogen Removal by Application

of the Linpor-CN Process. Water Science and


Odegaard H, Rusten B, Westrum T (1994). A New

Moving Bed Biofilm Reactor - Applications

and Results. Water Science and Technology,

29: 157-165.

Pastorelli G, Canziani R, Pedrazzi L, Rozzi A (1999).

Phosphorus and Nitrogen Removal in Moving-

bed Sequencing Batch Biofilm Reactors. Water


Pochana K, Keller J (1999). Study of Factors

Affecting Simultaneous Nitrification and

Denitrification (SND). Water Science and


Qiao S, Kawakubo Y, Koyama T, Furukawa K

(2008). Partial Nitritation of Raw Anaerobic

Sludge Digester Liquor by Swim-bed and

Swim-bed Activated Sludge Processes and

Comparison of Their Sludge Characteristics.

Journal of Bioscience and Bioengineering, 106:

433-441.

Ruiz G, Jeison D, Rubilar O, Ciudad G, Chamy R

(2006). Nitrification-denitrification via Nitrite

Accumulation for Nitrogen Removal from

Wastewaters. Bioresource Technology, 97: 330-

335.

Seifi M, Fazaelipoor MH (2012). Modeling

Simultaneous Nitrification and Denitrification

(SND) in a Fluidized Bed Biofilm Reactor.

Applied Mathematical Modelling, 36: 5603-

5613.

Sirianuntapiboon S, Yommee S (2006). Application of

a New Type of Moving Bio-film in Aerobic

Sequencing Batch Reactor (Aerobic-SBR).

Journal of Environmental Management, 78:

149-156.

Third KA, Burnett N, Ruwisch RC (2003).

Simultaneous Nitrification and Denitrification

Using Stored Substrate (PHB) as the Electron

Donor in an SBR. Biotechnology and

Bioengineering, 83: 706-720.

Uemura S, Takahashi K, Takaishi A, Machdar I,

Ohashi A, Harada H (2002). Removal of

Indigenous Coliphages and Fecal Coliforms by

a Novel Sewage Treatment System Consisting

of UASB and DHS Units. Water Science and


Valdivia A, Gonzalez-Martinez S, Wilderer PA

(2007). Biological Nitrogen Removal with

Three Different SBBR. Water Science and


Walters E, Hille A, He M, Ochmann C, Horn H

(2009). Simultaneous

Nitrification/Denitrification in a Biofilm Airlift

Suspension (BAS) Reactor with Biodegradable

Carrier Material. Water Research, 43: 4461-

4468.

Wobus A, Ulrich S, Roske I (1995). Degradation of

Chlorophenols by Biofilms on Semi-permeable

Membranes in Two Types of Fixed Bed

Reactors. Water Science and Technology, 32:

205-212.

Yao S, Ni J, Ma T, Li C (2013). Heterotrophic

Nitrification and Aerobic Denitrification at Low

Temperature by a Newly Isolated Bacterium,

Acinetobacter sp. HA2

Zeng W, Zhang Y, Li L, Peng Y, Wang S (2009).

Control and Optimization of Nitrifying

Communities for Nitritation from Domestic

Wastewater at Room Temperatures. Enzyme

and Microbial Technology, 45: 226-232.



©2014 IJSRPUB

96


Treatment Systems

Chapter 4: Electrochemical Methods



©2014 IJSRPUB

97

Electrochemical Oxidation Process Contribution in Remediating Complicated

Wastewaters

Mohammed J. K. Bashir1,*

, Jun-Wei Lim1, Shuokr Qarani Aziz

2, Salem S. Abu Amr

3

1,*Department of Environmental Engineering, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman,

31900 Kampar, Perak, Malaysia 3Department of Civil Engineering, College of Engineering, University of Salahaddin–Erbil, Iraq

2School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia,14300 Nibong Tebal, Penang, Malaysia


Abstract. In recent years, electrochemical oxidation process has gained increasing interest due to its exceptional technical

features to eliminate a wide range of pollutants exist in various types of wastewaters, e.g., refractory organic matter, nitrogen

species, microorganisms, etc. Serve as a clean, adaptable and powerful tool in removing pollutants, this review paper focuses

on the fundamental mechanisms of electrochemical oxidation process and provides discussions on the possible applications in

wastewater treatment. To top it off, special attention on the most recent developments and challenges are as well highlighted in

this review.

Keywords: Electrochemical, Wastewater, Oxidation Process

1. INTRODUCTION

Basically, wastewater treatment aims to improve the

quality of wastewater before discharging to the

receiving water bodies by using reliable technology.

The conventional sequence of wastewater treatment

starts with draining the wastewater in a central,

separated location and subjecting the wastewater to

several treatment processes. Wastewater treatment can

be generally categorized by the character of the

treatment process operation being used such as

biological, chemical or physical methods. Wastewater

treatment via biological technology is the most

economical means of treatment and normally utilizes

for the removal of biodegradable organic pollutants

presented in the wastewater. Nevertheless, the

presence of toxic and refractory substrates in the

wastewater would virtually foil the biological

treatment process as these substrates are potentially

inhibiting the bioactivity of microorganism (Grimm et

al., 1998). Among the various techniques, the use of

electro-chemical oxidation process in the wastewater

treatment has engrossed many researchers attention,

particularly in remediating industrial wastewater. To

date, electrochemical oxidation processes have been

shown to be a valuable option for the elimination of

refractory organic compounds from various types of

wastewaters (Bashir et al., 2013). Electrochemical

oxidation is highly capable and efficient in reducing

the organic compounds from various types of

wastewater as compared with other types of physio-

chemical technologies which only bring about phase

transfer of the contaminants in question with no

chemical destruction is taking place.

Similarly, Kapalka et al. (2009) stated that the

electrochemical oxidation process is a clean, versatile

and powerful tool for the destruction of organic

pollutants in wastewater. Furthermore,

electrochemical method presents many significant

gains since it does not require any ancillary chemical,

appropriate for large range of pollutants removal and

does not require high pressures and temperatures for

the reaction to commence. However, the efficiency of

the electro-oxidation techniques depends strongly on

the operation conditions and on the nature of the

electrode materials (Wang et al., 2008). Recently, the

strict wastewater discharge limits with health quality

standards obligation set by legislation may be met by

applying electrochemical oxidation. Wastewaters

generated from municipal landfill and a wide diversity

of industries including the food, textile, and tannery

productions have been successfully treated by this

process. Thus, due to its high competence together

with its disinfection capabilities, electro-oxidation is a

suitable technique for water reuse programs. On the

other hand, treatment costs have to be cut down prior

to full-scale application of this technology.

Accordingly, the employment of electrochemical

oxidation together with other technologies and the use

of renewable energy sources to operate this process

are two significant steps required to reduce the overall

operational cost (Anglada et al., 2009).

2. ELECTROCHEMICAL OXIDATION

PROCESS

Electrochemical oxidation process has been

recognized as one of the most effective techniques in

degrading pollutants present in textile wastewater,

landfill leachate, simulated wastewater, olive mill

wastewater, paper mill effluents, and industrial paint

wastewater (Körbahti and Tanyolaç 2003; Un et al.,

2008; Bashir et al., 2009 ). The electrochemical

reactor in the laboratory experiments is shown in


Bashir et al.

Electrochemical Oxidation Process Contribution in Remediating Complicated Wastewaters

98

Figure 1. Figure 2 shows the conceptual diagram of

electrochemical reactor for wastewater treatment,

which includes a direct current (DC) power supply, a

cathode, an anode, and the electrolyte (a medium that

provides the ion transport mechanism between the

anode and the cathode necessary to maintain the

electrochemical process).

Fig. 1: The electrochemical reactor in the laboratory experiments. (1) DC power supply, (2) magnetic stirrer, (3) cover, (4)

electrodes, (5) magnetic bar-stirrer, (6) wastewater and (7) electric wire (Source: Bouhezila et al., 2011).

Fig. 2: Conceptual diagram of an electrochemical reactor (Source: Anglada et al., 2009)

Electrochemical oxidation of impurities in

wastewater is accomplished through two different

mechanisms as demonstrated in Figure 3: (1) direct

anodic oxidation, where the pollutants are destroyed at

the anode surface and (2) indirect oxidation where

mediators (NaCL, HClO, H2S2O8, etc) are

electrochemically produced to achieve the oxidation.

It should be clear that during electro-oxidation of

aqueous effluents, both oxidation mechanisms may

coexist (Chiang et al., 1995). Generally, the

mechanism of electrochemical degradation of

wastewater is a complex phenomenon involving

coupling of electron transfer reaction with a dissociate

chemisorptions step.



99

Fig. 3: Schemes for direct and indirect electrolytic treatment of pollutants (Chiang et al., 1995).

2.1. Direct oxidation

Direct oxidation of pollutants takes place in two steps:

(i) diffusion of pollutants from the bulk solution to the

anode surface and (ii) oxidation of pollutants at the

anode surface. As a result, the effectiveness of the

electrochemical oxidation will depend on the

correlation between mass transfer of the substrate and

electron transfer at the electrode surface. The rate of

electron transfer is determined by the electrode

activity and current density. In general, there are two

different pathways of anodic oxidation of organic

substances as shown henceforth (Drogui et al., 2007):

• Electrochemical conversion. Organic

substances (R) are partially oxidized as presented in

Eq. 1. Thus, a following treatment is needed to

completely destroy the oxidized substrates.

R → RO + e− (1)

• Electrochemical incineration (combustion).

Organic substances are transformed into water, carbon

dioxide and other inorganic constituents as presented

in Eq. 2.

R → CO2 + H2O + Salts + e− (2)

2.1. Indirect oxidation

During indirect electrochemical oxidation, a strong

oxidizing agent is electro-generated at the anode

surface and subsequently destroys the organic

compounds in the bulk solution. The most widespread

electrochemical oxidant is chlorine which is produced

via the oxidation of chloride at the anode. Throughout

indirect oxidation, the agents produced on the anode

that are responsible for oxidation of inorganic and

organic matters could be chlorine and hypochlorite,

hydrogen peroxide, peroxodisulfuric acid, and ozone

(Li et al., 2010; Scialdone et al., 2009). Accordingly,

throughout the electrochemical oxidation of

wastewater, the impurities removal principally

occurred due to indirect oxidation, utilizing

chlorine/hypochlorite produced by anodic oxidation of

chlorine that existing or being added in the aqueous. A

chain of reactions that involve chlorine/hypochlorite

indirect oxidation are presented in Eqs. 3-9.

Anodic reactions:

2Cl−→ Cl2 + 2e

− (3)

6HOCl + 3H2O → 2ClO3− + 4Cl

− + 12H

+ + 1.5O2 +

6e− (4)

2H2O → O2 + 4H+ + 4e

− (5)

Bulk reactions:

Cl2 +H2O → HOCl + H+ + Cl

− (6)

HOCl → H+ + OCl

− (7)

Cathodic reactions:

2H2O + 2e- → 2OH

− + H (8)

OCl− + H2O + 2e

− → Cl

− + 2OH

− (9)

The hypochlorite (OCl−) generated in bulk solution

(Eqs. 6 and 7) is a strong oxidizing agent that can

oxidize aqueous organic substances (Scialdone et al.,

2009). In addition to the common oxidants that can be

electrochemically produced, metal catalytic mediators

(Ag+2

, Co+3

, Fe+3

, etc.) are also employed for the

generation of hydroxyl radicals, as seen in the electro-

Fenton system. Nevertheless, the use of metal ions

may result in the treated effluent to be more toxic than

that its initial state. Therefore, the system of this kind

needs a separation step to recover the metallic species

Bashir et al.


100

(Anglada et al., 2009), leading to the unfavorable

intricate treatment process.

2.3. Process Design Issues

Electrode materials, cell design (configuration),

working conditions and energy consumption have to

be taken into the consideration when it comes to the

building up of the electrochemical oxidation system.

2.3.1. Electrode material

The choice of electrode materials is very important

since it affects the selectivity and the efficiency of the

process. The complexity of electrode performances

and lack of adequate information insights make it

unfeasible to choose the optimum electrode for a

given process on a theoretical basis. The preliminary

selection is depending on process experience and this

is then tested and refined during an extensive

development program. In fact, it is complicated to

expect the achievement of an electrode material or to

characterize its lifetime without extended studies

under realistic operation conditions (Klamklang et al.,

2012).

Essentially, the electrode materials must have the

following properties (Anglada et al., 2009; Klamklang

et al., 2012):

(a) High physical stability; the electrode material

must have good mechanical strength, good resistance

to erosion and must be resistant to cracking.

(b) High chemical stability; the electrode material

must be resistant to corrosion, unwanted oxide or

hydride formation and the deposition of inhibiting

organic films under all conditions.

(c) Suitable physical shape; it should be feasible to

make the material into the required shape, to assist

sound electrical connections and also to allow simple

fixing and replacement at a variety of scales.

(d) Electrical conductivity; conductivity must be

practically high throughout the electrode system

including the current feeder, electrode connections

and the entire electrode surface exposed to the

electrolyte.

(e) Catalytic activity and selectivity; the electrode

material must sustain the desired reaction and in some

cases, significant electro-catalytic properties are vital.

The electrode material must encourage the desired

chemical change while inhibiting all competing

chemical changes.

(f) Low cost/life ratio; the use of reasonably priced

and durable electrode materials must be favored.

Competition between organics oxidation at the

anode and the side reaction of oxygen evolution

should be considered to assess the choice of an anode

material. The oxidation of water to oxygen (Eq. 5)

happens at about 1.2 V versus normal hydrogen

electrode. Yet, a higher voltage is required for

electrochemical oxidation of water to take place at the

anode. The oxygen evolution over potential of a

number of electrode materials is illustrated in Table 1

(Chen, 2004).

Table 1: Potential of oxygen evolution of different anodes, V versus normal hydrogen electrode (Chen, 2004)

Anode Potential (V) Conditions

Pt 1.3 0.5 mol L−1 H2SO4

Pt 1.6 0.5 mol L−1 H2SO4

IrO2 1.6 0.5 mol L−1 H2SO4

Graphite 1.7 0.5 mol L−1 H2SO4

PbO2 1.9 1.0 mol L−1 H2SO4

SnO2 1.9 0.5 mol L−1 H2SO4

TiO2 2.2 1.0 mol L−1 H2SO4

Si/BDD 2.3 0.5 mol L−1 H2SO4

Ti/BDD 2.7 0.5 mol L−1 H2SO4

There are some general guidelines to assist the

choice of an electrode material. In general, low O2

overvoltage anodes are distinguished by a high

electrochemical activity toward oxygen evolution and

low chemical reactivity toward oxidation of organic

compounds. Efficient pollutants oxidation at these

anodes may take place at low current densities. A

significant reduction of the current efficiency is

expected at high current densities due to the

production of oxygen. Conversely, at high O2

overvoltage anodes, higher current densities may be

used with minimal involvement from the oxygen

evolution side reaction. Thus, high O2 overvoltage

anodes are generally preferred. For example, boron-

doped diamond (BDD) anodes have been confirmed

to yield higher organic oxidation rates and superior

current efficiencies than other commonly used metal

oxides including PbO2 and Ti/SnO2-Sb2O5 (Anglada et

al., 2009).

2.3.2. Cell design

Maintaining high mass transfer rates as the main

reactions that occur in electrochemical process

transpire on electrode surfaces are the most important

issue in cell design. To improve mass transfer,



101

techniques such as gas sparging, high fluid velocity,

use of baffles and incorporation of several types of

turbulence promoters are frequently employed. In

obtaining a high mass transfer rate, the cell

construction should account for simple access to and

exchange of cell parts (Wendt and Kreysa, 1999).

Figure 4 summaries the various features that should be

considered in the design of an electrochemical reactor

(Anglada et al., 2009).

Fig. 4: Categorization of electrochemical reactors in regards to cell configuration, electrode geometry and flow type (Anglada

et al., 2009).

Two types of electrodes, principally of 2-

dimensional and 3-dimensional construction subsist.

The 3-dimensional assures a high value of electrode

surface to cell volume ratio. Both types can be

classified into static and moving electrodes as shown

in Figure 4. Accordingly, the utilization of moving

electrodes increases the mass-transport coefficient

owing to the turbulence promotion. However, among

the 2-dimensional electrodes, static parallel and

cylindrical electrode cells are used in the major

reactor designs in the latest studies. Cell designs using

the parallel plate geometry in a filter press

arrangement are generally used because of the

simplicity of scale-up to a larger electrode size by

merely adding electrodes or increasing number of cell

stacks (Rajeshwar and Ibanez, 1997). Furthermore,

cell configuration (divided and undivided) needs to be

considered. In divided cells, the anolyte and catholyte

are separated via a porous diaphragm or an ion

conducting membrane. The selection of the separating

diaphragm or membrane in divided cells is equally

vital as the selection of electrode materials. In general,

divided cells choice should be avoided whenever

possible, as separators are expensive and tightening of

a divided cell (reduction of electrode gap) is difficult

and encounters a host of mechanical and corrosion

problems (Wendt and Kreysa, 1999).

2.3.3. Operation conditions

(a) The current density (CD) is among the most

important factors that usually control electrochemical

oxidation processes through the reaction rate. It

should be clear that an increase in CD does not

necessarily result in the increase of oxidation

efficiency; the effect of current density on the

treatment level depends on the features of the effluent

to be treated. On the other hand, the use of higher CD

generally results in higher operating costs due to the

increase of energy use.

(b) An increase in the temperature leads to more

efficient processes by global oxidation. While direct

oxidation processes remain almost unaffected by

temperature, this fact may be explained in terms of the

presence of inorganic electro-generated reagents. An

enhancement with rising temperature of the mediated

Bashir et al.


102

oxidation processes by inorganic electro generated

reagents (active chlorine, peroxodisulfate) has been

reported. But, operation at ambient temperature is

preferred as it offers electrochemical processes with

less temperature requirements than those of the

equivalent non-electrochemical counterparts (i.e.,

incineration, supercritical oxidation) (Canizares et al.,

2006).

(c) The physicochemical features of the wastewater

(e.g., electrolyte nature and amount, pH value and

initial concentration of pollutants) also affect the

electrochemical oxidation process. The higher the

concentration of electrolyte is used, the higher the

conductivity and the lower cell voltage for a given

current density are recorded. Thus, treatment by

electrochemical oxidation is more suitable and cost

efficient when the wastewaters contain high salinity.

The effect of pH value is similar temperature, affects

mostly indirect oxidation processes (Anglada et al.,

2009). In chloride mediated reactions, the pH value

may influence the oxidation rate. During indirect

oxidation, chlorine evolution occurs at the anode (Eq.

3). At pH values < 3.3, the primary active chloro

species is Cl2 while at higher pH values its diffusion

away from the anode is coupled to its

disproportionation reaction to form HClO at pH<7.5

(Eq. 6) and ClO− at pH>7.5 (Eq. 7). Theoretically,

operation at acidic conditions could be the finest

option as chlorine is the strongest oxidant followed by

HClO. Accordingly, higher pH values would improve

the electro-oxidation of pollutants, as HClO and ClO−

are almost unaffected by desorption of gases and they

can act as oxidizing reagents in the total volume of

wastewater (Canizares et al., 2006).

2.3.4 Energy Consumption

The energy expenditure should be reduced to

minimize the power costs. The total power

requirement has contributions for both electrolysis and

movement of either the solution or the electrode. The

design of both electrodes and cell has a chief role in

reducing power needed. Therefore, a very open flow-

through porous electrode will have a low pressure

drop linked with it, giving rise to modest pumping

costs and facilitating reactor sealing. A high surface

area electrode which itself a turbulence promoter in

bed electrode, will give rise to a moderately high mass

transfer coefficient and active area without the need

for high flow rates through the cell; the pumping cost

will again be moderately low (Klamklang et al.,

2012). The maintenance of a low cell voltage requires

awareness to electrodes and cell design. The following

aspects should be considered:

(a) The counter electrode reaction should be

selected to reduce the reversible cell voltage. Thus, a

suitable and stable electrode material is required.

(b) The over-potentials at both electrodes should

be minimized through using electro catalysts.

(c) The electrodes, current feeders, and connectors

should be prepared from greatly conducting materials.

(d) Electrode and cell design should allow a small

inter-electrode or electrode membrane gap. The

electrode may touch the membrane as in zero-gap or

solid polymer electrolyte cells.

(e) A separator should be avoided by suitable

selection of the counter electrode chemistry or a thin

conductive membrane should be applied.

3. APPLICATIONS OF ELECTROCHEMICAL

OXIDATION IN WASTEWATER TREATMENT

Being touted as an effective treatment process, the

performance of electrochemical oxidation process in

treating various types of complicated wastewater

containing various pollutants has been studied. Also,

considerable efforts have been contributed recently to

elimination micro-contaminants using electrochemical

oxidation process. In general, microorganisms can be

deactivated via direct electrochemical process or by

the creation of ‘‘killer’’ agents, for example ·OH

(Lazarova and Spendlingwimmer 2008; Polcaro et al.,

2007). The combination of pollutants removal with

disinfection of wastewaters in a single treatment step

poses an attractive option, mainly in water recovery

and reuse where effectual removal of pathogens is

critical to protect public health. Table 2 presents the

effectiveness of electrochemical oxidation process in

treating variety of wastewaters.

Post-treatment of slaughterhouse wastewater via

electrochemical oxidation process was studied by

Awang et al. (2011). The most favorable conditions

were determined as 220 mg/L influent COD, 30

mA/cm2 current density and 55 min reaction time.

This resulted in 96.8% of color removal, 81.3% of

BOD removal and 85.0% of COD removal. Under the

optimal operation conditions (initial pH 6.9, current

density of 10 mA/cm2, conductivity of 3,990 micro

S/cm, and electrolysis time of 10 min), the removal

efficiencies of the textile wastewater by

electrochemical oxidation were 78% of COD and 92%

of turbidity. The energy and electrode consumptions

at the optimum conditions were calculated to be 0.7

kWh/kg COD (1.7 kWh/m3) and 0.2 kg Fe/kg COD

(0.5 kg Fe/m3), respectively (Kobya et al., 2009).

Landfill leachate treated electrochemically using

graphite carbon electrodes by Bashir et al. (2009), the

highest COD removal of 68% was achieved under the

operational conditions of 4 h reaction time and 79.9

mA/cm2 current density, while the initial COD was



103

1414 mg/L. In another study conducted by Moraes

and Bertazzoli (2005), about 73% of COD, 57% of

TOC, 86% of color removals at a current density of

116.0 mA/cm2 and 180 min of reaction were attained.

They used oxide-coated titanium as an anode

electrode. The electrochemical treatment of industrial

water-based paint wastewater was examined in a

continuous tubular reactor. The effects of reaction

time on COD, color and turbidity removals was

investigated at 30 °C, 35 g/L electrolyte and

7496 mg/L of initial COD concentrations with

66.8 mA/cm2 current density. The optimum residence

time in the reactor was fixed at 6 h for a cost driven

approach, enabling COD, color and turbidity removal

of 44.3%, 86.2% and 87.1%, respectively (Körbahti

and Tanyolaç, 2009).

Electrochemical treatment of organic pollutants

from paper mill effluent was investigated by El-

Ashtoukhy et al. (2009). The results showed that the

percentage of COD and color removals were 97% to

100%, respectively. Energy consumption calculation

shows that energy consumption ranges from 4 to

29 kWh/m3 of effluent depending on the operating

conditions. In another study, the electrochemical

oxidation of paper mill effluents was investigated via

a dimensionally stable anode of composition Ti/RuPb

(40%) Ox. The results indicated that about 99% of

COD and 95% of color and polyphenols were

removed after 15 min of electrolysis. The UV-Vis

spectrum illustration confirmed the formation of

hypochlorite ions (ClO-) during the electrolysis

process, indicating that the electrochemical oxidation

proceeds via an indirect mechanism with the

participation of hypochlorite ions (Zayas et al., 2011).

In the case of olive oil mill wastewater, the removal

rates of organics increased with the increase of

applied current density, sodium chloride level,

recirculation rate and temperature. The original COD

concentration of 41,000 mg/L was reduced to 167

mg/L, 99.85% of turbidity removal, 99.54% of oil-

grease removal were achieved after 7 h electrolysis at

the conditions of 135mA/cm2, 2M NaCl, 7.9 cm

3 /s,

and 40◦C (Un et al., 2008). The effect of current

density (40-120 A/m2) and initial pH (3-11) on the

Pharmaceutical wastewater treatment efficiency by

electro oxidation process was investigated

(Deshpande et al., 2012).Under optimum operating

conditions (CD 80 A/m 2; pH 7.2), the process used

aluminum electrodes resulted in 24% of COD

removal after 25 min, whereas the process used

carbon electrode achieved 35.6% of COD removal

after 90 min of treatment (Deshpande et al., 2012). An

investigation of tannery wastewater treatment using

graphite cathodes and Ti/SnO2/PdO2/RuO2 anode,

with a current density of 2.1 A/dm2 was carried out.

After 55 min of the process the catholyte was

transferred into the anodic space and the process was

continued. After 55 min of electro-Fenton process, the

COD was reduced by 52.0%. Electrooxidation

continued by the anodic process resulted in

elimination of ammonia in 55 min and a total

reduction of COD by 72.9% (Naumczyk and

Kucharska, 2011).

Due to its unique performance in treating various

types of wastewater especially industrial wastewater

and landfill leachate which contain large amount of

the toxic and non-biodegradable pollutants as

aforementioned, it can be concluded that

electrochemical oxidation process represents a useful

solution when the existence of refractory and toxic

pollutants prevents the use of conventional biological

treatments. Under suitable operation conditions, a

total removal of COD, color, ammonia and

microorganisms can be achieved.

4. OPORTUNITIES AND CHALLENGES

The appearance of pollutants that are unmanageable

by conventional biological and chemical treatments

together the means of stricter restrictions enforced by

new legislation have resulted in much research work

focus on wastewater treatment via electro-oxidation

processes. Electrochemical oxidation has been found

to be an environmentally caring technology with

capability to remove completely non-biodegradable

organic compounds and eliminate nitrogen species.

Recently, the researchers in this field directed their

work towards two lines: (i) replacement of

conventional processes by electrochemical oxidation

and (ii) integration of electrochemical oxidation into a

treatment plant. As electrical energy is mainly

consumed in electrochemical oxidation process, the

use of photovoltaic (PV) modules as a power supply is

also expected to reduce the operating costs

(Klamklang et al., 2012; Anglada et al., 2009).

Indeed, high energy consumption is generally

required, limiting the further full-scale marketable

application. Two steps have been taken to reduce

treatment costs; (i) the use of this technology in

combination with other techniques as either a pre-

treatment or as a polishing step and (ii) the use of

renewable energy sources to power electrochemical

oxidation (Anglada et al., 2009). In addition to the

energy consumption, during the process design some

critical issues are important to be considered

especially in the design of electrodes and cells. These

include cost, safety, simplicity of maintenance, and

ease to use. It is also necessary that the performance

of the electrodes is maintained during the expected

operating life of the cells (Klamklang et al., 2012).

Although it has been confirmed that

electrochemical oxidation is a technically practicable

Bashir et al.


104

option to eliminate organic pollutants, the partial

oxidation of ammonia to nitrate ions has been

reported. The deployment of electrochemical

oxidation in combination with other process such as

ion exchange (Cabeza et al. 2007) as a post treatment

step could be a plausible solution to this issue.

Consequently, Comninellis et al. (2008) had

demonstrated the promising results obtained from the

treatment of industrial wastewaters via combined

methods involving electrochemical oxidation have

built up foundation for upcoming works. Contriving a

sustainable process based on the combination of

efficient technologies is one of the key obstructions

that need to be overcome before full-scale

implementation of electrochemical oxidation.

Table 2: Application of electrochemical oxidation process in waste water treatment

Type of wastewater Electrode material Performance references

Slaughterhouse

Wastewater

aluminum

96.8% color, 81.3% BOD, and 85.0%

COD removals.

Awang et al.(2011)

Textile Wastewater iron electrode 78% COD, and 92% turbidity removals

Kobya et al. (2009)

Textile Wastewater

graphite electrodes 100% dye removal Kariyajjanavar et al.

(2011)

Landfill leachate graphite Carbone

68%COD, 84% color, and 70% BOD

removals.

Bashir et al. (2009)

Landfill leachate 30% RuO2 and 70% TiO2

coated titanium

73% COD, 57% TOC, 86% color

removals

Moraes and

Bertazzoli, (2005)

Industrial paint

wastewater

stainless steel

44.3% COD, 86.2% color, and 87.1%

turbidity removals

(Körbahti and

Tanyolaç, 2009)

paper

mill effluents

-A cylindrical lead sheet as

anode

- a cylindrical stainless steel

sheet as cathode

97% COD, and 100% color removals

El-Ashtoukhy et al.

(2009)

paper

mill effluents

-Ti/RuPb(40%)Ox as anode

-Ti/PtPd(10%)Ox as

cathode.

99% COD and 95% of color and

polyphenols removals

Zayas et al. (2011)

Olive oil mill

Effluents

RuO2 coated Ti 99.6% COD, 99.85% turbidity, and

99.54% oil-grease removals

Un et al.(2008)

Pharmaceutical

Wastewater

Carbon electrode 35.6% COD removal

Deshpande et al.

(2012)

Tannery

Wastewater

-graphite cathodes

-Ti/SnO2/PdO2/RuO2 anode

72.9 % COD removal Naumczyk and

Kucharska (2011)

5. CONCLUSION

Wastewater treatment by electrochemical oxidation

process was established in a laboratory scale for many

years. However, electrochemical oxidation

technologies have not reached real application

maturity in commercial scale perhaps due to the

limitation of comparatively high capital investment

and the cost of electricity supply. Consequently,

operating cost reduction and efficient electrode

materials manufacturing are the main problems need

to be overcome before the site-scale accomplishment

of electrochemical oxidation in wastewater treatment.

ACKNOWLEDGEMENTS

The authors are grateful for the financial support

provided by the Universiti Tunku Abdul Rahman

(UTAR) through grant No:

IPSR/RMC/UTARRF/2012-C2/M03.

REFERENCES

Anglada A, Urtiaga A, Ortiz I (2009). Contributions

of electrochemical oxidation to waste-water

treatment: fundamentals and review of

applications Emerging Technologies. J Chem

Technol Biotechnol, 84: 1747-1755.



105

Awang ZB, Bashir MJK, Kutty SRM, Isa MH (2011).

Post-Treatment of Slaughterhouse Wastewater

using Electrochemical Oxidation, Research

Journal of Chemistry and Environment, 15:

229-237.

Bashir MJK, Aziz HA, Aziz SQ, Abu Amr SS (2013).

An Overview of Electro-oxidation Processes

Performance in Stabilized Landfill Leachate

Treatment. Desalination and Water Treatment,

51: 2170–2184.

Bashir MJK., Isa MH, Awang ZB, Aziz HA, Kutty

SRM, Mohajeri S, Farooqi IH (2009). Landfill

leachate treatment by electrochemical

oxidation. Waste Management, 29: 2534-2541.

Bouhezila F, Hariti M, Lounici H, Mameri N (2011).

Treatment of the OUED SMAR town landfill

leachate by an electrochemical reactor.

Desalination, 280: 347–353.

Cabeza A, Primo O,Urtiaga AM, Ortiz I (2007).

Definition of a clean process for thetreatmentof

landfill leachates integration of electro-

oxidation and ion exchange technologies. Sep

Sci Technol, 42:1585–1596.

Canizares P, Paz R, Lobato J, Saez C, Rodrigo MA

(2006). Electrochemical treatment of the

effluent of a fine chemical manufactruring

plant. Journal Hazardous Materials, 138: 173–

181

Chen G (2004). Electrochemical technologies in

wastewater treatment. Separation and

Purification Technology, 38: 11-41.

Chiang LC, Chang JE, Wen TC (1995). Indirect

oxidation effect in electrochemical oxidation

treatment of landfill leachate. Water Research,

29: 671–678.

Comninellis C, Kapalka A, Malato S, Parsons SA,

Poulios I, Mantzavinos D (2008). Advanced

oxidation processes for water treatment:

advances and trends for R&D. J Chem Technol

Biotechno, l83:769–776.

Deshpande A, Ramakant, Satyanarayan S (2012).

”Treatment of Pharmaceutical Wastewater by

Electrochemical Method: Optimization of

Operating Parameters by Response Surface

Methodology.” J. Hazard. Toxic Radioact.

Waste, 16: 316–326.

Drogui P, Blais JF and Mercier G (2007). Review of

electrochemical technologies for environmental

applications. Recent Patents on Engineering,

1:257–272.

El-Ashtoukhy ESZ, Amin NK, Abdelwahab O (2009).

Treatment of paper mill effluents in a batch-

stirred electrochemical tank reactor. Chemical

Engineering Journal, 146: 205–210.

Grimm J. et al. (1998). Sol-gel film preparation

electrodes for the electrocatalytic oxidation of

organic pollutants in water. Desalination, 115:

295-302,

Kapałka A, Fóti G, Comninellis C (2010). Basic

Principles of the Electrochemical

Mineralization of Organic Pollutants for

Wastewater Treatment. Journal of Applied

Electrochemistry, 40: 2203.

Kariyajjanavar P, Narayana J, Nayaka YA (2011).

Degradation of Textile Wastewater by

Electrochemical Method. Hydrology Current

Research, 2:110. doi:10.4172/2157-

7587.1000110

Klamklang S, Vergnes H, Pruksathorn K,

Damronglerd S (2012). Electrochemical

Incineration of Organic Pollutants for

Wastewater Treatment: Past, Present and

Prospect,Organic Pollutants Ten Years After

the Stockholm Convention - Environmental and

Analytical Update, Dr. T. Puzyn (Ed.), ISBN:

978-953-307-917-2, InTech

Kobya M, Demirbas E, Akyol A (2009).

Electrochemical treatment and operating cost

analysis of textile wastewater using sacrificial

iron electrodes, Water Science and Technology,

60: 2261-2270.

Körbahti BK, Tanyolaç A (2009) Electrochemical

treatment of simulated industrial paint

wastewater in a continuous tubular reactor.

Chemical Engineering Journal, 148: 444-451.

Lazarova Z, Spendlingwimmer R, (2008). Treatment

of yellow water by membrane separations and

advanced oxidation methods. Water Science

and Technology, 58: 419–426.

Li W, Zhou Q, Hua T (2010). Removal of organic

matter from landfill leachate by advanced

Oxidation processes: A Review. Int. J. Chem.

Eng. Article ID 270532 (2010) 1-10.

Moraes PB, Bertazzoli R (2005). Electrodegradation

of landfill leachate in a flow electrochemical

reactor. Chemosphere, 3: 41–46.

Naumczyk J, Kucharska M (2011). Tannery

wastewater treatment by anodic elect oxidation

coupled with electro-fenton process.

Environment Protection Engineering, 37: 47-

54.

Polcaro AM, Vacca A, Mascia M, Palmas S, Pompei

R, Laconi S (2007). Characterization of a

stirred tank electrochemical cell for water

disinfection processes. Electrochim Acta, 52:

2595–2602.

Rajeshwar K, Ibanez J (1997). Environmental

Electrochemistry: Fundamentals and

Applications in Pollution. Abatement Academic

Press Inc, USA.

Scialdone O, Randazzo S, Galia A, Silvestri G

(2009). Electrochemical oxidation of organics

http://www.ncbi.nlm.nih.gov/pubmed?term=Kobya%20M%5BAuthor%5D&cauthor=true&cauthor_uid=19901457

http://www.ncbi.nlm.nih.gov/pubmed?term=Demirbas%20E%5BAuthor%5D&cauthor=true&cauthor_uid=19901457

http://www.ncbi.nlm.nih.gov/pubmed?term=Akyol%20A%5BAuthor%5D&cauthor=true&cauthor_uid=19901457

http://www.ncbi.nlm.nih.gov/pubmed/19901457

http://yadda.icm.edu.pl/baztech/contributor/a2520a8faf9add0c8e538ef7c7689b7f

Bashir et al.


106

in water: Role of operative parameters in the

absence and in the presence of NaCl, Water

Research, 43: 2260-2272.

Un UT, Altay U, Koparal AS, Ogutveren UB (2008).

Complete treatment of olive mill wastewaters

by electrooxidation. Chemical Engineering

Journal, 139: 445–452

Wang B, Chang X, Ma H (2008). Electrochemical

oxidation of refractory organics in the coking

wastewater and chemical oxygen demand

(COD) removal under extremely mild

conditions. Ind. Eng. Chem. Res., 47:8478–

8483.

Wendt H, Kreysa G (1999). Electrochemical

Engineering: Science and Technology in

Chemical and other Industries. Springer,

Germany.

Zayas T, Picazo M, Salgado L (2011). Removal of

Organic Matter from Paper Mill Effluent by

Electrochemical Oxidation, Journal of Water

Resource and Protection, 3(1): 32-40.



©2014 IJSRPUB

107


Treatment Systems

Chapter 5: Wastewater Treatment by Bioremediation

Technologies



©2014 IJSRPUB

108

Wastewater Treatment by Bioremediation Methods

Andrew N. Amenaghawon*, Kessington O. Obahiagbon

Department of Chemical Engineering, Faculty of Engineering, University of Benin, PMB 1154, Benin City, Nigeria

*Corresponding Author: E mail: [email protected]

Abstract. Increasing global population and the attendant increase in the level of industrialisation has resulted in the generation

of large amounts of effluents with wastewater chief amongst them. The deleterious effect the improper discharge of untreated

wastewater has on the environment and living things has prompted many nations of the world to take this problem very

seriously. This has resulted in the promulgation of stiffer environmental laws and regulations. This in turn has led to the search

for more efficient ways of treating wastewater. Because of the inherent limitations of the conventional wastewater treatment

methods, bioremediation has jumped to the fore front as a sustainable, acceptable, effective, low cost and environmentally

friendly wastewater treatment strategy. Bioremediation is a wastewater treatment process that utilises microorganisms either

autochthonous or exogenous to degrade numerous organic pollutants as a result of their metabolic activity and their capability

to adapt to extreme and often inhospitable environments. Bioremediation as a treatment strategy involves natural attenuation

which involves the use of indigenous microbial population without any interference, biostimulation which is the process of

stimulating viable indigenous microbial population, bioaugmentation which is the introduction of exogenous microbial

population to the remediation medium, bioaccumulation which is the use of live cells, biosorption which is the use of dead

microbial biomass, phytoremediation (the use of plants) and rhizoremediation which involves plant and microbe interaction.

This chapter explores the special features of bioremediation, its principles, strategies and forms of operation. It also takes a

look at the application of bioremediation to selected pollutants like petroleum hydrocarbons and derivatives, polycyclic

aromatic hydrocarbons and heavy metals as well as factors that affect bioremediation and the role of microorganisms in

bioremediation.

Keywords: Bioremediation, Wastewater, Treatment

1. INTRODUCTION

The population explosion currently experienced in the

world today especially in the urban centres of the

world poses a considerable threat to the environment

as a result of the generation of huge amounts of

wastewater (Amenaghawon et al., 2013; Akpor and

Muchie, 2010). The increasing population has resulted

in a corresponding increase in industrial and

agricultural activities from which these wastewaters

are produced (Godos et al., 2009; Morill et al., 2009).

The discharge of untreated wastewater generated from

these operations into natural water bodies represents a

huge environmental burden such as the accumulation

of hazardous and toxic chemicals in the receiving

water bodies with potentially deleterious

consequences on the immediate ecological

environment. This has necessitated the need to

monitor the quality and quantity of wastewater

generated and discharged into natural water bodies

(Amenaghawon et al., 2013; Kshirsagar, 2013). This

has also led to the development of wastewater

treatment strategies that focus on effective removal of

pollutants from water rather than disposal.

Wastewater treatment is presently carried out

mainly through chemical and biological means.

Irrespective of the advantages of chemical treatment

methods which include the mineralisation of

otherwise non-biodegradable substances and the use

of a smaller reactor size, the disadvantages of this

method are still significant (Agarry et al., 2008; Akpor

and Muchie, 2010; Otokunefor and Obiukwu, 2010).

As a result of the limitations of the chemical treatment

methods, biological methods have been encouraged.

Recently, a lot of attention has been given to

bioremediation as a suitable and sustainable biological

method for wastewater treatment. It is touted to be an

effective, environmentally friendly and economic

treatment option for wastewater (Crawford and

Crawford, 2005; Otokunefor and Obiukwu, 2010).

Bioremediation involves the use of naturally

occurring microorganisms to breakdown complex

pollutant molecules into simpler substances such as

carbon dioxide and water (Obahiagbon and Aluyor,

2009). It can be accomplished on its own, a process

known as natural attenuation or intrinsic

bioremediation or it can be stimulated through

nutrient supplementation to increase bioavailability

within the remediation medium, a process known as

biostimulation (Sharma, 2012). In order to be a

successful treatment strategy, bioremediation must

take into account, the multiphase and complex nature

of the remediation environment; thus resulting in the

need to incorporate other subjects such as chemistry,

microbiology, engineering etc. Recently, a lot of

attention is being given to the state and condition of

the environment culminating in the formulation of

new environmental laws and regulations. These

regulations advocates better protection for the

environment and mandates cleanup of polluted sites.

This has had a positive boomerang effect on

bioremediation. Bioremediation is now gaining more


Amenaghawon and Obahiagbon


109

acceptance as a suitable treatment method, it is

becoming a choice treatment scheme for the cleanup

of polluted sites particularly sites contaminated with

petroleum hydrocarbons, it is attracting funding from

agencies in governments and private corporations and

a lot of advancement is being made in improving on

the effectiveness of the technology (Crawford and

Crawford, 2005; Vidali, 2001).

1.1. Natural Attenuation

Natural attenuation also called intrinsic

bioremediation refers to the combination of natural

processes that occur, without human involvement, to

decrease or “attenuate” contaminant concentrations

and toxicity in wastewater, and thereby reduce the

hazards posed by the contaminants. The

Environmental Protection Agency (EPA) defines

natural attenuation to include biodegradation,

biosorption and biological stabilisation or destruction

of contaminants (EPA, 1999). Natural attenuation is

usually employed at the site of contamination and the

appropriate conditions must be put in place for the

process to function effectively. This is why the

process is typically monitored in which case the

process is referred to as monitored natural attenuation.

Cleanup of contaminants from wastewater by

natural attenuation occurs in various forms:

(a) Biodegradation: this occurs when

microorganisms metabolise pollutants in

wastewater leading to the mineralisation of these

pollutants into simpler and harmless substances

such as carbon dioxide and water. The wastewater

typically contains indigenous microorganisms that

effect the mineralisation of the pollutants through

the natural attenuation process (Adekunle and

Adebambo, 2007; Amenaghawon et al., 2013).

(b) Dilution: the concentration of pollutants

reduces as they move through the bulk of the

receiving water body and mix with it.

(c) Evaporation: some volatile compounds in

wastewater such as petroleum hydrocarbons and

their derivatives typically referred to as volatile

organic compounds (VOCs) can be mobilised into

the vapour phase thereby reducing their

concentration in the liquid phase. These gases

released are typically diluted by mixing with air

when they are mobilised into the vapour phase.

(d) Chemical reactions: pollutants in wastewater

may be converted to less harmful forms when they

react with other components in the wastewater.

A lot of studies have been done on the use of

natural attenuation for the remediation of polluted

wastewater. Most of the early work done in this area

was focused on remediation of benzene, toluene,

ethylbenzene, and xylene (BTEX) using natural

means. These are hydrocarbons that are commonly

found in refinery effluents (Hidayat and Tachibana

2012). Fono et al. (2006) investigated the application

of natural attenuation to the remediation of

wastewater derived chemical contaminants in an

effluent dominated river. The results they reported

indicate that natural attenuation can result in a

substantial reduction in the concentrations of the

wastewater derived contaminants. Yu et al. (2005)

investigated the bioremediation of a mixture of PAHs,

namely fluorene (Fl), phenanthrene (Phe) and pyrene

(Pyr) in mangrove sediment slurry using natural

attenuation. They reported that the indigenous

microorganism degraded over 99% Fl and Phe but

only around 30% of Pyr through natural attenuation

within a one month period suggesting that natural

attenuation may be a more appropriate method for

remediation of Fluorene and Phenanthrene

contaminated mangrove sediments.

Monitored natural attenuation is a low cost process

involving only the cost of monitoring and the time

required for the process to occur. It also does not lead

to the destruction of the immediate ecological

environment (Yu et al., 2005). This has made it an

acceptable choice for the cleanup of low risk oil

contaminated waters (Margesin and Schinner, 2001).

However, the process usually takes a long time to

effect satisfactory cleanup as a result of the small

population of indigenous microorganisms present

(Forsyth et al., 1995; Yu et al., 2005).

1.2. Biostimulation

The indigenous microbial population present in

wastewater implies that there is possibility for

remediation through natural attenuation. Nevertheless,

the remediation of wastewater through natural means

is still limited by several factors which might inhibit

microbial growth and activity. Some contaminated

wastewater might contain complex synthetic and

recalcitrant pollutants which are not readily amenable

to biodegradation. Also, there might be deficiency of

electron acceptors or donors and low availability of

nutrient sources such as nitrogen and phosphorus

(Nyyssönen et al., 2009; Qin et al., 2013). In

situations like these, it becomes imperative to improve

the conditions of remediation in the form of external

nutrient supplementation to improve bioavailability,

supplying air or adding electron acceptors or donors to

the substrate analogue (Cosgrove et al., 2010; El

Fantroussi and Agathos, 2005). The process of

externally stimulating microbial growth and activity

for the remediation of contaminants is referred to as

biostimulation. Biostimulation enhances the rate of

bioremediation since the addition of rate limiting

nutrients to the remediation medium promotes the


Chapter 5: Wastewater Treatment by Bioremediation Technologies

110

decontamination capacity of the indigenous microbial

population (Nikolopoulou and Kalogerakis, 2009).

Previous studies on bioremediation of contaminated

wastewater have investigated the role of

biostimulation during bioremediation. Roling et al.

(2002) investigated the effect of nutrient

supplementation on the bioremediation of oil spills.

They observed that nutrient supplementation

substantially enhanced the degradation of oil over a

wide range of concentrations. They further noted that

the addition of nutrients rather than the quantity of

nutrient added was the determining factor. Yu et al.

(2005) reported that biostimulation in the form of the

addition of mineral salt medium resulted in the

degradation of over 97% of fluorene (Fl),

phenanthrene (Phe) and pyrene (Pyr), in mangrove

sediment slurry. This was an improvement over the

results obtained for the case of natural attenuation.

Kim et al. (2005) reported that the addition of

inorganic nutrients stimulated the CO2 evolution rate

which was used as an indication of the performance of

the bioremediation treatment. They also observed that

the addition of surfactants enhanced the degradation

of oil though the effect was not very significant.

Obahiagbon and Aluyor, (2009) investigated the

potential use of sodium nitrate and sodium nitrite as

inorganic nutrient supplements for the biostimulation

of Aspergilllus niger for the bioremediation of

petroleum hydrocarbon polluted water. They recorded

over 97 and 98% reduction levels respectively in total

hydrocarbon content and biochemical oxygen demand

which are key indicators of biodegradation efficiency.

Otokunefor and Obiukwu, (2010) investigated the

treatment of phenol laden refinery effluent using

inorganic fertilisers as biostimulants. They noted that

the inorganic nutrients stimulated rapid growth of the

microbial population present in the remediation

medium. They further observed that the degree of

phenol degradation was dependent on the

concentration of fertilisers used. Amenaghawon et al.

(2013) studied the role of urea and NPK 15:15:15

fertilisers as biostimulants of microbes in the

bioremediation of domestic wastewater. By

monitoring bioremediation indicating parameters such

as pH, biochemical oxygen demand, and dissolved

oxygen over a period of 5 weeks, they were able to

obtain treated wastewater that met the requirements of

environmental regulatory agencies such as the Federal

Environmental Protection Agency (FEPA) and the

department of petroleum resources (DPR) (FEPA,

1999).

For marine environments, it is important to

monitor the use of biostimulants for the following

reason. Firstly, the stimulants should be directed at the

microbial population close to the point of

contamination. Secondly, they should be dispensed in

such a way that they don't get diluted or washed away

by the on rushing waves. Lastly, the concentration of

nutrients especially those containing nitrogen and

phosphorus should be closely monitored as too low a

concentration will not have the desired effect while

too high a concentration can result in eutrophication

which subsequently leads to algal bloom which in turn

can cause a reduction in the dissolved oxygen

concentration in the water body (Gonzalez, 2011).

1.3. Bioaugmentation

The major limitation of natural attenuation processes

for bioremediation of contaminants is that the

population of indigenous microorganisms is low. In

some cases, the indigenous microorganisms might not

even be able to utilize the pollutants for metabolic

activities possibly because they do not possess the

metabolic pathways for the mineralisation of the

pollutants into smaller and harmless compounds like

carbon dioxide and water (El Fantroussi and Agathos,

2005). It is also possible that the pollutants to be

removed can only be biodegraded by a particular

consortium of microorganisms. This observation is

typical of recalcitrant pollutants like polycyclic

aromatic hydrocarbons (PAHs), aromatic and aliphatic

halogenated hydrocarbons, pesticides and nitrated

compounds like 2,4,6-trinitrotoluene (TNT). For

instances like this, it becomes necessary to inoculate

the contaminated site with specific microbial

populations that can perform the desired

bioremediation functions. The process of introducing

external microbial populations into the remediation

medium is referred to as bioaugmentation (Yu et al.,

2005). Bioaugmentation can therefore be defined as

the process of introducing specific microorganisms

with the desired degradation capability for the purpose

of enhancing the removal of pollutants from

wastewater or any other contaminated site (Vogel,

1996).

Bioaugmentation has been the subject of much

research. Gentry et al. (2004) reviewed recent trends

and approaches in bioaugmentation strategies. They

highlighted the techniques and tools advanced in

recent times to improve the activity of externally

introduced microorganisms after their introduction

into the site of remediation. Mrozik and Piotrowska-

Seget, (2010) did a survey on bioaugmentation as a

strategy for the degradation of aromatic compounds.

They argued that a successful bioaugmentation

strategy requires knowledge on the type and

concentration of pollutants and suitable strains of

microorganism with the required degradation

capability. They further mentioned that certain

features of the microorganisms should determine their

selection. These features include the ability for grow



111

fast, ability to tolerate high concentrations of the

target pollutants and survive in a wide range of

environmental conditions. Lendvay et al. (2003) did a

side by side comparison of bioaugmentation and

biostimulation for the cleanup of a chloroethene

contaminated aquifer. Their results indicated that the

deployment of the bioaugmentation strategy yielded a

near stoichiometric dechlorination of the chloroethene

to ethane within six weeks. Smith et al. (2005) also

did a comparison study of biostimulation and

bioaugmentation with a bacterial strain for the

remediation of methyl tertiary butyl ether (MTBE)

contaminated groundwater. They reported a

substantial decrease in the concentration of MTBE in

the contaminated water after a treatment period of six

months. Obahiagbon and Owabor, (2009) utilised a

mixed microbial consortium for the remediation of

crude oil contaminated water. By monitoring key

indicators of bioremediation such as biochemical

oxygen demand and total hydrocarbon content, they

observed that almost 100% degradation was achieved

over a treatment duration of nine weeks. Obahiagbon

and Akhabue, (2009) investigated the effect of

microbial count of Pseudomonas aeruginosa bacteria

amended with NPK 15:15:15 fertiliser on the

biodegradation of crude oil-contaminated water. At

the end of nine weeks of bioremediation, they

obtained satisfactorily treated wastewater that met the

requirements of environmental regulatory agencies

such as the Federal Environmental Protection Agency

(FEPA) and the department of petroleum resources

(DPR) (FEPA, 1999).

Despite the results reported in the literature, there

is still some debate regarding the implementation of

bioaugmentation strategies particularly in relation to

the use of biostimulation. Gonzalez, (2011) reported

that the prevailing conditions at the site of

contamination should determine the choice of strategy

to be deployed.

Bento et al. (2005) did a comparative study on the

degradation of diesel oil by natural attenuation,

biostimulation and bioaugmentation. They noted that

the best option was bioaugmentation. They further

observed that to obtain best performance,

bioaugmentation should be implemented in such a

way that microorganisms chosen for bioaugmentation

should be isolated from the site of contamination. The

reasoning behind this is that the indigenous

microorganisms are more likely to survive and grow

when introduced into the site of contamination

compared with microorganisms foreign to that site

(Thompson et al., 2005).

Some contradictory reports have shown that the

introduction of enriched external cultures of

microorganisms did not affect the degradation

capacity of the microorganisms but biostimulation

enhanced the degradation of the pollutants

(Thomassin-Lacroix et al., 2002). When implemented

with bioaugmentation, biostimulation can provide the

nutrient requirement and the enabling environment for

both indigenous and externally introduced

microorganisms. In the light of this and the inherent

limitations of both methods, it is instructive to deploy

both as complementary strategies rather than

competing ones (Gonzalez, 2011).

2. PRINCIPLES AND OPERATION OF

BIOREMEDIATION

Perhaps the most significant principle of

bioremediation is that microorganisms possess the

ability to degrade contaminants in wastewater.

Wastewater contains biodegradable contaminants that

can be degraded by microorganisms such as bacteria,

fungi and even protists. These organisms are either

indigenous to the contaminated water or are added

exogenously to perform specific remediation

functions. Conventional wastewater treatment is

accomplished in three steps namely primary,

secondary and tertiary. The primary treatment step is

designed to remove suspended and coarse solids from

the wastewater stream. It typically involves the use of

screens and sedimentation vessels to remove solids

from wastewater by the action of gravity. It has been

reported that the primary treatment step can reduce the

total suspended solid particles by as much as 60%.

The secondary step is a biological treatment step

which removes dissolved organic materials. The

tertiary step is often optional. It can remove more than

99% of all contaminants in wastewater resulting in

treated wastewater of almost drinking quality.

Bioremediation finds its place during the

secondary treatment stage. The principles and

technologies applied here typically includes the basic

activated sludge process, trickling filters, rotating

biological contactors and other forms of treatment that

involves the use of biological means to degrade

pollutants.

2.1. The Activated Sludge Process

The activated sludge process developed by Ardern

and Lockett, (1914) represented the highpoint of

major advances in the application of aeration

technology to wastewater treatment. In its

conventional form, it consists of a multi chamber

reactor which makes use of an activated population of

aerobic microorganisms (bacteria, fungi, yeast,

protozoa etc) with the capability to degrade organic

pollutants in wastewater (EPA, 1997). Despite the fact

that a lot of configurations of the process currently

exists, the essential features has remained the same as



112

shown in Figure 3.1. Typically, the set up consists of

an aeration tank and a secondary settling tank or

clarifier. Wastewater that has gone through primary

treatment is charged into the aeration tank along with

the microorganisms to form what is referred to as an

activated sludge or mixed liquor. The aeration in the

tank is accomplished by using submerged diffused or

surface mechanical aeration or a combination of both

to keep the activated sludge as a suspension

(Amenaghawon and Aisien, 2011; EPA, 1997). The

compressed air used for aeration is constantly

transported through the wastewater as it flows through

the aeration tank. The compressed air serves to

provide a source of oxygen for the aerobic microbial

population floc that forms in the tank and the

turbulence and agitation required to bring the

wastewater and the microorganisms into proper

contact. During the process of aeration, the microbial

population will aggregate together to form small

clumps or flocs. After a sufficient time of aeration, the

activated sludge which would have become flocculent

is separated from the wastewater in the secondary

settling tank or clarifier where the flocs settle out. The

clarified wastewater is then sent for further

processing. A part of the effluent sludge from the

clarifier is returned to the aeration tank to maintain the

required microbiological balance in the tank and the

process is repeated. The concentration of sludge

returned to the aeration tank is specified by the ratio

of mixed liquor volatile suspended solids (MLVSS) to

the biochemical oxygen demand of the influent

wastewater that will result in the degradation of the

most amounts of organics from the influent

wastewater. The excess sludge due to microbiological

growth is wasted for further handling. The value of

this ratio affects the efficiency of the treatment

process (Amenaghawon and Aisien, 2011; Guyer,

2011).

Fig. 1: Conventional activated sludge process configuration with aeration and settling tanks

Over the years, different configurations of the

conventional activated sludge process have surfaced.

These adaptations have been made to meet acceptable

effluent standards for biochemical oxygen demand,

nitrogen and phosphorus. By properly calibrating the

aerobic condition, providing microorganisms capable

of degrading the target pollutants, proper nutrient

supplementation, recycle design etc, the activated

sludge process can achieve significantly high

treatment efficiencies.

2.2. Trickling Filter Systems

A trickling filter is a fixed packed bed, biological

filter that operates under aerobic conditions. The fixed

bed is packed with high specific surface area solid

materials like rocks, gravel, shredded PVC materials

or more conventional filter materials (Guyer, 2011).

Typically, materials to be used in designing trickling

filters should have a specific surface area in the range

of 30 to 900 m2/m

3. The filter usually has a depth of

1–2 m however filters packed with lighter synthetic

plastic materials can have depths as high as 12 m. The

use of deeper filters can enhance nitrification potential

and can serve as a second stage in a two-stage

biological system design for nitrification (Guyer,

2011). Wastewater that has gone through primary

treatment is trickled or sprayed over the filter set up

using a rotating sprinkler. The filter arrangement

provides support for the growth of microorganisms

and the wastewater flowing downward through the

filter provides nutrients for the microorganisms. It is

important to first take the wastewater through the

primary treatment step to prevent clogging and ensure

an efficient treatment process. As the wastewater

trickles through the filter, the organic pollutants in it

are degraded by the microorganisms growing on the

filter to simple and non toxic compounds like carbon

dioxide and water and in the process more microbial

cells are produced. A secondary clarifier is typically

needed to clarify the effluent from the trickling filter.

For a completely functional trickling filter system,

reduction levels in biochemical oxygen demand in the

range of 80 to 90 percent have been obtained.

The primary justification of trickling filter systems

has been their low start up, operating and maintenance



113

cost as well as their ease of operation. With modern

designs, the performance of trickling filter systems is

comparable to that of conventional activated sludge

processes. However, trickling filter systems holds the

advantage of providing good performance with little

operational and energy requirements.

2.3. Rotating Biological Contactors

The rotating biological contactor (RBC) is a fixed film

equipment used for secondary wastewater treatment

and operates similar to the trickling filter system

(Guyer, 2011). In its basic form, it consists of a series

of circular disks arranged side by side on a shaft

which is rotated through the wastewater stream. The

surface of the disks is usually covered with

populations of microorganisms such as bacteria and

protists. As the disks slowly rotate on the shaft, it

extracts a film of wastewater into the air which

trickles down and absorbs oxygen. During this

process, the aerobic microorganisms degrade the

organic pollutants in the wastewater to simpler and

less toxic products (Amenaghawon et al., 2013). Any

excess solid and waste product are sloughed off and

transported with the wastewater into the secondary

clarifier for separation. This is usually done to prevent

clogging of the media and to maintain the population

of the microorganisms (Grady et al., 2011). For

operational purposes, a group of contactors are usually

used in series as a single contactor cannot achieve the

desired level of treatment. The speed of rotation of the

RBC is typically in the range of 1 to 2 rpm. This is

enough to provide a reasonable level of degradation.

Some advantages of the RBC system include: high

level of nitrification, short contact times as a result of

the large surface available, ability to handle a wide

range of flows, the excess solids are readily separated,

low operating cost and power requirement. However,

there are some limitations to the use of the RBCs

which include the fact that the RBC units must be

covered in temperate regions to prevent freezing. Also

there is need for frequent maintenance of the moving

parts of the unit which include the shaft bearings and

mechanical drive units.

2.4. Waste Stabilization Ponds

Waste stabilisation ponds are a sustainable and

economic means of treating a wide range of

wastewater with a reasonably clean effluent. The

ponds are usually 0.8 to 1.2m in depth and can be

used after primary treatment as standalone treatments

facilities, in series or parallel. The degree of treatment

obtainable is dependent on the type and number of

ponds used. Wastewater treatment is effected by the

indigenous microorganisms in the pond which

metabolise the organic substances in the wastewater to

simpler and less toxic products (Amenaghawon et al.,

2013).The exposure of the surface of the pond to the

atmosphere creates the aerobic environment need for

the microorganisms to function hence they are

typically referred to as oxidation ponds. This method

of treatment has an advantage over other aeration

treatment methods in that there is no need to install an

aerator to add oxygen to the water as the aeration is

accomplished naturally (Grady et al., 2011). However,

aeration by the natural method is much slower

compared to when an aerator is used. The

consequence of this is that stabilisation ponds treat

wastewater slower than other conventional treatment

plants.

3. FORMS AND STRATEGIES OF

BIOREMEDIATION

3.1. In situ bioremediation

In situ bioremediation is a biological treatment

process where microorganisms metabolise organic

contaminants to stable and less toxic substances such

as carbon dioxide, methane, water and inorganic salts,

either in natural or engineered conditions (Farhadian

et al., 2008). This technique for the cleanup of

contaminants is applied at the site of contamination.

In making the decision to implement in situ

bioremediation for the cleanup of pollutants, it is

important to carry out a bioremediation feasibility

study. Answers to the following questions must be

provided: biodegradability of the contaminants,

distribution of the contaminants in the wastewater

stream, chemical reactivity of the contaminants, extent

of contamination, presence or absence of substances

that inhibit the growth of microorganisms and the

ability of the microorganisms to degrade the target

contaminants. Answers to these questions will

determine the suitability of in situ bioremediation for

a given contaminated site (Cauwenberghe and Roote,

1998; Farhadian et al., 2008).

In situ bioremediation is usually adopted when

economic considerations are of paramount importance

and since the wastewater is treated at the site of

contamination, there is no further environmental

burden posed as a result of transportation and disposal

of the wastewater. However, this strategy may not

succeed in sites that are saturated with contaminants

as a result of the inadequate mass transfer of electron

acceptors typically oxygen. This limitation is also

evident in sites with limited nutrient availability

particularly nitrogen and phosphorus. In situ

bioremediation processes include phytoremediation,

bioventing, biosparging etc (Sharma, 2012).



114

Phytoremediation: this is a bioremediation

strategy that involves the use of plants and vegetation

in the cleanup of contaminated water and soil. It is an

emerging treatment strategy which has been receiving

a lot of attention in terms of research, innovation, and

funding. It has been reported to be a more cost

effective alternative to more conventional treatment

methods (Sharma, 2012). This is the main advantage

of phytoremediation. Phytoremediation techniques

include the following:

(a) Phytoextraction is a phytoremediation process

that involves the use of plants to remove contaminants

particularly metals from the environment. The

contaminants are accumulated in the roots, shoots and

leaves of the plants. This is a cost effective method for

the remediation of low concentrations of contaminants

(Macek et al., 2000; Zhuang et al., 2007).

(b) Phytotransformation is a phytoremediation

process that involves the uptake of contaminants from

wastewater and soil and the subsequent transformation

of these contaminants to more stable and less toxic

forms (Subramanian et al., 2006).

(c) Phytodegradation is a phytoremediation process

that involves the degradation of contaminants by

taking advantage of the symbiotic relationship

between plants and microorganisms. The plants

provide the nutrients needed by the microorganisms to

survive while the microorganisms in turn provide the

necessary soil environment for the plants (Garrison et

al., 2000; Newman and Reynolds, 2004).

(d) Rhizofiltration is similar to phytoextraction in

that it is mainly used for the removal of heavy metals

from the environment. The plant to be used for

phytoremediation is grown on clean water initially

and after the development of a large root system, the

plant is transferred and planted in the polluted site

where the roots take up the polluted water and the

contaminants along with it. When the roots become

saturated with contaminants, they are harvested and

disposed off safely (Dushenkov et al., 1995; Verma et

al., 2006).

In situ bioremediation has been applied with some

success to the remediation of ground water

contaminated with various organic pollutants such as

carbon tetrachloride, trichloroethylene,

tetrachloroethylene, pentachlorophenol, gasoline,

diesel, methyl tert-butyl ether (MTBE), petroleum

hydrocarbons, xenobiotics, mono aromatic

hydrocarbons, benzene, toluene, ethylbenzene and

xylenes isomers (BTEX) and so on (Bradley and

Landmeyer, 2006; Cunningham et al., 2001; Curtis

and Lammey, 1998; Da Silva et al., 2005; Ferguson

and Pietari, 2000; Kao et al., 2006; Schmidt et al.,

1999; Widdowson, 2004; Yerushalmi et al., 1999).

Biosparging:

Biosparging is an in situ bioremediation strategy

that utilises indigenous microorganisms to degrade

contaminants. For contaminated sites that suffer from

inadequate mass transfer of oxygen, biosparging

becomes a useful strategy to adopt. It involves the

injection of air under pressure and nutrients (if

necessary) to stimulate the microorganisms and

increase the microbial activity of the indigenous

microorganisms. It can be applied in the remediation

of groundwater contaminated with petroleum

hydrocarbons as well as contaminated soil. The ease

and low cost of installing small-diameter air injection

points allows significant flexibility in the

implementation of this strategy.

3.2. Ex situ bioremediation

This is a bioremediation strategy in which the

contaminated material is removed from the site of

contamination and treated in another location. It is

most suited for the decontamination of polluted soil.

The techniques that can be used include land farming,

composting, biopiles and bioreactors (Sharma, 2012).

4. BIOREMEDIATION OF PETROLEUM AND

CRUDE OIL POLLUTED WASTEWATER

Crude oil is a complex biodegradable substance

containing a large variety of hydrocarbons (Hidayat

and Tachibana, 2012; Obahiagbon et al., 2009). It is

the major pollutant in marine environments as a result

of its release from activities such as offshore drilling,

natural oil seepage, washing of oil tankers as well as,

transportation and ruptured pipeline accidents

(Elshafie et al., 2007; Hasanuzzaman et al., 2007).

The effect of petroleum pollution on the environment

depends on the type and quantity of crude oil

involved. The water soluble fraction of the oil has

been reported to reduce the growth of biomass in the

affected environment as a result of the reduction in

dissolved oxygen, increase in turbidity, and the

toxicity of the crude oil components (Edema, 2012).

Bioremediation has been identified as the most

rational choice in the decontamination of water

polluted with hydrocarbon derivatives. It involves

treating the petroleum pollutants with hydrocarbon

degrading microorganisms possessing the kind of

enzymes required for such a process. Many

microorganisms (Pseudomonas, Escherichia coli,

clostridium, Candida, Aspergillus niger, Yeasts,

Penicillium etc) are known to grow on and utilize

petroleum and its derivatives for metabolic activities.

These organisms are able to actively degrade fractions

of petroleum oil to less toxic and stable product

(Adekunle and Adebambo, 2007; Chaillana et al.,

2004; Obahiagbon and Owabor, 2008). Khan et al.



115

(2006) isolated, identified and characterised efficient

oil degrading bacterial strains to study the effect of

crude oil loading on the growth of the bacterial

strains. They observed that Bacillus sp. showed

promise for the biodegradation of crude oil and

consequently, its application for the cleanup of oil

spills. Obahiagbon and Owabor, (2009) investigated

the bioremediation of crude oil polluted water using a

mixed microbial culture. Using biochemical oxygen

demand and total hydrocarbon content as indicators of

bioremediation, they recorded almost 100%

degradation over a treatment period of nine weeks.

Thavasi et al. (2006) investigated the potential use of

Azotobacter chroococcum a marine nitrogen fixing

hydrocarbon degrading bacterium and its

biosurfactants for the bioremediation of crude oil

polluted marine environment. Obahiagbon and

Akhabue, (2009) studied the effect of microbial count

of Pseudomonas aeruginosa stimulated with NPK

fertiliser on the biodegradation of crude oil-

contaminated water. The effluent water obtained at the

end of nine weeks of bioremediation satisfied the

treatment requirements of environmental regulatory

agencies such as the Federal Environmental Protection

Agency (FEPA) and the department of petroleum

resources (DPR) (FEPA, 1999).

Mukred et al. (2008) examined the growth of

microbial populations and effectiveness of

bioremediation of crude oil polluted water. They

reported up to 98% reduction in total hydrocarbon

content with a bacteria consortium of Acinetobacter

faecalis, Staphylococcus sp. and Neisseria elongate.

Bioremediation by natural attenuation was a major

part of the cleanup effort of the Exxon Valdez oil spill

in Prince William Sound, Alaska (Samanta et al.,

2002). Arotiowa et al. (2007) studied the

bioremediation of petroleum diesel polluted water

using an ex situ strategy. They investigated the

potential of Bacillus subtilis, Pseudomonas

aeruginosa and Penicillium funiculosum isolated from

refinery wastewater for the degradation of diesel in

water. They reported that of the three microorganisms,

Penicillium funiculosum had the highest degradation

ability. Otokunefor and Obiukwu, (2010) investigated

the treatment of refinery wastewater using a microbial

consortium consisting of Bacillus sp, Pseudomonas

sp, Staphylococcus sp, Klebsiella sp., and Citrobacter

sp. stimulated with inorganic fertilisers. They

observed that inorganic nutrient supplementation

enhanced the growth of the microbial population.

They further observed that the degree of degradation

of the target pollutant was dependent on the

concentration of fertilisers used. Gargouri et al. (2011)

applied a continuous stirred tank bioreactor for the

bioremediation of hydrocarbon laden industrial

wastewater. They developed a successful

bioremediation strategy using an efficient

acclimatised microbial consortium. After a treatment

period of 225 days, they observed that the process was

highly efficient in remediating the wastewater as seen

in the performance of the bioaugmented reactor

demonstrated by the reduction levels of up to 95 and

97.5% respectively for chemical oxygen demand

(COD) and total hydrocarbon content. They further

reported that the use of the mixed cultures resulted in

a high degradation performance for hydrocarbons

range of n-alkanes (C10–C35). Li et al. (2005) treated

wastewater produced from an oil field using Bacillus

sp. immobilised on polyvinyl alcohol (PVA). They

reported over 90% reduction in chemical oxygen

demand for continuous wastewater treatment using

immobilised bacteria cells. They further reported that

the efficiency of COD removal was improved when

the remediation medium was supplemented with a

nitrogen source such as (NH4)2SO4.

The effectiveness of the application of

bioremediation technology to contaminated sites

varies from site to site and it requires information on

the characteristics of the site, the type of contaminant

and the factors that affect the growth of pollutant

degrading microorganisms (Obahiagbon and Aluyor,

2009). Okoh, (2006) reported that a number of factors

such as the composition of crude oil contaminant,

availability of nutrients especially nitrogen and

phosphorus, and the nature of the contaminated

environment could affect the biodegradation of

petroleum hydrocarbons.

5. BIOREMEDIATION OF POLYCYCLIC

AROMATIC HYDROCARBON POLLUTED

WASTEWATER

Polycyclic aromatic hydrocarbons (PAHs) are a group

of compounds that consists of two or more benzene

rings fused together in various configurations (Woo et

al., 2009). These compounds enter the environment

through two major avenues which are natural and

anthropogenic. Natural sources of PAHs include

volcanic eruptions and forest fires (Bamforth and

Singleton, 2005). They also enter the environment

through a host of anthropogenic activities such as

incomplete combustion of gasoline and diesel in

internal combustion engines, combustion of coal and

oil for power generation, wood burning, tobacco

smoking, fumigants, and many other sources (Jia and

Batterman, 2010; Li et al. 2010; Wilson et al. 2003).

They are typically hydrophobic and have low

solubility in water; however they can be mobilised

into the aqueous phase through discharges from

industrial and domestic effluents, leaks of PAHs

containing materials, runoff from paved roads,

parking lots among other sources (Alamo-Nole et al.



116

2011; Yu et al., 2005). PAHs are highly recalcitrant

and resistant to degradation and typically persist in the

environment (Bamforth and Singleton, 2005).

Exposure to these compounds can cause serious

deleterious effects to humans such as central nervous

system and respiratory diseases as well as damage to

the kidneys and liver hence they are classified to as

priority pollutants by many countries (ATSDR 2005;

NPT 2004; Owabor and Aluyor, 2008).

The vast amounts of evidences and results reported

in literature have identified bioremediation as one of

the major ways of treating PAH contaminated

wastewater (Owabor and Aluyor, 2008). The principle

behind the use of this method is the degradation of the

pollutants into less harmful products by the

indigenous microorganisms in the wastewater. This

process is referred to as natural attenuation or intrinsic

bioremediation (Amenaghawon et al., 2013; Yu et al.,

2005). Reports have also shown that biostimulation

which is the addition of nutrients and provision of

enabling environment for biodegradation as well as

bioaugmentation which is the introduction of

exogenous microorganisms with specific

bioremediation capabilities to the remediation

medium have enhanced the efficiency of

bioremediation (Obahiagbon and Akhabue, 2009;

Cosgrove et al., 2010; Yu et al., 2005). To this end,

bioremediation has become successful over the years

in harnessing the natural activity of microorganisms.

Bioremediation of PAH contaminated wastewater can

be effected by in situ and ex situ methods. Though

more expensive than in situ methods, PAH

contaminated can also be treated using bioreactors

(Bamforth and Singleton, 2005).

The first stage in the biodegradation of PAHs

involves the incorporation of oxygen at two carbon

atoms of a benzene ring of the PAH molecule by

dioxygenase resulting in the formation of cis –

dihydrodiol. This intermediate compound undergoes

rearomatisation in the presence of dehydrogenase to

form dihydroxylated intermediates which

subsequently undergo ring cleavage (Samanta et al.,

2002). An appreciable number of PAH degrading

microorganisms including Alcaligenes denitrificans,

Mycobacterium sp., Pseudomonas putida, P.

fluorescens, P. paucimobilis, P. vesicularis, P.

cepacia, P. testosteroni, Rhodococcus sp.,

Corynebacterium venale, Bacillus cereus, Moraxella

sp., Streptomyces sp., Vibrio sp. and Cyclotrophicus

sp. have been isolated and tested for mineralisation

(Hedlund and Staley, 2001).

Yu et al. (2005) compared the efficiency of

degrading a PAH mixture by three bioremediation

strategies, namely natural attenuation,

bioaugmentation and biostimulation, in a mangrove

microcosm. The study evaluated the population sizes

of PAH degraders, the fate and mass balance of PAH

compounds under these three strategies. They reported

that natural attenuation may be a more suitable

remediation strategy for the degradation of Fluorene

and Phenanthrene while biostimulation was more

suited to the degradation of pyrene. They further

suggested that the nature of the target pollutant should

be considered when choosing a suitable remediation

strategy. Owabor and Aluyor, (2008) applied a

combination of adsorption and biodegradation in the

abatement of a PAH, anthracene. They reported

significant reductions in the concentration of

anthracene in the course of treatment. They further

observed that the percentage reduction in the

concentration of anthracene was proportional to the

temperature of activation of the adsorbent with almost

100% reduction recorded at a temperature of 900oC.

San Miguel et al. (2009) investigated the

bioremediation of naphthalene in water by bacterial

populations of Sphingomonas paucimobilis using a

new biodegradable surafactant based on poly (ε -

caprolactone). They reported up to 90% reduction in

the concentration of naphthalene in solution after 140

hours of incubation time. Furthermore, they observed

that the addition of poly caprolactone based surfactant

did not interfere with the mineralisation of

naphthalene in solution. Ting et al. (2011) studied the

biodegradation of two polycyclic aromatic

hydrocarbons, phenanthrene and pyrene by a white rot

fungus, Ganoderma lucidum. They reported that over

90 and 85 % of phenanthrene and pyrene respectively

were degraded. They further noted that the addition of

copper sulphate (CuSO4,) citric acid, gallic acid,

tartaric acid, veratryl alcohol, guaiacol, 2,2-azino-bis-

(3- ethylbenzothazoline-6-sulfonate) (ABTS)

enhanced the degradation of both PAHs and laccase

activities; whereas the supplement of oxalate, di-n-

butyl phthalate (DBP), and nonylphenol (NP)

decreased the degradation of both PAHs and inhibited

laccase production. Janbandhu andJanbandh, (2011)

isolated and characterised a high efficiency PAH

degrading microbial consortium from a 3 decade old

petrochemical refinery field. Results of

biodegradation studies revealed 100, 56.9 and 25.8%

degradation at concentrations of 100, 250 and 500

mg/L respectively within 14 days. They further

indicated that the microbial consortium holds great

promise for the bioremediation of petrochemical

contaminated environments. Lin et al. (2010) studied

the biodegradation of naphthalene in cultured medium

using bacterial strains isolated from oil refining

sludge. They recorded a degradation efficiency of

more than 99% during a treatment period of 96 hours

under optimum conditions reported as initial

naphthalene concentration of 50 mg/L, temperature of

30 ◦C, pH of 7.0, 0.2% inoculum size and C/N ratio of



117

1.0. Bamforth and Singleton, (2005) reported that

several environmental factors including pH and

temperature of the medium, nutrient and oxygen

availability and the bioavailability of the contaminant

can affect the bioremediation process by inhibiting the

growth of PAH degrading microorganisms.

6. BIOREMEDIATION OF HEAVY METALS

POLLUTED WASTEWATER

Heavy metals are elements such as cobalt, nickel,

chromium, copper, iron, lead, manganese, arsenic,

molybdenum, vanadium, strontium, and zinc having

atomic weights between 63.5 and 200.6. They are

referred to as heavy because the metal ions have

specific gravity greater than 5 g/L (Chen and Wang,

2008; Srivastava and Majumder, 2008). Living

organisms require trace amounts of some of these

metals such as cobalt, copper, iron, manganese,

molybdenum, zinc etc for specific metabolic

activities. However, some of these metals particularly,

lead; arsenic, chromium, nickel, cadmium, mercury,

antimony etc are very toxic and carcinogenic to living

organism (Srivastava and Majumder, 2008).

Increased use of metals and chemicals in process

industries has led to the generation of large quantities

of wastewater containing high concentrations of these

toxic heavy metals. The most common sources of

heavy metal polluted wastewater include

electroplating and metal finishing plants, mining,

nuclear and electronics industries (Ahluwalia and

Goyal, 2007; Costley and Wallis, 2001). Heavy metals

contamination has become a serious environmental

problem today because of the health risks posed to

humans and animals from exposure as a result of their

toxicity. This problem is heightened by the fact that

the toxicity of these metals can last for a long time in

the environment. Some of the metals can undergo

transformation from relatively less toxic forms to

more toxic forms. The concentration of some of the

metals can increase in the environment over the

course of time through bioaccumulation and

bioaugmentation and more importantly some of the

metals are toxic even in low concentrations (Chen and

Wang, 2008).

Environmental scientists and researchers are

therefore faced with the task of developing suitable

and sustainable strategies for treating heavy metals

contaminated effluents. Conventional methods for

treating heavy metals contaminated effluents include

ion exchange, precipitation, reverse osmosis,

adsorption, electro dialysis, ultra filtration, chemical

oxidation or reduction (Adebayo et al., 2012;

Ahluwalia and Goyal, 2007; Congeevaram et al.,

2007; Kadirvelu et al., 2002; Zouboulish et al., 2004).

The application of these methods is limited on

economic grounds as they are expensive when used

for the treatment of low concentration heavy metal

solutions. In addition, these methods are not very

effective and they often lead to the production of toxic

sludge and its disposal becomes another challenge

both environmentally and economically (Adebayo et

al., 2012; Chen and Wang, 2008; Schiewer and Patil,

2008). This has led to the search for effective,

economically viable and sustainable alternatives.

Over the last two decades, biological approaches to

the decontamination of heavy metals polluted sites

have been examined. Biosorption involves the use of

biological materials of microbial and plant origin to

remove heavy metals from contaminated water. These

materials are able to effectively interact with heavy

metals in solution and as a result of their unique

chemical composition; they are able to remove these

metals from solution. Biosorption is a cost effective

strategy for the treatment of high volume wastewaters

contaminated with low concentrations of heavy metals

(Chen and Wang, 2008; Costley and Wallis, 2001). In

addition, the biomaterials are usually readily available

and cheap and the process does not lead to the

generation of further waste products like conventional

physicochemical methods. Varieties of

microorganisms like bacteria, algae, yeasts, and fungi

have been used as biosorbent and studied extensively

(Chen and Wang, 2008). In the same vein, a lot of

agricultural materials have been examined for their

potential to remove heavy metals from solutions

(Babarinde et al., 2008).

Davis et al. (2003) and Wang and Chen. (2006)

reviewed the biosorption of heavy metals by algae and

Saccharomyces cerevisiae respectively. Whitehead et

al. (2005) investigated the potential of natural

attenuation for the bioremediation of acid mine

drainage. They reported iron oxidation and removal of

other important toxic metals using the indigenous

microbial populations. Costley and Wallis, (2001)

utilised a rotating biological contactor for the

bioremediation of heavy metals in a synthetic

wastewater. They used immobilised microorganisms

to treat heavy metal contaminated waters using

multiple sorption-desorption cycles. The results they

obtained suggested that the rotating biological

contactor can be used for the successful treatment of

high strength contaminated wastewaters.

Congeevaram et al. (2007) investigated the

biosorption of chromium and nickel using heavy

metal resistant fungal and bacterial populations

isolated from soil samples in an electroplating

industry environment. The isolated microorganisms

were characterised to evaluate their applicability for

heavy metal removal from industrial wastewater.

Their results indicated that extended residence times

in the stationary phase can be recommended while



118

using the fungal and bacterial chromium resistant

isolates for removing chromium. In the case of Nickel

resistant bacterial isolate, a non-extended residence

time was recommended for designing continuous-flow

completely stirred (CFCS) bioreactor so that a mid-

log phase of cellular growth can be kept during the

bioaccumulation process.

Li et al., (2004) studied uranium biosorption using

powdered biomass of lake harvested water bloom

cyanobacterium Microcystis aeruginosa. They

reported batch equilibrium time of 1 hour and an

optimum biosorption pH of 4-8. The biosorption

process was well described by the Freunlich isotherm

model. The study showed that the abundant otherwise

nuisance constituting biomass show a great promise

for application in removal of uranium from

wastewaters. Padilha et al. (2005) used waste biomass

of Sargassum sp. for the removal of copper from

simulated semiconductor wastewater using a

continuous system made up of four biomass filled

column reactors. Starting with different volumes of

different copper solutions with the same copper

concentration of 500mg/L, they reported that the

adopted treatment method resulted in an effluent

containing less than 0.5g/L of copper. Adebayo et al.

(2012) investigated the batch biosorption of lead from

aqueous solution using Streblus asper leaves. They

determined the optimum conditions of biosorption in

terms of initial lead concentration, contact time,

temperature, biosorbent dose, and pH. Results of

thermodynamic studies indicated that the biosorption

process was spontaneous, endothermic and there was

increased entropy at the solid-solution interface.

Results of isotherm studies indicated that the process

was well described by the Langmuir, Tempkin, and

Dubinin–Radushkevich isotherms. Kinetic studies

revealed that the process was well described by the

pseudo second-order, intra-particle diffusion and

Elovich kinetic models. Babarinde et al. (2012)

reported the biosorption of nickel, chromium and

cobalt on cocoyam leaves. They observed that the

biosorption process was pH dependent. The process

was well described by the pseudo-second order kinetic

model. Thermodynamically, the biosorption of each of

nickel and chromium was found to be endothermic

while that of cobalt was determined to be exothermic.

The biosorption of each metal ion was also

determined to be spontaneous and the order of

spontaneity of the biosorption process was

cobalt>nickel>chromium. They further reported

positive change in entropy for each metal and the

order of disorder was nickel>cobalt >chromium.

Lodeiro et al. (2005) investigated the potential use

of five different brown seaweeds, Bifurcaria

bifurcata, Saccorhiza polyschides, Ascophyllum

nodosum, Laminaria ochroleuca and Pelvetia

caniculata for the removal of cadmium from aqueous

solutions. They observed that the biosorption process

was relatively fast with about 90% removal of

cadmium occurring within 1 hour. Chen and Wang,

(2008) investigated the removal of lead, silver,

caesium and strontium from aqueous solution using

brewery's waste biomass. Their results revealed that

the biosorption process was rapid and was well

described by the pseudo second order kinetic and

Langmuir isotherm models. Ho, (2005) investigated

the biosorption of lead using tree fern in a baffled

agitated system. The optimum pH for lead removal

was determined to be 4.9. The pseudo second order

kinetic model sufficiently described the kinetics of the

biosorption process. Bishnoi et al. (2004) studied the

removal of chromium from aqueous solutions using

activated rice husk and activated alumina while Garg

et al. (2004) studied the removal of chromium from

aqueous solution using formaldehyde treated saw dust

and saw dust carbon activated with sulphuric acid. In

both studies, results obtained indicated that the degree

of chromium removal was proportional to the dosage

of the adsorbent used and their contact time. All the

works done by these researchers show that biosorption

utilizing microorganism and agricultural materials

offer an ideal alternative for the treatment of heavy

metal polluted water.

7. MICROORGANISMS IN BIOREMEDIATION

The ability of microorganisms to utilise natural and

synthetic pollutants as substrate for growth is a very

important quality upon which bioremediation is based.

A lot of work is still ongoing in the area of isolation,

identification and characterisation of microorganisms

and their potential for bioremediation. Reports

suggests that more work still needs to be done to

explore microbial diversity with a view to identifying

microorganisms with specific and unique qualities

vital to bioremediation.

Microorganisms indigenous to the site of

contamination have been utilised in various

bioremediation processes. Information on microbial

populations relevant to bioremediation is building up

at a fast pace as a result of recent advances in

molecular microbial ecology (Watanabe, 2001). This

has made available new tools that makes it possible to

carry out molecular analyses of microbial populations

at contaminated and bioiremediation sites.

Microorganisms can be isolated from virtually any

environmental condition as they are able to adapt even

in very extreme conditions of temperature, oxygen,

water, pH etc. The major requirement for growth is an

energy and a carbon source. The ability of

microorganisms to adapt is what makes them very

versatile in the bioremediation of contaminated sites



119

(Sharma, 2012). These microorganisms can be divided

into the following groups:

(a) Aerobic microorganisms or aerobes carry out

metabolic activities in the presence of oxygen. They

require oxygen to oxidise susbtrates through cellular

respiration. Examples of aerobic microorganisms with

the capacity for biodegradation include Pseudomonas,

Alcaligenes, Sphingomonas, Rhodococcus, and

Mycobacterium (Giavasis et al., 2006). These

microorganism have been reported to possess the

capacity to degrade organic pollutants such as

aliphatic hydrocarbons, polycyclic aromatic

hydrocarbons, pesticides etc. (Vidali, 2001).

(b) Anaerobic microorganisms can carry out

metabolic activities in the absence of oxygen. They

are not as common in use compared to aerobic

microorganisms. However, there is an increasing

interest in the use of anaerobic microorganisms for the

biodegradation of polychlorinated biphenyls (PCBs)

in river sediments and the dechlorination of the

solvent like trichloroethylene (TCE), and chloroform

(Sharma, 2012).

(c) Ligninolytic fungi such as white rot fungus like

Phanaerochaete chrysosporium have been reported to

have the ability to degrade an extremely diverse range

of recalcitrant and toxic contaminants. (Adenipekun

and Fasidi, 2005).

(d) Methylotrophs are aerobic bacteria that utilise

methane for metabolic activities. They have the ability

to degrade a wide range of organic contaminants such

as chlorinated aliphatic trichloroethylene and 1,2-

dichloroethane.

For efficient biodegradation, it is important that the

microorganism and the target contaminant be in

intimate contact. . This can be enhanced by making

use of some surfactants such as sodium dodecyl

sulphate (SDS).

8. FACTORS OF BIOREMEDIATION

A host of factors can affect the extent and

effectiveness of bioremediation. These factors are

environmental in consideration and include the

availability of nutrients and oxygen. These can be

readily manipulated using effective biostimulation

strategies. Other factors include temperature and pH

of the remediation medium. These however are not

easily controllable.

8.1. Nutrients

Microorganisms need nutrients to survive. These

nutrients are the basic building blocks of living things

and enable microorganisms to carry out metabolic

activities needed for the breakdown of contaminants

during bioremediation. All microorganisms need

carbon, nitrogen and phosphorus and some others in

lesser amounts but carbon is needed in greater

proportions than the others. These nutrients are often

present in wastewater stream but not in the proportion

required by the cells for optimum metabolic activities.

The lack of nitrogen and phosphorus limits the rate of

biodegradation. In the light of this, it becomes

important to ensure adequate supply of these

important nutrients to enhance biodegradation rates.

This is usually accomplished through biostimulation

which involves the addition of limiting nutrient such

as nitrogen and phosphorus to the wastewater stream.

Biostimulation has been reported to enhance the

biodegradation of organic pollutants (Obahiagbon et

al., 2009; Otokunefor and Obiukwu, 2010)

8.2. Oxygen

This is one of the most important requirements for

microbial degradation. Most wastewater treatment

facilities adopt aeration based treatment strategies. In

such cases, the availability of oxygen becomes a

critical factor. Oxygen is generally necessary for the

initial degradation of oil, and subsequent reactions

may also require direct incorporation of oxygen.

Typically, 3 to 4 parts of dissolved oxygen are

necessary to completely oxidize 1 part of oil into

carbon dioxide and water (Giavasis et al., 2006).

Though anaerobic degradation of oil in wastewater

can occur, it is however in very small degrees. For oil

spills on the ocean surface, oxygen is not usually a

factor that limits the rate of biodegradation as there is

plentiful supply of oxygen close to the surface of the

ocean. However, inadequate supply of oxygen limits

the extent of biodegradation. This is the reason why it

takes longer to degrade oil that has sunk below the

surface of the water.

8.3. Temperature

Temperature is another important environmental

factor that affects the rate of bioremediation. In the

same way that chemical reactions are affected by

temperature, biochemical reactions upon which the

process of bioremediation is based are also

temperature dependent. A temperature increase results

in a decrease in viscosity of liquid organic pollutants,

consequently affecting the degree of distribution and

increasing diffusion rates of the compounds.

Typically, an increase in temperature favours the

biodegradation reaction. However, above a certain

optimum temperature which is organism specific, the

activity of the microorganism begins to slow and they

subsequently die. Hence it is important to identify this

optimum and ensure that bioremediation operations

are maintained at that temperature.



120

REFERENCES

Adebayo MA, Adebayo AA, Adediji JF, Adebayo TO

(2012). Isotherm, Kinetic, and Thermodynamic

Studies of Lead (II) Biosorption by Streblus

asper. Pacific Journal of Science and

Technology. 13(2): 283-293.

Adekunle AA, Adebambo OA (2007). Petroleum

hydrocarbon utilization by fungi isolated from

Detarium senegalense seeds. Journal of

American Science, 3(1): 69-76.

Adenipekun CO, Fasidi IO (2005). Bioremediation of

oil-polluted soil by Lentinus subnudus, a

Nigerian white-rot fungus. African journal of

biotechnology, 4(8): 796-798.

Agarry SE, Durojaiye AO, Solomon BO (2008).

Microbial degradation of phenols: a review.

International Journal of Environment and

Pollution, 32(1): 12-28.

Ahluwalia SS, Goyal D (2007). Microbial and plant

derived biomass for removal of heavy metals

from wastewater. Bioresource technology, 98:

2243-2257.

Akpor OB, Muchie M (2010). Bioremediation of

polluted wastewater influent: Phosphorus and

nitrogen removal. Scientific Research and

Essays, 5(21): 3222-3230.

Alamo-Nole LA, Perales-Perez O, Roman-Velazquez

FR (2011). Sorption study of toluene and xylene

in aqueous solutions by recycled tires crumb

rubber. Journal of hazardous materials, 185(1):

107-111.

Amenaghawon NA, Aisien FA (2011). Investigation

of the effect of influent substrate concentration

on the dynamic response of an activated sludge

reactor system by dynamic modelling and

simulation. Journal of Nigerian Institution of

Production Engineers, 13: 9-20.

Amenaghawon AN, Asegame PA, Obahiagbon KO

(2013). Potential Application of Urea and NPK

15: 15: 15 Fertilizers as Biostimulants in the

Bioremediation of Domestic

Wastewater. American Journal of

Environmental Protection, 1(4): 91-95.

Ardern E, Lockett WT (1914). Experiments on the

oxidation of sewage without the aid of

filters. Journal of the Society of Chemical

Industry, 33(10): 523-539.

ATSDR. (2005). Toxicological Profile for

Naphthalene, 1-methylnaphthalene, and 2-

methylnaphthalene; U.S. Department of Health

and Human Services, Agency for Toxic

Substances and Disease Registry: Atlanta, GA,

USA, 1-291.

Babarinde A, Babalola JO, Adegoke J, Osundeko AO,

Ibidapo TJ, Nwabugwu CA, Ogundimu OF

(2012). Biosorption of Ni(II), Cr(III), and Co(II)

from Aqueous Solutions using Cocoyam

(Colocasia esculenta) Leaf: Kinetic,

Equilibrium, and Thermodynamic Studies.

Pacific Journal of Science and Technology.

13(2): 272-282.

Bamforth SM, Singleton I (2005). Bioremediation of

polycyclic aromatic hydrocarbons: current

knowledge and future directions. Journal of

chemical technology and biotechnology, 80:

723-736.

Bishnoi NR, Bajaj M, Sharma N, Gupta A (2004).

Adsorption of Cr(vi) on activated rice husk

carbon and activated alumina. Bioresource

Technology, 91: 305–307.

Bradley PM, Landmeyer JE (2006). Low-temperature

MTBE biodegradation in aquifer sediments with

a history of low, seasonal ground water

temperatures. Ground Water Monitoring and

Remediation, 26(1): 101–105.

Cauwenberghe LV, Roote DS (1998). In situ

bioremediation. Technology overview report,

Ground water remediation technologies analysis

center, p.4.

Chaillana F, Flècheb A, Burya E, Phantavonga Y-hui,

Saliot A, Oudot J (2004). Identification and

biodegradation potential of tropical aerobic

hydrocarbon-degrading microorganisms.

Research in Microbiology. 155(7): 587-595.

Chen C, Wang J (2008). Removal of Pb2+

, Ag+, Cs

+

and Sr2+

from aqueous solution by brewery’s

waste biomass. Journal of hazardous materials,

151: 65-70.

Congeevaram S, Dhanarani S, Park J, Dexilin M,

Thamaraiselvi K (2007). Biosorption of

chromium and nickel by heavy metal resistant

fungal and bacterial isolates. Journal of

hazardous materials, 146: 270-277.

Cosgrove L, McGeechan PL, Handley PS, Robson

GD (2010). Effect of biostimulation and

bioaugmentation on degradation of polyurethane

buried in soil. Applied and environmental

microbiology, 76(3): 810-819.

Costley SC, Wallis FM (2001). Bioremediation of

heavy metals in a synthetic wastewater using a

rotating biological contactor. Water research,

35(15): 3715-3723.

Crawford RL, Crawford DL Eds. (2005).

Bioremediation: Principles and Applications,

Biotechnology Research Series: 6, Cambridge

University Press, UK.

Cunningham JA, Rahme H, Hopkins GD, Lebron C,

Reinhard M (2001). Enhanced in situ

bioremediation of BTEX-contaminated



121

groundwater by combined injection of nitrate

and sulphate. Environmental Science

Technology, 35: 1663–1670.

Curtis F, Lammey J (1998). Intrinsic remediation of a

diesel fuel plume in Goose Bay, Labrador,

Canada. Environmental Pollution, 103: 203–

210.

Da Silva ML, Ruiz-Aguilar GM, Alvarez PJ (2005).

Enhanced anaerobic biodegradation of BTEX-

ethanol mixtures in aquifer columns amended

with sulfate, chelated ferric iron or nitrate.

Biodegradation, 16: 105–114.

Davis TA, Volesky B, Mucci A (2003). A review of

the biochemistry of heavy metal biosorption by

brown algae. Water Research 37: 4311–4330.

Dushenkov V, Nanda Kumar PBA, Motto H, Raskin I,

(1995). Rhizofiltration: the use of plants to

remove heavy metals from aqueous streams.

Environmental Science and Technology, 29:

1239- 1245.

Edema (2012). Effects of Crude Oil Contaminated

Water on the Environment, Crude Oil

Emulsions-Composition Stability and

Characterization, In: Crude Oil Emulsions-

Composition Stability and Characterization

(Manar El-Sayed Abdul-Raouf, Eds), InTech,

169-180

EPA (1997). Wastewater treatment manuals: Primary,

secondary and tertiary treatments.

Environmental Protection Agency Ardcavan,

Wexford, Ireland.

EPA (1999). Use of Monitored Natural Attenuation at

Superfund, RCRA Corrective Action, and

Underground Storage Tank Sites. Directive

number 9200.4-17P. Washington, D.C.: EPA,

Office of Solid Waste and Emergency

Response.

El Fantroussi S, Agathos SN (2005). Is

bioaugmentation a feasible strategy for pollutant

removal and site remediation? Current Opinion

in Microbiology, 8(3): 268-275.

Elshafie A, Alkindi AY, Al-Busaidi S, Bakheit C,

Albahry SN (2007). Biodegradation of crude oil

and n-alkanes by fungi isolated from Oman.

Marine Pollution Bulletin, 54: 1692-1696.

Farhadian M, Vachelard C, Duchez D, Larroche C

(2008). In situ bioremediation of monoaromatic

pollutants in groundwater: A review.

Bioresource technology, 99: 5296-5308.

Federal Environmental Protection Agency (FEPA)

(1997). Guidelines and `Standards for

Environmental Impact Assessment in Nigeria,

87-95.

Ferguson JF, Pietari JMH (2000). Anaerobic

transformations and bioremediation of

chlorinated solvents. Environmental Pollution

107: 209–215.

Fono LJ, Kolodziej EP, Sedlak DL (2006).

Attenuation of wastewater-derived contaminants

in an effluent-dominated river. Environmental

science & technology, 40(23): 7257-7262.

Forsyth JV, Tsao YM, Bleam RD (1995).

Bioremediation: when is augmentation needed?

In: Hinchee, RE., Fredrickson J, Alleman BC

(Eds.), Bioaugmentation for Site Remediation.

Battelle Press, Columbus, 1–14.

Garg VK, Gupta R, Kumar R, Gupta RK (2004).

Adsorption of chromium from aqueous solution

on treated sawdust. Bioresource Technology,

92: 79–81.

Gargouri B, Karray F, Mhiri N, Aloui F, Sayadi S

(2011). Application of a continuously stirred

tank bioreactor (CSTR) for bioremediation of

hydrocarbon-rich industrial wastewater

effluents. Journal of hazardous materials,

189(1): 427-434.

Garrison AW, Nzengung VA, Avants JK, Ellington

JJ, Jones EW, Rennels D, Wolfet NL, (2000).

Phytodegradation of p,p’ - DDT and the

enantiomers of o,p’ – DDT. Environmental

Science and Technology, 34:1663-1670

Gentry T, Rensing C, Pepper I (2004). New

approaches for bioaugmentation as a

remediation technology. Critical Reviews in

Environmental Science and Technology, 34(5):

447-494.

Giavasis I, Harvey LM, McNeil B (2006). The effect

of agitation and aeration on the synthesis and

molecular weight of gellan in batch cultures of

Sphingomonas paucimobilis. Enzyme and

microbial technology, 38(1): 101-108.

Godos ID, Blanco S, Garcia-Encina PA, Becares E,

Munoz R (2009). Long-term operation of high

rate algal ponds for the bioremediation of

piggery wastewaters at high loading rates.


Gonzalez,YS (2011). Bioremediation of oil spills.

B.Eng Degree Thesis, Department of Chemical

Engineering and fuels. Polytechnic University

of Madrid, 16-19.

Grady Jr CL, Daigger GT, Love NG, Filipe CD,

Grady CP (2011). Biological wastewater

treatment (No. Ed. 3). IWA Publishing.

Guyer JP (2011). Introduction to secondary

wastewater treatment. Continuing Education and

Development, Inc., Course No: C04-022, p. 18.

Hassanuzzaman M, Ueno A, Ito H, Ito Y, Yamamoto

Y, Yumoto I, Okuyama H (2007). Degradation

of long-chain n-alkanes (C36-C40) by

Pseudomonas aeruginosa strain WatG.



122

International Biodeterioration and

Biodegradation Journal, 59: 40-43.

Hedlund BP, Staley JT (2001). Vibrio cyclotrophicus

sp. nov., a polycyclic aromatic hydrocarbon

(PAH)-degrading marine bacterium.

International journal of systematic and

evolutionary microbiology, 51(1): 61-66.

Hidayat A, Tachibana S (2012). Bioremediation of

Aliphatic Hydrocarbon in Three Types of Crude

Oil by Fusarium sp. F092 under Stress with

Artificial Sea Water. Journal of Environmental

Science and Technology, 5(1): 64-73.

Ho YS (2005). Effect of pH on lead removal from

water using tree fern as the sorbent. Bioresource

Technology, 96: 1292–1296.

Janbandhu A, Fulekar MH (2011). Biodegradation of

phenanthrene using adapted microbial

consortium isolated from petrochemical

contaminated environment. Journal of hazardous

materials, 187: 333-340.

Jia C, Batterman S (2010). A Critical Review of

Naphthalene Sources and Exposures Relevant to

Indoor and Outdoor Air. International Journal of

Environmental Research and Public Health, 7:

2903-2939.

Kadirvelu K, Senthilkumar P, Thamaraiselvi K,

Subburam V (2002). Activated carbon prepared

from biomass as adsorbent: elimination of Ni(II)

from aqueous solution. Bioresource Technology

81: 87–90.

Kao CM, Huang WY, Chang LJ, Chen TY, Chien

HY, Hou F (2006). Application of monitored

natural attenuation to remediate a petroleum-

hydrocarbon spill site. Water Science and

Technology, 53: 321–328.

Kim SJ, Choi DH, Sim DS, Oh YS (2005). Evaluation

of bioremediation effectiveness on crude oil-

contaminated sand. Chemosphere, 59(6): 845-

852.

Khan K, Naeem M, Arshed MJ, Asif M (2006).

Extraction and characterization of oil degrading

bacteria. Journal of Applied Sciences, 6: 2302-

2306.

Kshirsagar AD (2013). Bioremediation of wastewater

by using microalgae: an experimental study.

International Journal of Life Sciences

Biotechnology and Pharma Research, 2(3): 339-

346.

Lendvay JM, Löffler FE, Dollhopf M, Aiello MR,

Daniels G, Fathepure BZ, Adriaens P (2003).

Bioreactive barriers: a comparison of

bioaugmentation and biostimulation for

chlorinated solvent remediation. Environmental

Science & Technology, 37(7): 1422-1431.

Li PF, Mao ZY, Rao SJ, Wang XM, Min MZ, Qiu

LW, Liu ZL (2004). Biosorption of uranium by

lake harvested biomass from cyanobacterium

bloom. Bioresource Technology, 94: 193–195

Li Q, Kang C, Zhang C (2005). Waste water produced

from an oilfield and continuous treatment with

an oil-degrading bacterium. Process

Biochemistry, 40(2): 873-877.

Li Z, Mulholland JA, Romanoff LC, Pittman EN,

Trinidad DA, Lewin MD, Sjödin A (2010).

Assessment of non-occupational exposure to

polycyclic aromatic hydrocarbons through

personal air sampling and urinary

biomonitoring. Journal of Environmental

Monitoring, 12(5): 1110-1118.

Lin C, Gan L, Chen ZL (2010). Biodegradation of

naphthalene by strain Bacillus fusiformis

(BFN). Journal of hazardous materials, 182:

771-777.

Lodeiro P, Cordero B, Barriadh JL, Herrero R,

deVicente MES (2005). Biosorption of

cadmium by biomass of brown marine

macroalgae. Bioresource Technology, 96: 1796–

1803.

Macek T, Mackova M, Kas J (2000). Exploitation of

plants for the removal of organics in

environmental remediation. Biotechnology

Advances, 18: 23-34.

Margesin R, Schinner F (2001). Bioremediation

(Natural Attenuation and Biostimulation) of

Diesel-Oil-Contaminated Soil in an Alpine

Glacier Skiing Area. Applied and

Environmental Microbiology, 67(7): 3127-3133.

Morill JA, Antizar-Ladislao B, Monteoliva-Sanchez

M, Ramos-Cormenzana A, Russell NJ (2009).

Bioremediation and biovalorisation of olive-mill

wastes. Applied Microbiology and


Mrozik A, Piotrowska-Seget Z (2010).

Bioaugmentation as a strategy for cleaning up of

soils contaminated with aromatic

compounds. Microbiological Research, 165(5):

363-375.

Mukred AM, Hamid AA, Hamzah A, Yusoff WMW

(2008). Development of three bacteria

consortium for the bioremediation of crude

petroleum oil in contaminated water. Online

Journal of Biological Sciences. 8(4): 73-79.

Newman LA, Reynolds CM (2004). Phytodegradation

of organic compounds. Current Opinion in


Nikolopoulou M, Kalogerakis N (2009).

Biostimulation strategies for fresh and

chronically polluted marine environments with

petroleum hydrocarbons. Journal of Chemical

Technology and Biotechnology, 84(6): 802-807.

NTP (2004). Report on Carcinogens - Naphthalene,

Eleventh Edition; U.S. Department of Health



123

and Human Services, Public Health Service,

National Toxicology Program: 1-2.

Nyyssönen M, Kapanen A, Piskonen R, Lukkari T,

Itävaara M (2009). Functional genes reveal the

intrinsic PAH biodegradation potential in

creosote-contaminated groundwater following in

situ biostimulation. Applied microbiology and


Obahiagbon KO, Akhabue CE (2009). Effect of

microbial count of P. aeruginosa on

biodegradation of crude oil contaminated

water. Petroleum Science and

Technology, 27(13): 1402-1412.

Obahiagbon KO, Aluyor EO (2009). Comparison of

the efficiency of sodium nitrate and sodium

nitrite as nutrients in the bioremediation of

petroleum hydrocarbon polluted water.

Scientific Research and Essay, 4(8): 728-732.

Obahiagbon KO, Owabor CN (2008). Biotreatment of

Crude Oil polluted water using mixed microbial

populations of P. aureginosa, Penicillium

notatum, E. coli and Aspergillus niger.

Proceedings of the 2nd

International Conference

on Engineering Research and Development:

Innovations, Benin City, Nigeria.

Obahiagbon KO, Owabor C.N (2009). Bio-Treatment

of Crude Oil Polluted Water Using Mixed

Microbial Populations of P. Aureginosa,

Penicillium Notatum, E. Coli and Aspergillus

Niger. Advanced Materials Research, 62: 802-

807.

Okoh A, Ajisebutu S, Babalola G, Trejo-Hernandez M

(2001). Potential of Burkholderia cepacia RQ1

in the biodegradation of heavy crude oil.

International Microbiology, 4(2): 83-87.

Olu-Arotiowa OA, Aremu MO, Alade AO (2007). Ex

situ bioremediation of diesel polluted

wastewater in tropical hot climate. Asian

Journal of Information Technology 6(9):961-

963.

Otokunefor TV, Obiukwu C (2010). Efficacy of

inorganic nutrients in bioremediation of a

refinery effluent. Scientia Africana, 9(1): 111-

125.

Owabor CN, Aluyor EO (2008). Application of

adsorbent as a novel technique during

biodegradation of a polycyclic aromatic

hydrocarbon (anthracene). African journal of


Padilha FP, de Franca FP, daCosta ACA (2005). The

use of waste biomass of Sargassum sp. for the

biosorption of copper from simulated

semiconductor effluents. Bioresource

Technology, 96: 1511–1517.

Qin G, Gong D, Fan MY (2013). Bioremediation of

petroleum-contaminated soil by biostimulation

amended with biochar. International

Biodeterioration & Biodegradation, 85: 150-

155.

Röling WF, Milner MG, Jones DM, Lee K, Daniel F,

Swannell RJ, Head IM (2002). Robust

hydrocarbon degradation and dynamics of

bacterial communities during nutrient-enhanced

oil spill bioremediation. Applied and

Environmental Microbiology, 68(11): 5537-

5548.

Samanta SK, Singh OV, Jain RK (2002). Polycyclic

aromatic hydrocarbons: environmental pollution

and bioremediation. Trends in biotechnology,

20(6): 243-248.

San Miguel V, Peinado C, Catalina F, Abrusci C

(2009). Bioremediation of naphthalene in water

by Sphingomonas paucimobilis using new

biodegradable surfactants based on poly (ɛ-caprolactone). International Biodeterioration &

Biodegradation, 63(2): 217-223.

Schiewer S, Patil SB (2008). Modeling the Effect of

pH on Biosorption of Heavy Metals by Citrus

Peels. Journal of Hazardous Materials. 157:8-

17.

Schmidt LM, Delfino JJ, Preston JF, Laurent GS

(1999). Biodegradation of low aqueous

concentration pentachlorophenol (PCP)

contaminated groundwater. Chemosphere, 38:

2897–2912.

Sharma S (2012). Bioremediation: Features, Strategies

and applications. Asian Journal of Pharmacy

and Life Science, 2(2): 202-213.

Smith AE, Hristova K, Wood I, Mackay DM, Lory E,

Lorenzana D, Scow KM (2005). Comparison of

biostimulation versus bioaugmentation with

bacterial strain PM1 for treatment of

groundwater contaminated with methyl tertiary

butyl ether (MTBE). Environmental health

perspectives, 113(3): 317-322.

Srivastava NK, Majumder CB (2008). Novel

biofiltration methods for the treatment of heavy

metals from industrial wastewater. Journal of

hazardous materials, 151: 1-8.

Subramanian M, Oliver DJ, Shanks JV (2006). TNT

Phytotransformation Pathway Characteristics in

Arabidopsis: Role of Aromatic Hydroxylamines.

Biotechnology Programme, 22: 208 -216.

Thavasi R, Jayalakshmi S, Balasubramanian T, Banat

IM (2006). Biodegradation of crude oil by

nitrogen fixing marine bacteria Azotobacter

chroococcum. Research Journal of

Microbiology 1(5): 401-408.

Ting WTE, Yuan SY, Wu SD, Chang BV (2011).

Biodegradation of phenanthrene and pyrene by

Ganoderma lucidum. International



124

Biodeterioration & Biodegradation, 65(1): 238-

242.

Thomassin-Lacroix E, Eriksson M, Reimer K, Mohn

W (2002). Biostimulation and bioaugmentation

for on-site treatment of weathered diesel fuel in

Arctic soil. Applied microbiology and

biotechnology, 59(4-5): 551-556.

Thompson IP, Van Der Gast CJ, Ciric L, Singer AC

(2005). Bioaugmentation for bioremediation: the

challenge of strain selection. Environmental

Microbiology, 7(7): 909-915.

Verma P, George K, Singh H, Singh S, Juwarkar A,

Singh R (2006). Modeling rhizofiltration: heavy

metal uptake by plant roots. Environmental

Modeling and Assessment, 11: 387-394.

Vidali M (2001). Bioremediation. An overview. Pure

and Applied Chemistry, 73(7): 1163-1172.

Vogel TM (1996). Bioaugmentation as a soil

bioremediation approach. Current Opinion and

Biotechnology 7: 311–316.

Wang JL, Chen C (2006). Biosorption of heavy metal

by Saccharomyces cerevisiae: a review.

Biotechnology Advances, 24: 427–451.

Watanabe K (2001). Microorganisms relevant to

bioremediation. Current opinions in

biotechnology, 12: 237-241.

Whitehead PG, Hall G, Neal C, Prior H (2005).

Chemical behaviour of the Wheal Jane

bioremediation system. Science of the Total

Environment, 338: 41–51.

Widdowson MA, (2004). Modeling natural

attenuation of chlorinated ethenes under

spatially varying redox conditions.

Biodegradation, 15: 435–451.

Woo OT, Chung WK, Wong KH, Chow AT, Wong

PK (2009). Photocatalytic oxidation of

polycyclic aromatic hydrocarbons:

Intermediates identification and toxicity

testing. Journal of hazardous materials, 168(2):

1192-1199.

Wilson NK, Chuang JC, Lyu C, Menton R, Morgan

MK (2003). Aggregate exposures of nine

preschool children to persistent organic

pollutants at day care and at home. Journal of

Exposure Science and Environmental

Epidemiology, 13(3): 187-202.

Yerushalmi L, Manuel MF, Guiot SR (1999).

Biodegradation of gasoline and BTEX in a

microaerophilic biobarrier. Biodegradation, 10:

341–352.

Yu KSH, Wong AHY, Yau KWY, Wong YS, Tam

NFY (2005). Natural attenuation, biostimulation

and bioaugmentation on biodegradation of

polycyclic aromatic hydrocarbons (PAHs) in

mangrove sediments. Marine pollution

bulletin, 51(8): 1071-1077.

Zhuang P, Yang QW, Wang HB, Shu WS (2007).

Phytoextraction of heavy metals by eight plant

species in the field.Water. Air and Soil

Pollution, 184: 235-242.

Zouboulish AI, Loukidou M, Matis KA (2004).

Biosorption of Toxic Metals from Aqueous

Solutions by Bacteria Strains Isolated from

Metal Polluted Soils. Process Biochemistry, 39:

909-916.



©2014 IJSRPUB

125

Supplementation of Novel Solid Carbon Source Prepared from Dried Attached-

Growth Biomass for Bioremediation of Wastewater Containing Nitrogen

Jun-Wei Lim, Mohammed J.K. Bashir*, Choon-Aun Ng, Xinxin Guo

Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul



Abstract. The main objective of this study is to validate the feasibility of using dried attached-growth biomass from the

polyurethane (PU) foam cubes as a solid carbon source for the enhancement of denitrification process in the intermittently

aerated moving bed sequencing batch reactor (IA-MBSBR). The IA-MBSBR packed with PU foam cubes coated with dried

attached-growth biomass could maintain approximately 80% of total nitrogen (TN) removal efficiency for 8 consecutive cycles

of operation. Subsequently, the exhausted stored carbon source within the PU foam cubes could be replenished by merely

drying the fresh attached-growth biomass formed when the cubes were used as a carbon source. Thus, the reuse/recycle of

biomass-coated PU foam cubes is possible, making it a sustainable solid carbon source for the enhancement of denitrification

process in bioremediating wastewater containing nitrogen-cum-low COD/N via IA-MBSBR.

Keywords: Bioremediation, Wastewater, Solid Carbon Source

1. INTRODUCTION

It is undeniable that water is an indispensable element

for survival of all living creatures. Although the

Earth's surface is virtually covered with 70% of water,

its presence is not limitless. Of all the water, 97%

consist of salt water which is unacceptable for the

direct human consumption. To top it off, of the

remaining 3% of potable water, only about 15% is

easily accessible, e.g., rivers, streams, creeks, ponds,

etc., and about 85% is found in ice floes and glaciers,

neither of which are readily accessible (Sills, 2003).

Owing to the tremendous increase of human growth in

recent decades, water scarcity has emerged as one of

the dire issue for communities across the country. In

United States, almost all the region of the country has

experienced water shortages in the last five years

(USEPA, 2008). As per United Nations, every day

approximately 4400 children under the age of five die

because of diseases cause by contaminated water

ingestion and sanitation (Ghaitidak and Yadav, 2013).

In a third world’s population, one in every six persons

has no access to clean water within a kilometer of

reach (Ghaitidak and Yadav, 2013). In parts of Asia

and Africa, it was estimated that the people under a

threshold of water stress, i.e., accessible of renewable

water resources <1700 m3/person/year, would surge to

three billion by 2025 (FAO, 2012; WHO, 2012;

Ghaitidak and Yadav, 2013).

Globally, although the storage of potable water is

getting lesser due to the excessive consumption, the

problem of potable water contamination particularly

the groundwater and surface water with nitrate is

inevitable. The nitrate pollution is mainly caused by

the intensive use of nitrogen-based fertilizers in

agriculture and irrigation with ammonia-rich effluents

discharged by wastewater treatment plants and

improper sewage treatment and disposal. In some

industrial activities such as fossil fuels combustion

and fertilizers, explosive, glass, plastics and foods

productions also can contribute to nitrate pollution

(Robinson-Lora and Brennan, 2009; Wang et al.,

2013). In nature, the concentrations of nitrate in

groundwater are usually less than 2 mg/L (Mueller et

al., 1995). However, in contaminated areas, nitrate

concentrations can exceed 200 mg/L (ITRC, 2002).

The Water Quality Assessment program of the US

Geological Survey reports that nitrate is the pollutant

that most frequently exceeds its standard limits

(Squillace et al., 2002). The main health effect

associates to the ingestion of water contaminated with

high concentration of nitrate are the occurrence of

methemoglobinemia notably in infants or “blue-baby

syndrome”. Some studies have demonstrated that

nitrate can be endogenously reduced to nitrite, which

can then undergo nitrosation reactions in the stomach

with amines and amides to form various N-nitroso

compounds, most of which are extremely

carcinogenic (WHO, 2004; Yang et al., 2007). On that

account, appropriate standards have been set by

various agencies. The USEPA (2000) has set the

maximum contaminant levels of 10 and 1 mg/L for

nitrate-nitrogen (NO3--N) and nitrite-nitrogen (NO2

--

N), respectively, in drinking water. The World Health

Organization and European Economic Community

have set the standards of 11.3 mg/L for NO3--N which

are later adopted as national standard for drinking

water by most of the countries in the world (Wang,

2013). Therefore, for the sake of fulfilling the

standards requirement, the discharge wastewater

containing nitrate must be stringently treated before

releasing to the environment in order to minimize the

possibilities of contamination of potable water with

nitrate.


Lim et al.

Supplementation of Novel Solid Carbon Source Prepared from Dried Attached-Growth Biomass for Bioremediation of

Wastewater Containing Nitrogen

126

Various technologies have been used to treat

wastewater containing nitrate. These include ion-

exchange, electrodialysis, photocatalytic reduction of

nitrate, reverse osmosis, etc (Kesore et al., 1997;

Schoeman and Steyn, 2003; Kim and Benjamin, 2004;

Yang et al., 2013). Of these, biological denitrification

has been proven to be the most cost effective and

environmentally sound technique by many researchers

particularly in treating large quantity of wastewater

containing nitrate (Robinson-Lora and Brennan, 2009;

Wang and Wang, 2009). Biological denitrification is a

natural process that is part of the nitrogen cycle, and is

commonly exploited in the wastewater treatment plant

for the removal of nitrate. The term denitrification

was first used in France in 1886 to describe the use of

nitrate and nitrite (oxidized nitrogen) to biodegrade

substrate (Gerardi, 2002). The process of

denitrification can be accomplished by a wide range

of facultative anaerobes which make up an

approximately 80% of the bacteria in the activated

sludge system. The largest numbers of facultative

anaerobes that are capable of performing the

denitrification process are from Alcaligenes, Bacillus

and Pseudomonas genera (Gerardi, 2002). Besides,

the facultative anaerobes that denitrify are also termed

by several names including denitrifying bacteria,

denitrifiers, heterotrophs and organotrophs. During

the denitrification process, nitrate is reduced to nitrite

and subsequently to nitrogen gas by means of the

action of denitrifying bacteria in accordance with the

following sequence (Gerardi, 2002):

Nitrate ion (NO3-) → Nitrite ion (NO2

-) → Nitric

oxide (NO) → Nitrous oxide (N2O) → Ntrogen gas

(N2)

Each step of the denitrification process is

fundamentally regulated by the specialized reductase

enzymes of the denitrifying bacteria. The nitrate is

initially reduced to nitrite by the nitrate reductase

enzyme. This enzyme is a highly soluble membrane-

bound molybdoprotein which is only produced in the

presence of nitrate and its concentration synthesized is

directly proportional to the concentration of nitrate

(Downey, 1966; Bryan, 1981; Payne, 1985; Gerardi,

2002; Gardner, 2008). In the second step of

denitrification process, the nitrite is further reduced to

nitric oxide which is regulated by nitrite reductase

enzyme found in the periplasm of the denitrifying

bacteria (Gerardi, 2002; Kumar and Lin, 2010). The

nitric oxide is then swiftly reduced to nitrous oxide by

the nitric oxide reductase enzyme, a membrane bound

protein (Bryan, 1981; Payne, 1985; Gerardi, 2002;

Kumar and Lin, 2010). In the final step of

denitrification process, the nitrous oxide reductase

enzyme, a periplasmic copper-containing protein,

reduces the nitrous oxide to nitrogen gas before it is

being released to the atmosphere (Payne, 1985;

Gerardi, 2002; Kumar and Lin, 2010).

The prerequisite for the commencement of

denitrification process is the availability of strictly

anoxic environment. This is because the energy

harvested through the aerobic respiration of substrate

is greater than the energy obtained through anoxic

respiration for the growth of denitrifying bacteria. The

examples of quantity of energy produced via the

aerobic and anoxic oxidations of glucose are shown in

Eqs. 1 and 2, respectively (Gerardi, 2002):

C6H12O6 +6O2 → 6CO2 + 6H2O + 686 kcal (1)

C6H12O6 + 4.8NO3- + 4.8H

+ → 6CO2 + 2.4N2 +

8.4H2O + 636kcal (2)

In terms of energy point of view, i.e., more

negative Gibbs free-energy value, oxygen is a more

favorable electron acceptor than either nitrate or

nitrite during the respiration of similar substrate

(Lens, 2005), vindicating the dominant selection of

oxygen species over the oxidized nitrogen. In

addition, the synthesis and activity of all

denitrification enzymes are also adversely affected in

the presence of oxygen with nitrous oxide reductase

being the most sensitive denitrification enzymes and it

is inhibited by the dissolved oxygen (DO)

concentrations of less than 0.2 mg/L (Kumar and Lin,

2010). Thus, it is important to ensure the anoxic

environment exists with the redox potential of less

than +50 mv (negligible detection of DO

concentration in the measured environment) so as to

kindle the denitrification process (Gerardi, 2002). This

obligation is crucial since the presence of oxygen has

a direct impact on the usage of substrate in which the

oxygen is selected as an electron acceptor during the

substrate respiration instead of nitrate, leading to the

wastage of the added substrate for the enhancement of

denitrification process. Also, the objective of

removing nitrate from the wastewater remains

unfulfilled as the denitrification process is retarded in

the aerobic environment.

Another factor which is also playing an important

role during the denitrification process is the substrate

used as the reducing agent in reducing nitrate to

nitrogen gas. In this case, organic carbon source is

commonly used as an electron donor and its

characteristic has an imperative effect on denitrifying

bacteria performing the denitrification process. In

general, the denitrifying bacteria will use organic

carbon source found in the wastewater in performing

the denitrification process. This category of carbon

source is termed as internal carbon source. However,

in treating wastewaters containing low COD/N ratio,

e.g., supernatants from sludge digesters and

stabilization ponds as well as pretreated industrial



127

wastewaters by anaerobic fermentation, carbon source

is frequently added in order to stimulate the

denitrifiation process; and this category of carbon

source is termed as external carbon source. The

external carbon source can be further subdivided

depending on their physical phases, namely liquid

carbon source and solid carbon source. In recent

years, the research on the solid carbon source used for

the denitrification process has gained increasing

momentum among the researchers and this carbon

source can be classified based on their origin, namely

synthetic and natural polymer carbon sources. Figure

1 shows all the organic carbon source categories as

above-mentioned in a tree diagram.

Fig. 1: Tree diagram presenting category of carbon sources used for denitrification process

1.1. Solid Carbon Source for Denitification Process

Liquid carbon sources such as ethanol, methanol,

acetate or glucose are normally added during the

denitrification process. However, the disadvantages of

this approach are the risk of overdosing which would

deteriorate the effluent quality, leading to the

requirement of using a sophisticated and costly

process control (Hiraishi and Khan, 2003; Zhou et al.,

2009; Shen and Wang, 2011). In recent years, the use

of solid carbon sources as an alternative to liquid

carbon sources has gained increasing momentum of

interests among the researchers (Robinson-Lora and

Brennan, 2009; Wang and Wang, 2009; Zhou et al.,

2009; Zhou et al., 2009a; Shen and Wang, 2011; Fan

et al., 2012). The solid carbon sources packed in the

bioreactors perform two important tasks, namely to

serve as a reducing agent in denitrification process

and to act as a support media for biofilm formation

(Wang and Wang, 2009; Zhou et al., 2009a). The

presence of constant carbon sources and anoxic zones

within the deeper layers of the biofilm as well as in

the porous structure of the solid carbon sources would

ensure a stable reduction of the oxidized nitrogen

(Walters et al., 2009; Wang and Wang, 2009; Zhou et

al., 2009a). Hence, the use of an expensive and

sophisticated system control can be avoided as the

addition of liquid carbon sources is no longer

necessary. Generally, two types of solid carbon

sources, namely synthetic and natural polymers have

been studied. Synthetic polymers include

polycaprolactone (PCL) (Boley et al., 2000; Honda

and Osawa, 2002; Zhou et al., 2009), polylactic acid

(PLA) (Fan et al., 2012), polyhydroxyalkanoates

(PHA) (Boley et al., 2000; Hiraishi and Khan, 2003)

and bionolle (Boley et al., 2000) whereas, natural

polymers include wheat straw (Soares and Abeliovich,

1998; Fan et al., 2012), cotton (Rocca et al., 2005),

biodegradable meal box (Wang and Wang, 2009) and

crab-shell chitin (Robinson-Lora and Brennan, 2009).

Since the rate of denitrification is closely related to

the biodegradability of the solid carbon source

(Hiraishi and Khan, 2003), the use of natural polymers

which are more likely to be biodegraded than

synthetic polymers is expected to attain higher

denitrification rates than synthetic polymers. In fact,

Wang and Wang (2009) had proved that the rate of

denitrification by using biodegradable meal box was

higher than PCL.

Fan et al. (2012) had also revealed that faster

biofilm development and higher denitrification rate

could be achieved when wheat straw was utilized as a

solid carbon source as compared to PLA. From an

economic point of view, the relatively high cost of

using synthetic polymers such as PCL and bionolle

limits its extensive application especially in treating

Lim et al.



128

large quantity of wastewater (Walters et al., 2009;

Chu and Wang, 2011). Zhou et al. (2009) estimated

that a denitrification system using

polyhydroxybutyrate (PHB) costs almost ten times

higher than a traditional system using methanol. As

natural polymers such as cotton and wheat straw are

easily available and they can achieved higher

denitrification rate than synthetic polymers, the

application of solid carbon source using natural

polymers have gained priority over that of synthetic

polymers.

Nevertheless, in order to maintain the efficiency of

denitrification, the addition of fresh natural polymers

after certain period of operation is necessary. Soares

and Abeliovich (1998) reported that a weekly addition

of fresh wheat straw could be seen preventing the

deterioration of denitrification performance.

Moreover, for some natural polymers such as crab-

shell chitin, the content of soluble components such as

volatile fatty acids and chloride, sulphate and

ammonium ions are very high. Robinson-Lora and

Brennan (2009) illustrated that an initial flushing

period of 3 days was needed before the crab-shell

chitin could be used as a solid carbon source to reduce

the concentrations of soluble components in the

treated effluent. Consequently, as a result of frequent

addition or replacement of such solid carbon sources,

more time is required to be spent on flushing. In

addition, post-treatment system such as trickling sand

filter or powdered activated carbon adsorption have to

be installed to remove colour and organic matter

released particularly from fresh natural polymers

(Soares and Abeliovich, 1998; Rocca et al., 2005).

Thus, it is essential to find an alternative natural

material which can serve as a solid carbon source

without releasing soluble components.

Qiao et al. (2008) reported that the potential of

attached-growth biomass in oxidizing NH4+-N was

lower than the suspended-growth biomass as the

Ammonium Oxidizing Bacteria (AOB)’s mobility was

more restricted in the attached state, leading to

lessened contact possibility between AOB and NH4+-

N. For this reason, the attached-growth biomass

formed onto and into the support media in the

bioreactor could be utilized to serve as a solid carbon

source for the denitrification process. These days, the

research activities have only focused on the reuse of

suspended-growth biomass or sludge as a carbon

source (Ra et al., 2000; Kampas et al., 2007; Soares et

al., 2010). The use of dried attached-growth biomass

as a solid carbon source has not been thoroughly

reported in the literature. Therefore, the possibility of

using dried attached-growth biomass formed onto and

into the support media as a solid carbon source for the

denitrification process in the bioreactor deserves more

research attentions. On that account, the main

objective of this study is to investigate the feasibility

of using this dried attached-growth biomass as a solid

carbon source for the denitrification process

enhancement in an intermittently aerated moving bed

sequencing batch reactor (IA-MBSBR).

2. METHODOLOGY

2.1. Set-up of Bioreactor and Operation

A sequencing batch reactor (SBR) was initially set-up

and operated with a cycle time of 24 h in the

following sequencing periods: instantaneous FILL, 0

h; REACT, 12 h; SETTLE, 1.5 h; DRAW, 1 h and

IDLE, 9.5 h. The REACT phase was operated with

cyclical intermittent aeration (IA) strategy which

began with 1 h of aeration period followed by 1 h of

non-aeration period. The activated sludge collected

from municipal sewage treatment plant was cultured

in this SBR and fed with synthetic wastewater

simulating the municipal wastewater composition

with NH4+-N and COD concentrations of

approximately 48 and 200 mg/L, respectively. The

instantaneous addition of adequate amount of ethanol

solution to serve as a carbon source was carried out at

the beginning of the last non-aeration period in every

cycle to reduce the oxidized nitrogen to N2. At the end

of the REACT period, mixed liquor was wasted to

maintain the sludge age of suspended-growth biomass

at 40 days. During the DRAW period, the supernatant

or treated effluent was drawn out with an exchange

volume of the reactor being retained at 70.3%. The

residual settled solids in the reactor were left to rest

throughout the IDLE period in preparation for the next

cycle.

Upon achieving the quasi-steady state, the SBR

with integrated IA strategy was converted to IA-

MBSBR by packing with 8% (v/v) of 8-mL

polyurethane (PU) foam cubes as the support media.

Figure 2 shows the laboratory set-up of IA-MBSBR

packed with PU foam cubes. After the PU foam cubes

were cultured for 4 consecutive cycles in the IA-

MBSBR, they were taken out of the reactor and

replaced by a new batch of 8% (v/v) of 8-mL PU

foam cubes. The PU foam cubes with the attached-

growth biomass was dried in the oven at 60 oC and

weighed. The dried foam cubes were then introduced

back into the reactor whilst the batch in the reactor

was taken out for drying and weighing. This process

was repeated until the weight of the PU foam cubes

with dried attached-growth biomass reached a

constant value and this cubes with dried attached-

growth biomass were later used in the studies

hereafter this section.



129

Fig. 2: Laboratory set-up of IA-MBSBR packed with PU foam cubes, (a) side view and (b) top view of the bioreactor.

2.2. Assessment of IA-MBSBR Performance Using

PU foam Cubes with Attached-Growth Biomass

The evaluation of IA-MBSBR using PU foam cubes

with dried attached-growth biomass involved 4-cycle,

6-cycle, 8-cycle, 10-cycle and 12-cycle operational

modes and every mode was repeated twice. In the 4-

cycle mode, the batch of PU foam cubes which had

achieved a constant dried weight was put back into the

IA-MBSBR for 4 cycles. At the end of the operational

mode, the foam cubes were taken out for drying at 60 oC and could be reused again as recycled PU foam

cubes with dried attached-growth biomass to complete

another 4-cycle operational mode. During the first and

fourth cycles of this operational mode, time courses of

nitrogen species (NH4+-N, NO2

--N and NO3

--N)

concentrations during the REACT period were

analyzed. Similarly, for the 6-, 8-, 10- and 12-cycle

modes, the PU foam cubes with dried attached-growth

biomass of constant dried weight were also replaced

by the recycled PU foam cubes with dried attached-

growth biomass at the end of the respective mode. The

performance data were collected for the first and last

cycles of each operational mode.

2.3. Investigation of the Optimum IA Strategy and

Selection of the Optimum Operational Mode

The study as described in Section 2.2 allows the

optimum operational mode to be ascertained based on

the IA strategy of consecutive 1 h of aeration period

followed by 1 h of non-aeration period (1:1 IA

strategy). This operational mode was adopted for the

determination of the optimum IA strategy in the

experiment conducted in this section. The time

courses of the concentrations of nitrogen species

during the REACT period of the IA-MBSBR without

the instantaneous addition of ethanol solution were

determined using the 1:1, 1:2 and 1:3 IA strategies.


The total weight of attached-growth biomass in the

PU foam cubes saturated with attached-growth

biomass was found to be 58 g (constant dried weight

of total attached-growth biomass). The Scanning

Electron Microscope (SEM) images show that the

attached-growth biomass shrank and adhered closely

to the surface of PU foam cubes when it was dried to

constant weight as shown in Figure 3 (c). Figure 3 (b)

shows that the fresh attached-growth biomass could

be instantly formed on the surface of dried biomass-

coated PU foam cubes after only one cycle in the IA-

MBSBR.

Lim et al.



130

Fig. 3: SEM images of (a) fresh PU foam cube, (b) fresh attached-growth biomass formed on the surface of dried biomass-

coated PU foam cubes after one cycle in the IA-MBSBR and (c) dried biomass-coated PU foam cubes all with 150X

magnification and captured at WD of 7 mm using EHT and signal of 5.00 kV and SE2, respectively.

3.1. Optimum Operational Mode Based on 1:1 IA

Strategy

The selection was based on the performance of IA-

MBSBR packed with dried biomass-coated PU foam

cubes during the REACT period under different

operational modes based on 1:1 IA strategy. The

concentration profiles of the nitrogen species during

the REACT period of the IA-MBSBR operating at 8-

cycle, 10-cycle and 12-cycle modes and using the 1:1

IA strategy with the addition of ethanol solution for

complete denitrification are shown in Figure 4.

Fig. 4: Profiles of nitrogen species concentrations during the REACT period for (a) first cycle of the 8-cycle operational mode,

(b) last cycle of the 8-cycle operational mode, (c) last cycle of the 10-cycle operational mode and (d) last cycle of the 12-cycle

operational mode under 1:1 IA strategy.

It should be noted that the results were the same

irrespective of whether dried biomass-coated PU foam

cubes of constant dried weight in the first run or

recycled dried biomass-coated PU foam cubes in

subsequent runs were used. Figure 4 (a) shows the

concentration profiles of nitrogen species for the first

cycle of the 8-cycle operational mode and it was

found that those of all the studied operational modes

of first cycle were the same. Figure 4 (b) to Figure 4

(d) show the concentration profiles of the monitored

parameters for the last cycles of the 8-cycle, 10-cycle

and 12-cycle operational modes, respectively. The

results for the last cycles of the 4-cycle and 6-cycle

operational modes were basically the same as those of

the 8-cycle mode. Figure 4 shows that, irrespective of

the first or last cycle of the operational modes, NH4+-



131

N was only completely removed in the fourth aeration

period. During the first three non-aeration periods, the

oxidation of NH4+-N was hindered due to the limited

supply of DO in the mixed liquor resulting in a

relatively flat region of the NH4+-N concentration

profile. In all cases, after the complete removal of

NH4+-N, the concentration of NO3

--N was seen

decreasing gradually until the last non-aeration period

due to the occurrence of the denitrification process.

The NO3--N was then removed rapidly when the

ethanol solution was added during the last non-

aeration period to accelerate the denitrification

process. However, the rate of decrease of NO3--N

concentration was less for the last cycles of the 10-

cycle and 12-cycle modes [Figures. 4-c and 4-d)

compared to that of the 8-cycle mode (Figure 4-b).

Figure 5 shows the calculated total nitrogen (TN)

removal percentages in the REACT period prior to the

addition of ethanol solution for various operational

modes. It was observed that the TN removal

efficiency was fairly constant up to the 8-cycle

operational mode but decreased sharply after that.

This indicates that the stored carbon source in the

prepared biomass-coated PU foam cubes had been

depleted to the level which could not sustain the same

TN removal efficiency after 8 consecutive cycles of

operation. Thus, replacement with new batch of

biomass-coated PU foam cubes after every 8

consecutive cycles in the IA-MBSBR was necessary

to avoid the deterioration of the denitrification

process. The depleted stored carbon source in the

biomass-coated PU foam cubes could be replenished

and reused again by drying the attached-growth

biomass formed when these cubes were used in the

IA-MBSBR. Of all the solid carbon sources reported

in the literature which were used to improve the

denitrification process (Soares and Abeliovich, 1998;

Walters et al., 2009; Fan et al., 2012), it was not

possible to reuse the solid carbon sources as these

biodegradable materials would be eventually

consumed by the microorganisms in long term

operation period.

Fig. 5: TN removal efficiency for different operational modes under 1:1 IA strategy. Error bars indicate standard deviations.

3.2. Optimum IA Strategy Based on 8-Cycle

Operational Mode

Figure 6- a to 6-c show the concentration profiles of

nitrogen species during the REACT period of the IA-

MBSBR operating under 1:1, 1:2 and 1:3 IA

strategies, respectively. For all the three IA strategies,

the results of the monitored parameters concentration

profiles in the first cycle were basically similar to

those in the last cycle of the 8-cycle operational mode

Figure. 6. It was observed that the solid carbon source

in the prepared biomass-coated PU foam cubes was

insufficient to completely reduce the oxidized

nitrogen in cases involving 1:1 and 1:2 IA strategies.

As a result, the NO3--N was detected in the treated

effluents of these cases (Table 1). In the IA-MBSBR

operated with 1:1 IA strategy, two aeration periods

after the complete removal of NH4+-N (Figure. 6-a)

was unnecessary. In addition, the application of

excess aeration period would bring a negative impact

on the reduction of NO3--N as the NO2

--N formed was

likely to be oxidized to NO3--N again rather than

being reduced to N2. As a consequence, the highest

effluent NO3--N concentration was detected with this

IA strategy resulting in the lowest percentage of TN

removal as compared to the other IA strategies (Table

1). Although the percentage of TN removal was the

highest in the IA-MBSBR operated with 1:3 IA

strategy (Table 1), the removal of NH4+-N was

incomplete due to inadequate aeration period Figure

6-c). In comparison, NH4+-N was completely removed

during the last aeration period in the IA-MBSBR with

1:2 IA strategy Figure -b; thus, avoiding the

subsequent unnecessary aeration period. In this case,

the NO3--N formed could be easily removed by adding

adequate ethanol solution during the last non-aeration

period. Therefore, the IA-MBSBR under 1:2 IA

strategy appears to be the optimum alternative.

Lim et al.



132

Fig. 6: Profiles of nitrogen species concentrations during the REACT period for the last cycle of the 8-cycle operational mode

under (a) 1:1, (b) 1:2 and (c) 1:3 IA strategies.

Table 1: Concentrations of NH4

+-N and NO3

--N in the treated effluent and percentages of TN removal in the IA-MBSBR

operated with 8-cycle mode.

Parameter IA strategy

1:1 1:2 1:3

Effluent NH4+-N concentration

(mg/L)

0.0 0.0 6.4

Effluent NO3--N concentration

(mg/L)

10.7 7.1 0.0

TN removal (%) 71 79 83

3.3. Proposed Methodology for Real Wastewater

Treatment Plant

This study has shown the potential of utilizing the

dried attached-growth biomass from the PU foam

cubes as a novel solid carbon source in enhancing the

denitrification process. When the PU foam cubes with

dried attached biomass was introduced into the reactor

to serve as the solid carbon source, the attached-

growth biomass would start forming on the surface of

the biomass-coated PU foam cubes; which slowly

limited the penetration of suspended-growth biomass

containing denitrifiers as well as oxidized nitrogen

into the interior layers of the cubes. Consequently, the

solid carbon source located in deeper layers of the

biomass-coated PU foam cubes was foreseen to be not

completely exploited due to the limitation of

denitrification process. Thus, in a practical scale, the

procedure for the preparation of the dried attached-

growth biomass can be simplified by air drying or

drying under the sunlight rather than oven drying

which will reduce the cost of treatment. This is

because only the surface of the wet biomass-coated

PU foam cubes is needed to be dried as merely the

superficial portion of the dried biomass is

predominantly used as a solid carbon source during

the denitrification process. In addition, the application

of air drying or drying under the sunlight also permits

the real scale reactor to be packed with higher

percentage of PU foam cubes as the drying process of

large quantity of wet biomass-coated PU foam cubes

can be easily carried out in an open space instead of

limited space when it is performed in the oven. The

presence of higher percentage of packing volume of



133

biomass-coated PU foam cubes is more preferable as

it can further enhance the denitrification process due

to the availability of more solid carbon source.

Furthermore, as merely the superficial portion of the

PU foam cubes with dried attached biomass is

predominantly used as a solid carbon source during

the denitrification process, attaining a constant weight

of the dried biomass-coated PU foam cubes is

unnecessary in the real application. In fact, once the

surface of the fresh PU foam cubes is covered by the

attached-growth biomass, which takes about a week of

culturing, these PU foam cubes could be dried and are

ready to serve as a solid carbon source. Therefore,

only a short start-up period is required which added its

advantages of using PU foam cubes with dried

attached biomass as a solid carbon source for the

denitrification process enhancement.

Generally, low DO concentration and accessibility

of carbon source are found to be the crucial factors

affecting the denitrification process. Since the IA

strategy was applied in this study, the denitrification

process was deemed more preferable to take place

during the non-aeration period as the high DO

concentration in aeration period would inhibit the

reduction of oxidized nitrogen. When the dried

biomass-coated PU foam cubes were introduced into

reactor, it would serve as the solid carbon source to

facilitate the denitrification process. The freshly

formed attached-growth biomass was as well capable

of harvesting the carbon source from these biomass-

coated PU foam cubes to ensure the commencement

of the denitrification process. Once the supplied of

carbon source was exhausted, the formed attached-

growth biomass onto and into the biomass-coated PU

foam cubes could be dried to replenish the depleted

carbon source and subsequently reused/recycle in the

IA-MBSBR. Thus, lesser amount of external carbon

source was required to completely remove the

oxidized nitrogen as large portion of this oxidized

nitrogen was already being reduced using the carbon

source from the prepared dried biomass-coated PU

foam cubes which served as a supplementary solid

carbon source for the denitrification process.

4. CONCLUSIONS

The dried attached-growth biomass from the PU foam

cubes could be exploited as a novel solid carbon

source to enhance the denitrification process in the

IA-MBSBR. With the packing volume of 8% (v/v) of

prepared dried biomass-coated PU foam cubes, the

TN removal efficiency of approximately 80% could

be achieved up to 8 consecutive cycles of operation

when 1:2 IA strategy during the REACT period was

adopted during the treatment of low COD/N

containing wastewater. Once the stored carbon source

within the prepared biomass-coated PU foam cubes

was exhausted due to the continuously used in every

cycle to maintain the denitrification process, the

formed attached-growth biomass could be dried again

to enable the reuse/recycle of biomass-coated PU

foam cubes.

REFERENCES

Boley A, Muller WR, Haider G (2000). Biodegradable

Polymers as Solid Substrate and Biofilm

Carrier for Denitrification in Recirculated

Aquaculture Systems. Aquacultural


Bryan BA (1981). Physiology and Biochemistry of

Denitrification. John Wiley and Sons, Inc, New

York.

Chu L, Wang J (2011). Nitrogen Removal Using

Biodegradable Polymers as Carbon Source and

Biofilm Carriers in a Moving Bed Biofilm

Reactor. Chemical Engineering Journal, 170:

220-225.

Downey RJ (1966). Nitrate Reductase and Respiratory

Adaptation in Bacillus Stearothermophilus.

Journal of Bacteriology, 9: 634-641.

Fan Z, Hu J, Wang J (2012). Biological Nitrate

Removal Using Wheat Straw and PLA as

Substrate. Environmental Technology, 33:

2369-2374.

FAO (2012). Coping with Water Scarcity, an Action

Framework for Agriculture and Food Security.

Food and Agriculture Organization of the

United Nations (FAO).

Gardner LM (2008). Denitrification Enzyme Activity

as an Indicator of Nitrate Loading in a Wetland

Receiving Diverted Mississippi River Water.

Thesis - Graduate Faculty of the Louisiana

State University and Agricultural and

Mechanical College, Ohio State

University,USA.

Gerardi MH (2002). Nitrification and Denitrification

in the Activated Sludge Process, John Wiley

and Sons, Inc, New York.

Ghaitidak DM, Yadav KD (2013). Characteristics and

Treatment of Greywater - A Review.

Environmental Science and Pollution Research,

20: 2795-2809.

Hiraishi A, Khan ST (2003). Application of

Polyhydroxyalkanoates for Denitrification in

Water and Wastewater Treatment. Applied

Microbiology and Biotechnology, 61: 103-109.

Honda Y, Osawa Z (2002). Microbial Denitrification

of Wastewater Using Biodegradable

Polycaprolactone. Polymer Degradation and

Stability, 76: 321-327.

Lim et al.



134

ITRC (2002). A Systematic Approach to In Situ

Bioremediation in Groundwater Including

Decision Trees on In Situ Bioremediation for

Nitrate, Carbon Tetrachloride and Perchlorate.

Interstate Technology and Regulatory Council

In Situ Bioremediation Team.

Kampas P, Parsons SA, Pearce P, Ledoux S, Vale P,

Churchley J, Cartmell E (2007). Mechanical

Sludge Disintegration for the Production of

Carbon Source for Biological Nutrient

Removal. Water Research, 41: 1734-1742.

Kesore K, Janowski F, Shaposhnik VA (1997). Highly

Effective Electrodialysis for Selective

Elimination of Nitrates from Drinking Water.

Journal of Membrane Science, 127: 17-24.

Kim J, Benjamin MM (2004). Modeling a Novel Ion

Exchange Process for Arsenic and Nitrate

Removal. Water Research, 38: 2053-2062.

Kumar M, Lin JG (2010). Co-existence of Anammox

and Denitrification for Simultaneous Nitrogen

and Carbon Removal-strategies and issues.

Journal of Hazardous Matererials, 178: 1-9.

Lens P, Westermann P, Haberbauer M, Moreno A

(2005). Biofuels for Fuel Cells: Renewable

Energy from Biomass Fermentation. IWA

Publishing, TJ International (Ltd.), Padstow,

United Kingdom.

Mueller DK, Hamilton PA, Helsel DR, Hitt KJ,

Ruddy BC (1995). Nutrients in Ground Water

and Surface Water of the United States; An

Analysis of Data Through 1992. USGS Water-

Resources Investigations Report: 95-4031, US

Geological Survey.

Payne WJ (1985). Diversity of Denitrifiers and Their

Enzyme. Plenum Press, New York.

Qiao S, Kawakubo Y, Koyama T, Furukawa K

(2008). Partial Nitritation of Raw Anaerobic

Sludge Digester Liquor by Swim-bed and

Swim-bed Activated Sludge Processes and

Comparison of Their Sludge Characteristics.

Journal of Bioscience and Bioengineering, 106:

433-441.

Ra CS, Lo KV, Shin JS, Oh JS, Hong BJ (2000).

Biological Nutrient Removal with an Internal

Organic Carbon Source in Piggery Wastewater

Treatment. Water Research, 34: 965-973.

Robinson-Lora MA, Brennan RA (2009). The Use of

Crab-shell Chitin for Biological Denitrification:

Batch and Column Tests. Bioresource

Technolohy, 100: 534-541.

Rocca CD, Belgiorno V, Meric S (2005). Cotton-

supported Heterotrophic Denitrification of

Nitrate-rich Drinking Water with a Sand

Filtration Post-treatment. Water SA, 31: 229-

236.

Schoeman JJ, Steyn A (2003). Nitrate Removal with

Reverse Osmosis in a Rural Area in South

Africa. Desalination, 155: 15-26.

Shen Z, Wang J (2011). Biological Denitrification

Using Cross-linked Starch/PCL Blends as Solid

Carbon Source and Biofilm Carrier.


Sills AD (2003). Earth Science the Easy Way.

Barron’s Educational Series, Inc, Hauppauge

New York.

Soares A, Kampas P, Maillard S, Wood E, Brigg J,

Tillotson M, Parsons SA, Cartmell E (2010).

Comparison between Disintegrated and

Fermented Sewage Sludge for Production of a

Carbon Source Suitable for Biological Nutrient

Removal. Journal of Hazardous Materials, 175:

733-739.

Soares MIM, Abeliovich A (1998). Wheat Straw as

Substrate for Water Denitrification. Water

Research, 32: 3790-3794.

Squillace PJ, Scott JC, Moran MJ, Nolan BT, Kolpin

DW (2002). VOC’S, Pesticides, Nitrate and

Their Mixtures in Groundwater Used for

Drinking Water in the United States.

Environmental Science and Technology, 36:

1923-1930.

USEPA (2000). Drinking Water Standards and Health

Advisories. United States Environmental

Protection Agency (USEPA).

USEPA (2008). Water Supply and Use in the United

States. United States Environmental Protection

Agency (USEPA).

Walters E, Hille A, He M, Ochmann C, Horn H

(2009). Simultaneous

Nitrification/Denitrification in a Biofilm Airlift

Suspension (BAS) Reactor with Biodegradable

Carrier Material. Water Research, 43: 4461-

4468.

Wang X, Xing L, Qiu T, Han M (2013). Simultaneous

Removal of Nitrate and Pentachlorophenol

from Simulated Groundwater Using a

Biodenitrification Reactor Packed with

Corncob. Environmental Science and Pollution

Research, 20: 2236-2243.

Wang XM, Wang JL (2009). Removal of Nitrate from

Groundwater by Heterotrophic Denitrification

Using the Solid Carbon Source. Science in

China Series B, 52: 236-240.

WHO (2004). Guidelines for Drinking-water Quality.

3rd edn. Chapter 12: Chemical Fact Sheets 417-

420.

WHO (2012). The Challenge of Extending and

Sustaining Services. UN-Water Global Annual

Assessment of Sanitation and Drinking-Water

(GLAAS) 2012 Report.



135

Yang CY, Wu DC, Chang CC (2007). Nitrate in

Drinking Water and Risk of Death from Colon

Cancer in Taiwan. Environment International,

33: 649-653.

Yang T, Doudrick K, Westerhoff P (2013).

Photocatalytic Reduction of Nitrate Using

Titanium Dioxide for Regeneration of Ion

Exchange Brine. Water Research, 47: 1299-

1307.

Zhou H, Zhao X, Wang J (2009). Poly (ɛ -

Caprolactone) as Substrate for Water

Denitrification. International Journal of

Environment and Pollution, 38: 349-359.

Zhou H, Zhao X, Wang J (2009a). Nitrate Removal

from Groundwater Using Biodegradable

Polymers as Carbon Source and Biofilm

Support. International Journal of Environment

and Pollution, 38: 339-348.



©2014 IJSRPUB

136


Treatment Systems

Chapter 6: Wastewater Treatment by Membrane

Techniques



©2014 IJSRPUB

137

Supported Liquid Membrane in Wastewater Treatment

Tjoon Tow Teng1*, Amir Talebi

1 and Govindaraju Muthuraman

2

1School of Industrial Technology, Universiti Sains Malaysia, 11800 Penang, Malaysia

2Department of Chemistry, University of Madras, Presidency College, Chennai 600 005, India


Abstract. Supported liquid membrane (SLM) has been investigated by many researchers recently. The application of SLM in

wastewater treatment, pharmaceutical purposes, metallurgy, biological process and medical science has made it an interesting

topic for researchers globally. Different methods of SLM systems such as simple, facilitated and active transport have been

described and discussed, as well as solute transport mechanism from the aqueous feed phase through the polymeric

microporous support into the aqueous stripping phase. SLM stability, selectivity, recovery and reuse are the other topics which

in this chapter have been discussed and described. For future trend, green supported liquid membrane and the application of

ionic liquids in SLM system have been suggested.

Keywords: Supported Liquid Membrane, Wastewater Treatment

1. INTRODUCTION

Liquid membrane (LM) is water insoluble

(immiscible) liquid with the feed and stripping phases.

Once a specific solute of a mixture moves through the

liquid membrane from feed phase towards the

stripping phase, extraction can be achieved.

Liquid membrane systems are being investigated

widely in various fields such as chemistry (organic,

inorganic and analytical), biotechnology, biomedical

technology, wastewater treatment, etc. (Muthuraman

et al., 2009; Talebi et al., 2013). LM can be applied

for different purposes such as selective removal and

recovery of heavy metals, separation of aromatics

from hydrocarbons, antibiotics purification,

purification of aromatics such as benzene, xylene and

toluene, protein extraction using aqueous two-phase

systems, dyes and pigments removal, metallurgical

purifications, etc. (Chang et al., 2011).

The term solvent extraction refers to the

distribution of a solute between two immiscible liquid

phases, which are somehow in contact with each

other. For this process, the International Union of Pure

and Applied Chemistry (IUPAC) recommends the

term liquid-liquid extraction (LLE) (Rydberg et al.,

2004). The mechanism of the contact of two

immiscible liquid phases with each other makes three

different liquid membrane types: Bulk Liquid

Membrane (BLM), Supported Liquid Membrane

(SLM) and Emulsion Liquid Membrane (ELM).

1.1. Bulk Liquid Membrane

Bulk liquid membrane (BLM) contains two bulk

aqueous phases (feed and stripping) separated by a

bulk organic and water immiscible liquid phase. In its

simplest type, the extraction and partition take place in

U-tube (high density solvent, Figure. 1A) or H-tube

(low density solvent, Figure 1B.) configuration

1.2. Supported Liquid Membrane

In supported liquid membrane (SLM) a thin

micropourous filter is installed as a support between

feed and receiving (stripping) phases. The support is

impregnated by an organic carrier (mobilizer) or ionic

liquid to modify the extraction process. Fig.2. shows a

schematic of an SLM reactor in which the LM is

sandwiched between feed and stripping phase.

1.3. Emulsion Liquid Membrane

In 1968, Li invented a different type of liquid

membrane in which the stripping phase was

emulsified in an immiscible liquid membrane (Li,

1968). ELM may be in water-organic-water (W/O/W)

or organic-water-organic (O/W/O). In emulsion liquid

membrane (ELM) mass transfer takes place by

dispersion of emulsion in the feed solution. Fig. 3

shows a configuration of ELM.

Teng et al.


138

2. LIQUID-LIQUID EXTRACTION

Liquid-Liquid Extraction (LLE) is considered as the

principal and first step for liquid membrane

application. Recently, separation techniques such as

solvent extraction or liquid-liquid extraction (LLE),

liquid membrane, etc. have been widely used in the

separation of heavy metals from aqueous solutions

(Cox, 2004; Talebi et al. 2012). The principle of

liquid-liquid extraction (LLE) is based on distribution

of a solute between an aqueous phase and an organic

phase (Thornton, 1992). LLE is a process where a

solute can distribute itself in a certain ratio between

immiscible solvents and extraction process depends

on its mass transfer rate (Lee et al., 2000).

LLE is considered as the preliminary investigation

of LM application in order to find the best and

optimum condition for various parameters affecting

the extraction efficiency, such as equilibrium pH

(pHeq), mixing time, extractant concentration, salt

concentration, and organic-to-aqueous phase ratio

(O/A).


Chapter 6: Wastewater Treatment by Membrane Techniques

139

If for example solute A dissolved in aqueous phase

distributes into the organic phase in a way to reach

equilibrium, the distribution coefficient is

D = [A]org/[A]aq (1)

where [A]org is the solute concentration in organic

phase and [A]aq is the solute concentration in aqueous

phase at equilibrium stage.

The percentage of extraction (%E) is given by:

%E= [D/(1+D)]*100 (2)

3. Diffusion and Transport Mechanism

Considering LLE as a liquid membrane basic

principle, diffusion transport plays a major role in

liquid membrane technique. Diffusion flux J (M,

g/cm2/s) for a particular species (S) is basically the

amount of that particular species diffuses through unit

area per unit time.

Based on Fick’s first law:

J = −D

(3)

where D is diffusion coefficient and d is membrane

thickness. For steady state diffusion across a thin

membrane, only one dimension can be considered. By

applying the mass transfer coefficient k

k = D/d (4)

Equation (3) can be simplified and integrated to give

J = − k ([Ss] – [Sf]) = k ([Sf] – [Ss]) (5)

where [Sf] and [Ss] are the concentrations of the

species in feed phase and stripping phase, respectively

(Kislic, 2010).

Fig. 4 shows a schematic of solute transfer from

the aqueous feed phase, through the LM and extracted

in the aqueous stripping phase in BLM system. The

basic concept of liquid membrane can be described as

follows:

Feed phase and receiving (stripping) phases are

divided by an immiscible organic solution either with

or without carrier (modifier). Different types of LM

are based on this organic phase installation and

position: if it is a thin layer of emulsion sphere

(globule), then it is called ELM, while in BLM the

organic phase is an independent bulk layer and in

SLM, the impregnated microporous membrane

support is considered as the organic phase.

The diffusion of a solute from the feed phase, into

the LM and consequently the extraction of it in

stripping phase, is the main direction of liquid

membrane extraction process. There is not any routine

and unique model for all different types of LM

diffusions (Kislic, 2010); however, the basic

diffusion-chemical reactions in LM can be described

based on diffusion, partition and interdiffusion of the

solute in different stages of LM process, as shown in

Fig. 5.

First the diffusion inside the aqueous feed phase:

Jf = kf ([Sf] – [Sfm1]) (6)

This is the first step of all three LM configurations

(BLM, SLM and ELM). [Sf] is solute concentration at

feed phase and [Sfm1] is solute concentration at the

feed phase-LM interface.

After this step, solute partition takes place due to

thermodynamic conditions and the solute

concentration changes from [Sfm1] to [Sfm2].

The second step is after solute partition and

complexations, which is solute diffusion through the

liquid membrane (interdiffusion):

Jm = km ([Sfm2] – [Sms1]) (7)

where [Sfm2] is solute concentration at feed phase-LM

interface and [Sms1] is the solute concentration at

LM−stripping phase interface.

Similar to the first step, after this step the solute

concentration changes from [Sms1] to [Sms2] as a result

of different thermodynamic conditions.

The third step is diffusion through the stripping phase:

Js = ks ([Sms2] – [Ss]) (8)

where [Ss] is the solute concentration at the stripping

phase.

Teng et al.


140

Fig. 4: Solute transport configuration from aqueous feed phase to aqueous stripping phase through BLM

3.1. Liquid Membrane Types Based on Transport

Mechanism

Three different types of liquid membranes (BLM,

SLM and ELM) can be divided by three different

types based on transport mechanism:

3.1.1. Simple Transport

The solubility of the solute in the liquid membrane

plays an important role. Because the LM is not

impregnated with any modifier then no chemical

reaction occurs between the solute and the LM and

hence it is in the same form in stripping phase as it has

been in feed phase and LM. The permeation stops at

equilibrium level (Schlosser et al., 1993; Schlosser

and Sabolova, 1999; Wodzki and Nowaczyk, 2002).

Fig.6. shows the mechanism of solute transport

through the LM; [Sf] is solute concentration in the

feed phase, [Sm] is the solute concentration in the

liquid membrane and [Ss] is the solute concentration

in the stripping phase. In simple transport mechanism,

the solute permeation stops when concentration

equilibrium is reached.



141

Fig. 5: Solute concentration [S] Vs distance (H) profile from the aqueous feed phase through the LM to the aqueous stripping

phase

Fig. 6: Simple transports of Species (S) from Feed phase (F), through LM (M) to Stripping phase (S)

3.1.2. Facilitated Transport

The most important transport mechanism in LM

technique is based on carrier facilitated transport.

Partitioning, complexations and diffusion are the

major parts of facilitated transport mechanism. This

type of LM accelerates the flux rate and extraction

process. However, subsequent mechanisms like

coupled counter transport or coupled co-transport can

be derived from the facilitated transport mechanism.

Fig.7. shows the schematic of the solute facilitated

transport through LM. The first stage is the solute (S)

partitioning in LM, diffusing from feed phase to feed-

LM interface due to chemical potential, then a

chemical reaction between the solute and the carrier

(C) at the interface occurs to form solute-carrier

complex (S=C) and the complex moves through the

LM (due to the solubility of the solute) towards the

LM-stripping interface (reverse reaction and

partitioning the solute in stripping phase and

decomplexation) and stripping in the stripping phase

due to different thermodynamic conditions at LM-

stripping phase interface and diffusion to the stripping

phase (SA formation due to chemical reaction with

anion (A−) in stripping phase (Kislic, 2010).

Teng et al.


142

Fig. 7: Facilitated transport of species S through LM

3.1.3. Active Transport

Active transport is based on oxidation-reduction

process and is recommended for the time that only one

and no other species is transported and is suitable for

selectivity. Catalytic reactions and also biochemical

conversions at the LM interface are the other major

parameters in active transport mechanism. Fig.8.

shows the active transport process where S2+

→ S+

reduction takes place in the feed phase, S=C=A2

complexation in LM (with the carrier C) and S+→S

2+

oxidation in the stripping phase.

Fig. 8: Active transport of species S through LM

4. SUPPORTED LIQUID MEMBRANE (SLM)

In SLM a thin microporous hydrophobic (or

hydrophilic) support is impregnated by an organic

solvent (mixed with an extractant or suitable carrier)

and is placed between aqueous feed and stripping

phases. For laboratory purpose, the most common

design is flat sheet geometry but due to low surface

area to volume ratio this design is not suitable for

industrial application.

Usually hydrophobic organic solvents like

hydrophobic ethers and esters, long chain alcohols etc,

are the most common immobilizers in separating the

feed and stripping phases. Although this will lead to

instability of the SLM system, using ionic liquids can

be a solution to this matter. Ionic liquids (IL) are

eutectic molten salts which consist of organic cation

or anion. Large distribution coefficients in metal ion

extraction for example are the advantage of using IL

as an alternative for common organic solvents in LM

technique (Lertlawasin et al., 2010).

The advantages of SLM can be summarized as

being energy saving and low cost, and also due to

small scale of operation the possibility of using

expensive materials.

In SLM the overall driving force for solute

diffusion is concentration difference of the solute in

both feed and stripping phases.

For the solid support, hydrophobic microfilters are

most recommended. Choosing a right polymeric

support can improve the SLM stability which is the

main disadvantage of using SLM in practice.

The driving force for a solute to be dissolved in the

organic phase (the impregnated hydrophobic

membrane) is chemical potential across the

membrane. For the solute “i” the chemical potential is:

Dμi = RT d ln ci + RT d ln μi (9)

where μi is chemical potential of the species

(solute) i, R is universal gas constant, T is absolute

temperature and Ci is the concentration of the species

i, μi is activity coefficient of the species i.

The solute transport from the feed phase to the

stripping phase can be simplified into the five

sequential steps as follows:



143

1. Solute diffusion across the feed phase

2. Solute extraction on the Feed-LM phase

interface

3. Diffusion across the LM

4. Re-extraction on the LM-Stripping phase

interface

5. Diffusion across the stripping phase

The membrane flux can be calculated by Fick’s first

law:

J = km ΔS/d (10)

where km is mass transfer coefficient, ΔS the

solute concentration difference between stripping

phase and feed phase.

For an SLM, considering that mass transfer

coefficient is equal to diffusion coefficient multiple by

partition coefficient, the flux can be given by:

J = DSLM Kf ΔS/d

where DSLM is the SLM diffusion coefficient and

Kf is the partition coefficient, ΔS the concentration

difference between stripping phase and feed phase of

the solute (Reid et al., 1977).

Diffusion coefficient itself is derived by Stokes-

Einstein (kinetic theory) equation:

D = kT/(6μ η r) (11)

where D is diffusion coefficient (cm2/s), k is

Boltzmann constant, T is absolute temperature

(Kelvin), η is the organic phase viscosity and r is the

pore radius of the solute.

4.1. Permeation in SLM

Permeation is the process in which a solute dissolves

in the organic phase and diffuses towards the stripping

phase. Substance solubility plays the major role in

simple transport permeation and equilibrium level is

the final step of permeation. In other words, the

transported compound is in the same condition either

in feed phase or stripping phase (Cussler et al., 1989).

The advantage of SLM system is that by using

facilitated organic phase, the transport of solute

through the LM and as a consequence solute

permeability and selectivity can be enhanced

drastically. In this method, the reaction between solute

and the selected carrier only takes place at the

membrane or in better words at the interface of feed

phase and membrane (Juang et al., 1998).

If C is the carrier in the LM which is able to form a

complex with the solute S:

S + C ↔ S=C (12)

then the dissociation constant is:

Ka = [S=C]/[S][C] (13)

From the Fick’s first law, the flux is:

J = Ds=c ΔSs=c/d (14)

where Ds=c is diffusion coefficient of carrier-solute

complex, ΔSs=c, is the concentration gradient of the

carrier-substance complex and is equal to the initial

concentration of the transported substance multiple by

the extraction constant:

ΔSs=c, = Ss Kext (15)

and Kext (extraction constant) can be written as:

Kext = Kf Ka (16)

where Kf is partition coefficient between organic

phase and feed phase, and Ka is dissociation constant

(Park et al., 2006).

4.2. Ionic Liquids Application in SLM

Application of ionic liquids (IL) either as carrier or as

organic solvent in SLM is considered as a new

approach. IL consists of two parts: cationic organic

part and anionic part. IL mostly remain in liquid form

between 0 and 400 °C and shows very high thermal

stability (Han and Armstrong, 2007).

Ionic liquids are capable of amino acids and their

derivatives transportation through liquid membranes.

The needed driving force for this transportation is

counter ions gradient from the stripping (receiving)

phase to the feed (source) phase. Dzygiel et al. (1998)

used Aliquat 336 as a carrier to transport amino acids

by supported liquid membrane in which the extraction

process performed from an aqueous donor phase with

pH > 11, a gradient of chloride ions from the strip to

the feed phase provides a driving force for the mass

transport. According to Wieczorek et al. (1997), on

their investigation on concentration of amino acids

using supported liquid membranes with di-2-

ethylhexyl phosphoric acid as a carrier, the extractions

are made from an aqueous donor phase with pH 3 to a

more acidic acceptor phase and the mass transfer is

driven by the proton gradient between these phases.

4.3. Organic Solvents Selection in SLM

Hydrophobicity is the main characteristic for an

organic solvent to be used in SLM to ensure

immiscibility with aqueous feed and stripping phases.

Low viscosity of an organic solvent also plays a major

role in mass transfer of solute through the LM but at

the same time it has a negative effect on SLM

stability. Low volatility, low interfacial tension

between aqueous and organic phases in the support

pores which lead to the higher mass transfer can be

considered as important parameters for organic

solvent selection in SLM. Table 1 Shows selected

parameters of the commonly used organic solvents in

SLM (Rydberg et al., 2004).

Teng et al.


144

Table 1: selected parameters of the commonly used organic solvents in SLM (Rydberg et al., 2004)

Organic Solvent Density x 10-3

(kg/m3)

Viscosity x 103 (Pa s) Surface tension x 10 -

3 (N/m)

Solubility in water x

10-3 (kg/m3)

Dodecane 0.75 1.50 24.9 0.07

Heptane 0.68 0.38 19.6 -

Toluene 0.78 0.54 27.9 -

Kerosene 0.79 1.24 25.3 -

Diphenyl methane 1.00 2.96 38.4 6.51

Diheyl ether 0.79 1.87 - -

1-octanol 0.83 7.47 27.1 -

4.4. SLM and Species Recovery

One of the most important issues in extraction and

separation process is substance recovery which is

directly related to pertraction efficiency and/or

recovery.

The solute concentration difference ΔS over the

membrane can be expressed as:

ΔS = αfSf - αsSsKs/Kfm (17)

where αf and αs as are the fractions of the

transported substance, which are pertractable from the

feed to the strip phases, respectively. Ksm is the

partition coefficient for the solute between the strip

and membrane phase, and Kfm is the partition

coefficient for the solute between the feed and

membrane phase. While the feed and strip phases are

mostly aqueous, both partition coefficients are similar

(Kislic 2010).

Other parameters like diffusion coefficients in feed

and membrane phases, partition coefficients or

membrane thickness can also affect the extraction

rate. According to Jonsson and Mathiasson (1999),

there are two possibilities for pertraction controlling:

(i). Membrane controlled pertraction, when there is

a limiting step for diffusion of the transported solute

through the liquid membrane.

(ii). Feed controlled pertraction, when the diffusion

through the feed phase to the feed-membrane interface

appears as the limiting step.

4.5. SLM and Selective Extraction

The selective extraction in SLM is the ability to

transfer the desired compounds only and not the

interfering or unwanted compounds. The selectivity

depends mostly on the species capture method and

also the used transport mechanism (Thornton, 1992).

When the simple permeation is applied, the selectivity

is not high and is governed by solubility differences

between the solutes in the membrane phase; however,

when carrier is used the transport efficiency and

consequently the selectivity increases. Various carrier

molecules or ions can be incorporated in the

membrane phase to enhance the selectivity and mass

transfer (Dzygiel and Wieczorek, 2010).

A good carrier that can be used to enhance the

selectivity should have the following characteristics:

(i). Formation and decomposition of the complex

on membrane interfaces should be fast and rapid.

(ii). Side reactions can decrease the selectivity and

extraction process.

(iii). Irreversible reaction and degradation are

considered as limiting parameters in selectivity.

(iv). Low solubility in the aqueous feed and strip

phases of the carrier has a key role for choosing a

suitable carrier.

(v). Should not be hazardous or toxic to the

environment and should be cost effective specially in

industrial applications.

As the ionic carriers, mostly amines or carboxylic

and phosphoric acids for metals, organic acids and

amines are typically used. For metal extraction, the

addition of thiocyanate ions to the donor is needed to

form a negatively charged metal thiocyanate complex,

which can give an ion pair with the carrier (Papantoni

et al., 1995).

The other parameter that can affect the selectivity

is the diffusion coefficient which depends on the

molecular radius of the solute. Changing pH in feed

and stripping phases for the purpose of activating and

deactivating the compounds in these phases can

increase the selectivity as well. For example, the basic

feed phase and acidic strip phase is useful for selective

amines extraction (Dzygiel and Wieczorek, 2010). If

the pH of the feed (donor) phase is adjusted to a

sufficiently high value, the transported amines are

uncharged and are transported over the organic liquid

that is used as a membrane phase. The strip (acceptor)

phase on the other side of the membrane is an acidic

solution or buffer with low pH.

4.6. SLM Unit Design

Pertraction, extraction and transport processes in SLM

like all LM types, highly depend on the membrane

design and constructor.

Microporous polymeric membrane is typically

used for the membrane phase in SLM design and

modification. The type of polymeric microporous

membrane has a direct impact on the membrane

lifetime, stability, performance and efficiency.

Nowadays, the new generation of developed inorganic



145

membranes like ceramics, porous metals and zeolites

are used in SLM reactor modification and design and

have shown a range of advantages like thermal and

mechanical stability, being resistant to chemical and

organic solvents and being recyclable.

The utilization of advanced inorganic membrane

materials is nowadays very important. Cot et al.

(2000) worked on preparation of inorganic membrane

materials innovative concepts like templating effect,

nanophase materials, growing of continuous zeolite

layers, and hybrid organic–inorganic materials with

permselective properties for gas separation and

facilitated transport of solutes in liquid media which

have been successfully adapted to membrane

applications. Resina et al. (2008) used hybrid and

activated composite membranes containing Aliquat

336 for the transport of Pt(IV) and McCleskey et al.

(2002) used a thin layer of gold (700 Å) and have

reported high selectivity of U over Eu until [U] is

<0.84 mM in the feed solution on manufactured

alumina pourous supports to yield nanopores with

openings of <7 nm.

To choose an effective and practical organic

solvent for the membrane, it should be taken into

account that the organic liquid should be hydrophobic

enough to ensure immiscibility with aqueous phases.

Moreover, the low viscosity of the solvent can

decrease the SLM stability but increase the mass

transfer through the membrane.

4.7. SLM Stability and Reuse

One of the disadvantages of SLM in industrial

application is the low stability of this kind of liquid

membrane. Leaching carrier and emulsification in

liquid membrane phase reduce the SLM lifetime

(Neplenbroek, 1992). The other factor affecting SLM

stability is the operating temperature. According to

Saito (1992), the increasing operating temperature has

shown a direct effect on the solubility of both

membrane solvent and carrier in the aqueous phases

and has reduced membrane lifetime; but due to the

lower viscosity of the membrane phase, flux rate has

increased. Increasing viscosity leads to lifetime

increase but this will badly affect the flux rate and

deduce it drastically (Deblay et al., 1991).

Physicochemical characteristics and molecular

structure of the carriers are very important in SLM

stability. According to Chiarizia (1991), the more the

surface-active compound carrier usage, the less the

SLM stability would become.

The microporous polymeric membrane with

smaller pore size is more stable than larger pore size

type; although the surface porosity should not be

lower than certain levels that can decrease the flux.

The SLM could be reused by reloading the

membrane supports with fresh liquid membrane

solution after they have been used or continuous re-

impregnated which provides the same extraction

efficiency as a newly prepared SLM (Dzygiel et al.,

1998; Kocherginsky and Yang, 2007).

Membrane emulsification is considered as an

important degradation parameter of SLM stability.

Membrane emulsification occurs due to the lateral

shear forces and can be prevented by barrier formation

on the membrane interface or interfacial

polymerization (Wijers et al., 1998; Wang et al., 1998;

Wang et al., 1999).

Yang et al., (2000) worked on SLM stability by

applying polymerization surface coating, using

hexamethyldisiloxane and heptylamine as monomers

and hydrophobic microporous microfiltration

membranes with pore sizes of 0.05–0.2 μm were used

as substrate.

5. SUPPRTED LIQUID MEMBRANE and

WASTEWATER TREATMENT

SLM has been used widely in researches and

investigation, in various fields of environmental

technologies, pharmaceutical, food technology,

biotechnology and environmental science and sample

preparation in related fields. Metal ions such as Cu,

Cd, Co, Ni and Zn have been enriched in SLM for

sample preparation by a simple diffusion and pH

adjustment (Papantoni et al., 1995).

A counter coupled transport using D2EHPA as

carrier has been investigated for Pb sample

preparation and lead determination in urine (Djane et

al., 1997). Selective separation and pre-concentration

application of SLM for Cr (VI) removal from

wastewater (Ashraf and Mian, 2006) or Cu recovery

from spent ammoniacal electronic industry using

kerosene as solvent, can be mentioned as reference

(Kocherginsky and Grishchenko, 2003). Kedari et al.

(2013) reported uranium(VI) and thorium(IV)

transport across SLM containing

trioctylphosphineoxide as carrier. Ríos et al. (2013)

worked on the selective separation of metal ions using

supported ionic liquid membranes. Raut et al. (2012)

worked on selective strontium separation and Azzoug

et al. (2014) on metallic ions extraction.

6. FUTURE TREND: GREEN SUPPORTED

LIQUID MEMBRANE and IONIC LIQUIDS

The conventional solid membranes and related

separation processes have disadvantages such as low

flux rate, low selectivity (because of polymeric

membrane characteristic) and fouling. On the other

hand, liquid membrane separation methods use

Teng et al.


146

conventional petroleum-based organic solvents and

carriers which are hazardous to the environment.

According to Teng and Talebi (2012), liquid

membrane can be modified to “green liquid

membrane” by using environmentally friendly

materials. Vegetable oils (such as coconut or palm oil)

as non-toxic and clean solvents can be used instead of

regular organic solvents and are capable of reducing

the amount of common toxic and hazardous chemicals

using in liquid membrane methods.

Ionic liquids also can be considered as promising

materials that recently researchers have been working

on in order to enhance the ability of solute

transportation in liquid membrane technology

(Meinderisma et al., 2005; Sun et al., 2011 and

González et al., 2012).

Ionic liquids are basically salts that are liquid

below 100°C and consist entirely of ions. The unique

characteristics of ionic liquids such as an extremely

low vapour pressure, high thermal stability and certain

physico-chemical properties can be used in modifying

of organic phase in liquid membrane basicl structure

(Fischer et al., 2011) despite the fact that using a large

amount of ionic liquids as solvent in liquid–liquid

extraction process might not be cost effective due to

the high cost of ionic liquids compared to

conventional organic carriers (Matsumoto et al.,

2007).

REFERNCES

Ashraf MW, Mian A (2006). Selective separation and

preconcentration studies of chromium(VI) with

Alamine 336 supported liquid membrane.

Toxicological and Environmental Chemistry,

88: 187-96.

Azzoug S, Arous O, Kerdjoudj H (2014). Metallic

ions extraction and transport in supported liquid

membrane using organo-phosphoric compound

as mobile carriers. Journal of Environmental

Chemical Engineering, 2: 154–162

Chang SH, Teng TT, Norli I (2011). Optimization of

Cu(II) Extraction from Aqueous Solutions by

Soybean-Oil-Based Organic Solvent Using

Response Surface Methodology. Water Air Soil

Pollution, 217:567–576

Chiarizia R (1991). Stability of supported liquid

membranes containing long-chain aliphatic

amines as carriers. Journal of Membrane

Science, 55: 65-77.

Cot L, Ayral A, Durand J, Guizard C, Hovnanian N,

Julbe A, Larbot A (2000). Inorganic membranes

and solid state sciences. Solid State Sciences, 2:

313-334.

Cox M (2004). Solvent extraction in hydrometallurgy,

In : J. Rydberg, M. Cox, C. Musikas and G.R.

Choppin (Eds.), Solvent extraction principles

and practice, Marcel Dekker., New York,USA.

Cussler EL, Aris R, Bhown A (1989). On the limits of

facilitated diffusion. Journal of Membrane

Science, 43: 149-164.

Deblay P, Delepine S, Minier M, Renon H (1991).

Selection of organic phases for optimal stability

and efficiency of flat-sheet supported liquid

membranes. Separation Science and


Djane NK, Bergdahl IA, Ndung’u K, Schutz A,

Johansson G, Mathiasson L (1997). Supported

liquid membrane enrichment combined with

atomic absorption spectrometry for the

determination of lead in urine. Analyst, 122:

1073-7.

Dz˙ygiel P, Wieczorek P, Mathiasson L, Jo¨nsson JA

° (1998). Enrichment of amino

acids by supported liquid membrane extraction using

Aliquat 336 as a carrier. Analytical Letters, 31:

1261-74.

Dzygiel P, Wieczorek P (2010).Supported liquid

membranes and their modifications. In: Liquid

membranes, Kislic, Elsevier, oxford, UK.

Fischer L, Falta T, Koellensperger G, Stojanovic A,

Kogelnig D, et al. (2011). Ionic liquids for

extraction of metals and metal containing

compounds from communal and industrial

wastewater. Water Research, 45: 4601-4614.

González JMR, Pérez RG, López JCR, Martínez MA

(2012). Extraction of Copper(II) ions from

aqueous solutions with the ionic liquid 3-

butylpyridiniumbis (trifluoromethanesulfonyl)

imide. Separation and Purification Technology,

89: 320-328.

Han X, Armstrong DW (2007). Ionic liquids in

separations. Accounts of Chemical Research,

40: 1079-86.

Jonsson JA, Mathiasson L (1999). Liquid membrane

extraction in analytical sample preparation.

International Principles and Trends in

Analytical Chemistry, 18: 318-325.

Juang RS, Lee SH, Huang RH (1998). Modeling of

amine-facilitated liquid membrane transport of

binary organic acids. Separation and


Kedari, Pandit SS, Gandhi PM (2013). Separation by

competitive transport of uranium(VI) and

thorium(IV) nitrate across supported renewable

liquid membrane containing trioctylphosphine

oxide as metal carrier. Journal of Membrane

Science, 430: 188–195

Kislic VS (2010) Liquid Membranes. Elsevier, oxford,

UK.



147

Kocherginsky NM, Grishchenko A (2003). Method

for metal recovery from aqueous solutions. US

Patent, 6: 521,117.

Lee DW, Hong WH, Hwang KY (2000). Removal of

an organic dye from water using a predispersed

solvent extraction. Separation Science and

Technology, 35: 1951-1962.

Lertlapwasin R, Bhawawet N, Imyim A, Fuangswasdi

S (2010). Ionic liquid extraction of heavy metal

ions by 2-aminothiophenol in 1-butyl-3-

methylimidazolium hexafluorophosphate and

their association constants. Separation and


Li NN (1968). Separating hydrocarbons with liquid

membranes, US patent, 3: 410-794.

Matsumoto M, Ohtani T, Kondo K (2007).

Comparison of Solvent Extraction and

Supported Liquid Membrane Permeation Using

an Ionic Liquid for Concentrating Penicillin G.


McCleskey TM, Ehler DS, Young JS., Pesiri GD,

Jarvinen GD, Sauer NN (2002). Asymmetric

membranes with modified gold films as

selective gates for metal ion separations.


Meinderisma GW, Podt AJG, de Haan AB (2005).

Selection of ionic liquids for the extraction of

aromatic hydrocarbons from aromatic/aliphatic

mixtures. Fuel Processing Technology, 87: 59-

70.

Muthuraman G, Teng TT, Leh CP, Norli I (2009).

Extraction and recovery of methylene blue from

industrial wastewater using benzoic acid as and

extractant. Journal of Hazardous Materials,

163(1): 363-369.

Neplenbroek AM, Bargeman D, Smolders CA (1992).

Supported liquid membranes: Instability effects.


Papantoni M, Djane NK, Ndung’u K, Jo¨nsson J A,

Mathiasson L (1995). Trace enrichment of

metals using a supported liquid membrane

technique. Analyst, 120: 1471-77.

Park SW, Choi BS, Kim SS, Lee JW (2006).

Facilitated transport of organic acid through a

supported liquid membrane with a carrier.


Raut, Mohapatra PK, Manchanda VK (2012). A

highly efficient supported liquid membrane

system for selective strontium separation

leading to radioactive waste remediation. Journal of Membrane Science, 390-391: 76-83.

Reid RC, Prausnitz JM, Sherwood TK (1977). The

properties of gases and liquids. McGraw-Hill,

New York, USA.

Resina M, Fonta´s C, Palet C, Mun˜oz M (2008).

Comparative study of hybrid and activated

composite membranes containing Aliquat 336

for the transport of Pt(IV). Journal of

Membrane Science, 311: 235-242.

Ríos AP, Hernández-Fernández FJ, Lozano LJ,

Sánchez-Segado S, Ginestá-Anzola A, Godínez

C, Tomás-Alonso F, Quesada-Medina J (2013).

On the selective separation of metal ions from

hydrochloride aqueous solution by pertraction

through supported ionic liquid membranes. Journal of Membrane Science, 444: 469–481

Rydberg J, Cox M, Musikas C, Choppin GR, Dekker

M (2004). Solvent Extraction, principles and

Practice, 2n edition, New York, USA, 2004.

Saito T (1992). Deterioration of liquid membrane and

its improvement in permeation transport of

Zn(II) ion through a supported liquid membrane

containing a Bathocuproine. Separation Science

and Technology, 27: 119- 132.

Schlosser S, Rothova I, Frianova H (1993). Hollow-

fibre pertractor with bulk liquid membrane.


Schlosser S, Sabolova E (1999). Transport of butyric

acid through layered bulk liquid membranes.

Sun X, Ji Y, Guo L, Chen J, Li D (2011). A novel

ammonium ionic liquid based extraction

strategy for separating scandium from yttrium

and lanthanides. Separation and Purification


Talebi A, Teng TT, Abbas FMA, Norli I (2012).

Optimization of nickel removal using liquid–

liquid extraction and response surface

methodology. Desalination and Water

Treatment, 47 (1-3): 334-340.

Talebi A, Tjoon Tow Teng TT, Abbas AFM, Su Y

(2013). Methyl Violet Removal from Synthetic

Wastewater by Liquid-Liquid Extraction using

Vegetable Oils as Solvent. International Journal

of Scientific Research in Environmental

Sciences, 1(12): 357-364.

Teng TT, Talebi A (2012). Green Liquid Membrane:

Development and Challenges. Journal of

Membrane Science and Technology, 2:3, e107.

Thornton JD (1992). Science and practice of liquid-

liquid extraction, vol 2, Oxford university press,

New York.

Visser HC, Reinhoud DN, De Jong F (1994). Carrier-

mediated transportthrough liquid membranes.

Chemical Society Reviews, 23: 75-81.

Wang Y, Thio YS, Doyle FM (1998). Formation of

semi-permeable polyamide skin layers on the

surface of supported liquid membranes. Journal

of Membrane Science, 14: 109-126.

Wang Y, Doyle FM (1999). Formation of epoxy skin

layers on the surface of supported liquid

membranes containing polyamines. Journal of


Teng et al.


148

Wodzki R, Nowaczyk J (2002). Propionic and acetic

acid pertraction through a multimembrane

hybrid system containing TOPO or TBP.

Separation and Purification Technology, 26:

207-220.

Wieczorek P, Jo¨nsson JA , Mathiasson L (1997).

Concentration of amino acids using supported

liquid membranes with di-2-ethylhexyl

phosphoric acid as a carrier. Analytica Chimica

Acta, 346: 191-97.

Wijers MC, Jin M, Wessling M, Strathmann H (1998).

Supported liquid membranes modification with

sulphonated poly (ether ether ketone),

Permeability, selectivity and stability. Journal


Yang XJ, Fane AG, Bi J, Griesser HJ (2000).

Stabilization of supported liquid membranes by

plasma polymerization surface coating. Journal




©2014 IJSRPUB

149

Role of Emulsion Liquid Membrane (ELM) in Separation Processes

Tjoon Tow Teng1*

, Munisamy Soniya2, Govindaraju Muthuraman

2 and Amir Talebi

1




Abstract. Liquid membranes (LM) have received much attention over the last decades. In recent years, a remarkable increase

of the applications of emulsion liquid membranes in separation processes is observed. Emulsion liquid membrane (ELM) is

reputed to be a standout amongst the most guaranteeing detachment techniques for extraction of metal contaminants. Emulsion

liquid membrane systems in which two immiscible phases are separated by a third phase which is immiscible with the other

two phases, are divided into two types: (1) facilitated mass transfer, (2) mass transfer without any reaction involved.

Keywords: Emulsion Liquid Membrane, Separation Process

1. INTRODUCTION

Liquid membranes (LM) have received much

attention over the last decades: liquid membrane

applications having different configurations and have

recently become of substantial practical

consequentiality in separation technology,

macromolecular chemistry and membrane technology.

The benefits of utilizing LM are high selectivity,

utilizing carriers for the transport mechanism, and

concrete molecular recognition. Therefore, the

synthesizing of incipient type of carriers having high

selectivity for categorical applications becomes very

consequential and a number of researchers have been

working in the area (Muthuraman et al., 2009; Chang

et al., 2011; Talebi et al., 2012).

LM in general plays a paramount role in disunion

science. Their efficiency and economic advantages

distinguish them as a solution for some vital

environmental quandaries such as metals recovery,

hazardous species elimination (metals, organic

molecules) from wastewaters or selective applications.

LM can be modified as bulk liquid membranes

(BLM), supported liquid membranes (SLM), emulsion

liquid membranes (ELM), polymer inclusion

membranes (PIM) and activated composite

membranes (ACM).

In this deference, emulsion liquid membranes have

shown great potential, especially in cases where solute

concentrations are relatively low and other techniques

cannot be applied efficiently, since they cumulate the

process of extraction and divesting in a single unit

operation. The extraction chemistry is rudimentally

identical tantamount to that found in solvent

extraction, but the transport is supported by kinetic

rather than equilibrium parameters, that is, it is

governed by a non-equilibrium mass transfer (León

and Guzmán, 2004; Teng et al., 2013).

Since emulsion liquid membrane (ELM) as an

enhanced solvent extraction method was at first

proposed by Li (1968) ELM engineering has been

improved quickly. It has been used in dye removal

(Dâas and Hamdaoui, 2010; Kumar and Das, 2010;

Othman et al., 2011), phenols extraction (Park et al.,

2006; Ng et al., 2010) and metal ions removal and

recovery (Park et al., 2010; Nosrati et al., 2011) from

aqueous solutions. Typically, under best conditions,

dye removal can achieve up to 90–99% (Othman et

al., 2011). The reported extraction rate for phenol was

98–99% (Park et al., 2006; Ng et al., 2010) while for

aniline removal, the recovery rate was 99.5%

(Devulapalli and Jones, 1999) and for chromium

extraction, the most noteworthy recovery rate could be

97.5–99.6% (Rajasimman et al., 2009; Nosrati et al.,

2011). The emulsion is made by scattering the inner

phase in the membrane stage under high

emulsification speed in the vicinity of a surfactant,

which enhances the emulsion stability by averting the

droplets of inside stage from merging. Regularly, the

size circulation of the scattered internal phase droplets

is in the vicinity of 1 to 100 μm (Devulapalli and

Jones, 1999).

In contrast to liquid–liquid extraction, ELM holds

a lot of researcher’s preferences, for example,

effortlessness, enhanced kinetics and high selectivity.

Also, ELM method of action permits exceptionally

high mass transfer rate because of its expansive

surface range inside the emulsion globules and inner

droplets (Rajasimman et al., 2009) and permits both

extraction and stripping synchronously in one and

only step. Along these lines it is picking up more

vitality in the fields like metallurgy, solution, organic

chemistry and environmental industries.

2. DEFINITION and CLASSIFICATION

Emulsion liquid membrane (ELM) is reputed to be a

standout amongst the most guaranteeing detachment

techniques for extraction of metal contaminants

(Chakraborty et al., 2003; Ortiz et al., 2003;

Teng et al.


150

Kumbasar and Sahin, 2008) and hydrocarbons

(Correia et al., 2003; Park et al., 2006) due to the high

mass exchange rate, high specificalilty, low solvent

stock and low cost. Frankenfeld et al. (1981) reported

that the ELM could be dependent upon 40% shabbier

than other solvent extraction techniques. This process

consolidates both extraction and stripping stages to

perform a synchronous purification and concentration.

Be that as it may, this strategy has been constrained

by the emulsion unsteadiness (Florence and Whitehill,

1981; Li et al., 1988; Xuan-cai and Fu-quan, 1991;

Hou and Papadopoulos, 1996; Zihao et al., 1996;

Bandyopadhyaya et al., 1998; Wan and Zhang, 2002).

The absence of emulsion stability reduces extraction

efficiency. In the ELM process, three steps are

emulated incorporating an emulsification ready by

blending the layer and the internal phase, extraction,

and demulsification. In the first step, water is scattered

into the oil phase as fine globules. The second stage

takes after by penetration of solutes from the feed

phase, through the LM, to the receiving phase. In the

third stage, the emulsions are settled and demulsified

to discharge the internal phase within the loaded

solutes. This step is connected with the recuperation

and recovery of the LM.

ELM removes the limitation of solvent extraction

equilibrium by consolidating extraction and stripping

in a solitary operation, consequently accomplishing

decrease of metal concentration in the feed stream to

quite low levels. Besides, it diminishes the stock of

the organic solvent and metal extractant significantly.

ELM has been utilized to treat aqueous phases with

metals like copper, zinc, cadmium, chromium, and so

on. (Frankenfeld and Li, 1977; Marr and Kopp, 1982;

Gu et al., 1992; Winston and Li, 1996) ELM for metal

extraction is made by structuring a water in oil (W/O)

emulsion, stabilized by a surfactant, the W/O

emulsion holds the metal extractant (carrier) in the oil

phase and the stripping acid in the interior aqueous

receiving phase. This emulsion is then scattered by

mellow unsettling into a feed phase containing the

metal to be extracted. After extraction, the

concentrated emulsion is differentiated from the feed

phase, demulsification yields an oil phase that could

be reused.

Emulsion liquid membrane (ELM), also called

―surfactant liquid membranes‖ or "Double Emulsion

Membrane (DEM)", is basically twofold emulsions,

i.e., water/oil/water (W/O/W) systems or oil/water/oil

(O/W/O) systems. For the W/O/W frameworks, the oil

stage differentiating the two fluid stages is the liquid

membrane. For the O/W/O systems, the liquid

membrane is the water phase that is between the two

oil stages. Since their disclosure over two decades

prior, emulsion liquid membranes have shown

extensive potential as successful apparatuses for wide

separation applications.

2.1. Classification

Emulsion liquid membrane systems in which two

immiscible phases are separated by a third phase

which is immiscible with the other two phases, are

divided into two types: (1) facilitated mass transfer,

(2) mass transfer without any reaction involved. The

basic idea of mass transfer mechanism in emulsion

liquid membrane (ELM) systems can be described as:

solute transfer from the bulk external phase to the

external phase-membrane interface, followed by an

equilibrium reaction between the solute and the carrier

to form the solute carrier complex at the interface;

then, diffusion of the solute-carrier complex in the

membrane phase to the membrane-internal phase

interface; consequently, second equilibrium reaction

of the solute-carrier complex to strip the solute at the

membrane-internal phase interface into the internal

phase.

3. TRANSPORT MECHANISM

Emulsion liquid membrane (ELM) has been widely

used to investigate the ion transport against its

concentration gradient by the coupled transport

mechanism (uphill transport). The ion transport

through an ELM plays an important role in separation

technologies because of high transport efficiency,

excellent selectivity and economic advantage. Liquid

membrane extraction is the use of a carrier species

incorporated in the organic solvent to increase the

solute solubility: by introducing a 'carrier' molecule

into the membrane phase, the solute solubility is

increased by the reversible formation of a membrane-

soluble carrier-solute complex as shown in Fig 1

(Dâas and Hamdaoui, 2010).



151

Fig.1: Emulsion Liquid Membrane

The use of a carrier enhances selectivity by the

formation of a reversible complex between the carrier

and the solute, which is only soluble in the organic

solvent. This is particularly effective for the recovery

of charged solutes which may be poorly soluble in the

organic solvent. Many of the carriers so far employed

in liquid membrane processes are extractants used in

conventional liquid-liquid extraction, e.g. secondary

and tertiary amines, and phosphorus-containing

extractants.

The efficacy of emulsion liquid membrane (ELM)

process is a result of the facilitated mechanism that

maximizes both the extraction rate, i.e., the flux

through the membrane phase, and the capacity of the

receiving phase (the internal phase in the case with an

external feed phase) for the diffusing species. There

are fundamentally, however, two types of facilitated

transport in emulsion liquid membrane system, i.e.,

Type I and Type II facilitation. In the first type, the

concentration gradient of the membrane soluble

solute/permeate is maximized by irreversibly reacting

the solute with the reagent into an impermeable form

in the receiving phase thereby maintaining the

permeate concentration at efficaciously zero in this

phase. The diffusing species first dissolve in the layer

stage, which is made out of some organic solvents and

after that diffuse through the membrane and respond

with the interior stage (Ho et al., 1982; Fales and

Strove, 1984).

One approach in W/O/W systems for instance, is to

transport the solute from the outside fluid stage over

the oil medium of the globule (membrane), and

afterward in this manner react it with a reagent in the

internal water drops. This is regarded as type I. In

Type II category, a carrier (complexing executor or

extractant) is consolidated in the membrane phase and

it transports the diffusing solute over the membrane to

the stripping phase, a mechanism ordinarily reputed to

be "carrier mediated" facilitated transport. In this kind

of facilitation, the reaction in the internal phase

upholds a solute centralization of adequately zero. The

reaction of diffusing species with the chemical reagent

in the stripping phase structures an item that cannot

diffuse back again through the membrane (Stroeve

and Varanasi, 1982; Teramotoet al., 1983a). Reaction

happens both at the external interface between the

internal and membrane phase, and at the inner

interface between the membrane and internal phase.

This transporter (carrier) may be recovered after it

reacts with the inner reagent at the interface between

the films. Then again, for solutes insoluble in oil (e.g.

metal particles), an extractant must be utilized within

the interceding membrane phase, which ties and

discharges the solute at the outside and inward

interfaces progressively, permitting dispersion of the

solute-extractant complex through the liquid

membrane. This approach constitutes type

II(Teramotoet al., 1983b; Burge and Noble, 1984;

Chan and Lee, 1987).

4. MODELING of LIQUID MEMBRANES

Emulsion liquid membrane (ELM) is one of the

potential methods for treatment of industrial

wastewater aiming at recovery of various organic and

inorganic solutes (Borwankaret al., 1988; Kataokaet

al., 1989; Yan et al., 1992; Yan, 1993; Bhowal and

Datta, 1998; Kargariet al., 2006).Numerous

mathematical models have been developed. These

models can be categorized into two generic groups,

namely, carrier mediated transport models for type II

facilitation and diffusion-type mass transfer models

for type I facilitation. Applications of these models in

carrier mediated type II emulsion liquid membrane are

tested for extraction of various metal ions, namely,

silver (Lee et al., 1998), chromium (El-Said et al.,

2003)cesium (Chakraborty et al., 2003), nickel (Reis

and Carvalho, 2004), zinc, gold (Teramotoet al.,

1983b). Advancing front model is a much cited and

useful model for both types of transport models.

Therefore, it may be noted that in the modeling study,

concerted research efforts are directed to obtain the

closed form analytical solution. It may be worth

mentioning that all the modeling works in ELM are

based on single component system.

Teng et al.


152

4.1. Advancing Front Model

In this model the solute reacts at the same time and

irreversibly with the internal receiving reagent at a

reaction surface which transfers into the globules as

the reagent reacts. This model leads to nonlinear

partial differential equations which can be solved by a

perturbation method the zero-order solution generally

provides reasonable prediction for the diffusion

process. According to Yan et al. (1992) diffusion

controlled mass transfer model for type I facilitation

overcomes the shortcomings associated with

advancing front model. The model considers mass

transfer inside and outside of the globule as well as

the reaction between solute and internal phase reagent.

Yan’s model assumes that this reaction is not

instantaneous but irreversible. When Biot number and

Damkohler number approach infinity, Yan’s model

converges with the advancing front model (Kargariet

al., 2006).Lorbach and Marr (1987) developed a

model for type II facilitation that takes into account

the diffusion of the carrier and the carrier–solute

complex in emulsion globules and reversible reactions

at the externaland internal interfaces (Lorbach and

Marr, 1987). Some recent studies on ELM models

include: Bhowal and Datta (1998); Lee et al. (1998);

El-Said et al. (2003); Chakraborty et al. (2003) and

Reisand Carvalho (2004).

4.2. Carrier Facilitated Transport Model

Six major stages for the carrier facilitated transport

model for type II can be described as:(1) external

phase mass transfer resistance from external phase to

the external–membrane interface, (2) mass fluxes at

the external–membrane interface, (3) chemical

equilibrium of extraction reaction on the external–

membrane interface, (4)simultaneous diffusion of the

solute–carrier complex inside the emulsion globule

phase, (5) stripping of the complex at the membrane–

internal interface, and (6) chemical equilibrium of the

stripping reaction at the membrane–internal interface.

The advantage of carrier facilitated transport model is

capability of predicting theoretically the effects of

individual parameters on overall extraction rate. This

model not only predicts the concentration of the solute

in the external phase but gives the concentration

profile inside the membrane phase and the interfacial

concentration at the external phase–membrane

interface as well (Ho and Sirkar, 2001).

4.3. Mass Transport Model of ELM Systems

The model describes diffusion of the solute in the

external phase, mass transfer of the solute across the

film between the external phase and membrane phase,

chemical equilibrium of the extraction reaction on the

external-membrane interface, simultaneous diffusion

of solute-carrier complex inside the globule

membrane phase, and chemical equilibrium of the

stripping reaction on the membrane-internal interface.

Simultaneous partial differential equations can be

solved analytically by the Laplace transform method.

Some dimensionless groups are found with special

physical meanings to characterize the emulsion liquid

membrane system. The analytical solutions of the

model give concentration profiles of solute in the

external phase, concentration profile of the metal-

carrier complex in the membrane phase and surface

concentration of solute on the external membrane

interface (Ho and Sirkar, 2001).

4.4. Reversible Reaction Model

Extraction models for type-I facilitation clearly

demonstrate the superiority of the reversible reaction

model of Bunge and Noble (1984) in comparison to

the advancing front treatment of Ho et al., (1982)

which assumes irreversible reactions, especially

towards the end of batch extraction. Both these

models are based on diffusion of solute through the

membrane. However, the approach in reversible

model of depicting the reaction between solute and

internal reagent to be reversible is more realistic (Kim

et al., 1983; Stroeve and Varanasi, 1984; Chan and

Lee, 1987). A convenient method towards description

of interactions in the dispersed phase is the framework

of population balance equations (PBE). However, the

reversible model itself requires computation of the

solute concentration profile in a globule, typically of

around 0.1-0.2 mm in diameter that cannot be

regarded as well mixed because of slow membrane

phase diffusion of solute. A population balance

approach therefore has to incorporate both the size

distribution of globules as also the existing

concentration profile in each of them. Solution of such

a multivariate PBE would be too difficult to attempt.

Consequently, recourse to Monte Carlo simulation is

taken to describe interaction in a system of globules,

and in conjunction, the reversible model is solved for

each individual globule. A further aspect of

importance in membrane extraction is leakage of

internal droplets into the external phase, resulting in

decreased extraction efficiency. The leakage

phenomenon has been incorporated in existing

diffusion reaction models as a continuous flux of

internal droplets from a single globule (Borwankaret

al., 1963; Chan and Lee, 1987). However, in a model

of a single, stable globule, this should be accounted

for only during intermittent globule break-up.

Therefore, the task of the model is to include the

effect of globule interaction by combining membrane



153

phase mixing and exchange of globule contents along

with leakage of internal droplets, in order to interpret

batch extraction data in an ELM system.12

1.

5. ELM DESIGN CONSIDERATIONS

5.1. Emulsification

Ultrasound has been proved to be suitable for

emulsification (Higgins et al., 1972; Ensminger, 1973;

Alegria et al., 1989; Price, 1996; Chanamai et al.,

1998; Abismail et al., 1999; Asano and Sotoyama

1999; Ooi and Biggs, 2000; Behrend et al., 2000;

Sivakumar et al., 2002). However, these studies dealt

with oil-in-water (O/W) emulsions that are applied in

food, cosmetic, and pharmaceutical industries as well

as in polymerization processes. Little work has been

reported on the utilization of ultrasound for the

production of water-in-oil (W/O) emulsions used in

ELM processes. Juang and Lin (2004) have examined

the production of W/O emulsions in LM processes

using the model system of

water/kerosene/D2EHPA/Span 80 using a low

frequency ultrasound (20 kHz).

5.2. Membrane Stability

Figure 2 shows an emulsion liquid membrane process.

Some problems remain to be solved in order to apply

the ELM to a practical process. The major problem

associated with emulsion liquid membranes is

emulsion stability. The stability of the emulsion

globules (due to membrane rupture) is known as one

of the most serious problems in the application of the

liquid surfactant membrane to industrial separation.

The main factors affecting the emulsion stability

encompass membrane formulation, method of

emulsion preparation, and the condition under which

the emulsion is contacted with a reactant phase.

Stability of W/O/W emulsions is generally understood

as the resistance of the individual globules against

coalescence (Hou and Papadopoulos, 1996). The

breakdown of w/o/w type dispersions is described

through several possible mechanisms (Florence and

Whitehill, 1981), which include: (i) coalescence of the

internal aqueous droplets into larger internal droplets;

(ii) coalescence of the oil droplets suspended in the

aqueous phase; (iii) the expulsion of the internal

droplets following rupture of the thin oil films during

the interaction of the internal and external aqueous

phases (Li et al., 1988; Bandyopadhyaya et al., 1998)

and (iv) swelling or contraction due to water

permeation through the oil membrane by diffusion

(Ding and Xie, 1991; Wang et al., 1996; Wan and

Zhang, 2002).

5.3. Demulsification

Generally, an emulsion prepared with a high energy

density input (such as by an electrostatic method) will

have very small droplets. This will enhance membrane

stability if the surfactant concentration is high enough.

Meanwhile, the small droplet size gives a very large

interfacial area for mass transfer, but an ultra-stable

emulsion should be avoided because of possible

difficulties later during the demulsification step. Two

principal approaches for the demulsification of the

loaded emulsion are chemical and physical treatments.

Chemical treatment involves the addition of a

demulsifier to the emulsion. This method seems to be

very effective. However, the added demulsifier will

change the properties of the membrane phase and thus

inhibits its reuse. In addition, the recovery of the

demulsifier by distillation is rather expensive.

Therefore, chemical treatment is usually not suitable

for breaking emulsion liquid membrane, although few

examples of chemical demulsification have been

reported for certain liquid membrane systems (Zhang

et al., 1988). Physical treatment methods include

heating, centrifugation, ultrasonics, solvent

dissolution, high shear, and use of high voltage

electrostatic fields. The method of demulsification by

high shear includes the use of centrifugation as the

first step, followed by pumping the half-broken

emulsion through a high shear device (Kato and

Kawasaki, 1988). Demulsification with electrostatic

fields appears to be the most effective and economic

way for breaking W/O emulsion in ELM processes

(Lu et al., 1997). Since this type of technique is

strictly a physical process, it is most suitable for

breaking emulsion in liquid membranes to recover the

membrane phase for reuse.

Teng et al.


154

Fig. 2: The emulsion liquid membrane process

6. SUMMARY

6.1. Applications

In recent years, a remarkable increase of the

applications of emulsion liquid membranes in

separation processes is observed. The main

advantages of the ELM system are: (a) high interfacial

area for mass transfer, especially at the inner

membrane–water interface, due to the small size of

the aqueous phase droplets; (b) high diffusion rate of

metal ions through the membrane; (c) simultaneous

performance of extraction (at the outer interface) and

stripping (at the inner interface) in the same system,

and (d) capability of treating a variety of elements and

compounds in industrial setting at a greater speed and

with a high degree of effectiveness, with varying

contaminant concentrations and volume requirements.

6.2. Disadvantages

This system has several disadvantages, all having to

do with the formation of the emulsion.

1. Anything effecting emulsion stability must be

controlled. i.e. ionic strengths, pH, etc.

2. If, for any reason, the membrane does not

remain intact during operation, the separation

achieved to that point is destroyed.

3. In order to recover the receiving phase, and in

order to replenish the carrier phase, one has to break

down the emulsion. This is a difficult task, since in

order to make the emulsion stable; one has to work

against the ease of breaking it back down.

REFERENCES

Abismail B, Canselier JP, Wilhelm AM, Delmas H,

Gourdon C (1999). Emulsification by

ultrasound: drop size distribution and stability.


Alegria AE, Lion Y, Kondo T, RieszP (1989).

Sonolysis of aqueous surfactant solutions:

probing the interfacial region of cavitation

bubbles by spin trapping. Journal of Physical

Chemistry, 93: 4908–4913.

Alguacil FJ, Coedo AG, Dorado MT, Sastre AM

(2001). Uphill permeation of chromium (VI)



155

using Cyanex 921 as ionosphere across

immobilized liquid membrane.

Hydrometallurgy, 61 (1): 13–19.

Asano Y, Sotoyama K (1999). Viscosity change in

oil/water food emulsions prepared using a

membrane emulsification system. Food

Chemistry, 66: 327–331.

Bandyopadhyaya R, Bhowal A, Datta S, Sanyal SK

(1998). A new model of batch extraction in

emulsion liquid membrane: Simulation of

globule–globule interaction and leakage.

Chemical Engineering Science, 53: 2799–2807.

Behrend O, Ax K, Schubert H (2000). Influence of

continuous phase viscosity on emulsification by

ultrasound. Ultrasonics Sonochemistry, 7

(2000) 77–85.

Bhowal A, Datta S (1998). A mathematical model for

the analysis of extraction data of weak

acids/bases in liquid surfactant membranes.

Journal of Membrane Science, 139: 103–108.

Borwankar RP, Chan CC, Wasan DT, Kurzeja RM,

Gu ZM, Li NN (1988). Analysis of the effect of

internal phase leakage on liquid membrane

separation. AIChE Journal, 34 (5): 753–762.

Bunge AL, Noble RD (1984). A dilusion model for

reversible consumption in emulsion liquid

membranes. Journal of Membrane Science, 21:

55-71.

Chakraborty M, Bhattacharya C, Datta S (2003).

Effect of drop size distribution on mass

transfer analysis of the extraction of nickel (II)

by emulsion liquid membrane, Colloids Surf.

A: Physicochemical and Engineering Aspects,

224: 65–74.

Chan CC, Lee CJ (1987). Amass transfer model for

the extraction of weak acids/bases in emulsion-

liquid membrane systems. Chemical

Engineering Science, 42 (1): 83–95.

Chanamai R, Coupland JN, McClements DJ (1998).

Effect of temperature on the ultrasonic

properties of oil-in-water emulsions. Colloids

and Surfaces A: Physicochemical and

Engineering Aspects, 139: 241–250.





Pollution, 217:567–576.

Correia PF, Carvalho JMR (2003). Recovery of

phenol from phenolic resin plant effluents

emulsion liquid membranes. Journal of

Membrane Science, 225: 41–49.

Das C, Rungta M, Arya G, Das Gupta S, De S (2008),

Removal of dyes and their mixtures from

aqueous solution using liquid emulsion

membrane, Journal of Hazardous Materials,

159: 365–371.

Dâas A, Hamdaoui O (2010). Extraction of anionic

dye from aqueous solutions by emulsion liquid

membrane. Journal of Hazardous Materials,

178:973–981.

Devulapalli R, Jones F (1999). Separation of aniline

from aqueous solutions using emulsion liquid

membranes. Journal of Hazardous Materials,

70: 157–170.

Ding X, Xie F (1991). Study of the swelling

phenomena of liquid surfactant membranes.


FalesJL, Strove P (1984). A perturbation solution for

batch extraction with double emulsion: role of

continuous-phase mass transfer resistance,

Journal of Membrane Science, 21 (1): 35–53.

El-Said N, El-Sheref E, Borai E (2003). Modeling of

transport of Cs (137) by emulsion liquid

membrane (18C6) in xylene promoted by

ephedrine hydrochloride in stripping phase.


Ensminger D (1973). Ultrasonic: The Low- and High-

Intensity Applications. Marcel Dekker, New

York, USA.

Florence AT, Whitehill D (1981). Some features of

breakdown in water-in-oil-in water multiple

emulsions. Journal of Colloid and Interface

Science, 79 (1): 243–256.

Frankenfeld JW, Li NN (1977). Wastewater treatment

by liquid ion exchange in liquid membrane

system. In: Li NN (Ed.), Recent Developments

in Separation Science, vol. 3. CRC Press, pp.

285–292.

Frankenfeld JW, Chan RP, Li NN (1981). Extraction

of copper by liquid membranes. Separation

Science and Technology, 16: 385–402.

Gu Z, Winston Ho WS, Li NN (1992). Emulsion

liquid membranes—design considerations. In:

Winston Ho WS, Sirkar KK (Eds.), Membrane

Handbook. Chapman and Hall, pp. 656–700

New York, USA.

Higgins DM, Skauen DM (1972). Influence of power

on quality of emulsions prepared by ultrasound.

Journal of Pharmaceutical Sciences, 61: 1567–

1576.

Ho WS, Hatton TA, Lightfoot EN, Li NN (1982).

Batch extraction with liquid surfactant

membrane: a diffusion control model. AIChE

Jouranl, 28 (4): 662–670.

Ho WSW, Li NN (1992) In: Ho WSW, Sirkar KK

(eds) Membrane handbook, 1st edn. Van

Nostrand Reinhold, New York.

Ho WSW, Sirkar KK (2001). Emulsion liquid

membranes, in: Ho WSW, Sirkar KK (Eds.),

Teng et al.


156

Membrane Handbook, Kluwer Academic

Publishers, Dordrecht, Netherlands.

Hou W, Papadopoulos KD (1996). Stability of water-

in-oil-in-water type globules. Chemical

Engineering Science, 51 (22): 5043–5051.

Juang RS, Lin KH (2004). Ultrasound-assisted

production of W/O emulsions in liquid

surfactant membrane processes. Colloids Surf.

A: Physicochemical and Engineering Aspects

238: 43–49.

Kargari A, Kaghazchi T, Soleimani M (2006).

Mathematical modeling of emulsion liquid

membrane pertraction of gold (III) from

aqueous solutions. Journal of Membrane

Science, 279: 380–388.

Kato S, Kawasaki J (1988). Proceedings of the First

Sino-Japanese Symposium Liquid Membrane

24–27 September, China.

Kataoka T, Nishiki T, Kimura S, Tomioka Y (1989).

Batch permeation of metal ions using liquid

surfactant membrane. Journal of Membrane

Science, 46 (1): 67–80.

Kim KS, Chol SJ, Ihm SK (1983). Simulation of

phenol removal from waste water by liquid

membrane emulsion. Industrial & Engineering

Chemistry Fundamentals, 22: 167-172.

Kumar AA, Das C, De S (2010). Modeling of

extraction of dyes and their mixtures from

aqueous solution using emulsion liquid

membrane. Journal of Membrane Science, 360:

190-201.

Kumbasar RA, Sahin I (2008). Separation and

concentration of cobalt from ammoniacal

solutions containing cobalt and nickel by

emulsion liquid membranes using 5,7-

dibromo-8- hydroxyquinoline (DBHQ). Journal

of Membrane Science 2008, 164: 712–718.

Lee SC, Chang JH, Ahn BS, Lee WK (1998).

Mathematical modeling of penicillin G

extraction in an emulsion liquid membrane

system containing only a surfactant in the

membrane phase. Journal of Membrane

Science, 149: 39–49.

León G, Guzmán MA (2004). Facilitated transport of

cobalt through bulk liquid membranes

containing diethylhexyl phosphoric acid.


Li NN (1968). Separation of hydrocarbons with liquid

membranes. US patent 3: 410-794.

Li NN, Borwankar RP, Chan, Wassan DT, Kurzeja

RM, Gu ZM (1988). Analysis of the effect of

internal phase leakage on liquid membrane

separations. AIChE Journal, 34: 753–762.

Lorbach D, Marr R (1987). Emulsion liquid

membranes. Part II. Modeling of mass transfer

of zinc with bis (2-ethylhexyl) dithiophosphoric

acid. Chemical Engineering and Processing, 21:

83.

Lu G, Lu Q, Li P (1997). Break-down of liquid

membrane emulsion under high electric field.

Journal of Membrane Science, 128:1-6.

Marr R, Kopp A (1982). Liquid membrane

technology—a survey of phenomena,

mechanisms, and models. International

Chemical Engineering, 22: 44–60.

Muthuraman G, Teng TT , Leh CP, Norli I (2009).

Use of bulk liquid membrane for the removal of

chromium (VI) from aqueous acidic solution

with tri-n-butyl phosphate as a carrier.


Ng YS, Jayakumar NS, Hashim MA (2010).

Performance evaluation of organic emulsion

liquid membrane on phenol removal. Journal of

Hazardous Materials, 184: 255–260.

Nosrati S, Jayakumar NS, Hashim MA (2011).

Extraction performance of chromium (VI) with

emulsion liquid membrane by Cyanex 923 as

carrier using response surface methodology.


Ooi SK, Biggs S (2000). Ultrasonic initiation of

polystyrene latex synthesis. Ultrasonics


Ortiz MF, San Roman SM, Corvalan A, Eliceche M

(2003). Modeling and optimization of an

emulsion pertraction process for removal and

concentration of Cr(VI). Industrial and

Engineering Chemistry Research, 42: 5891–

5899.

Othman N, Zailani SN, Mili N (2011). Recovery of

synthetic dye from simulated wastewater using

emulsion liquid membrane process containing

tri- dodecyl amine as a mobile carrier. Journal

of Hazardous Materials, 198: 103–112.

Park Y, Skelland AHP, Forney LJ, Kim JH (2006).

Removal of phenol and substituted phenols by

newly developed emulsion liquid membrane

process. Water Research, 40: 1763–1772.

Price GJ (1996). Ultrasonically enhanced polymer

synthesis. Ultrasonics Sonochemistry 3: 229–

238.

Rajasimman M, Sangeetha R, Karthik P (2009).

Statistical optimization of process parameters

for the extraction of chromium(VI) from

pharmaceutical wastewater by emulsion liquid

membrane, Chemical Engineering Journal ,

150: 275–279.

Reis MTA, Carvalho JMR (2004). Modeling of zinc

extraction fromsulphate solutions with bis(2-

ethylhexyl) thiophosphoric acid by emulsion

liquid membranes. Journal of Membrane

Science 237: 97–107.



157

Sivakumar M, Senthikumar P, Majumdar S, Pandit

AB (2002). Ultrasound mediated alkaline

hydrolysis of methyl benzoate––reinvestigation

with crucial parameters. Ultrasonics


Stroeve P, Varanasi PP (1982), Transport process in

liquid membrane: double emulsion separation

systems, Separation and Purification Methods,

11 (1): 29–69.

Stroeve P, Varanasi PP (1984). Extraction with double

emulsion in a batch reactor: Effect of

continuous phase resistance. AIChE Journal,

30: 1007-1009.





Treatment, 47 (1-3): 334-340.

Tang B, Yu G, Fang J, Shi T (2010). Recovery of

high-purity silver directly from dilute effluents

by an emulsion liquid membrane-crystallization

process, Journal of Hazardous Materials1, 77:

377–383.

Teng TT, Muthuraman G, Mubeena K, Sathya M

(2013). Emulsion Liquid Membrane: Removal

and Recovery of Organic and Inorganic Ions.

Journal of Membrane Science and Technology,

3:2, e117.

Teramoto M, Sakai T, Yanagann K, Ohsuga M,

Miyake Y (1983a). Modeling of the permeation

of copper through liquid surfactant membrane,

Separation Science and Technology 18 (8)

(1983) 735–764.

Teramoto M, Takihana H, Shibutani M, Yuasa T,

Hara N (1983b). Modeling of the permeation of

copper through liquid surfactant membrane by

continuous operation, Separation Science and

Technology 18 (11): 985–997.

Wan Y, Zhang X (2002).Swelling determination of

W/O/W emulsion liquid membranes. Journal of

Membrane Science, 196: 185–201.

Wang Z, Jiang Y, Fu J (1996). The entrainment

swelling of emulsion during lactic acid

extraction by LSMs. Journal of Membrane

Science, 109: 25–34.

Winston Ho WS., Li NN (1996). Recent advances in

emulsion liquid membranes. In: Bartsch R.A.,

Douglas Way, J. (Eds.), Chemical Separations

with Liquid Membranes. ACS Symposium

Series, vol. 642. American Chemical Society,

pp. 208–221, Washington, USA.

Yan N, Shi Y, Su YF (1992). A mass transfer model

for type I facilitated transport in liquid

membranes, Chemical Engineering Science, 47

(17/18): 4365–4371.

Yan N (1993). A mass transfer model for type-II-

facilitated transport in liquid membranes.

Chemical Engineering Science. 48 (22): 3835–

3843.

Zhang XJ, Huang PY, Chen XJ (1988). Proc First

Sino-Jap SympLiqMembr 24–27 Sept, Chona,

pp 83–85.



©2014 IJSRPUB

158

Bulk Liquid Membrane and Its Applications in Wastewater Treatment

Tjoon Tow Teng1*, Sampath Elumalai

2, Govindaraju Muthuraman2, Amir Talebi

1




Abstract. Liquid membrane separation is a process which is depending on rate process and chemical potential gradient. The

theory of membrane-based solvent extraction suggests that overall mass transfer of solute consists of several steps: diffusion of

the solute through the aqueous layer from the bulk source aqueous solution to the phases’ interface (nonequilibrium process),

interaction of the solute with extractant and formation of the solute-extractant complex (as a rule, the process is rapid and

reaches equilibrium at the interface), diffusion of the solute-extractant complex through the membrane support itself

(nonequilibrium process), and diffusion of the solute-extractant complex through the organic layer to the bulk organic solution

(nonequilibrium process).

Keywords: Bulk Liquid Membrane, Wastewater Treatment

1. INTRODUCTION

After the novel study in liquid membrane (LM)

extraction method by Li (1968), LM technology was

further investigated by various researchers for the

recovery of metal ions (Gyves and Rodriquez, 1999; Chang et al., 2011) and organic compounds

(Muthuraman et al., 2009). LM process combines

extraction and stripping into one single stage and thus

involves non-equilibrium mass transfer characteristic

where the separation is not limited by the conditions

of equilibrium. Membrane can be defined as selective

barrier between two phases with mass transfer taking

place from the feed phase to the stripping phase.

There are mainly three types of liquid membrane

(LM) namely bulk liquid membrane (BLM),

supported liquid membrane (SLM) and emulsion

liquid membrane (ELM). BLM consists of three

phases, namely an organic phase that is sandwiched

between an aqueous feed phase and an aqueous

stripping phase. The organic solvent contains a carrier

that is insoluble in both aqueous solutions. The

samples diffuse from the aqueous feed solution

through the organic liquid membrane into the aqueous

receiving solution (Yang et al., 2009).

The BLM term includes several similar LM

systems, developed by different research groups such

as hybrid liquid membrane (HLM) (Majumdar and

Sirkar, 1992; Kislik et al., 1996 a&b; Gega et al.,

2001), hollow-fiber contained liquid membrane

(HFCLM) (Boyadzhiev, 1987; Bovadzhiev and

Lazarova, 1987; Sengupta et al., 1988; Bovadzhiev,

1990; Boyadzhiev and Alexandrova, 1992; Lazarova

and Bovadzhiev, 1992; Lazarova and Bovadzhiev,

1993; Schlosser et al., 1993; Boyadzhiev and

Dimitrov, 1994; Schlosser and Rothova, 1994;

Kawasaki et al., 1996; Qin and Cabral, 1998; Schlosser et al., 1999; Dai et al., 2000; Cara et al.,

2001; Cichy et al., 2001a; Dimitrov et al., 2002;

Schlosser and Sabolova, 2002; Teramoto et al., 2002; Wodzki and Szczepanski, 2002; Wodzki et al., 2002;

Boyadzhiev et al., 2003; Bhaumik et al., 2004;

Wodzki et al., 2004; Zhivkova et al., 2004), flowing

liquid membranes (FLMs) (Teramoto et al., 1987;

1989; Matsuyama et al. 1990; Teramoto et al., 1990;

Prasad and Sirkar 1992; Teramoto et al., 1994; 2001),

membrane based solvent extraction and stripping

(Baniel et al., 1992; Bromberg et al., 1992; Kedem et

al., 1992; Eyal and Bressler, 1993; Kedem and

Bromberg, 1993; Vajda et al, 2003; Kubisova et al.,

2004) and multimembrane hybrid system (Wodzki

and Sionkowski, 1996 a&b; Wodzki et al., 1999;

2000; Wodzki and Nowaczyk, 2001; 2002).

In this chapter, BLM principles such as

interaction mechanisms and theories of transport,

solvent used in the membrane phase, carrier in

membrane phase and kinetics studies are analyzed by

comparison of the modifications by different research

groups.

2. THEORY and MECHANISMS

Recovery and concentration of solutes, as well as

separation of samples, have attracted interest of

researchers, especially in connection with their

recovery from fermentation broths, reaction mixtures

and waste solutions. Several reviews including

membrane based solvent extraction (MBSE),

pertraction, solvent extraction and extractive

fermentations or bioconversions have been published

(Daugulis, 1988; Mattiasson and Holst, 1991; Roffler

et al., 1991; Schugerl, 1994; Schlosser, 2000 a&b;

Schugerl, 2000; Malinowski, 2001). The solvent can

be regenerated by membrane based solvent stripping

(MBSS) where the solute is re-extracted into the

stripping solution.

Several mechanisms have been proposed to

achieve transport of solute(s) through the L/L

Teng et al.


159

interface or through a liquid membrane. The

separation mechanism could be based on differences

in physical solubility of the solutes or their

solubilisation into the solvent or reverse micelles or

on the chemistry and rate of chemical or biochemical

reactions occurring on L/L interface(s). The

complexing or solubilisation agent extractant (carrier

in the liquid membrane) forms by reversible reaction

complex(es) or aggregate(s) with the solute, which are

soluble in the solvent or membrane. The chemistry of

reactive extraction and stripping in MBSE and MBSS,

as well as in PT, is identical with the classical solvent

extraction or stripping and is presented in several

books (e.g. Lo 1983; Schugerl et al., 1994; Bart,

2001).

To avoid direct contact of biomass with the liquid

membrane, whose components are not seldom toxic, a

multi-membrane hybrid system (MHS) with both

extraction and stripping L/L interfaces immobilised in

ion-exchange polymer membranes was suggested

(Kedem et al., 1992; Eyal and Bressler, 1993. MHS

was studied for the separation of carboxylic acids by

Wodzki and Nowaczyk, 1997; Wodzki et al., 2000).

3. MEMBRANE CONTACTORS

There are two main types of hollow fiber (HF)

contactors, those with parallel flow or cross-flow of

phases. Cylindrical HF contactor with cross-flow of

phases is shown in Fig. 1.

HF contactors have a large interfacial area per unit

volume of the contactor without requirement of

desperation of one phase that can be advantageous in

systems sensitive to emulation (Tong et al., 1999;

Cichy et al., 2001b). The volume ratio of phases could

be varied practically without limitations. The

disadvantage of HF contactors is connected with

additional mass-transfer resistance introduced by

porous wall(s) immobilizing L/L inter-face(s). Some

problems with swelling of HF and especially of

potting material of HF in solvents may occur.

4. KINETICS of TRANSPORT

In a liquid membrane setup, variation of solute

concentration versus time can be directly measured in

both feed solution (feed phase, [Sf]), stripping solution

(stripping phase, [Ss]) and the solute concentration in

the membrane phase is established from the material

balance.

The transport of solute can be expressed by the

following equations.

d[Sf]/dt = -k1 Sf* = Jf (1)

d[SLM]/dt = ξk1 SLM* - λk1 SLM* (2)

d[Ss]/dt = k1 Ss* = Js (3)

where ξ and λ are volume ratio of the feed phase to

the liquid membrane phase and the stripping phase to

the liquid membrane phase, respectively. (ξ =λ=2).

Sf*, SLM* and Ss* are the dimensionless concentration

of solute in feed, liquid membrane and stripping

phase, respectively. They are defined as,

Sf* = [Sf]/[Sf,0]

SLM* = [SLM]/[Sf,0]

Ss* = [Ss]/[Sf,0]

where [Sf,0] is the initial concentration of solute in

the feed phase.

If the differential Equations (1) to (3) are

integrated, the following equations are obtained

(Gyves and Miguel, 1999).

Sf* = exp(-k1t) (4)

SLM* = 2k1/2(k2-k1) [exp(-k1t) – exp (-2k2t)] (5)

Ss* = 1- 1/2k2-k1 [2k2 exp(-k1t) – k1 exp(2k2t)] (6)

According to Equation (5) SLM has a maximum

value (dSLM*/dt) =0, and the corresponding maximum

time is:

Fig. 1: Hollow fiber contactor with cross-flow of phases (Schlosser et al., 2005)



160

SLMmax

= (k1/k2)-k2/(k1-k2) (7)

tmax = ln(2k1/k2)/2k1-k2 (8)

Combining Equation (7) and (8) gives

k2 = ln (1/SLMmax

)/tmax (9)

First order time differentiation of Equation (4) to

(6) leads to the final form of flux Equation

dSf*/dt = -k1 exp( k1t) (10)

dSLM*/dt = k1/k2-k1×[-k2exp(-k2t)-k1exp(k1t)] (11)

dSs*/dt=k1 k2/ k2-k1 [exp(-k1 exp(-k1t)-exp-k2t)] (12)

The complexity of these equations prevents simple

comparison of kinetics observed for different

membrane materials. Therefore, it is useful to

examine and compare maximum release rates which

can be attained in a given experimental condition.

dSf/dt|max = -k1 (k1/k2)-k

1/(k

1-k

2) =Jf

max (13)

dSLM/dt|max = 0 (14)

dSs/dt|max = k2 (k1/k2)-k

2/(k

1-k

2) =Js

max (15)

5. SOLVENTS and EXTRACTANTS used in LM

The nature of solvent in liquid membranes has also

great influence on the efficiency of solute transport

through membrane to received phase. Some

commonly used solvents are listed in Table 1. The

solvents chosen belong to different solvent classes

(Kislik and Eyal, 1996a). Class I: open-chain and

cyclic hydrocarbons (methylcyclohexane, n-octane, n-

decane, n-dodecane, n-tetradecane, n-hexadecane and

kerosenes of & different origine); class II: π-donor

molecules (isopropylbenzene (cumene) and

1,3,Strimethylbenzene (mesitylene) ); class III:

halogenated molecules (2-chlorobutane); class IV, π-

donor molecules (ethylcarbonate); protic solvents (2-

ethylhexanol). The solvents selection is based on the

fact that they ought to have lower density than water

and a vanishingly small mutual solubility in the latter.

The paraffins used as membrane were chosen for

physicochemical and technological reasons. They are

structurally related (homologous series) and lighter

than water, they have low volatility, high flash point

and low melting point (all liquids at T= 25’ C). Their

viscosity varies considerably in the series and their

mutual solubility in water is practically zero.

The solvents generally used for liquid membrane

processes are typically flammable, volatile, toxic and

their use leads to environmental and safety risks. The

room temperature ionic liquids (RTIL) are considered

green solvent due to non-flammable nature and

negligible vapour pressure. However, their toxicity

data are not available (Swatloski et al., 2003).

Nevertheless, the choice of a solvent also depends on

economic considerations, which has not been

evaluated in laboratory scale studies. Vegetable oils

are considered to be the non-hazardous, cheap

chemicals, which can be used in BLM, as the novel

and greener LM. These vegetable oils are naturally

occurring, easily available, non-toxic, low cost and

renewable sources (Muthuraman and Palanivelu,

2006; Chang et al., 2010; Talebi et al., 2012).

Teng et al.


161

Table 1: Some commonly used solvents for liquid membrane

Extracted Material Solvent (extractant/diluent) Reference

Acetic acid TOA/(MIBK; octanol; n -alkanes) Lee et al., 2001

Amines/n –alkanes Solichien et al., 1995

(TOA; TOPO; TBP)/hexane Eyal and Bressler, 1993; Wodzki and

Nowaczyk , 2002

(Aliquat 336; TBP; TOPO; Alamine Juang and Wu, 1999

Tertiary amines/diluents Senol, 1999

Propionic acid Amines/n –alkanes Solichien et al., 1995

TOA/xylene Juang et al., 1993

(TOA; TOPO; TBP)/hexane Wodzki et al., 2000; Wodzki and

Nowaczyk, 2002

Butyric acid TOA/n –alkanes Sabolova and Schlosser, 2000, 2001,

Kertesz and , Schlosser, 2005)

Amines/(corn oil, oleyl alcohol) Wu and Yang, 2003

TOA/n –alkanes Sabolova and Schlosser, 1998; 2000

TOA/n –alkanes Sabolova and Schlosser, 1999

Valeric acid Amberlite LA-2/toluene Rodriguez et al., 1997; Viegas et al.,

1997; Coelhoso et al., 2000; Gonzalez

et al., 2001

(Amines; Aliquat 336; TBP)/(kerosene; n - heptane;

toluene)

Luque et al., 1995

Dimethylcyclopropan-

carboxylic acid (DMCCA)

TOA/n –alkanes Schlosser et al., 2001

TOA/n –alkanes Schlosser et al., 2001; Sabolova et al.,

1999; Schlosser and Sabolova, 2002

5-Methyl-2-pyrazinecarboxylic

acid (MPCA)

TOA/xylene Sabolova et al., 2001; Kubisova et al.,

2002a, 2004

TOA/xylene Kubisova et al., 2002b

Succinic acid n –Butanol Prasad et al., 1988

Aconitic, oxalic, malic acids TBP/Shellsol 2046 McMurray et al., 2002

Lactic acid Tertiary amines/(n -alkanes, isodecanol, isotridecanol) Kubisova and Schlosser, 1996

Aliquat 336/Shellsol A Coelhoso et al., 1997; Coelhoso et al., 2000

TOPO/kerosene Scheler et al., 1999

(TOA; TOPO)/(oleyl alcohol, n -hexane) Hano et al., 1996

Alamine 336/(kerosene, oleyl alcohol) Chen and Lee, 1997

Alamine 336/2-octanol Huang et al.,2004

(Amines; Aliquat 336)/(n -alkanes; oleyl alcohol) Giorno et al., 1996


TOMAC/oleyl alcohol Tong et al., 1998

TOMAC/oleyl alcohol Tong et al., 1999

TOMAC/1-decanol Gonzalez et al., 2004

TOPO, TOA, TBP Kondo et al., 2004

Tertiary amines/(n -alkanes, isodecanol, isotridecanol) Kubisova and Schlosser, 1996; Gonzalez´-

Munoz et al., 2004

TBP/(isooctane, SPAN80) Demirci et al., 2007

Amines/n –alkanes Scholler et al., 1993


(Amines; TOA; TOPO; TOMAC; TBP)/(kerosene; hex- Hano et al., 1993

ane; toluene; oleyl alcohol)

(Amines; trialkylphosphinoxides)/(kerosene; oleyl alco -hol) Siebold et al., 1995

Alamine 336/oleyl alcohol Tik et al., 2001

Citric acid TOA/xylene Juang et al., 2000; Juang and Chen, 2000

TOA/MIBK Basu and Sirkar, 1991

Amines/(hydrocarbons, alcohols) Juang et al., 1998; Friesen et al., 1991

Alamine 336/(n -alkanes and chloroform) Yordanov and Boyadzhiev, 2004

Phenylalanine Aliquat 336/(kerosene, isodecanol) Escalante et al., 1998; Escalante and

Irabien, 2001

L-Isoleucine D2EHPA/kerosene Ma et al., 2002

Phenylalanine, L-isoleucine Reversed micelles with polyoyalkylene Wang et al., 2004

L-Lysine D2EHPA/n –alkanes Boyadzhiev and Atanassova, 1991

Tryptophan, dipeptide AOT/oleyl alcohol Hossain, 2000

Tryptophan Aliquat 336/Shellsol A Coelhoso et al., 2000

N -(Benzyloxycarbonyl)-L- tert –Amylalcohol Rindfleisch et al., 1997; Lazarova et al.,



162

2002

Antibiotics: Penicillin G Amberlite LA2/(kerosene, isodecanol) Kedem and Bromberg, 1993; Isono et

al.,1997

Cephalosporine Aliquat 336/n –heptanes Sahoo et al., 1999

Erythromycin Decanol Kawasaki et al. 1996

Tylosin Isodecanol; octanol Boyadzhiev and Kirilova, 2000;

Boyadzhiev et al., 2003

Bacteriocins (nisin, variacin, Alkanes, toluene, decanol, butylacetate Kelly et al., 2000

Mevinolinic acid (MV-819) Isopropyl acetate Prasad and Sirkar, 1989

4-Methyltiazole, 4-

cyanothiazole

Toluene; benzene Prasad and Sirkar, 1990

Diltiazem Decylalcohol Basu and Sirkar, 1992

7-Aminocephalosporanic acid Aliquat 336/butylacetate Sahoo et al., 1996

p -Aminobenzenesulfonic acid TOA/(kerosene, octanol) Wang et al., 2002

2-Aminoethanesulfonic acid Ionic liquids Gu et al., 2004

Nucleotides (adenosine

derivatives)

Quaternary ammonium salt/isooctane Kubota et al., 2002

Oxygenates (aromas) from

citrus oil

Cyclodextrine derivatives Brose DJ et al., 1995

Phenol Cyanex 923/kerosene Urtiaga et al., 1994

Phenol Cyanex 923/kerosene Urtiag et al., 1994

Nitrophenol 1-Octanol Tompkins et al., 1992

5.1. Hybrid Membrane System

From simple bulk liquid membranes (BLM) shown in

Fig. 2A, construction of multimembrane hybrid

systems (MHS) such as those shown in Fig.2B and

Fig. 2C have been proposed (Kedem and Bromberg,

1993; Wodzki and Sionkowski, 1995; Kislik and Eyal,

1996 a&b; Wodzki and Nowaczyk, 1997; Eyal and

Kislik, 1999, 2000 a&b; Wodzki et al., 2000; Gega et

al., 2001; Wodzki and Nowaczyk, 2002; Wodzki and

Nowaczyk, 2002; Wodzki et al., 2002a).

Fig. 2: Schemes of pertractors and their operation: (A) simple bulk liquid membrane pertractor (BLM), (B) multimembrane

hybrid system (MHS), (C) multimembrane hybrid system coupled to Donnan dialysis (DD-MHS). f: feed solution, s: stripping

solution, m: liquid membrane, CEM: cation-exchange membrane, ms: mediating solution. Specific subprocessesÑ1: interfacial

extraction or back-extraction, 2: loaded (C2 M) and unloaded (CH) carrier diffusion, 3: ion-exchange sorption or desorption of

cations from/into aqueous solution, 4: cation-exchange dialysis, 5: interfacial cation-exchange between the carrier and CEM

functional groups, 6: Donnan dialysis between f and ms through CEM.

Teng et al.


163

The general idea of MHS is based on coupling the

operation of the BLM to the operation of other,

physically different, dense polymer membranes that

are able to support the properties of BLM and

maintain its stable functioning in time. A simple MHS

can be constructed in the form of a serial arrangement

of cation (CEM) or anion exchange polymer

membranes with a liquid membrane placed in between

(Fig. 2B). The MHS performances can be regulated

both by changing the properties of a liquid membrane

and the proper selection of polymer membranes

(Wodzki et al., 1999). Moreover, the ―sandwiching‖

of a liquid membrane between two reactive polymer

membranes allows an additional coupling of the

overall system with other membrane processes, e.g.

Donnan dialysis (DD) and pertraction in the MHS

(DD-MHS), Fig. 2C (Wodzki and Szczepanski, 2001),

or MHS pertraction with the process of water

pervaporation from the liquid membrane (Wodzki and

Szczepanski, 2000, 2002; Wodzki et al., 2002b). On

the other hand, a parallel arrangement of two different

MHS leads to a double multi membrane hybrid system

(d-MHS) presented schematically in Fig. 3A. A

general idea of pertraction (Zn2+

and Cu2+

) throughout

two different supported liquid membranes arranged in

parallel was described in the pioneering paper by

Drioli and coworkers (Loiacono et al., 1986). A

similar system made of two different bulk liquid

membranes, with noncyclic polyether carriers, for

simultaneous separation of Li+ and K

+ was reported by

Hiratani et al. (1987). The same idea was applied by

Sengupta and Sirkar (1988) for simultaneous

separation of two gases from ternary mixtures in a

system composed of two different polymer

membranes. The d-MHS was designed in order to

separate simultaneously two species of metal cations

from a multicomponent feed followed by their active

concentration in two separate receivers. The

pertraction process in each of the two MHS

subsystems involves such steps as:

• cation-exchange between the feed and CEM,

• cation-exchange dialysis throughout the

CEM,

• cation-exchange between CEM and the ionic

carrier in a liquid membrane,

• diffusion of the carrier in a loaded and

unloaded form,

• back-cation-exchange between the carrier and

the CEM at the stripping side,

• cation-exchange dialysis,

• stripping-cation-exchange into one of the two

receiving solutions.

Depending on the carrier used for the liquid

membrane preparation, the selective separation of a

given cation from the feed can be achieved. On the

other hand, both the cation-exchange dialysis (CEM),

and pertraction (LM with ionic carrier) operate as

counter-transport processes which enable the up-hill

transport of cations (chemical pumping or secondary

active transport) to occur (Wodzki et al., 1999) . In

practice, liquid membrane systems operate slowly

when the concentration of the feed phase or the

concentration of the carrier in the membrane is very

low. To overcome this problem, the DD-d-MHS

integrated system (Fig. 3B) can be applied. In this

case, both MHS modules functioning is additionally

supported by preconcentration of cations by the DD

process. The respective membrane device (pertractor)

can be constructed by adding one additional CEM and

a mediating solution containing counter-transported

cations (usually H+) to d-MHS. According to the

scheme of operation in Fig. 3B, during Donnan

dialysis (Wodzki and Szczepanski, 2001), the cations

from the feed are continuously preconcentrated in the

mediating solution which is used simultaneously as

the feed solution for the MHS modules.



164

Fig. 3: Schemes of new pertractors and their operation: (A) double multimembrane hybrid system (d-MHS), (B) double

multimembrane hybrid system supported by Donnan dialysis unit (DD-d-MHS). Detailed description: the same as in Fig. 2.

6. CONCLUSION

Liquid membrane separation is a process which is

depending on rate process and chemical potential

gradient. This is important to take notice that

equilibrium between phases has not any effect on the

separation process.

The theory of membrane-based solvent extraction

suggests that overall mass transfer of solute consists

of several steps: diffusion of the solute through the

aqueous layer from the bulk source aqueous solution

to the phases interface (nonequilibrium process),

interaction of the solute with extractant and formation

of the solute-extractant complex (as a rule, the process

is rapid and reaches equilibrium at the interface),

diffusion of the solute-extractant complex through the

membrane support itself (nonequilibrium process),

and diffusion of the solute-extractant complex through

the organic layer to the bulk organic solution

(nonequilibrium process). Here, only one of the

components may be equilibrium based in two phases:

(1) if the kinetics of solute-extractant interaction is a

rate-controlling process, which is not true in the

majority of separations published in the literature, and

(2) if the overall mass transfer of the solute from the

bulk source solution to the bulk membrane solution

reaches equilibrium.

REFERENCES

Baniel A, Eyal A, Bressler E (1992). Extraction of

electrolytes from aqueous solutions. Israeli

Patent, Appl. No: 101906

Bart HJ (2001). Reactive Extraction, Springer, New

York, USA.

Basu R, Sirkar KK (1992). Pharmaceutical product

recovery using a hollow fiber contained liquid

membrane, a case-study. Journal of Membrane

Science, 75: 131.

Bhaumik D, Majumdar S, Fan O, Sirkar KK (2004).

Hollow fiber membrane degassing in

ultrapurewater and microbiocontaminationn.


Bromberg L, Levin G, Libman J, Shanzer A (1992). A

novel tetradentate hydroxamate as ion carrier in


Science, 69: 143-153.

Brose DJ, Chidlaw MB, Friesen DT, Lachapelle ED,

P. Vaneik-eren P (1995). Fractionation of citrus

oils using a membrane-based extraction

process. Biotechnology Progress, 11: 214.

Boyadzhiev L (1987). Recovery of valuable metals

from diluted aqueous solutions by creeping film

pertraction. In: Devis GA, ed. Separation

Process in Hydrometallurgy. Society of

Chemical Industry, 1987; Pt. 3, Chapter 26:

259-287, London, UK.

Boyadzhiev L, Lazarova Z (1987). Study on Creeping

film pertraction. Recovery of copper from

diluted aqueous solutions. Chemical

Engineering Science, 42(5): 1131-1139.

Teng et al.


165

Boyadzhiev L (1990). Liquid pertraction or liquid

membranes – State of the art. Separation

Science and Technology, 25(3): 187-190.

Boyadzhiev L, Atanassova I (1991). Recovery of L-

lysine from dilute water solutions by liquid

pertraction. Biotechnology and Bioengineering,

38: 1059.

Boyadzhiev L, Alexandrova S (1992). Dephenolation

of phenol-containing waters by rotating film

pertraction. Separation Science and

Technology, 27(10): 1307-1312.

Boyadzhiev L, Dimitrov K (1994). Recovery of silver

from nitrate solution by means of rotating film

pertraction. Journal of Membrane Science, 86:

137-143.

Boyadzhiev L, Kirilova N (2000). Extraction of

tylosin from its aqueous-solutions by rotating

film-pertraction. Bioprocess Engineering, 22:

373.

Boyadzhiev L, Alexandrova S, Kirilova N, Saboni A

(2003). Pertraction continue de tylosine dans un

contacteur a films tournants. Chemical


Cara G, Schlosser S, Munoz M, Valiete M (2001).

Pertraction of neodymium and holmium in a

hollow fiber contactor. In: Cox M, Hidalgo M,

Valiente M, eds. Solvent Extraction for the 21st

Century. Proceedings of the ISEC’99,

Barcelona, Spain, July 1999. Society of

Chemical Industry; 2: 1023-1028, London, UK.

Chang SH, Teng TT, Norli I (2010). Extraction of

Cu(II) from aqueous solutions by vegetable oil-

based organic solvents. Journal of Hazardous

Materials, 181: 868–872.





Pollution, 217:567–576.

Chen RF, Lee YY (1997). Membrane-mediated

extractive fermentation for lactic-acid

production from cellulosic biomass. Applied

Biochemistry and Biotechnology, 63–65: 435.

Cichy W, Schlosser S, Szymanowski J (2001a).

Transport of phenol through bulk liquid

membranes. In: Cox M, Hidalgo M, Valiente

M, eds Solvent Extraction for the 21st Century.

Proceedings of the ISEC’99, Barcelona, Spain,

July 1999. Society of Chemical Industry, 2001;

2: 1065-1070, London, UK.

Cichy W, Schlosser S, Szymanowski J (2001b).

Recovery of phenol with cyanex(R )-923 in

membrane extraction-stripping systems.

Solvent Extraction. Ion Exchange 19: 905.

Coelhoso IM, Silvestre P, Viegas RMC, Crespo JPSG,

Carrondo MJT (1997). Membrane based

solvent extraction and stripping of lactate in

hollow fiber contactors. Journal of Membrane

Science, 134: 19.

Coelhoso IM, Cardoso MM, Viegas RMC, Crespo

JPSG (2000). Transport mechanism and

modelling in liquid membrane contac-tors.


183.

Dai XP, Yang ZF, Luo RG, Sirkar KK (2000). Lipase

–faciliatated separation of organic acidsin a

hollow fiber contained liquid membrane

module. Journal of Membrane Science, 171:

183-196.

Daugulis AJ (1988). Integrated reaction and product

recovery in bioreactor systems. Biotechnology

Progress, 4:113.

Demirci A, Cotton JC, Pometto AL, Harkins KR,

Hinz PN (2007). Resistance of in plastic

composite Lactobacillus casei support biofilm

reactor during liquid membrane extraction and

optimization of the lactic acid extraction

system. Biotechnology and Bioengineering 83:

493.

Dimitrov K, Alexandrova S, Saboni A, Debray E,

Boyadzhiev L (2002). Recovery of zinc from

chloride media by batch pertraction in a rotating

film contactor. Journal of Membrane Science.

207: 119-127.

Escalante H, Alonso AI, Ortiz I, Irabien A (1998).

Sepa-ration of L-phenylalanine by

nondispersive extraction and backextraction—

equilibrium and kinetic-parameters. Separation

and Science Technology, 33: 119.

Escalante H, Irabien A (2001).

Separation/concentration of L-phenylalanine in

hollow fibre modules. Decrease of the

extraction rate, in: M. Valiente, M. Hidalgo

(Eds.), Solvent Extraction for the 21st Century,

Society of Chemical Industry, London, U.K.

Eyal A, Kislik V (1999). Aqueous hybrid liquid

membrane: A novel system for separation of

solutes using water-soluble polymers as

carriers. Journal of Membrane Science, 161:

207-221.

Eyal A, Bressler EJ (1993). Industrial separation of

carboxylic and amino acids by acids by liquid

membranes: Applicability, process

considerations and potential advantages.

Biotechnology and Bioengineering, 41: 287-

293.

Friesen DT, Babcock WC, Brose DJ, Chambers AR

(1991). Recov-ery of citric-acid from

fermentation beer using supported-liquid

membranes. Journal of Membrane Science, 56:

127.



166

Gega J, Walkowiak W, Gajda B (2001). Separation of

Co (II) and Ni (II) ions by supported and hybrid

liquid membranes. Separation and Purification

Technology, 22-23: 551-558.

Giorno L, Spicka P, Drioli E (1996). Downstream

processing of lactic-acid by membrane-based

solvent-extraction. Separation and Science

Technology, 31: 2159.

Gonzalez Munoz MJ, Luque S, Alvarez JR (2001).

Coca membrane assisted solvent extraction in

the recovery of valuable model compounds, in:

Proceedings of Engineering with Membranes,

Granada, June 4–6, Spain.

Gonzalez Munoz MJ, Luque S, Alvarez JR, Coca J

(2004). Simulation of an integrated extraction

and stripping process using membrane

contactors. Desalination, 163: 1-12.

Gu Y, Shi F, Yang H, Deng Y (2004), Leaching

separation of taurine and sodium sulfate solid

mixture using ionic liquids. Separation and


Gyves J, Rodriquez EM (1999). Metal ion separations

by supported liquid membranes. Industrial and

Engineering Chemistry Research, 38: 2182-

2202.

Hano T, Matsumoto M, Uenoyama S, Ohtake T,

Kawano Y, Miura S (1993). Separation of lactic

acid from fermentation broth by solvent

extraction. Bioseparation, 3: 321.

Hano T, Matsumoto M, Hirata M, Tohda T, Kubuta F,

Goto M, Nakashio F (1996). Extraction of

fermentation organic acids with hollow fiber

membranes, in: D.C. Shallcross, R. Paimin,

L.M. Prvcic (Eds.), Value Adding Through

Solvent Extraction, The University of

Melbourne, Melbourne, Ausralia.

Hiratani K, Taguchi K, Sugihara H, Iio K (1987).

Synthesis and properties of noncyclic polyether

compounds. XV. Noncyclic polyether carriers

exhibiting potassium-ion-selective transport

through liquid membranes: their structures and

cation selectivity. Journal of Membrane

Science, 35: 91-102.

Hossain MM (2000). Mass transfer studies of amino

acids and dipeptides in AOT-oleyl alcohol

solution using a hollow fiber module.

Separation and Purification Technology, 18: 71.

Huang HJ, Yang ST, Ramey DE (2004). A hollow-

fiber membrane extraction process for recovery

and separation of lactic acid from aqueous

solution, Applied Biochemistry and

Biotechnology, 113–116: 671.

Isono Y, Fukushima K, Kawakatsu T, Nakajima M

(1997). Integration of charged membrane into

perstraction system for separation of amino-

acid derivatives. Biotechnology and

Bioengineering, 56: 162.

Juang RS, Lee SH, Huang RH (1998). Modeling of

amine-facilitated liquid membrane-transport of

binary organic-acids. Separation and Science


Juang RS, Wu RT (1999). Extraction of acetate from

simulated waste solutions in chloromycetin

production. Separation and Purification


Juang RS, Chen JD (2000). Mass transfer modeling of

citric and lactic acids in a microporous

membrane extractor. Journal of Membrane

Science, 164: 67.

Juang RS, Chen JD, Huan HC (2000). Dispersion-free

membrane extraction—case-studies of metal-

ion and organic-acid extraction, Journal of

Membrane Science, 165: 59.

Kaul R, Mattiasson B (1991). Extractive

Bioconversions in aqueous Two phase systems.

In Entractive Bioconversions, Mattiasson B and

Holst O, Eds, M. Dekker, New York, USA.

Kawasaki J, Egashira R, Kawai T, Hara H,

Boyadzhiev L (1996). Recovery of

erythromycin by a liquid membrane. Journal of


Kedem O, Eyal AM, Bromberg L, Bressler E (1992).

Membrane Extraction in the Fermentation of

Carboxylic Acids, Israel Patent No. 101905,

1992.

Kedem O, Bromberg L (1993). Ion-exchange

membranes in extraction processes. Journal of


Kelly NA, BReuben BG, Rhoades J, Roller S (2000).

Solvent extrac-tion of bacteriocins from model

solution and fermentation broth. Journal of

Chemical Technology and Biotechnology, 75:

777.

Kertesz R, Schlosser S (2005). Design and simulation

of two phase hol-low fiber contactors for

simultaneous membrane based solvent ex-

traction and stripping of organic acids and

bases. Separation and Purification Technology,

41: 275.

Kislik V, Eyal A (1996a). Hybrid liquid membrane

(HLM) system in separation technologies,


Kislik V, Eyal A (1996b). Hybrid liquid membrane

(HLM) and supported liquid membrane (SLM)

based transport of titanium (IV). Journal of


Kislik V, Eyal A (2000a). Aqueous hybrid liquid

membrane process for metal separation. Part I.

A model for transport kinetics and its

experimental verification. Journal of Membrane

Science, 169: 119-132.

Teng et al.


167

Kislik V, Eyal A (2000b). Aqueous hybrid liquid

membrane process for metal separation. Part II.

Selectivity of metals separation from wet-

process phosphoric acid. Journal of Membrane

Science, 169: 133-146.

Kubisova L, Sabolova E, Schlosser S, Martak J,

Kertesz R (2002a). Membrane based solvent

extraction and stripping of a heterocyclic

carboxylic acid in hollow fiber contactors.

Desalination 148: 205.

Kubisova L, Martak J, Schlosser S (2002b), Transport

of 5-methyl-2-pyrazinecarboxylic acid through

a layered bulk liquid membrane. Chemical

Papers, 56: 418.

Kubisova L, Sabolova E, Schlosser S, Mikulova E

(2004). Mass transfer in membrane based

solvent extraction and stripping of 5-methyl-2-

pyyrazinecarboxylic acid and co-transport of

sulphuric acid in HF contactors. Desalination,

163: 27-38.

Kubota F, Yamada H, Goto M, Furusaki S (2002).

Liquid membrane transport of nucleotides by a

quaternary ammonium salt. Solvent Extraction

Research and Development, Japan. 9: 81.

Kondo K, Otono T, Matsumoto M (2004). Preparation

of microcapsules containing extractants and the

application of the microcapsules to the

extractive fermentation of lactic acid. Journal of

Chemical Engineering of Japan, 37: 1.

Lazarova Z, Bovadzhiev L (1992). Liquid film

pertraction A liquid membrane

preconcentration technique. Talanta, 39(8):

931-935.

Lazarova Z, Bovadzhiev L (1993). Kinetic aspects of

copper (II) transport across liquid membrane

containing LIX-860 as a carrier. Journal of

Membrane Science, 78(3): 239-245.

Lazarova Z, Syska B, Schugerl K (2002). Application

of large scale hollow fiber membrane

contactors for simultaneous extractive removal

and stripping of penicillin G. Journal of


Lee YM, Kang JS, Nam SY, Choi CH (2001),

Removal of acetic-acid with amine extractants

from fermentation broth using hydrophobic

hollow-fiber membrane contactor, Separation

and Science Technology, 36: 457.

Li NN (1968). Separation of hydrocarbons with liquid

membranes. US patent 3: 410-794.

Lo TH, Baird MHI, Hanson C (1983). Handbook of

Solvent Extraction, Wiley.

Loiacono O, Drioli E, Molinari R. (1986). Metal ion

separation and concentration with supported


Science, 28: 123-138.

Luque S, Alvarez JR, Pazos C, Coca J (1995),

Recovery of valeric acid from aqueous

solutions by solvent extraction. Solvent

Extraction and Ion Exchange 13: 923.

Ma M, He DS, Liao SH, Zeng Y, Xie QJ, Yao SZ

(2002). Kinetic study of L-isoleucine transport

through a liquid membrane containing di(2-

ethylhexyl) phosphoric acid in kerosene.

Analytica Chimica Acta, 456: 157.

Majumdar S, Sirkar KK (1992). Hollow-fiber

contained liquid membrane. In: Ho WSW,

Sirkar DD, des. Membrane Handbook. Van

Nostrand Reinhold, New York, NY, USA.

Malinowski JJ (2001). Two-phase partitioning

bioreactors in fermentation technology.

Biotechnology Advances19: 525.

Matsuyama H, Teramoto M (1990). Separation and

concentration of heavy metal ions by spiral type

flowing liquid membrane module. Water

Treatment, 5: 237-252.

Mattiasson B, Holst O (1991). Extractive

Bioconversions, M. Dekker, New York, USA.

McMurray SH, Griffin GJ (2002). Extraction of

aconitic acid from mix-tures of organic acids

and cane molasses solutions using sup-ported

liquid membranes. Journal of Chemical

Technology and Biotechnology, 77: 1262.

Muthuraman G, Palanivelu K (2006). Transport of

textile dye in vegetable oils based supported

liquid membrane. Dyes Pigments, 70: 99-104.

Muthuraman G, Teng TT, Leh CP, Norli I (2009).

Extraction and recovery of methylene blue from

industrial wastewater using benzoic acid as and

extractant. Journal of Hazardous Materials,

163(1): 363-369.

Prasad R, Frank GT, Sirkar KK (1988). Nondispersive

solvent extraction using microporous

membranes. AICHE, 84 (1988) 42.

Prasad R, Sirkar KK (1989). Hollow fiber solvent

extraction of phar-maceutical products: a case

study. Journal of Membrane Science, 47: 235.

Prasad R, Sirkar KK (1990). Hollow fiber solvent

extraction: perfor-mances and design. Journal

of Membrane Science, 50: 153.

Prasad R, Sirkar KK (1992). Membrane-based solvent

extraction, in: W.S.W. Ho, K.K. Sirkar (Eds.),

Membrane Handbook, Van Nos-trand

Reinhold, New York, USA.

Qin Y, Cabral J (1998). Theoretical analysis on design

of hollow fiber modules and modules cascades

for the separation of diluted species. Journal of


Rindfleisch D, Syska B, Lazarova Z, Schugerl K

(1997). Integrated membrane extraction,

enzymic conversion and electrodialysis for the

synthesis of ampicilin from penicilin G. Process

Biochemistry, 32: 605.



168

Rodriguez M, Viegas RMC, Luque S, Coelhoso IM,

Crespo JPSG, Alvarez JR (1997). Removal of

valeric acid from wastew-aters by membrane

contactors, Journal of Membrane Science137:

45.

Roffler SR, Randolph TW, Miller DA, Blanch HW,

Praus-nitz JM (1991). Extractive

bioconversions with nonaqueous solvents, in:

Mattiasson B, Holst O (Eds.), Extractive

Bioconversions, New York, USA.

Sabolova E, Schlosser S (1998). Pertraction of butyric

acid through liquid membrane in hollow fiber

contactors, in: Proceedings of the 13th

International Congress CHISA 98 on CD ROM,

August 23–28, Praha, Czech Republic.

Sabolova E, Schlosser S, Martak J (1999). Pertraction

of DMCCA through bulk liquid membranes, in:

Proceedings of Membrany i procesy

Membranowe w Ochrorone Srodowiska.

October, 21–23, Szczyrk, Poland.

Sabolova E, Schlosser S (2000). Pertraction and

membrane based solvent extraction of butyric

acid in hollow fiber contactors, in: Proceedings

of 14th International Congress CHISA, August

28–31, Praha, Czech Republic.

Sabolova E, Schlosser S (2001). Analysis of mass-

transfer resistances in extraction and

pertraction of butyric acid in HF contac-tors

considering kinetics of complex decomposition,

in: Proceed-ings of 28th International

Conference SSCHI, full text of lec-ture on CD

ROM, May 22–26, Tatranske Matliare,

Slowakei.

Sabolova E, Schlosser S, Kubisova L (2001).

Membrane based solvent extraction of

heterocyclic acid in hollow fiber contactors, in:

Pro-ceedings of 28th International Conference

SSCHI, full texts on CD ROM, (SK), May 22–

26, Tatranske´ Matliare, Slovakia.

Sahoo GC, Ghosh AC, Dutta NN (1996). Mathur RK,

Facilitated transport of 7-aminocephalosporanic

acid in a bulk liquid mem-brane. Journal of


Sengupta A, Sirkar KK (1988). Ternary gas separation

using two different membranes. Journal of


Scheler C, Popovic MK, Iannotti EL, Bajpai RK

(1999). Mass transfer coefficients in reactive

extraction of lactic-acid from fermentation

broths in hollow fiber membranes. Canadian

Journal of Chemical Engineering, 77: 863.

Schlosser S, Rothova I, Frianova H (1993). Hollow

fiber pertractor with bulk liquid membrane.


Schlosser S, Rothova I (1994). A new type of hollow

fiber pertractor. With bulk liquid membrane.


Schlosser S (1997), Method and Equipment for Mass

and Heat Transfer Among Several Liquids (in

Slovak). Slovak Patent No. 278547, 1997.

Schlosser S (1999). Pertraction through liquid and

polymeric membranes. In: Belafi-Bako K,

Gubicza L, Mulder M, eds. Integration of

Membrane Process into Bioconversions.

Proceesings of the 16th European Membrane

Society Annual Summer School, Veszprem,

Hungary, Kluwer Academic/Plenum

Publizhers, New York, NY, USA.

Schlosser S, Sabolova E (1999). Transport of butyric

acid through layered bulk liquid membrane.

Chemical Papers, 53: 403.

Schlosser S, Sabolova E, Martak J (1999). Pertraction

and membrane based extraction of carboxylic

acids in hollow fibre contactors. In: Cox M,

Hidalgo M, Valiente M, eds. Solvent Extraction

for the 21st Century. Proceedings of the

ISEC’99, Barcelona, Spain, Society of

Chemical Industry, London, UK.

Schlosser S (2000a). Membrane based processes with

immobilized inter-face, in: Bako K, Gubicza L,

Mulder M (Eds.), Integration of Membrane

Processes into Bioconversions, Kluwer

Academic, New York, USA.

Schlosser S (2000b). Pertraction through liquid and

polymeric membranes, in: K. Bako, L. Gubicza,

M. Mulder (Eds.), Integration of Membrane

Processes into Bioconversions, Kluwer

Academic publishers, New York.

Schlosser S, Sabolova E, Martak J (2001). Pertraction

and membrane based solvent extraction of

carboxylic acids in hollow fiber contactors in:

M. Valiente, M. Hidalgo (Eds), Solvent

Extraction for the 21st Century, SCI, London,

UK.

Schlosser S, Sabolova E (2002). Three-phase

contactor with distributed U-shaped bundles of

hollow jibers for petraction. Journal of


Schlosser S, Ketesz R, Martak J (2005). Recovery and

separation of organic acids by membrane-based

solvent extraction and pertraction An overview

with a case study on recovery of MPCA.

Separation and Purification Technology, 41(3):

237–266

Scholler C, Chaudhuri JB, Pyle DL (1993). Emulsion

liquid membrane extraction of lactic-acid from

aqueous-solutions and fermentation broth.

Biotechnology and Bioengineering, 42: 50.

Schugerl K (1994). Solvent Extraction in

Biotechnology, Springer-Verlag, Berlin,

Teng et al.


169

Germany.

Schugerl K (2000). Integrated processing of

biotechnology products. Biotechnology

Advances, 18: 581.

Sengupta A, Basu R, Sirkar KK (1988). Separation of

solutes from aqueous solutions by contained

liquid membranes. AIChE Journal, 34: 1698-

1713.

Senol A (1999). Extraction equilibria of formic,

levulinic and acetic-acids using (Alamine

336/diluent) and conventional solvent

systems—modeling considerations. Journal of

Chemical Engineering of Japan. 32: 717.

Siebold M, Vonfrieling P, Joppien R, Rindfleisch D,

Schugerl K, Roper H (1995). Comparison of the

production of lactic-acid by 3 different

lactobacilli and its recovery by extraction and

electrodialysis, Process Biochemistry, 30: 81.

Solichien MS, Obrien D, Hammond EG, Glatz CE

(1995). Membrane-based extractive

fermentation to produce propionic and acetic-

acids—toxicity and mass-transfer

considerations, Enzyme and Microbial

Technology, 17: 23.

Swatloski PR, Holbrey DJ, Rogers DR (2003). Ionic

liquids are not always green: hydrolysis of 1-

butyl-3-methylimidazolium

hexafluorophosphate. Green Chemistry, 5: 363-

363.





Treatment, 47 (1-3): 334-340.

Teramoto M, Matsuyamaa H, Takayaa H (1987a).

Development of spiral type supported liquid

membrane module for separation and

concentration of metal ions. Separation Science

and Technology, 22: 2175-2201.

Teramoto M, Matsuyama H, Ohnishi N (1989).

Development of a spiral-type flowing liquid

membrane module with high stability and its

application to the recovery of chromium and

zinc. Separation Science and Technology, 24:

981-999.

Teramoto M, Matsuyama H, Ohnishi N (1990).

Selective facilitated transport of benzene across

supported and flowing liquid membranes

containing silver nitrate as a carrier. Journal of

Membrane Science and Technology, 50: 269-

278.

Teramoto M, Matsuyama H, Nakai K, Uesaka T,

Ohnishi N (1994). Facilitated uphill transport of

eicosapenaenoic acid ethyl ester through bulk

and supported liquid membranes containing

silver nitrate as carrier: A new type of uphill

transport. . Journal of Membrane Science, 91:

209-213.

Teramoto M, Takeuchi N, Maki T, Matsuyama H

(2001), Gas separation by liquid membrane

accompanied by permeation of membrane

liquid through membrane physical transport.


101-112.

Teramoto M, Takeuchi N, Maki T, Matsuyama H

(2002), Ethylene/ehane separation by facilitated

transport membrane accompanied by

permeation of aqueous silver nitrate solution.


117-124.

Tik N, Bayraktar B, Mehmetoglu U (2001). In situ

reactive extraction of lactic acid from

fermentation media. Journal of Chemical

Technology and Biotechnology, 76: 764.

Tompkins CJ, Michaels AS, Peretti SW (1992).

Removal of p - nitrophenol from aqueous

solution by membrane-supported solvent

extraction. Journal of Membrane Science, 75:

277.

Tong YP, Hirata M, Takanashi H, Hano T, Kubota F,

Goto M, Nakashio F, Matsumoto M (1998).

Extraction of lactic acid from fermented broth

with microporous hollow fiber membranes.

Journal of Membrane Science, 143: 81.

Tong YP, Hirata M, Takanashi H, Hano T (1999).

Back-extraction of lactic-acid with microporous

hollow-fiber membrane. Journal of Membrane

Science, 157: 189.

Urtiaga AM, Alonso AI, Galan B, Irabien JA, Ortiz

MI (1994). De-sign parameters of a hollow

fibre contactor for the non-dispersive extraction

of phenol, in: Proceedings of Extraction 94,

October 6–7 Edinburgh, UK.

Vajda M, Sabolova E, Schlosser S, Mikulova E

(2003). Membrane-based extraction joined with

membrane-based stripping in a circulating

arrangement II. Extraction of organic acids.

Chemical Papers, 57: 3-10.

Vajda M, Kosuthova A, Schlosser S (2004),

Membrane-based extraction joined with

membrane-based stripping in a circulating

arrangement III. Extraction of zinc. Chemical

Papers, 58: 1-8.

Viegas RMC, Rodriguez M, Luque S, Alvarez JR,

Coel-hoso IM, Crespo JPSG (1998). Mass-

transfer correlations in membrane extraction

analysis of Wilson plot methodology. Journal of

Membrane Science145: 129.

Wang YJ, Luo GS, Cai WB, Wang Y, Dai YY (2002).

Membrane extraction for sulfanilic acid

removal from waste water. Separation and

Science Technology, 37: 1163.



170

Wang Y, Shi Ch, Gan Q, Dai Y (2004), Separation of

amino acids by polymeric reversed micelle

extraction. Separation and Purification

Technology, 35: 1.

Wodzki R, Sionkowski G (1995). Recovery and

concentration of metal ions. II. Multimembrane

hybrid system. Separation Science and

Technology, 30: 2763-2778.

Wodzki R. Sionkowski G (1996a). Recovery and

concentration of metal ions III. Temperature

and concentration effects in multimembrane


Technology, 31: 1451-1553.

Wodzki R. Sionkowski G (1996b). Recovery and

concentration of metal ions II. Multimembrane


Technology, 31: 2763-2778.

Wodzki R, Nowaczyk J (1997). Extraction and

separation of propionic and acetic acid by

permeation in a hybrid membrane system

composed of liquid and ion-exchange polymer

membranes. Solvent Extraction and Ion

Exchange, 198: 1085-1106.

Wodzki R, Sionkowski G, Pozniak G (1999).

Recovery and concentration of metal ions. IV.

Uphill transport of Zn(II) in a multimembrane



Wodzki R, Szczepanski P (2000). Coupled membrane

processes and their biomimetic fundamentals.

Chemical Papers, 54: 430-436.

Wodzki R, Nowaczyk J, Kujawski M (2000).

Separation of propionic and acetic acid by

pertraction in a multimembrane hybrid system.


39-54.

Wodzki R, Szczepanski P (2001). Integrated process

of Donnan dialysis and pertraction in a

multimembrane hybrid system. Separation and

Purification Technology, 22-23: 697-706.

Wodzki R, Szczepanski P (2002). Integrated hybrid

liquid membrane systems – Membrane

extraction and pertraction coupled to a

pervaporation process. Journal of Membrane

Science, 197: 297-308.

Wodzki R, Nowaczyk J (2002). Propionic and acetic

acid pertraction through a multi-membrane

hybrid system containing TOPO or TBP.


207-220.

Wodzki R, Swiatkowski M, Kałuzynski K, Pretula J

(2002a). Pertraction of cations in a hybrid

membrane system containing soluble polymeric

ionophores, Journal of Applied Polymer

Science, 84: 99-109.

Wodzki R, Swiatkowski M, Łapienis G (2002b).

Properties of star-shaped polymer with

poly(oxyethylene) branches and monoesters of

phosphoric acid end groups in pertraction of

alkali, alkaline earth, and transient metal

cations. React. Funct. Polym. 52: 149-161.

Wodzki R, Szczepanska G, Szczepanski P (2004).

Unsteady state pertraction and separation of

cationsin a liquid membrane system: Simple

network and numerical model of competitive

M2+

/H+ counter-transport. Separation and


Wu ZT, Yang ST (2003). Extractive fermentation for

butyric acid pro-duction from glucose by

Clostridium Tyrobutyricum. Biotechnology and

Bioengineering, 82: 93.

Yang L, Chen XQ, Jiao FP, Braz J (2009). Extractive

resolution of racemic mendelic acid through a

bulk liquid membrane containing binary metal

carrier. Journal of Chemical Society, 20: 1493–

1498.

Yordanov B, Boyadzhiev L (2004). Pertraction of

citric acid by means of emulsion liquid

membranes. Journal of Membrane Science,

238: 197.

Zhivkova S, Dimitrov K, Kyuchoukov , Boyadzhiev L

(2004). Separation of zinc and iron by

pertraction in rotating film contactor with Kelex

100 as a carrier. Separation and Purification




©2014 IJSRPUB

171

Challenges in Fabricating Suitable Membrane for Water Treatment Application

Ling-Yong Wong 1, Choon-Aun Ng

1, Mohammed J. K. Bashir

1,*, Thiam-Leng Chew

2

1Department of Environmental Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul


2Department of Petrochemical Engineering, Faculty of Engineering and Green Technology (FEGT), University Tunku Abdul



Abstract. In this study, a multi component dope solutions with the range of polymer concentration 13 wt.% to 17 wt.% were

prepared using polyethersulfone, 1-methyl-2-pyrrolidinone and water. The PES-UF membranes were prepared based on a

dry/wet phase inversion technique. Membrane performances in terms of pure water permeability, salt water permeation, salt

water rejection and bacteria removal had been evaluated using low operating pressure (100 kPa to 500 kPa). The results

showed that membrane with polymer concentration of 13.60 wt.% had the best performance in terms of flux production and

salt water rejection. High salt water permeation from 1027.4 L/m2.h to 4109.6 L/m

2.h and also 99.7% of salt water rejection

were achieved. In addition, a total rejection for E. coli and E. faecalis with flux rate of 320 L/m2.h and 400 L/m

2 were obtained

,

respectively. SEM image obtained also showed a fine asymmetrical membranes structures.

Keywords: Membrane, Water, Treatment

1. INTRODUCTION

Water is an important essence for every living

creature throughout the world. We hardly separate our

daily life with the water. As water is so important to

us, various water treatment plants had been built up in

order to make sure the water we use is safe and free

from any threats. Water that passes through all the

treatments such as coagulation, flocculation,

sedimentation, filtration, and disinfection had been

guaranteed to be clean enough for us to drink, but

there is doubt about the safety to drink the water.

Since cases of waterborne breakout caused by

pathogens had been detected in many countries all

over the world, it indicates that water treatment plants

had face difficulties in providing water that is

perfectly safe for us (Drinking Water Contamination,

2006). One of the reasons is the water treatment plant

involves complicated and complex system in treating

process (Xia et al., 2004). With many systems

involved, some avoidable technical error will occur

causing severe effect to the users. In addition, the

disinfection product, usually chlorine had been long

time ago proven to cause side effects to the users

(Rook et al., 1974; Monarca et al., 2004). Over the

last decades, the tightening of water quality

regulations and the increased attention given to trace

contaminants in surface water and drinking water has

been urging of alternative treatment technologies in

order to improve conventional water treatment

processes (Hassan et. al., 2013). In order to serve the

purpose, membrane technology has become an

alternative since it can provide clean water and

rejecting all threats during the treatment process.

Moreover, the technology also has easier handling

system than the commercial water treatment plant.

Studies showed that membrane technology has the

potential to remove all pathogens in the water as well

as guaranteed a satisfactory safety level for the users

(Lo et al., 1996 and Cheryan, 1998). It has been the

most promising water treatment technique as it

provides good quality of treated water (Gupta and Ali,

2013).

In order to serve the application, membrane has to

be fabricated accordingly. There are generally several

important factors which will influence the

performance of produced membranes. Those factors

are materials selection (polymer, solvent, non-solvent

and additive), casting condition (shear rate and

surrounding temperature), and system components

(binary, ternary, and quaternary). Since there are too

many factors to be concerned, this study will mainly

focus on the polymer concentration which found to be

affecting the UF membrane performance most.

2. MEMBRANE TECHNOLOGY

Generally a membrane is defined as a selective barrier

between two phases; the term ‘selective’ being

inherent to a membrane or a membrane process

(Mulder, 1996). There are several types of membrane,

such as microfiltration membrane (MF), ultrafiltration

membrane (UF), nanofiltration membrane (NF), and

reverse osmosis (RO). These membranes are listed in

Table 1 below according to their type, pore radius,

material, pressure, and application.

Membrane processes including microfiltration,

ultrafiltration, and nanofiltration experienced rapid

growth in drinking water purification applications in

past two decades (Liu, 2014). The low pressure

membrane processes, which are microfiltration (MF)

and ultrafiltration (UF) play an important role in the


Wong et al.


172

water cycle. They are used in surface water treatment

to remove turbidity, pathogens and natural organic

matter (NOM). Often in these cases a chemical

coagulant is applied to generate a fine flocculants to

aid removal before membrane filtration. The low

pressure processes are also being used for

pretreatment of sea water or secondary treated effluent

(used water) prior to the high pressure of reverse

osmosis membranes (Fane et al., 2005). The low

pressure membranes also potentially simplify

treatment process trains by eliminating coagulation,

flocculation and sedimentation processes, and have

been considered as a substitute for conventional

drinking water treatment. Since membrane technology

has been applied for water and wastewater treatment,

it has grown steadily coincident with public demand

for high water quality and strict regulations (Lee et al.,

2004), until membrane considered as a promising

process to provide better drinking water quality for

water supply.

Table 1: Characterization of different membranes (Cheryan, 1998)

Process Membrane Type

and Pore Radius

Membrane

Material

Process Diving

Force Applications

Microfiltration

Symmetric

microporous, 0.1-

10 microns

Cellulose nitrate or

acetate,

Polyvinylidene

difluoride (PVDF),

Polyamides,

Polysulfone, etc.

Hydro-static

pressure

difference at

approx.

10-500 kPa

Sterile filtration,

Clarification

Ultrafiltration

Asymmetric

microporous,

1-10 nm

Polysulfone,

Polypropylene,

Nylon 6, PTFE,

PVC, Acrylic

Copolymer

Hydrostatic

pressure

difference at

approx.

0.1-1.0 Mpa

Separation of

macromolecular

solutions

Reverse

Osmosis

Asymmetric skin-

type,

0.5-1.5 nm

Polymers,

Cellulosic acetate,

Aromatic

Polyamide

Hydrostatic

pressure

difference at

approx.

2-10 Mpa

Separation of salts

and microsolutes

from solutions

Gas Separation

Asymmetric

homogeneous

polymer

Polymers &

copolymers

Hydrostatic

pressure and

concentration

gradients

Separation of gas

mixtures

Nanofiltration Thin-film

membranes

Cellulosic Acetate

and Aromatic

Polyamide

9.3-15.9 bar

Removal of

hardness and

desalting

Many researchers have suggested that the use of

low-pressure membrane filtrations that are MF and

UF have rapidly increased in the last decade due to

stricter regulations for finished water quality,

decreased cost, improved membrane materials and

modules, relative simplicity of installation, and

improved reliability compared to conventional

treatment processes such as sedimentation and rapid

filtration (Choi and Dempsey, 2004).

The rapid development of membrane technology in

recent years has boosted MF membrane to be an

alternative to overcome the problem. MF membranes

are widely utilized in the pharmaceutical industry, in

chemical engineering, biochemistry, and medicine as

well as in water and waste water engineering (Brock,

1983). It is more effective and safer compared with

the commercial treatment, but it brings about

drawback where microorganisms may pass through

the pores of microfiltration membranes. In certain

environment, during nutrient deprivation, bacteria

may become smaller than the previous size (Morita,

1993), with the diameter less than 0.3 µm and pass

through 0.2 µm membranes (Sadr Ghayeni et al.,

1999). In order to overcome this problem, UF

membrane which have relatively smaller pore size

than UF will be a need in the system.

2.1. Introduction to Asymmetric Membrane

An asymmetric membrane consists of very thin dense

top layer which has a thickness less than 0.5μm. The

top layer is supported by porous layer where the

thickness is in a range of 50 to 200μm (Mulder, 1996).

Its separation characteristics are determined by the



173

nature of membrane material or pore size, and skin

thickness mainly determines the mass transport rate.

Porous sub-layer acts as a support for the thin, fragile

skin and has little effect on the separation

characteristics (Hendricks, 2005). Asymmetric

membrane has been widely used for gas separation

and liquid separation. The thin top layer that plays a

role as a selective barrier film and the porous

sublayer, in which includes macrovoids, pores and

micropores, offers unique properties in terms of high

mass transfer rates and where good mechanical

stability can be best utilized (Ismail and Hassan,

2005). According to Ismail and Lai (2004),

asymmetric membrane consists of a skin of uniform

thickness supported on a porous sub layer whereby the

active skin layer acts as the actual separating barrier

while the sub layer serves only as a mechanical

support and do not involve in the separation process.

Mulder (1996) also stated that the transport rate is

determined by the top layer whereas sub-layer only

acts as a support. The permeation rate is inversely

proportional to the thickness of the actual barrier layer

and thus asymmetric membrane shows a much higher

permeation rate (water flux) than symmetric

membrane of comparable thickness. In addition to

high filtration rates, asymmetric membranes are most

fouling resistant. Asymmetric membranes are surface

filters and retain all rejection materials at the surface

where they can be removed by shear forces applied by

the feed solution moving parallel to the membrane

surface (Mark, 1988).

3. FABRICATION OF ASYMMETRIC

MEMBRANE

Asymmetric structure is usually prepared with dry/wet

phase inversion method. This method was preferred

since it was claimed to be the most versatile method to

produce high performance asymmetric membrane for

separation process (Ismail and Rahman, 2004). The

phase inversion process consists of the induction of

phase separation in a previously homogeneous

polymer solution either by changing the temperature,

by immersing the solution in a non-solvent bath (wet

process) or exposing it to a non-solvent atmosphere

(dry process) (Nunes and Peinemann, 2001). In the

formation process of a membrane, two types of phase

inversion can be distinguished. The dry phase

inversion takes place in the atmosphere by

evaporation of the volatile solvent during the casting

procedure forming a nascent membrane (more

obvious if induced with forced convection

evaporation). As for the wet phase inversion, it is

carried out by immersing the polymer solution into

coagulation bath of a non-solvent, where an exchange

of solvent and non-solvent occurred for membrane

mechanism formation or counter-diffusion of solvents

and non-solvent and extraction of the remaining

components occurs (Ismail and Hassan, 2005). The

dry/wet phase inversion normally involves multi-

component casting solution such as polymer, solvent,

non-solvent, and additive. In the research reported by

Ismail and Lai (2004), the solution was tailored to be

close to thermodynamic instability limit and

approaching phase separation. Binary system, ternary

system and quaternary system refer to casting solution

with respectively two, three and four components.

Figure 1 shows the framework for this study.

Membranes with ternary system were produced in

this study in order to differentiate the performance and

morphology among the membranes. For ternary

system, materials being used were polyethersulfone

(PES) as polymer, N-methyl-2-pyrrolidinone (NMP)

as solvent, and water as non-solvent. Water was also

used as coagulant bath, and methanol was used as post

treatment.

3.1. Binary System Solution

Binary system appears to be the fundamental solution

before it can be advanced to ternary and quaternary

system. To prepare a binary system, composition for

each chemical needed to be equipped as Equation 1.

Binary System Solution =

neededdopeofVolumexcomponentofwt

%100

.%(1)

3.2. Turbidimetric Titration Method

This is the method used to find out the real

composition of non-solvent additives which are

needed for preparing ternary or quaternary system

solution. Binary solution was poured into the burette

and the system was closed to minimize the

evaporation process. The solution was stirred with a

stirrer while non-solvent additives being added into it.

The speed for stirrer was set at 100rpm to minimize

the present of bubbles in the solution. 0.5ml of non-

solvent was added each time and waited until it was

totally dissolved by the solution before next 0.5ml

was added in. This process was continued until the

cloud point was reached, where the water

concentration in the polymer solution caused it to

remain turbid for more than 24 hours. This indicates

that phase inversion separation had occurred.

3.3. Ternary System Solution

The amount of non-solvent needed was determined

using turbidimetric titration method. The process for

ternary system preparation was similar to the process

Wong et al.


174

to produce binary system. The only difference was the

addition of non-solvent into the solution during the

preparation for the ternary system. Non-solvent was

added into the solution together with solvent, and then

only polymer was added into it.

Fig. 1: Framework flowchart

3.4. Quaternary System Solution

The process for quaternary system preparation was

similar to the ternary system, with only addition of

additive into the ternary solution. In order to obtain

the amount of additive needed, titration method was

referred again. After polymer was totally dissolved in

solvent, additive was mixed into the solution little by

little until all of it was dissolved in the solution. The

solution was again kept agitated for at least 30

minutes to make sure that it was homogenized. After

that, casting solution was transfer into a glass bottle

and put into ultrasonic bath for about 2 hours in order

to remove the bubbles within the solution. Bottle

taken out from the ultrasonic bath was closed tightly

and placed at room temperature.

3.5. Membrane Casting

Membrane was prepared by casting solution with the

pre-determined composition. The membranes were

produced by dry/wet phase inversion technique using

casting machine. Firstly, solution was cast on a clean

smooth glass plate and supported with knife gap

setting of 150µm. Ambient temperature was around

27ºC – 30ºC. Time used in membrane casting was 10

seconds. After that, the glass plate with casting

solution was immersed into the coagulation bath

(water) immediately at ambient temperature for at

least 10 minutes where the phase inversion occurred

in the immersion. A thin polymeric film was form

immediately and after a few minutes it will separate

from the glass itself. This phase was responsible for

membrane formation by the diffusion exchange of

solvent and non-solvent across the interface between

casting solution and non-solvent (water). The formed

membranes were then overturned immersed with a

large amount of water for the next 24 hours (washing

bath). Finally, the membrane was put into methanol

for at least 8 hours for post treatment. This stage was

to remove excess solvent that was still left in the

membrane. Then the membrane was hanged for

drying at ambient temperature for 1 day.

4. Permeation Test

The membrane permeation tests were carried out with

Sterlitech stirred cell (Model HP4750) with volume

capacity of 300 ml. It has an active area of 14.6 cm2

for the membrane. Maximum operating pressure for

the cell is 69 bars. Membranes were characterized in

terms of membrane flux and salts rejection in the cell.



175

Prior to testing for inorganic electrolytes or salts

(NaCl, 0.01M), the pure water flux using distilled

water was measured to ensure the stability of the

membrane. The pure water flux was measured at

different pressure (1, 2, 3, 4, and 5 bar) with the

stirring speed fixed at 400 rpm to determine the water

permeability. The pressure was first increased then

decreased when testing the sample. Pure water flux

was calculated by taking the readings at a regular

interval by noting the time used to collect 5 ml of

permeate for each operating condition. Then,

experiments were carried out with salt (NaCl) to

obtain the rejection performance of the membrane.

The flux and rejection were measured using standard

0.01M NaCl for different feed solutions. The flux and

observed rejection were determined for each operating

condition. To reduce concentration polarization, the

stirring speed was fixed at 400 rpm or 41.881 rad/sec.

During the testing process, for example, when NaCl of

0.01M was passed through the membrane under

operating pressure of 1 bar, the rejection and flux was

measured after 10 ml of permeate had been collected

and the time to collect the mentioned volume was also

recorded.

Three sets of membrane samples were made from

the same casting condition specifies for this study and

the average flux and solute separation data was

reported. Under room temperature and different

operating pressure, the permeate flux was calculated

as Equation 2:

)()(,

)(,,

ttimeAareaMembrane

VratepermeationVolumeJvFlux

(2)

Where,Jv = the permeate flux of salt aqueous

solution or pure water flux (m3/m

2 s); A = the effective

area of membrane (m2); t = the time (s); V = Volume

of permeate solution collected, (m3)

For pure water flux, the permeability of each

membrane was determined by the measurement of

water flux as a function of applied pressure, shown as

Equation 3.

P

JvPm

(3)

Where, Pm = Permeability (m3/m

2 s Pa); Jv = Flux

(m3/m

2 s); ∆P = Applied Pressure (Pa)

5. RESULT AND DISCUSSION

Polyethersulfone UF membranes were formulated by

varying the polymer concentration in the range of 13

wt% to 17 wt%. This concentration range is usually

used for the preparation of ultrafiltration membranes.

High concentration of the polymer that is more than

19 wt.% may produce nanofiltration while polymer

concentration lower than 11 wt.% may produce

membranes in micofiltration. To achieve high

performance membrane (high rejection to solute with

acceptable flux rate), the membrane cannot be too

loose for the structure as it will decrease the rejection

ability, neither too tight which will decrease the flux

production. Range of polymer concentration used in

this study was in the middle of the UF range, which

means higher possibility to achieve the objectives. All

the compositions were obtained through turbidimetric

titration measurement. Composition for membrane

solution is presented in Table 2.

Table 2: Casting solution formulation for ternary system

5.1. Pure Water permeation (PWP) measurement

According to Figure 2 below, flux for produced

membranes increased when the pressure increased

from 1 bar to 5 bar. The flux increased linearly with

the increase in pressure. This is an important step to

shows that the produced membranes were stable in

producing flux and were suitable application. Figure 2

also shows the flux production increased as the

concentration of PES decreased in solution. Detail of

permeability coefficient of each membrane is as

shown in Table 3.

From Table 3, membrane with polymer

concentration of 17 wt.% produced the least flux if

compared to other membranes. It might cause by the

thickness of the membrane posts greater hindrance

effect on the permeate through the membrane. By

decreasing 2 wt.% of polymer, produced membrane

Components

Before titration, in wt.%

(binary system)

After titration, in wt.%

(ternary system)

Casting

Solution 1

Casting

Solution 2

Casting

Solution 3

Casting

Solution 1

Casting

Solution 2

Casting

Solution 3

PES 15 17 19 13.60 15.40 17.70

NMP 85 83 81 77.40 75.30 75.70

Water 0 0 0 9.00 9.30 6.60

Wong et al.


176

was showed a good result by increasing the flux rate

to around 100 times. This shows that membrane was

more porous and permeate could pass through it

faster. A deduction of polymer by 2 wt.% again

produced a membrane with a permeability coefficient

which was more than 4 times higher than the 15 wt.%

PES membrane, and 500 times than the 17 wt.% PES

membrane. It was found that PES 13 wt.% was the

best flux production membrane among the produced

membranes. Although having a great flux, there is

another important factor for membrane performance,

which is rejection. Produced membrane was tested

with 0.01 M NaCl and the result is explained in the

following section.

Fig. 2: PWP at different pressure

5.2. Electrolyte Rejection by ternary system

Based on Figure 3, once again the membrane with

PES concentration of 13 wt.% produced the highest

flux compared to other membranes. It was found that

flux produced has increased as the concentration of

PES in solution decreased. It shows a huge difference

in the amount of flux produced compared to the others

at the same pressure given (1 bar to 3 bar). Membrane

with 15 wt.% of PES produced second highest flux,

followed by membrane with 17 wt.% of PES which

produced the least flux at the same pressure provided.

Table 3: Permeability coefficient for membrane



177

0

300

600

900

1200

1500

1800

2100

2400

2700

3000

3300

3600

3900

4200

4500

0 1 2 3 4 5 6

Pressure (Bar)

Flux

[L/m

2 .h]

PES 13 wt.%

PES 15wt.%

PES 17wt.%

Fig. 3: Fluxes of NaCl (0.01M) vs pressure for fabricated membrane with different polymer concentration

Detail for permeation produced by membranes in

salt water (NaCl 0.01 M) rejection is shown in Table 4

below. Membrane with the concentration of PES 17

wt.% showed a low flux rate of 5.5 – 21.1 L/m2.h

when 1-5 bar of pressure was provided. Decreasing 2

wt.% of polymer increased significantly the flux to

260-1120.8 L/m2.h, which is around 50 times higher

than the flux for membrane with 17 wt.% PES. By

decreasing 2 wt.% more of polymer concentration,

which brought it to 13 wt.%, the flux produced has

increased to a range of 1027.4 – 4109.6 L/m2.h when

low pressure (1-5 bar) was provided.

Table 4: Salt water flux produced by membranes at pressure given

It shows a great increase with more than 200% for

flux produced with decreasing PES concentration

from 17 wt.% to 13 wt.%. For 15 wt.% of PES

concentration, even it was showing better increase in

flux produced if compared to membrane with 17

wt.%, decrease of 2 wt.% more will provide the

membrane with greater performance in flux

production.

A high performance membrane should be able to

offer high flux and high solute rejection Salt rejection

(0.01 M NaCl) studies was tested on the membranes

and results are as shown in Figure 4.

Wong et al.


178

0

20

40

60

80

100

120

0 1 2 3 4 5 6Pressure (Bar)

Rej

ectio

n (%

)

PES 13wt.%

PES 15wt.%

PES 17wt.%

Fig.4: Electrolyte rejection vs pressure for fabricated membrane with different polymer concentration

It was shown in Figure 4 that the membranes

performed differently in the solute rejection tests. The

lowest rejection range (40-60%) was provided by 17

wt.% membrane. Meanwhile, membrane with 15 wt.%

showed good rejection even at low pressure where the

performance increased with increasing pressure. For

membrane with 13 wt.% of polymer, it again achieved

highest rejection (100%) among the membranes even

only low pressure (1 and 2 bar) was provided. As

presented by Table 5, solute rejection for membranes

shows an increase in the sequence of membrane with

17 wt.%, 15 wt.% to 13 wt.%. This shows that the

concentration of PES had great effect on the rejection

performances.

Table 5: Rejection of NaCl (0.01M)

5.3. SEM Cross Section Analysis

Generally membrane with higher concentration of

polymer gives higher rejection, but in this study,

membrane with 17 wt.% PES performed poorly

compared to other membranes, which gave rejection

of 50% to 60%. By reducing the concentration of

polymer to 15 wt.%, the rejection increased to higher

range of 87% to 99%. When it came to membrane

with the PES concentration of 13 wt.%, again it gave

the total rejection ability to salt water, where it

achieved 99% rejection at low pressure (1-2 bar), and

even achieved 100% rejection when higher pressure

was applied (3-5 bar). The experimental data were

then confirmed by the cross sectional images taken by

scanning electron microscope (SEM) at magnification

of 700X. SEM images shown in Figure 5 revealed the

membranes cast at low polymer concentration tend to

posses thicker membrane layer, large pore radius and

larger macrovoids, thus exhibiting a highest rejection

and highest flux rate.



179

A thick skin layer, nicely formed finger-like

structure, and large macrovoids was observed in the

cross section SEM view for membrane with 13 wt.%

PES. The structure provided a good condition to

produce high flux rate and high rejection. When 2

wt.% more polymer being added, macrovoid was

found to be absent and skin layer became thinner. This

situation explain the reduction in flux rate and

rejection ability of PES 15 wt.% compare to PES 13

wt.% membrane. SEM images of the membranes

sample showed that a finger-like structure provided

good membrane porosity and higher flux rate. As

polymer concentration increased, the larger macro

voids disappeared. This is a factor for the reduction of

flux rate. At 17% of polymer concentration, the

membrane pore size and membrane active layer were

reduced which resulted in the lowest flux rate.

However, the increasing number of membrane pores

did not influence the higher flux and rejection due to a

thinner active layer.

The result for the ternary system showed that

membrane with 13 wt.% of polymer concentration

was the best among the produced membranes. It not

only showed the outstanding performances in flux

produced, but also for the rejection of 0.01 M NaCl.

5.3. Bacteria Filtration Application

One of the important parameters in water treatment is

bacteria rejection for the system. If the system failed

to do so, then all the users are exposed to the bacteria

contaminated water, which will definitely jeopardize

their health. In this study, the bacteria removal studies

at low pressure (1-3 bar) were carried out for the

ternary system membrane with PES concentration of

13 wt.%, since it was found that the membrane with

13 wt.% PES gave the best performance. Pure cultures

for E. coli and E. Faecalis were applied in feeding

solution and result for flux production and removal

ability were shown as Table 6 and Table 7.

Table 6: Filtration result in log form for E. coli

Wong et al.


180

Fig. 5: SEM cross-sections of novel AULP membranes at different composition; (a) 13 % polymer; (b) 15 % polymer and (c)

17 % polymer



181

Table 7: Filtration result in log form for E. faecalis

According to the result obtained, removal of

bacteria which was higher than 6 logs (totally

removed) was achieved by the PES 13 wt.%

membrane even high concentration was used as feed

solution. Changing pressure did not post significant

effect on change of the removal ability of the

membrane. However, increasing pressure during

application enhance significantly the flux production.

As totally rejection obtained, it means that produced

UF membrane had a smaller pores size than the

bacteria. According to Tsummi et al. (1990), in the

cell separation by such porous membrane, molecular

size exclusion effect is dominant factor. When the size

of the solute is much larger than pore size of the

membrane, effective cutoff occurs. Once it achieved

totally rejection to the bacteria, it shows that produced

membrane was safe to be used for bacteria filtration

process as it was impossible for bacteria which have

bigger size than the membrane pores can pass through

it.

6. CHALLENGES IN FABRICATING

SUITABLE MEMBRANE FOR WATER

TREATMENT APPLICATION

There are several parameters must be considered in

order to produce high performance membranes, such

as dope formulation, casting condition, ambient

temperature during membrane casting, shear rate,

membrane thickness and so on. Changes on each

parameter, may affect membrane properties, and

hence changing the membrane performance in

different application. In order to reproduce a high

performance membrane, all the parameters involved

during membrane fabrication must be recorded in

detail.

Another challenge is to produce membranes which

give good fluxes and rejection. These always become

a limitation when it comes to the flux and selectivity

problems. Generally a membrane with bigger pore

size and loose structure will have higher flux, but

lower rejection ability. By increasing polymer

concentration in the solution, it will make the

membrane become denser and usually rejection ability

will be improved as well, but flux production will be

decreased. This is supported by Li (1993) where in his

study, he found out that as flux increase, membrane

selectivity will decrease and as the selectivity

increased, fluxes will be decreased.

Even though there is no perfect membrane, which

can serve all application with perfect rejection and

high flux production, but still there is optimum

membrane which can serve in specific application

with the acceptable range of removal ability and flux

production.

7. CONCLUSION

Based on the experimental data, this study had proven

that membrane with ternary system and concentration

of PES 13.60 wt.% was the best among the produced

membranes. It showed excellent performance in terms

of flux (2700 L/m2.h) and achieved 99.97% for salt

rejection (NaCl 0.01 M). In addition, the membrane

also achieved 100% rejection for the bacteria (E. coli

and E. faecalis) with an average flux of 320 L/m2.h

and 400 L/m2.h at low pressure provided (1 bar). SEM

cross sectional images clearly observed that this

asymmetric membrane provides great properties and

excellent separation performances. As the produced

membrane able to totally remove the bacteria in water,

this may help to provide a safe and clean drinking

water for those needed.

Pressure

(Bar)

Feed

Concentration,

Cf (cfu/ml)

Permeate

Concentration,

Cp (cfu/ml)

Flux

(L/m2.h)

BRI

(%)

BRLI

(log)

1

7x106 1 342.47 100 >6

7x108 1 286.72 100 >6

2

7x106 1 616.44 100 >6

7x108 1 440.31 100 >6

3

7x106 1 725.22 100 >6

7x108 1 513.70 100 >6

Wong et al.


182

ACKNOWLEDGEMENTS

We would like to extend our gratitude to Ministry of

Science, Technology and Innovation (MOSTI) for

the fund with project No. 03-02-11-SF0161.

REFERENCES

Brock TD (1983). Membrane Filtration: A User’s

Guide and reference Manual. Springer, New

York.

Cheryan M (1998). Ultrafiltration and microfiltration

handbook. Technomic Publishing, Lancaster,

2nd

Edition, Florida.

Choi K Y J, Dempsey BA (2004). In-line coagulation

with low-pressure membrane filtration. Water

Research, 38: 4271–4281.

Drinking Water Contaminants (2006). Basic

information about E. coli O157:H7 in drinking

water.

http://www.epa.gov/safewater/contaminants/eco

li.html.

Fane AG, Yeo A, Law A, Parameshwaran K,

Wicaksana F, Chen V (2005). Low pressure

membrane processes ~ doing more with less

energy. Desalination, 185: 159-165.

Gupta V K, Ali I (2013). Water

Treatment by Membrane Filtration Techniques.

Environmental Water. 135-154.

Li S G (1993). Preparation of Hollow Fiber

Membranes for Gas Separation. Enschede.

Netherlands. 1-22

Hasan A A, Victor K, Nidal H (2013). Hybrid ion

exchange – Pressure driven membrane

processes in water treatment: A review. Vol

116, 253–264.

Hendricks D (2005). Water Treatment Unit Processes

Physical and Chemical. Taylor & Francis

Group, Boca Raton, 913-915.

Ismail AF, Hassan AR (2005). Formation and

Characterization of Asymmetric Nanofiltration

Membrane: Effect of Shear Rate and Polymer

Concentration. Journal of Membrane Science,

270: 57 – 72.

Ismail AF, Lai PY (2004). Development of defect-free

asymmetric polysulfone membranes for gas

separation using response surface methodology.

Separation and Purification Technology, Vol.

40 (2): 191-207.

Ismail AF, Rahman AR (2004). The deduction of fine

structural details of asymmetric nanofiltration

membrane using theoretical models. Journal of


Lee N, Amy G, Philippe J, Brisson CH (2004).

Identification and understanding of fouling in

low-pressure membrane (MF/UF) filtration by

natural organic matter (NOM). Water Research,

38: 4511-4523.

Liu C (2014). Advances in Membrane Technologies

for Drinking Water Purification.

Comprehensive Water Quality and

Purification. Vol 2, 75-97.

Lo YM, Yang ST, Min DB (1996). Kinetic and

feasibility studies of ultrafiltration of viscous

xanthan gum fermentation broth. Journal of

Membrane Science, 117 (1–2): 237–249.

Mark CP (1988). Handbook of industrial membrane

technology. Noyes Publication. Park Ridge,

New Jersey, U-S-A.

Monarca S, Zani C, Richardson S D, Thruston Jr A D,

Moretti M, Feretti D, Villarini M (2004). A

New Approach to Evaluating the Toxicity and

Genotoxicity of Disinfected Drinking Water.


Morita RY (1993). Bioavailability of energy and the

starvation state. In: Starvation in

Bacteria (ed. Kjelleberg. ) S, pp. 1 23. New

York: Plenum Press.

Mulder M (1996). Basic principles of membrane

technology. Dordrecht, Kluwer Academic

Publishers, Netherlands.

Nunes SP, Peinemann KV (2001). Introduction, in

Membrane Technology: in the Chemical

Industry (eds S. P. Nunes and K.-V.

Peinemann), Wiley-VCH Verlag GmbH,

Weinheim, FRG.

Rook J J (1974). Formation of haloforms during

chlorination of natural waters. Journal of Water

Treatment Examine, 23: 234–243.

Sadr Ghayeni SB, Beatson PJ, Fane AJ, Schneider RP

(1999). Bacteria Passage Through

Microfiltration Membranes In Wastewater

Applications. Journal of Membrane Science,

153: 71-82.

Tsummi T, Sato T, Osawa N, Hitaka H, Hirosawa T,

Yamaguch K, Hamamoto Y, Manabe S,

Yamasiki T, Yamamoto N (1990). Structure

and filtration performance of improved

cuprammonium regenerated cellulose hollow

fiber (Improved BMM hollow fiber) for virus

removal. Polymer Journal, 22: 1085-1100.

Xia S, Nan J, Liu R, Li G (2004). Study of Drinking

Water Treatment by Ultrafiltration of Surface

Water and its Application to China.




©2014 IJSRPUB

183

Removal of Copper from Aqueous Solution by Using Thermo-responsive Polymer

Hydrogel as Copper Chelator via Polymer-enhanced Ultrafiltration (PEUF)

J.J. Chen, A.L. Ahmad, B.S. Ooi*

School of Chemical Engineering, Engineering Campus, Universiti Sains Malaysia, Seri Ampangan, Nibong Tebal 14300 Pulau

Pinang, Malaysia, Tel: +6045996418; Fax :+6045941013;

* Corresponding Author: [email protected]

Abstract. Membrane based separation is another efficient method for separating inorganic substances from aqueous effluents

due to its simplicity to scale-up and low energy consumption. Nevertheless, application of this method to remove heavy metals

from wastewaters is still a challenge to most researchers owing to the fact that heavy metals exist in aqueous solution as

hydrated ions with low molecular weights that can easily pass through most membranes except membranes for reverse osmosis

(RO) and nanofiltration (NF). However, RO and NF membranes with denser structures and narrower pores have very low

selectivity where nearly all types of solutes in the effluent are retained instead of separating only the targeted metal ions.

Furthermore, high energy is consumed and high operating pressure is required to achieve adequate permeation of solvent

through these types of membrane. Advanced materials combined with smart membrane systems could provide low energy and

highly sustainable metal recovery process. For example, polymer enhanced ultrafiltration combined with thermo-responsive

hydrogel could offer such attractive process. Compared to the conventional adsorbents, the adsorption capacity of the thermo-

responsive hydrogel is a function of solution pH, ligand density as well as temperature. The separation of hydrogel from the

waste stream can be realized through temperature induced agglomeration as well as steric retention by the membrane. The

separation process could be almost instantaneous with the metal recovery obtained even at very trace concentrations. Lower

temperature is more favorable due to the swollen hydrogel that exerts less diffusional resistance for metal to bind within the

interior marix. Low temperature is also favorable as it promises less membrane fouling phenomenon.

Keywords: Copper removal, Polymer-enhanced Ultrafiltration, Aqueous Solution

1. INTRODUCTION

Heavy metals are inorganic pollutants commonly

found in wastewaters discharged from industries like

metal-plating, tanneries, batteries, automobile and

fertilizer manufacturing. The pollution of natural

water-bodies especially rivers by the metal-

contaminated industrial effluents is becoming one of

the most serious environmental problems. Several

conventional methods had been widely adopted to

scavenge metal ions from aqueous solution including

ion exchange resins, electrodialysis, chemical

precipitation, adsorption and membrane separation.

However, the efficiency of these methods has been

constraint by some unsolved challenges such as

production of secondary sludge, low selectivity and

low metal-recovery efficiency (Lazaridis et al., 2004,

Dermentzis et al., 2009, Ghodbane et al., 2008).

In terms of low energy consumption and high

solute-rejection efficiency, ultrafiltration (UF) is

deemed as a promising membrane separation process

for the treatment of industrial wastewaters. The

average pore diameters of UF membranes fall within

the range of 10 – 1000 Ǻ (1 – 100 nm). The pore

structures at this range of pore diameters allow

permeation of microsolutes with molecular weight

(MW) less than 300 g/mol. Nevertheless, copper

divalent ion (Cu2+

) as a transition metal ion with

molar mass of merely 63.55 g is too low to be retained

by a UF membrane. The shortcoming of UF process

on metal removal is currently resolved by integration

of both metal-polymer complexation and UF

(Micheals, 1990). This combination is known as

polymer-enhanced ultrafiltration (PEUF).

In PEUF operation, water-soluble polymer with

metal-chelating ability is used as chelating agent to

complex with a targeted metal ion. The metal ion is

chelated by functional ligands of the polymer such as

carboxylic (-COOH) and amide (-CONH) groups via

electrostatic interaction to form metal-polymer

complexes. The complex can then be retained by UF

membrane effectively (Almutairi and Lovitt, 2012,

Camarillo et al., 2012, Zerze et al., 2013).

Complexation of metal in aqueous solution is deemed

as a major type of metal ion adsorption by polymeric

substances through chelation process. Polymer with

metal-chelating ligands are known as chelating agent

(chelator) that bind with free metal ions in aqueous

solution to form complex molecules known as

‘chelates’. A chelator contains one or more ligands

that are ionic or neutral such as –SO3H, -COOH, -NH2,

-OPO3H and -C=O. Each ligand could possess mono-,

bi- or even polydentate enabling it to form ring

structure with a metal ion (Flora and Pachauri, 2010,

Butvin et al., 1988).

Chen et al.

Removal of Copper from Aqueous Solution by Using Thermo-responsive Polymer Hydrogel as Copper Chelator via

Polymer-enhanced Ultrafiltration (PEUF)

184

1. COMPLEXATION OF COPPER IONS BY

POLYMER WITH METAL-CHELATING

LIGANDS

The metal-ligand complex formation between a metal

ion and ligands of a polymer can be described by the

following reaction:

(1.0)

where M refers to a free metal ion; L is a ligand of

polymer, and n is known as the number of ligands

involved in chelation with the metal ion.

The chelation of metal ion by acidic ligands of a

polymer involves dissociation of acidic functional

groups like carboxylic groups (-COOH) into both

anionic carboxylates (-COO-) and H

+ protons at pH

higher than pKa of the acidic ligand involved, for

example, pKa of acrylic acid (AA) is around 4.3.

Hence, it is important to determine their dissociation

equilibrium prior to analyzing the copper binding

equilibrium. Camarillo et al. (2012) elucidated that the

dissociation equilibrium constant, Ka of protonated

ligands is expressed as:

(1.1)

where L- is the deprotonated ligands of polymers,

whereas HL refers to protonated ligands and β

represents a constant depending on each polymer

ligand which accounts for nearest neighbouring

interaction (Camarillo et al., 2012, Porasso et al.,

2000). This expression is also applicable to other

polymers with acidic ligands in which their chelating

ability is mainly contributed by their acidic functional

groups.

Since the dissociation of carboxylic groups is

dependent on pH, therefore solution pH in the absence

of metallic ions is expressed as a function of Ka, β and

the degree of protonation, α. The expression is known

as modified Henderson-Hasselbach equation as shown

in Equation 1.2:

(1.2)

The protonation degree, α is usually expressed as

molar ratio of added base (NaOH) to carboxylic acid

(-COOH) of AA moieties in a PNIPAM-co-AA

polymer hydrogel:

(1.3)

where [NaOH] represents molar concentration of

NaOH added into the reaction solution; CH is the total

molar concentration of carboxylic (-COOH) and

carboxylate (-COO-) of hydrogel (Camarillo et al.,

2012, Flora and Pachauri, 2010). Hence in the

presence of metal ions, the concentration balance of

protonated ligands, [LH], deprotonated ligands [L] and

chelated ligands [MLn] is expressed as:

(1.4)

where [L]total is the total ligand concentration in the

solution. In addition, the extent of metal-ligand

complexation is described by a complex formation

constant, Kf which is defined as: (1.5)

where: [MLn] = concentration of metal-ligand

complex formed; [M] = concentration of metal ion; [L]

= concentration of protonated ligand

The value of this formation constant is highly

dependent on solution pH, and it is also expected to be

influenced by temperature for thermo-responsive

PNIPAM-co-AA polymer hydrogels.

For ease of experimental measurements and

calculation, the chelating ability of a polymer is

usually quantified by its chelating capacity in

equilibrium state, which is always expressed as a

parameter known as equilibrium adsorption capacity,

qe in most of the metal adsorption studies as shown in

Equation 1.6:

(1.6)

where V (L) is the total volume of reaction solution,

M (g) is the dry mass of chelating polymer.

2. APPLICATION OF THERMO-RESPONSIVE

PNIPAM-co-AA POLYMER HYDROGEL AS

NOVEL CHELATING AGENT

Selection of a suitable chelating agent is important for

optimizing the efficiency of metal ions-complexing in

PEUF. Water-soluble polymers are usually chosen as

potential metal chelators due to the presence of

hydrophilic functional groups that act as ligands for

metal chelation such as amide (-CONH), carboxylate

(-COO-) and carbonyl (-C=O). These functional

ligands are Lewis acids or bases that donate electron

pair to fill the unoccupied d-orbital of a metal ion to

form a chelated complex (Flora and Pachauri, 2010,

Hancock and Martell, 1989).

There are numbers of studies reported on the use of

different water-soluble polymers as metal chelating

agents in PEUF processes. Among them,

polyethyleneimine (PEI), poly(vinyl sulfonic acid)

(PVSA) and poly(acrylic acid) (PAA) are the most

common chelators for the evaluation studies of PEUF.

Caῆ izares et al. (2002) applied water-soluble PEI

(MW = 25,000 g mol-1

) and PAA (MW = 250,000 g

mol-1

) for recovering Cu2+

, Ni2+

, Pb2+

and Cd2+

cations.

The affinity of these two types of polymer toward

metal ions are attributed to their metal-chelating

ligands like amines (-NH) groups in PEI and

carboxylic (-COOH) in PAA (Caῆ izares et al. 2002).

PEI also attracted interest from İslamoğlu et al. (2006)



185

to study the effect of ionic strength on its

complexation with Cd2+

and Ni2+

in PEUF process.

They found that in the absence of NaNO3 salt, the

retention value of Cd2+

was higher than that for Ni2+

.

In contrast, as 0.5 N NaNO3 was added to the binary

metal solution, the retention of Ni2+

became higher

instead (İslamoğlu et al. 2006).

In 2008, Llanos and partners synthesized a novel

derivative of PEI known as partially ethoxylated

polyethylenimine (PEPEI) as chelator for Cu2+

recovery in PEUF. The results show a maximum

retention percentage over 97 % obtained with feed

loading ratio of 208 mg Cu per g. of PEPEI at pH 6

(Llanos et al., 2008). Labanda et al. (2009) compared

the ethoxylated PEI (EPEI) with polyvinyl alcohol

(PVA), polyacrylic acid-co-maleic acid (PACM) and

PEI on the recovery of Cr3+

in PEUF. They discovered

that among these four types of chelating agents,

PACM with carboxylic groups formed relatively

stable irreversible complex with Cr3+

, whereas

alcoholic PVA with hydroxyl groups did not chelate

with Cr3+

at all. The amines of PEI and EPEI formed

hydroxo-complexes with Cr3+

ions under varying pH

(Labanda et al., 2009).

Since water-soluble polymers like PEI form highly

stable metal complex in aqueous solution, it is not

easy to be dissociated in low pH medium to achieve

objective of metal recovery. A type of ‘intelligent’

polymeric microgel with thermo-responsive property

known as poly(N-isopropylacrylamide-co-acrylic acid)

(PNIPAM-co-AA) polymer hydrogel is applied as a

novel chelating agent in PEUF for removal of divalent

copper ions (Cu2+

) from aqueous solution. This

polymer hydrogel consists of a cross-linked network

structure with a shape of globule conformation which

makes each hydrogel structure a discrete nano-size gel

particle in aqueous phase. It undergoes a temperature-

induced volume phase transition (VPT) process

manifested as swelling-shrinking behaviour under

varying temperature (Yang et al., 2004, Burmistrova

et al., 2011, Chen et al., 2013, Zhang and Wang, 2009,

Yamashita et al., 2003).

3. THERMO-RESPONSIVENESS OF PNIPAM-

co-AA POLYMER HYDROGEL

Physically, the hydrogel swells below its volume

phase transition temperature (VPTT) (~32 °C)

because its hydrophilic functional ligands like amides

(-CONH) and carboxylic (-COOH) and carbonyl (-

C=O) tend to form hydrogen bonds with water

molecules which leads to expansion of the hydrogel

network. As temperature increases above its VPTT,

the hydrophilic groups are hidden and forming strong

inner hydrogen bonds within hydrogel interior. The

formation of short-range hydrogen bonds among

ligands causes expulsion of water from hydrogel and

shrinkage of the entire hydrogel network

conformation. On the other hand, hydrophobic groups

like isopropyl and methyl groups are gradually

exposed to the exterior of hydrogel at temperature

above VPTT, and the shrinking mechanism causes

hydrogel to be separated from aqueous phase (Chen et

al., 2013, Zhang and Wang, 2009, Yamashita et al.,

2003).

The thermo-responsive behaviour of PNIPAM-co-

AA polymer hydrogel before and after copper

complexation is characterized by using dynamic light

scattering (DLS) technique. The change of

hydrodynamic diameter, Dh is plotted within

temperature range of 25 – 50 °C. Figure 1(a) and (b)

show the temperature-induced Dh curves for PNIPAM

before and after complexing with copper ions,

respectively. The Dh of hydrogel before complexation

with copper decreases from 610 to 380 nm with

increasing temperature from 25 to 50 ºC. After

copper-hydrogel complexation, the sharp reduction in

Dh above VPTT had disappeared. This indicates that

the binding of Cu2+

with functional groups had

replaced the water-associated hydrogen bonds. The

ordinary trend of Dh change with temperature as in

Figure 1(a) is disturbed. The curve in Figure 1(b)

shows that the rate of Dh change is lowered for both

heating and cooling cycles. More pronounced

hysteresis effect is observed between heating and

cooling curves in Figure 1(b). This is because copper

chelation had reduced water-associated ligands and

chelation of Cu2+

ion by different ligands causes

certain extent of irreversible shrinkage (Chen et al.,

2013, Zhang and Wang, 2009, Yamashita et al., 2003,

Zhang et al., 2005, Sun et al., 2011).

4. FACTORS THAT AFFECT COMPLEXATION

OF COPPER BY PNIPAM-co-AA

The affinity of chelating ligands in PNIPAM-co-AA

toward copper ions in aqueous solution are sensitively

influenced by several factors including pH, density of

ligands (chelating functional groups) and temperature

(thermo-responsive polymers) (Flora and Pachauri,

2010, Hancock and Martell, 1989).

Chen et al.



186

(a) (b) Fig. 1: Thermo-responsiveness of hydrogels: (a) before and (b) after copper complexation

4.1. Effect of pH

pH is a critical parameter for influencing the rate,

efficiency and stability of metal-polymer complex

formation by altering the charges of ligands. Most of

the polymers bearing functional groups like –NH

(amine) and -COOH (carboxylic acid) as chelating

ligands are inactive for chelating metal ions at low pH

(< pKa) due to protonation of the functional groups

and electron donors are not available for donating

electrons to metal ion. As pH increases above pKa of

the acidic ligand of polymer, the functional group

(ligand) is deprotonated and increases affinity of the

ligand toward metal ion. Nevertheless, at high pH (≥

pH 6), the formation of insoluble metal hydroxide

precipitates reduces free metal ions from being

chelated by polymer ligands leading to a decline in

metal-chelating efficiency of the polymer.

In the case of copper, it exists as a hydrated

tetracopper cationic complex, [Cu(H2O)4]2+

at pH ≤ 6

in aqueous solution. At pH range of 3 – 5,

mononuclear complexes of one ligand (LM+) or two

ligands (L2M) are formed. In condition where the

concentration ratio of ligands to copper cations is low

(<10), formation binuclear complexes (two central

copper ions) is induced at pH 4 – 8. Then at pH ≥ 6,

insoluble hydroxide species of copper, Cu(OH)2 is

precipitated. In the presence of high concentration of

carboxylate ligands, Cu(OH)2 tends have hydrophilic

interaction with the polymer ligands through hydrogen

bonds. The formation of binuclear, mononuclear

copper-ligand complexes and Cu(OH)2 species leads

to the increase in adsorption capacity of polymer

hydrogels with increasing solution pH at range of pH

4 – 7 (Yao et al., 2010, Powell et al., 2007, 2011).

Figure 2 shows the pH effect on copper removal

efficiency in percentage % in the presence of

PNIPAM-co-AA hydrogels. A sharp increase

observed from 26.6% to 53.4% when the solution pH

changed from 4.0 to 5.0. It is because increase in pH

value leads to increase in OH- hydroxyl ions from the

base added (NaOH) that consume H+ ions in the

solution. Consequently, lesser free H+ ions compete

with Cu2+

ions for the chelation sites of ligands. From

pH 6 to pH 9, copper removal was increased linearly

from 63.2% to 87.6% due to the fact that hydroxides

of copper started to form at pH 6 and it was increasing

at higher pH. The dominating species of copper at pH

3 – 5 is Cu2+

while at pH 6 and above, insoluble

Cu(OH)2 is dominating. At pH 6, Cu2+

is still

dominant in the solution, but the presence of

monohydroxo Cu(OH)+ which possess single charge

was adsorbed to a greater extent than divalent Cu2+

and slight Cu(OH)2 were also formed under this

condition. Thus, the increase in removal efficiency at

pH 5 is owed to deprotonation of AA ligands, whereas

at pH 6 and above, Cu(OH)+ and Cu(OH)

2 species had

significant effect towards enhancement of copper

removal (Yao et al., 2010, Powell et al., 2007, 2011).

4.2. Effect of Ligand Density

The density of chelating functionalities (ligands)

indicates the amount of dentates present in a polymer

hydrogel. As previously discussed, a ligand could be

mono-, bi- or poly-dentate (electron pairs) which acts

as donor groups donating electron pairs to a metal ion.

The arrangement and spacing among chelating

functionalities are also important in determining the

interactions among ligands. Increase in ligand density

results in narrower spacing among ligands and eases

more ligand-ligand interaction through hydrogen

bonds that may reduce active dentates for chelating

metal ion (Yang et al., 2004, Burmistrova et al., 2011,

Yamashita et al., 2003, Zhang et al., 2005).



187

Fig. 2: Effect of pH on copper removal %.

Figure 3 shows the equilibrium adsorption capacity,

qe of hydrogels increased from 0.0943 to 0.8397 mmol

g-1

as AA content incorporated in a hydrogel increased

from 0 mol% (PNIPAM monopolymer) to 30 mol%.

Nevertheless, as it was increased to 50 mol%, there is

a slight fall from 0.8397 to 0.7868 mmol g-1

. This is

ascribed at too high AA content (50 mol% AA),

hydrogels has high tendency to agglomerate and

aggregate into colloids owing to excessive ligand-

ligand interaction. The active chelating sites and

surface area of hydrogel networks are greatly reduced.

This results in the lowering ratio of copper-binding

sites to total Cu2+

concentration (Yamashita et al.,

2003, Zhang et al., 2005).

Fig. 3: Effect of AA content in PNIPAM-co-AA on qe.

4.3. Effect of Temperature

Figure 4 compares the curves of qe versus initial Cu2+

concentrations (C0) at three different temperatures:

303, 313 and 323 K. It shows that at lower C0 ranging

from 20 to 50 mg L-1

, temperature effect is apparently

significant on qe, where it decreases with increasing

temperature. This trend agrees with the postulates

proven by Morris et al. (1997) and Kaşgöz et al. (2006)

that diffusional resistance is a problem for metal ions

adsorption as the reductions in both interior volume

and active chelating sites (ligands) at higher

temperature (Morris et al., 1997, Kaşgöz et al., 2006,

Hou et al., 2008). However, as C0 rose from 60 to 80

mg L-1

, the increase in qe is getting independent of

temperature effect. At higher concentration of Cu2+

,

higher concentration gradient and osmotic pressure

created by the copper ions is prevailing and

overcomes diffusional resistance within hydrogel

interior.

5. MEMBRANE FOULING BY

CONCENTRATION POLARIZATION (GEL

LAYER FORMATION)

Most of the PEUF systems are operated at constant

applied pressure typically in the range of 1 – 5 bar,

where the feed is circulated at a fixed pressure across

a membrane and the permeate is collected for

measurements over time. As steady pressure is

maintained throughout an operation, convective

transport of complex solutes by the bulk feed stream

Chen et al.



188

towards membrane could cause the accumulation of

retained solute particles at the active layer of

membrane. These retained solutes remain stagnant at

the membrane surface because their large size and

solute-membrane interactions prevent them from

diffusing back into the bulk stream (Porter, 1990,

Baker, 2012, Wang et al., 2011, Palencia et al., 2009,

Camarillo et al., 2010).

As the concentration of solutes deposited at

membrane surface (Cs) rises to the point of incipient

gel precipitation, the solute particles consolidate by

solute-solute interaction and form a gel layer which

acts as if a ‘secondary membrane’ on the membrane

surface. This gel layer of retained solutes creates a

fouling resistance (Rf) as secondary barrier to the

permeation of solvent flow through membrane. This

type of fouling phenomenon is known as

concentration polarization (CP). Concentration

polarization is the major fouling mechanism in nearly

all PEUF studies. It is reversible in nature when

solute-solute interaction is dominating in the process

as the gel layer can easily be scavenged by

backwashing or feed flow scouring, whereas it could

also become irreversible if solute-membrane

interaction is dominant instead (Wang et al., 2011,

Palencia et al., 2009).

Fig. 4: Plot of qe of PNIPAM-co-AA as a function of C0 at different temperatures (303, 313 and 323 K).

Since the gel layer offers resistance to the

permeation of solvent through membrane, it leads to

decline in permeate flux. Hence, the formation of gel

layer is usually described in term of permeate flux

which is expressed mathematically as follows:

(1.7)

where µ is the solution dynamic viscosity (Pa·s);

∆P represents transmembrane pressure difference,

whereas Rm and Rc refer to hydraulic resistances of

membrane and gel layer, respectively (Porter, 1990,

Baker, 2012, Wang et al., 2011).

5.1. Effect of hydrogel (Chelator) Concentration

Figure 5(a) presents the permeate fluxes obtained for

different polymer PNIPAM-co-AA hydrogel

concentration, Ch. Increasing concentration of

hydrogel in the feed enhances the formation of gel

layer at membrane surface. As expected, the permeate

flux plunged sharply from 72.38 L m-2

h-1

for the feed

with 0.01 g L-1

hydrogel to 11.75 L m-2

h-1

in the

presence of 0.1 g L-1

hydrogel. Subsequently, the flux

decreases slowly from 11.75 to 6.90 L m-2

h-1

as the

concentration increases from 0.1 to 0.7 g L-1

. The

increase in hydrogel concentration provides more

available active site for chelating Cu2+

cations in the

feed. This factor in turn contributes to more copper-

hydrogel complexes formed and eases the formation

of gel layer at the membrane surface.

Figure 5(b) illustrates the percentage of copper

retained under different hydrogel concentrations. In

the absence of hydrogels, around 13.1% of Cu2+

was

retained by the functional groups at membrane active

layer. As hydrogel was added into the feed at

concentration of 0.1 g L-1

, retention rises up to

81.45% and slightly increases again until 84% as

hydrogel concentration increased further to 0.7 g L-1

.

Two phenomena responsible for the improved copper

rejection are: 1) gel layer formation due to CP driven

by the convective transport of hydrogels onto

membrane surface 2) The increase in hydrogel

concentration provides more ligands for chelating

Cu2+

(da Silva et al., 2007, Childress and Elimelech,

1996, Tokuyama et al., 2005, Wijmans et al., 1985).



189

(a) (b)

Fig. 5: Effect of P-30 hydrogel concentrations on: (a) permeate flux and (b) Cu

2+ rejection (50 ppm Cu

2+ in feed; T = 303 K;

pH =5; transmembrane pressure, ∆P = 1.5 bar).

The AFM images shown in Figure 6 illustrate the

effect of different hydrogel concentrations (0.01, 0.1,

0.5 and 0.7 g L-1

) toward the membrane (cellulose

acetate) surface characteristic. At 0.01 g L-1

, the

lowest density of hydrogels in its spherical structure

was found to be bound on the membrane surface with

a lot of unblocked membrane pores as shown in

Figure 6.0(a). These accounts for the highest permeate

flux (72.38 L m-2

h-1

) and the lowest Cu2+

rejection

(33.1 %). As hydrogel concentration increases, more

hydrogel particles are bound and consolidated at the

membrane surface as shown in Figure 6.0(b-d). At

higher concentration of 0.7 g L-1

, hydrogels tend to

agglomerate into lumps by intense hydrophobic

interaction. However, these lumps did not improve

permeate flux at 50°C owing to the fact that under

such high concentration, the feed is relatively more

viscous and hydrogels are more densely packed within

gel layer. As a result, the more rigid hydrophobic gel

layer hinders permeation of solvent.

6. CONCLUSION

PNIPAM-co-AA hydrogel is a potential chelating

agent for copper ion which present in abundance in

most of the industrial effluent. The chelating ability

of the hydrogel is very much depending on the

solution pH, ligand density as well as temperature due

to its thermoresponsive properties. At low copper

concentration, the adsorption process is highly

diffusional control therefore very sensitive to the

operating temperature. Lower temperature is more

favorable due to the swollen hydrogel that exerts less

diffusional resistance for copper to bind within the

interior marix. In a PEUF system, filtration at low

temperature (< 40 °C) encountered low permeation

flux due to formation of gel layer of polymer

hydrogels at the membrane surface. However, high

copper rejection was achieved owing to the copper-

ligand complexation and gel layer provides hydraulic

resistance to the permeation of copper through

membrane. The same phenomenon was observed for

high polymer loading which shows that hydrogel

agglomeration at the surface greatly reduces the

membrane flux. Continuos effort to enhance both

copper removal as well as permeation flux is required.

ACKNOWLEDGEMENT

The authors wish to thank the financial support

granted by Ministry of Higher Education Malaysia

(MOHE) FRGS (203/PJKIMIA/6071252)

Chen et al.



190

(a) (b)

(c) (d)

Fig. 6: AFM images of membrane surface after PEUF processes for different hydrogel concentration: (a) 0.01 g L

-1, (b) 0.1 g

L-1

, (c) 0.5 g L-1

and (d) 0.7 g L-1

.

REFERENCES

Lazaridis NK, Peleka EN, Karapantsios TD, Matis

KA (2004). Copper Removal from Effluents by

Various Separation Techniques.

Hydrometallurgy, 74: 149 – 156.

Dermentzis K, Davidis A, Papadopoulou D,

Christoforidis A, Ouzounis K (2009). Copper

Removal from Industrial Wastewaters by

Means of Electrostatic Shielding Driven

Electrodeionization. Journal of Engineering

Science and Technology Review, 2: 131 – 136.

Ghodbane I, Nouri L, Hamdaoui O, Chiha M (2008).

Kinetic and Equilibrium Study for the Sorption

of Cadmium(II) Ions from Aqueous Phase by

Eucalyptus Bark. Journal of Hazardous

Materials, 152: 148 – 158.

Micheals AS (1990). Membranes, Membrane

Processes, and Their Applications: Needs,

Unsolved Problems, and Challenges of the

1990's. Desalination, 77: 5 – 34.

Almutairi FM, Lovitt RW (2012). Polymer Enhanced

Ultrafiltration (PEUF) for the Removal of

Metals from Water: Influence of Competing

Solutes and Chelators. International Journal of

Chemical and Environmental Engineering, 3:

15 – 25.

Camarillo R, Pérez Á, Caῆizares P, de Lucas A (2012).

Removal of Heavy Metal Ions by Polymer

Enhanced Ultrafiltration. Batch Process

Modeling and Thermodynamics of

Complexation Reactions. Desalination, 286:

193 – 199.

Zerze H, Karagoz B, Ozbelge HO, Bicak N, Aydogan

N, Yilmaz L (2013). Imino-bis-propane Diol

Functional Polymer for Efficient Boron

Removal from Aqueous Solutions via

Continuous PEUF Process. Desalination, 310:

158 – 168.

Flora SJS, Pachauri V (2010). Chelation in Metal

Intoxication. International Journal of

Environmental Research and Public Health, 7:

2745 – 2788.

Butvin P, Skoumal M, Rádl Z, Peterka V, Majer J

(1988). Protonation Constants and Overall

Constants of Complex Formation of Copper(II)

Ions with Surface-active Ligands Derived from

Ethylenediamine. Chemical Papers, 42: 475 –

482.

Porasso RD, Benegas JC, van den Hoop MA, Paoletti

S (2000). Analysis of Potentiometric Titrations

of Heterogeneous Natural Polyelectrolytes in

Terms of Counterion Condensation Theory:



191

Application to Humic Acid. Biophysical

Chemistry, 86: 59 – 69.

Hancock RD, Martell AE (1989). Ligand Design for

Selective Complexation of Metal Ions in

Aqueous Solution. Chemical Reviews, 89: 1875

– 1914.

Cañizares P, Pérez Á, Camarillo R (2002). Recovery

of Heavy Metals by Means of Ultrafiltration

with Water-soluble Polymers: Calculation of

Design Parameters. Desalination, 144: 279 –

285.

İslamoğlu S, Yilmaz L (2006). Effect of Ionic

Strength on the Complexation of

Polyethyleneimine (PEI) with Cd2+

and Ni2+

in

Polymer Enhanced Ultrafiltration (PEUF).

Desalination, 200: 288 – 289.

Llanos J, Pérez Á, Cañizares P (2008). Copper

Recovery by Polymer Enhanced Ultrafiltration

(PEUF) and Electrochemical Regeneration.

Journal Membrane Science, 323: 28 – 36.

Labanda J, Khaidar MS, Llorens J (2009). Feasibility

Study on the Recovery of Chromium (III) by

Polymer Enhanced Ultrafiltration. Desalination,

249: 577 – 581.

Yang JX, Fang Y, Bai CL, Hu DD, Zhang Y (2004).

CuS-poly(N-isopropylacrylamide-co-acrylic

acid) Composite Microspheres with Patterned

Surface Structures: Preparation and

Characterization. Chinese Science Bulletin, 49:

2026 – 2032.

Burmistrova A, Richter M, Eisele M, Uzum C,

Klitzing RV (2011). The Effect of Co-monomer

Content on the Swelling / Shrinking and

Mechanical Behavior of Individually Adsorbed

PNIPAM Microgels. Polymers, 3: 1575 – 1590.

Chen JJ, Ahmad AL, Ooi BS (2013). Poly(N-

isopropylacrylamide-co-acrylic acid) Hydrogels

for Copper Ion Adsorption: Equilibrium

Isotherms, Kinetic and Thermodynamic Studies.

Journal of Environmental and Chemical

Engineering, 1: 339 – 348.

Zhang F, Wang CC (2009). Preparation of P(NIPAM-

co-AA) Microcontainers Surface-anchored

with Magnetic Nanoparticles. Langmuir, 25:

8255 – 8262.

Yamashita K, Nishimura T, Nango M (2003).

Preparation of IPN-type Stimuli-responsive

Heavy-metal-ion Adsorbent Gel. Polymers for

Advanced Technologies, 14: 189 – 194.

Zhang YW, Jiang M, Zhao JX, Ren XW, Chen DY,

Zhang GZ (2005). A Novel Route to

Thermosensitive Polymeric Core-shell

Aggregrates and Hollow Spheres in Aqueous

Media. Advanced Functional Materials, 15: 695

– 699.

Sun S, Hu J, Tang H, Wu PY (2011). Spectral

Interpretation of Thermally Irreversible

Recovery of Poly(N-isopropylacrylamide-co-

acrylic acid) Hydrogel. Physical Chemistry

Chemical Physics, 13: 5061 – 5067.

Yao ZY, Qi JH, Wang LH (2010). Equilibrium,

Kinetic and Thermodynamic Studies on the

Biosorption of Cu(II) onto Chestnut Shell.

Journal of Hazardous Materials, 174: 137 – 143.

Powell KJ, Brown PL, Byrne RH, Gajda T, Hefter G,

Sjöberg S, Wanner H (2007). Chemical

Speciation of Environmentally Significant

Metals with Inorganic Ligands. Part 2: The

Cd2+

, OH-, Cl

-, CO3

2-, SO4

2- and PO4

3- Systems

(IUPAC Technical Report). Pure and Applied

Chemistry, 79: 895 – 950.

Powell KJ, Brown PL, Byrne RH, Gajda T, Hefter G,

Leuz AK, Sjöberg S, Wanner H, (2011).

Chemical Speciation of Environmentally

Significant Metals with Inorganic Ligands. Part

4: The Cd2+

, OH-, Cl

-, CO3

2-, SO4

2- and PO4

3-

systems (IUPAC Technical Report). Pure and

Applied Chemistry, 83: 1163 – 1214.

Morris GE, Vincent B, Snowden MJ (1997).

Adsorption of Lead Ions onto N-

isopropylacrylamide and Acrylic Acid

Copolymer Microgels. Journal of Colloid and

Interface Science, 190: 198 – 205

Kaşgöz H, Kaşgöz A, Șahin Ü, Temelli TY, Bayat C

(2006). Hydrogels with Acid Groups for

Removal of Copper (II) and Lead (II) Ions.

Polymer-Plastics Technology and Engineering,

45: 117 – 124.

Hou YP, Matthews AR, Smitheman AM, Bulick AS,

Hahn MS, Hou HJ, Han A, Grunlan MA (2008).

Thermoresponsive Nanocomposite Hydrogels

with Cell-releasing Behavior. Biomaterials, 29:

3175 – 3184.

Porter MC (1990). Handbook of Industrial Membrane

Technology. Noyes Publications, New Jersey.

Baker RW (2012). Membrane Technology and

Applications. John Wiley and Sons, West

Sussex, UK.

Wang LK, Chen JP, Hung Y, Shammas NK (2011).

Membrane and Desalination Technologies.

Springer, New York.

Palencia M, Rivas BL, Pereira E (2009). Metal Ion

Recovery by Polymer-enhanced Ultrafiltration

Using Poly(Vinyl Sulfonic Acid): Fouling

Description and Membrane-metal Ion

Interaction. Journal of Membrane Science, 345:

191 – 200.

Camarillo R, Llanos J, García-Fernández L, Pérez Á,

Cañizares P (2010). Treatment of Copper (II)-

loaded Aqueous Nitrate Solutions by Polymer

Enhanced Ultrafiltration and Electrodeposition.

Chen et al.



192


320 – 328.

da Silva RMP, Mano JF, Reis RL (2007). Smart

Thermoresponsive Ccoatings and Surfaces for

Tissue Engineering: Switching Cell-material

Boundaries. Trends in Biotechnology, 25: 577 –

583.

Childress AE, Elimelech M (1996). Effect of Solution

Chemistry on the Surface Charge of Polymeric

Reverse Osmosis and Nanofiltration

Membranes. Journal of Membrane Science, 119:

253 – 268.

Tokuyama H, Kanazawa R, Sakohara S (2005).

Equilibrium and Kinetics for Temperature

Swing Adsorption of a Target Metal on

Molecular Imprinted Thermosensitive Gel

Adsorbents. Separation and Purification

Technology, 44: 152 – 159.

Wijmans JG, Nakao S, van den Berg JWA, Troelstra

FR, Smolders CA (1985). Hydrodynamic

Resistance of Concentration Polarization

Boundary Layers in Ultrafiltration. Journal of

Membrane Science, 22: 117 – 135.



©2014 IJSRPUB

194

Wastewater Treatment by Phytoremediation Methods

Hossein Farraji

School of Civil Engineering, Engineering Campus, University Sains Malaysia, 14300 Nibong Tebal, Penang, Malaysia

Email: [email protected]

Abstract. Nowadays many technologies are using for treatment of environmental pollutions and phytoremediation as a green

technology is going on to convert to one of the main ecofriendly technologies which scientist using in their researches. Aquatic

media as a fundamental and critical part of human environment have main role in water resources and food chain. In this

chapter we present different aspect and types of phytoremediation in aquatic media purification from metallic elements.

Keywords: Wastewater, Treatment, Phytoremediation

1. INTRODUCTION

The anthropogenic emissions of metallic elements

globally ends up to wastewater and industrialization

caused a lot of environmental problems .Heavy metals

spectacular Zn, Cu, Ni and Cr are most common

contain of industrial wastewater (Sun and Shi, 1998).

For example recently report from China indicate that

anthropogenic emissions of cadmium to atmosphere

from 1990 to 2010 increased from 474 to 2186 ton

(Shao et al., 2013). Most Cd is emitted to the

atmosphere at the first stage and then precipitate on to

water and soil (Yamagata, 1979). Meanwhile

vulnerability of ground water as drinking water

sources, is higher than surface water (Landmeyer,

2012). levels of arsenic in groundwater sources is

higher in comparing with surface-water sources

(Bissen and Frimmel, 2003). On the other hands,

concentrate ions and accumulation of some pollutants

same as 137

Cs in aquatic media are higher than

sediment or soils (Ashraf et al., 2013), so it shows

critical statue of decontamination of wastewater (Shao

et al., 2013).

2. PYTOREMEDIATION

Phytoremediation is vast, emerging term which has

been used in recent decades for a group of green

ecofriendly technologies that fundamentally based on

plants (aquatic, semiaquatic and terrestrial) and

related associated enzymes, microorganism and water

consumption, uptake, remove, retain, transform,

degrade or immobilize contamination (organic and/or

inorganic) with different origin, from soil, sediment

and aquatic media or atmosphere (Mueller et al., 1999,

Pivetz, 2001, Ghosh and Singh, 2005; Vishnoi and

Srivastava, 2007; Mulbry et al., 2008; Ridzuan et al.

2010, Dhir 2013). To sum it up, phytoremediation in

aquatic media is directly up take and accumulation of

contaminant from water media and assimilation by

plants (Ndimele and Ndimele, 2013). Two approaches

have been presented in literatures about

phytoremediation, natural or continuous

phytoremediation and chemically enhanced

phytoremediation (Lombi et al., 2001; Alkorta et al.,

2004).

2.1. Continuous or natural phytoremediation

This type of phytoremediation fundamentally based

on using natural hyperaccumulator plants with not

only exceptional metal accumulation in shoots but

also extraordinary tolerance to metal toxicity in root

around media (Assunção et al., 2003; Prasad, 2004).

2.2. Chemically enhanced phytoremediation

Natural phytoremediation have following drawbacks

(Hem, 1970; Prasad, 2004; Yoonet al., 2006):

(a) Low biomass and slow growth in

hyperaccumulator plants; (b) Low translocation of

metals in hyperaccumulator plants; (c) Insolubility or

immobility of metallic elements in media; (d) Low up

taking element by roots.

A chemical enhancement for phytoremediation

helps to overcome these limitations. Chelating ligands

like as ethylene diamine tetra-acetic acid (EDTA), 1,

3-propylene-diaminetetra-aceticacid (PTDA), nitrilo-

tri-acetic acid (NTA), diethyl triamine penta-acetic

acid (DTPA), etc. Which are microbial poorly

degradable and present in µg/L concentration in

aquatic media (Knepper, 2003), reduce net shoot and

root biomass production (Römkens et al., 2002).

Several aspect of phytoremediation improved and

established based on mechanism of decontamination,

kind of pollutions, affected factors and media of

remediation (Stout and Nüsslein 2010; Dhir, 2013).

Table 1 is summarized typical kind of this green

technology and its process.


Faraji


195

Table1: Phytoremediation process (Dhir, 2013)

2.3. Advantages of aquatic phytoremediation

These green technologies as an ecofriendly present

many advantages for decontamination of aquatic

polluted media comparing other treatment systems

(Ximénez-Embún et al., 2001; Bissen and Frimmel

2003; Ghosh and Singh, 2005; Kirkham, 2006; Verma

et al., 2007; Olguín and Sánchez-Galván, 2010;

Landmeyer, 2012; Wang and Calderon, 2012; Delmail

et al., 2013).

(a) Greatly cost effective; (b)

Environmentally Compatible, green aesthetically; (c)

Feasible globally technology (no need to high-tech

equipment or material); (d) Easy maintenance (supply

and energy) solar powered; (e) In situ and ex situ

operation available; (f) Inhibiting deployment of

contamination to atmosphere or soil by in situ

operation; (g) Suitable for shallow depth contaminated

water to hydrologic control of ground water; (h)

Periodic treatment against continued treatments; (i)

Different types of many contaminations can be treated

in one time; (j) Produce biomass for renewable energy

production; (k) Rapid mass propagation by tissue

culture available; (l) Almost treatable for all kind of

contaminations (organic, inorganic and radionuclides);

(m) Operable approximately in all media (aquatic,

sediment, soil and atmosphere); (n) Broad acceptable

pH (2-10) for treatment; (o) Chelating molecules

positively affected phytoremediation;

2.4. Disadvantages of aquatic phytoremediation

Same as each technology, phytoremediation suffer

from disadvantages which should considered through

applying (Mojiri et al. 2013; Assunção et al., 2003,

Alkorta et al., 2004; Ghosh and Singh, 2005; Shiyab

et al., 2009; Delmail et al., 2013, Kumar et al., 2013;

Nan et al., 2013).

(a) Harvested biomass contains hazardous

pollution; (b) Long time need for effective treatment;

(c) Deep polluted media are limited treating; (d)

Limited to climate growth condition (tropical,

subtropical and tempered zoon); (e) Restricted to low

concentration of polluted site; (f) Harvested biomass

managing required; (g) Limited hyperaccumulator

species introduced for aquatic media; (h) Leaves fall

may cause spread contaminant; (i) Most of

hyperaccumulator plants have limited roots and slow

growing; (j) Most of hyperaccumulator species up

take only limited elements; (k) Many of

hyperaccumulators propagation system is sexual; (l)

Recycling to soil and water by rain in volatilized

pollutions; (m) Possibility of entering contaminated

biomass to animal and human food chain

2.5. Phytovolatilization

This kind of phytoremediation involves the utilization

of plants to take up pollutions from contaminated

media, transforming them into volatile form and

finally transpiring them into the air.

phytovolatilization normally occurs in plants for up

taking water, organic and inorganic compounds so

some of contamination can pass through the plant

parts to the leaves and at low concentration ,volatilize

into the atmosphere (Mueller et al., 1999) In a

nutshell, the use of plant species for volatilize

contaminants from the leaves which can use for soil

and sediment pollutions (Mueller et al., 1999) air

contaminations (Burken and Schnoor, 1999) and water


Chapter 7: Wastewater Treatment by Phytoremediation Technologies

196

pollution (Baeder-Bederski-Anteda, 2003). The

phytovolatilization radionuclides such as Tritium(3H)

from soil (Dushenkov, 2003),significantly up taking

selenium compounds contain dimethyldiselenide and

dimethylselenide by Brassica species (Bañuelos et al.,

2000) are some sample studies on phytovolatilization.

Disadvantages of mercury phytovolatilization is

recycling by rain and a residue back into the

ecosystem (Alkorta et al., 2004).

2.6. Phytodegradation

Phyodegradation or phytotransformation refers to two

kind of plant reactions it is independent on

rhizospheric microorganisms (Vishnoi and Srivastava,

2007) contain phytoreduction which is the reductive

transformation of oxidized organic compounds by

reducing plant enzymes and phytooxidation which is

the oxidative transformation reactions catalyzed by

plant oxidizing enzymes (Nzengung and Jeffers,

2001).Plant enzymes contain five various

nitroreductaces (Schnoor et al., 1995), dehalogenase

and oxygenase (Vishnoi and Srivastava, 2007)

degrade organic compounds. The main

phytodegradation process for organic contamination in

plants are, uptake, translocation and metabolism

(Dzantor et al., 2002).Taken up of most of Polycyclic

aromatic hydrocarbons (PAHs)is hard for plants and

less suitable for phytodegradation (Reichenauer and

Germida, 2008). Some of the main organic compound

that successfully degraded by plant species are

hexachloroethane (HCA),

dichlorodiphenyltrichloroethane (DDT) (Nzengung

and Jeffers, 2001) and carbon tetrachloride (Wang et

al., 2004). Inorganics compounds like as sulfur oxides

and atmospheric nitrogen oxides can be taken up by

plants for degradation. Genetically modified plant

species also had been used for phytodegradation (Doty

et al., 2000).

2.7. Rhizodegradation

Rhizodegradation is breakdown organic

compounds(fuels and solvents) to plants nutrients in

rhizospher trough microbial (fungi, yeast , bacteria

and other microorganisms )activity which is very

slower than phytodegradation process and also called

phytostimulation (Hutchinson et al., 2003; Ghosh and

Singh, 2005; Ridzuan et al., 2010).This technology

widely applied in treatment of soils so that in

Germany there are more than 100 years background

for rhizoremediation (Wand et al., 2002). Generally,

plant exudates same as carboxylic , amino acids and

carbohydrates are capable to stimulate entire

rhizosphere microbial action and enhance

rhizodegradation proceeding (Dzantor, 2007) Some of

plants which applied for Rhizodegradation are

Hibiscus cannabinus for used lubricating oil (Abioye

et al., 2012), Sorghum bicolor L. for crude oil (Banks

et al., 2003), Broadleaf plantain plant (Plantago

major L.) for Imidacloprid (insecticide) in water and

soil (Romeh, 2009), willow (Salix babylonica)used for

perchlorate in soil and water (Mwegoha et al.

2007).Some treatments used for enhancing

rhizodegradation like as supplying organic carbon

(Yifru and Nzengung, 2008),chicken manure

(Mwegoha et al., 2007) and organic waste treatment

(Dadrasnia and Agamuthu, 2013).

2.8. Phytomining

Phytomining or bio-ore is a green technology which

could generate revenue from saleable metallic

elements that accumulated in ash of plants biomass

(Ghosh and Singh, 2005). Mining of nickele has been

already a patented technology (Chaney et al., 1998).

Normally phytomining occurs in situ at the

contaminated mine lands or sub-economical ore

(Anderson et al., 1998). Phytoextraction or

bioextraction of metals for commercial gain contain

cropping, harvesting, drying and ashing are main

process of phytomining (Sheoran et al., 2009). Cattails

(Typha latifolia) used for absorbing boron (B) from

the effluent of biggest borax mine of the world and

250mgkg-1 boron absorbed in constructed wetland

system (Türker et al., 2013).

2.9. Phytostabilization

This technology applied for decreasing bioavailability

of contamination from environment and stabilizing of

pollutants occurs more than removing

them(commonly metallic elements) by plants

(hydraulic control) ? (Padmavathiamma and Li, 2007).

Enhancing appropriate soil modification by plants

,caused to decreasing bioavailability of metallic

elements on the other hands plant cover decreased

leaching and enhance environmental protection

(Houben et al., 2011).Plants can help to stabilize

pollutants with up taking in adsorption system or

accumulate them in root system (Vangronsveld et al.,

2009). Plant selection for phytostabilizing is a critical

issue and perennial species, well local

environmentally adapted ,high biomass production

and high resistance to pollution recommended (Pilon-

Smits, 2005). Since there is no natural

hyperaccumulator phytoremediator for mercury

controlling this toxic metallic element is critical so

Indian mustard (Brassica juncea L.) reported as a

suitable plant species for stabilization of mercury in

soil and wastewater (Shiyab, Chen et al. 2009).

Aquatic plant Hydrilla verticillata reported as

Faraji


197

potential species for phytostabilization of wastewater,

in this plant has high translocation factor (TF) and low

bio concentration factor (BCF) for toxic metals

(Pb,Cr) (Ahmad et al., 2011). Rapeseeds (Brassica

napus), sunflowers (Helianthus annuus), tomatoes

(Solanum lycopersicum) and soapworts (Saponaria

officinalis) reported as capable plants for

phytostabilization with less than one bioaccumulation

coefficients obtained in all of these plants (Sung et al.,

2011). Major metallic element removal in plant

species phragmites (Phragmites australis), Typha spp,

Juncus spp, Spartina spp and Scirpus spp from

different wastewater contaminated area are

phytoextraction and phytostabilization (Bhatia and

Goyal, 2013).

2.10. Phytofiltration

Phytofiltration or rhyzofiltration is a green technology

for removing contaminations by plants roots in aquatic

media like as ground water, most of wastewaters and

extracted ground water (Pivetz, 2001; Mukhopadhyay

and Maiti, 2010). Terrestrial, aquatic and wetland

plants are suitable material for phytofiltration and

constructed wetlands are the best method for

removing metallic elements from wastewater (Cheng

et al., 2002). Limnocharis flava (L.) reported as

suitable plant species for phytofiltration of low

concentration Cd contaminated water (Abhilash et al.

2009) Wolffia globosa is a suitable nominated for

arsenic metabolism studying via phytofiltration

(Zhang et al. 2009).For effective accumulation of

cadmium and hyperaccumulation of arsenic,

micranthemum umbrosum introduced as suitable

macrophyte (Islam et al., 2013). Indian mustard

(Brassica juncea (L.) Czern) could uptake 95%

mercury from contaminated water via phytofiltration

(Moreno et al., 2008). Aquatic plants, Pteris

creticacvMayii (Moonlight fern) and Pteris.vittata

(Chinese brake fern) as arsenic hyperaccumulator

plants, used via phytofiltration (Alkorta et al., 2004;

Tu et al. 2004; Baldwin and Butcher, 2007). The

mechanism of phytofiltration is similar to

phytoextraction containe 1) chemisorption 2)

complexation 3) ion exchange 4)micro precipitation

5)hydroxide condensation 6) surface adsorption

(Cheng et al., 2002; Gardea-Torresdey et al., 2004).

Plant specification in phytofiltration should be 1)

tolerance to high level of metallic elements 2) facile

handling 3) high root biomass or surface area 4)

reasonable maintenance cost 5) least secondary waste

requirement (Dushenkov, 2003).

2.11. Phycoremediation

Phycoremediation is using macro and micro algae for

bio transforming or removing pollutions from

wastewater furthermore CO2 as air pollution from

waste air (Mulbry et al., 2008; Rawat et al., 2011).

Algae is a suitable plant for decontamination of

metallic elements, xenobiotic, nutrients in various

wastewater and consume of carbon dioxide from

exhausts ( gu , 2003). Micro algae is a capable

plant for treatment of various type of wastewater such

as industrial wastewater ,domestic wastewater and

solid wastes both aerobically and anaerobically

(Safonova et al., 2004). For this technology ,oldest

scientific research carried out in half century ago

(Oswald and Gotaas, 1957). In a pilot study on

chrome sludge, phycoremediation with Desmococcus

olivaceus could effectively decrease nitrate,

phosphate, ammonia, TDS ,TSS and chrome

(Sivasubramanian et al., 2010). Fresh water blue green

algae used successfully for treatment of dairy manure

effluent(Mulbry et al., 2008). Sewage water treated by

different algae and Chlorella vulgaris could remove

almost all of contaminations and after treating process

it can be thrown in water bodies 1n 2013.

2.12. Phytoextraction

Phytoextraction /phytoaccumulation (Khan et al.,

2000) can be considered as suitable green technology

for removing metallic elements from aquatic media

(Wang et al., 2008). This kind of remediation used for

accumulation of Zinc by duckwood (Lemna gibba)

(Khellaf and Zerdaoui, 2009), Cadmium by water

spinach (Ipomea aquatic) (Wang et al., 2008),

Chromium with small pondweed (Potamogeton

pusillus) in presence of Cu2+

(Monferrán et al., 2012).

recently reported study carried out for accumulation of

Pb by Ceratophyllum demersum and Myriophyllum

spicatum and finally introduced as phytoremediator

and bioindicator of Pb. (El-Khatib et al., 2014) Water

hyacinth (Eichhornia Crassipes) used for removing

heavy metals from coastal water (Agunbiade et al.,

2009), Crude oil from artificial wastewater amended

by urea fertilizer (Ndimele and Ndimele 2013) and

palm oil mill effluent treatment (Christwardana and

Soetrisnanto, 2013). Furthermore, heavy metals from

industrial wastewater have been removed by vetiver

(Chrysopogan zizanioides) (Roongtanakiat, 2009).

3. HEAVY METAL

Generally speaking, heavy metals which nowadays

known as metallic elements (Mbengue et al., 2014),

confined as element with density higher than 4.5- 5

mg mL-1

/kg dm -3

(Sarkar, 2002)but the collective



198

term includes arsenic(As), cadmium(Cd), lead(Pb),

Mercury (Hg), nickel(Ni), chromium (Cr), zinc (Zn),

copper (Cu), aluminum (Al), cesium (Cs), Ferrum

(Fe), manganese (Mn), molybdenum (Mo), radium

(Ra), uranium (U), strontium (Sr), platinum (Pt),

palladium (Pd), beryllium (Be) (McIntyre 2003)

Metallic elements are persistent and non-

biodegradable (Mahmood et al., 2010) .cooperation of

several disciplines engineering such as environmental,

civil, and chemical is fundamental requirement for

remediation of heavy metals (Bradl, 2005).

3.1. Metal removal rate

The rate of metallic elements removal from polluted

wetland depended on plant spices, climacteric

condition, statue of substrates, type of element

(Hg>Mn>Cd=Fe>Cr=Pb>Cu=Zn>Al>Ni>As), their

ionic forms (Marchand et al., 2010). Remediation

habit of plants species hyperaccumulator>

acuumulator> indicator> excluder (Figure 1) (Bradl,

2005).PH less than 5 is harmful for plants meanwhile

there is a report for a plant species (Lupinus) that can

growth in pH<2 and uptake 98% metallic elements

(Ximénez-Embún et al., 2001).Results of five aquatic

species indicated that all aluminum uptake increased

in low pH (Gallon et al., 2004).

3.2. Indexs of phytoremediation

There are some index that helps to calculate efficiency

of phytoremediation and assessment of plant ability in

up taking and translocation or mobilization (Kisku et

al., 2000) of metallic elements. these index use for

collecting hyperaccumulator plant species (McGrath

and Zhao, 2003).

3.2.1. Bioconcentration factor (BF)

The ratio of concentration of metallic elements in

plants tissue (roots, shoots) to that in contaminated

site (Tu et al., 2002),it also termed as bio

concentration factor (BCF) (Marchiol et al., 2004).

3.2.2. Enrichment Coefficient (EC)

It termed as degree of metallic elements accumulation

in plants (shoot, roots) to concentration of metallic

elements at contaminated site (Kisku et al., 2000). It

also termed as biological accumulation coefficient

(BAC).

3.2.3. Translocation (mobilization) factor (TF)

This factor calculated to determine the translocation of

metallic elements from the root of plant to shoot

(McGrath and Zhao, 2003).

where; TF in hyperaccumulator species is more than

one

3.2.4. Relative treatment efficiency index (RTEI)

Introduced by (Marchand et al., 2010)which is based

on comparing control treatment metal removal with

influent metal concentration and effluent treatment

metal removal ,meanwhile previous index are suitable

now (Kumar et al., 2013).

Fig. 1: Relative uptake and biological accumulation potential in plant species (Bradl, 2005)

Faraji


199

4. HYPERACCUMULATORH

Hyperaccumulator used as first time in 1948 by Italian

scientist (Fingerman and Nagabhushanam, 2005). The

plant spices which accumulate heavy metals more

than 1% dry weight, named hyperaccumulator

(Zavoda et al., 2001). Most common aquatic

hyperaccumulators are Spirodelta polyrhiza,

Eichhornia crassipes and Elodeu nuttali (Dhir et al.,

2009), meanwhile Marchand reported there is no

aquatic hyperaccumulator plant species(Marchand et

al., 2010). Tow pictures of native natural terrestrial

hyperaccumulator plants(Figure 2 and Figure 3)

collected in China and Malaysia(Xue et al. 2004; van

der Ent et al., 2013) .Effective phytoremediation

highly correlated to suitable plant selection

(Manousaki, Kadukova et al., 2008).High biomass

,deep and/or wide roots, easy harvesting easily

provide renewable energy and resistance to high

concentration of pollutant are common characters of

phytoremediators (Ghosh and Singh, 2005).

Collecting plants from high polluted area is scientist

preferred method for finding metallic elements

hyperaccumulators (Gleba et al., 1999).Through the

all plant species only 400 species (Brooks, 1998) or

500 species (Ma et al., 2001)reported as

hyperaccumulators. Heavy metals hyperaccumulators

belong to approximately 500 tax of angiosperm (less

than 0.2%) (Krämer, 2010). Recently study in China

introduced three aquatic hyperaccumulator species

for heavy metals Najas marina for arsenic (As) and

cadmium (Cd), Vallisneria natans for lead (Pb) and

Ceratophyllum demersum for cobalt (Co), chromium

(Cr) and ferrum (Fe) plant species selection for this

study carried out from native nominated plants from

studying site (Xing et al., 2013).Some of recently

carried out studies are listed in Table 2, and a number

of old researches added because of their importance.

5. FUTURE STUDIES

Micro propagation is the best way for mass production

of some aquatic machrophyte, since produce a lot of

unique genetically plants in short time for analytical

data collection (Kan et al. 1990),which may use for

phytoremediation of contamination particular metallic

elements, first report in this area reported by (Delmail

et al., 2013), There is no reports on natural Mercury

phytoremediator plants(Shiyab et al., 2009)and high

toxicity of this element for human (Harris et al.

2003)presents a critical circumstance for introducing

suitable candidate for phytoremediation of Mercury.

A genetically studying is trying to make Hg

hyperaccumulator by the Meagher laboratory

(Fingerman and Nagabhushanam, 2005).Nowadays

researches on different concentration of heavy metals

in sediments, soil and water, in unique case study also

(Falinski et al. 2014), introducing suitable spectacular

plant species for water, soil and sediment

phytoremediation(Nan, et al., 2013) and using

different enhancers like as microbial and Zeolite in

phytoremediation process(Karimzadeh et al., 2012) is

leading researches toward using constructed wetland

(CW) as a complete ecological system for treating

wastewater spectacular for continues

phytoremediation in practical treatment for example

removing metals from road runoff (Borne et al.

2014),electric industry wastewater (Yang et al., 2013)

fresh oilfield produced water (Alley et al., 2013).

Fig. 2: Nickel hyperaccumulator ultramafic area

Malaysia and the Philippine (van der Ent et al., 2013)

Fig. 3: Phytolacca acinosa Roxb.

Growing at Xiangtan manganese tailing ,China (Xue

et al., 2004)



200

Table 2: Summary of recent studies on aquatic hyperaccumulation Plant species Metallic

elements

Results Media Reference

Salvinia Minima(fern)

Pb2+

Phytochelatins cope with

pb2+ as one of mechanism

Artificially polluted

water

(Sánchez-Galván et al.,

2008)

Chara Australis (algae) Cd Chara is not a

hyperaccumulator

Sediment (Clabeaux et al., 2011)

Talinum Triangular Cu, Pb, Ni,

Cd

Stem cuttings are new Cu

hyperaccumulator

Hydroponic system (Rajkumar et al., 2009)

Cyperus rotundus L. Heavy

metals

New Sn

Hyperaccumulator

Soil and water media. (Ashraf et al., 2011)

Pteris vittata L.

As In pH:5.21 fern is

hyperaccumilator

hydroponic

experiment

(Tu et al., 2004)

Nasturtium officials

diplazium Esculentum

As These are not

Hyperaccumulator

Sediment soil

water

(Falinski et al., 2014)

Eichhornia crassipes

(Ash)

Heavy

metals

Pb,Zn,Cr,Cu and Ni

Hyperaccumulator

Artificially polluted

water

(Mahmood et al., 2010)

pistia stratiotens Cd Cd hyperaccumulator surface waters (Das et al., 2013)

Lemna minor Hg Phytoremediator Natural lake water (Isaksson et al., 2007)

Ceratophyllum demersum

Lemna gibba

Pb, Cr 2 Effective

phytoremediators

Artificially

contaminated water

(Abdallah, 2012)

Eleacharis acicularis Heavy

metals

Pb hyperaccumulators abandoned mining and

sediment

(Ha et al., 2009)

Salvinia minima (fern) Pb2+

Accumulation of Pb

increased Glutathiane

synthase

artificially

contaminated

water

(Estrella-Gómez et al., 2012)

Echonrnnia crassipes

Spirodela polyrhiza

Fe, Mn,

Cu, Cd,

Pb, Cr

Suitable

Bioindicators or

bioaccumulator

electric industries

waste water

(Sahu et al., 2007)

Eichhornia crassipes

Spirodela polyrrhiza

pistia stratiotens

Fe,Zn,Cu,

Cr,Cd

3 Suitable

phytoremediators

artificially

contaminated

water

(Mishra and Tripathi, 2008)

Myriophyllum

aterniflorum Cd,Cu Hyperaccumulator aquatic media (Delmail et al., 2013)

1-Najas marina

2-Ceratophyllum

demersum

3-Vallisneria natanas

Heavy

metals

1-As and Cd

hyperaccumulator

2- Co,Cr and Fe

hyperaccumulator

3-Pb hyperaccumulator

24 eutrophic lakes (Xing et al., 2013)

Alternantbera philoxeroids heavy

metals

Zn and Cd

hyperaccumulator

local river (Nan et al., 2013)

Myriophyllum vercillatum Pb,Cu,Cd

and Zn

May be hyperaccumulator

for Pb2+

hydroponic system (Ucer et al., 2013)

Pistia stratiotes L. heavy

metals

Cr,Cu,Fe,Mn,Ni,Pb, and

Zn hyperaccumulator

Surface water (Pant and Singh, 2013)

Eichhorna crassipes Cd,Ni The roots are effective in

absorbing most of Cd and

Ni

Artificially

contaminated water

(Elfeky et al., 2013)

Potamogeton natans

Alismaplantago aquatica

Filipendula ulmaria

Cu,Zn,Pb These plants are the most

efficient plant for up taking

heavy metals

Storm water (Fritioff and Greger, 2003)

Catharanthus roseus 137

Cs 90

Sr

Ideal hyperaccumulator

For radionuclides

Artificially

contaminated water

(Eapen et al., 2009)

Chrysopagann zizanioides Mn,

Fe,Cu,Zn,

Pb

Harvested plant safely can

use for compost or

handicrafts

Industrial wastewater (Roongtanakiat, 2009)

Leersiahexandra Cr Hyperaccumulator electroplating factory

wastewater

(Zhang et al., 2007)

Callitriche cophocarpa Cr Suitable bioaccumulator Artificially

contaminated water

(Augustynowicz et al., 2013)

Eichhorna crassipes

Ipomea aquatic

Typha angustata

Zn,Cu,Pb,

Ni, Co,Cd

Three native aquatic plants

are suitable species for

phytoremediation

Sediment and water

Gujarat wetland

(Kumar et al., 2008)

Lemna minor Pb Suitable phytoremediator

in low concentration

Artificially

contaminated water

(Bianconi et al., 2013)

Faraji


201

REFERENCES

Abdallah MAM (2012). Phytoremediation of heavy

metals from aqueous solutions by two aquatic

macrophytes, Ceratophyllum demersum and

Lemna gibba L. Environmental technology,

33(14): 1609-1614.

Abhilash P, Pand VC, Srivastava P, Rakesh P,

Chandran S, Singh N, Thomas A (2009).

Phytofiltration of cadmium from water by

Limnocharis flava (L.) Buchenau grown in free-

floating culture system. Journal of hazardous

materials, 170(2): 791-797.

Abioye O, Agamuthu P, Aziz A (2012).

Phytotreatment of soil contaminated with used

lubricating oil using Hibiscus cannabinus.

Biodegradation, 23(2): 277-286.

Agunbiade FO, Olu-Owolabi BI, Adebowale KO

(2009). Phytoremediation potential of Eichornia

crassipes in metal-contaminated coastal water.

Bioresource Technology, 100(19): 4521-4526.

Ahmad, A., R. Ghufran and A. Zularisam (2011).

"Phytosequestration of metals in selected plants

growing on a contaminated Okhla industrial

areas, Okhla, New Delhi, India." Water, Air, &

Soil Pollution, 217(1-4): 255-266.

Ahmad F, Kan AU, Yasar A (2013) Comparative

Phycoremediation of Sewage Water by Various

Species of Algae. Proceedings of the Pakistan

Academy of Sciences, 50(2): 131–139.

Alkorta I, Hernández-Allica J, Becerril J, Amezaga I,

Albizu I, Garbisu C (2004). Recent findings on

the phytoremediation of soils contaminated

with environmentally toxic heavy metals and

metalloids such as zinc, cadmium, lead, and

arsenic. Reviews in Environmental Science and

Biotechnology, 3(1): 71-90.

Alkorta I, Hernández-Allica J, Becerril J, Amezaga I,

Albizu I, Onaindia M, Garbisu C (2004).

Chelate-enhanced phytoremediation of soils

polluted with heavy metals. Reviews in

Environmental Science and Biotechnology,

3(1): 55-70.

Alkorta I, Hernández-Allica J, Garbisu C (2004).

Plants against the global epidemic of arsenic

poisoning. Environment international, 30(7):

949-951.

Alley BL, Willis B, Rodgers J, Castle JW (2013).

Water depths and treatment performance of

pilot-scale free water surface constructed

wetland treatment systems for simulated fresh

oilfield produced water. Ecological


Anamika S, Eapen S, Fulekar MH (2009).

Phytoremediation technology for remediation of

radiostrontium (90Sr) and radiocaesium

(137Cs) by Catharanthus roseus (L.) G. Don in

aquatic environment. Environmental

Engineering and Management Journal, 8(3):

527-532.

Anderson CW, Brooks RR, Stewart RB, Simcock R

(1998). Harvesting a crop of gold in plants.

Nature, 395: 553-554.

Ashraf M, Maah M, Yusoff I (2011). Heavy metals

accumulation in plants growing in ex tin mining

catchment. International Journal of


401-416.

Ashraf MA, Rehman MA, Maah MJ, Yusoff I (2013).

Cesium-137: Radio-chemistry, Fate and

Transport, Remediation and Future Concerns.

Critical Reviews in Environmental Science and

Technology (accepted).

Assunção AG, Schat H, Aarts MG (2003). Thlaspi

caerulescens, an attractive model species to

study heavy metal hyperaccumulation in plants.

New Phytologist, 159(2): 351-360.

Augustynowicz J, Kyzioł-Komosińska J, Smo eń S,

Waloszek A (2013). Study on Chromium-

Binding Capacity of Callitriche cophocarpa in

an Aquatic Environment. Archives of

environmental contamination and toxicology, 1-

9.

Baeder-Bederski-Anteda O (2003). Phytovolatilisation

of organic chemicals. Journal of Soils and

Sediments, 3(2): 65-71.

Baldwin PR, Butcher DJ (2007). Phytoremediation of

arsenic by two hyperaccumulators in a

hydroponic environment. Microchemical

journal, 85(2): 297-300.

Banks M, Kulakow P, Schwab A, Chen Z, Rathbone

K (2003). Degradation of crude oil in the

rhizosphere of Sorghum bicolor. International

Journal of Phytoremediation, 5(3): 225-234.

Bañuelos G, Zambrzuski S, Mackey B (2000).

Phytoextraction of selenium from soils irrigated

with selenium-laden effluent. Plant and Soil,

224(2): 251-258.

Bhatia M, Goyal D (2013). Analyzing remediation

potential of wastewater through wetland plants:

A review. Environmental Progress &

Sustainable Energy.

Bianconi D, Pietrini F, Massacci A, Iannelli M (2013).

Uptake of Cadmium by Lemna minor, a

(hyper?-) accumulator plant involved in

phytoremediation applications. E3S Web of

Conferences, EDP Sciences.

Bissen M, Frimmel FH (2003). Arsenic—a review.

Part I: occurrence, toxicity, speciation,



202

mobility. Acta hydrochimica et hydrobiologica,

31(1): 9-18.

Borne KE, Fassman-Beck EA, Tanner CC (2014).

Floating Treatment Wetland influences on the

fate of metals in road runoff retention ponds.

Water research, 48: 430-442.

Bradl H (2005). Heavy Metals in the Environment:

Origin, Interaction and Remediation: Origin,

Interaction and Remediation, Access Online via

Elsevier.

Brooks RR (1998). Plants that hyperaccumulate heavy

metals: their role in phytoremediation,

microbiology, archaeology, mineral exploration

and phytomining, CAB international.

Burken JG Schnoor JL (1999). Distribution and

volatilization of organic compounds following

uptake by hybrid poplar trees. International

Journal of Phytoremediation, 1(2): 139-151.

Chaney RL, Angle JS, Baker AJ, Li YM (1998).

Method for phytomining of nickel, cobalt and

other metals from soil, Google Patents.

Cheng S, Grosse W, Karrenbrock F, Thoennessen M

(2002). Efficiency of constructed wetlands in

decontamination of water polluted by heavy

metals. Ecological Engineering, 18(3): 317-325.

Christwardana M, Soetrisnanto D (2013).

Phytoremediations of Palm Oil Mill Effluent

(POME) by Using Aquatic Plants and

Microalge for Biomass Production. Journal of

Environmental Science & Technology, 6(2).

Clabeaux BL, Navarro DA, Aga DS, Bisson MA

(2011). Cd tolerance and accumulation in the

aquatic macrophyte, Chara australis: potential

use for charophytes in phytoremediation.

Environmental science & technology, 45(12):

5332-5338.

Dadrasnia A, Agamuthu P (2013). Bioremediation of

Diesel Fuel Contaminated Soil by Podocarpus

polystachyus Enhancedwith Organic Wastes.

Malaysian Journal of Science, 32(1): 9-14.

Das S, Goswami S, Talukdar AD (2013). A Study on

Cadmium Phytoremediation Potential of Water

Lettuce, Pistia stratiotes L. Bulletin of

environmental contamination and toxicology, 1-

6.

Delmail D, Labrousse P, Hourdin P, Larcher L,

Moesch C, Botineau M (2013).

Micropropagation of Myriophyllum

alterniflorum (Haloragaceae) for stream

rehabilitation: first in vitro culture and

reintroduction assays of a heavy-metal

hyperaccumulator immersed macrophyte.

International Journal of Phytoremediation,

15(7): 647-662.

Dhir B (2013). Phytoremediation: Role of Aquatic

Plants in Environmental Clean-Up. Springer.

Dhir B, Sharmila P, Saradhi PP (2009). Potential of

aquatic macrophytes for removing contaminants

from the environment. Critical Reviews in


754-781.

Doty SL, Shang TQ, Wilson AM, Tangen J,

Westergreen AD, Newman LA, Strand SE,

Gordon MP (2000). Enhanced metabolism of

halogenated hydrocarbons in transgenic plants

containing mammalian cytochrome P450 2E1.

Proceedings of the National Academy of

Sciences, 97(12): 6287-6291.

Dushenkov S (2003). Trends in phytoremediation of

radionuclides. Plant and Soil, 249(1): 167-175.

Dzantor EK (2007). Phytoremediation: the state of

rhizosphere ‘e gi eeri g’for acce erated

rhizodegradation of xenobiotic contaminants.

Journal of Chemical Technology and

Biotechnology, 82(3): 228-232.

Dzantor EK, Woolston J, Momen B (2002). PCB

dissipation and microbial community analysis

in rhizosphere soil under substrate amendment

conditions. International Journal of

Phytoremediation, 4(4): 283-295.

El-Khatib A, Hegazy A, Abo-El-Kassem AM (2014).

Bioaccumulation Potential and Physiological

Responses of Aquatic Macrophytes to Pb

Pollution. International Journal of


Elfeky SA, Imam H, Alsherbini AA (2013). Bio-

absorption of Ni and Cd on Eichhornia

crassipes root thin film. Environmental Science

and Pollution Research,1-7.

Estrella-Gómez NE, Sauri-Duch E, Zapata-Pérez O,

Santamaría JM (2012). Glutathione plays a role

in protecting leaves of Salvinia minima from

Pb2+ damage associated with changes in the

expression of SmGS genes and increased

activity of GS. Environmental and

Experimental Botany, 75: 188-194.

Falinski K, Yost R, Sampaga E, Peard J (2014).

Arsenic accumulation by edible aquatic

macrophytes. Ecotoxicology and environmental

safety, 99: 74-81.

Fingerman M, Nagabhushanam R (2005).

Bioremediation of aquatic and terrestrial

ecosystems. Science Publishers.

Fritioff Å, Greger M (2003). Aquatic and terrestrial

plant species with potential to remove heavy

metals from stormwater. International Journal

of Phytoremediation, 5(3): 211-224.

Gallon C, Munger C, Prémont S, Campbell PG

(2004). Hydroponic study of aluminum

accumulation by aquatic plants: effects of

fluoride and pH. Water, Air, and Soil Pollution,

153(1-4): 135-155.

Faraji


203

Gardea-Torresdey J, De La Rosa G, Peralta-Videa J

(2004). Use of phytofiltration technologies in

the removal of heavy metals: A review. Pure

and Applied Chemistry, 76(4): 801-813.

Ghosh M, Singh S (2005). A review on

phytoremediation of heavy metals and

utilization of it's by products. Asian J Energy

Environ, 6(4): 18.

Gleba D, Borisjuk NV, Borisjuk LG, Kneer R, Poulev

A, Skarzhinskaya M, Dushenkov S, Logendra

S, Gleba YY, Raskin I (1999). Use of plant

roots for phytoremediation and molecular

farming. Proceedings of the National Academy

of Sciences, 96(11): 5973-5977.

Ha H, Thi N, Sakakibara M, Sano S, Hori RS, Sera K

(2009). The potential of Eleocharis acicularis

for phytoremediation: case study at an

abandoned mine site. CLEAN–Soil, Air, Water,

37(3): 203-208.

Harris HH, Pickering IJ, George GN (2003). The

chemical form of mercury in fish. Science,

301(5637): 1203-1203.

Hem JD (1970). Chemical behavior of mercury in

aqueous media. Mercury in the environment.

United States Geological Survey Professional

Papers, 713: 19-24.

Houben D, Pircar J, Sonnet P (2011). Heavy metal

immobilization by cost-effective amendments

in a contaminated soil: Effects on metal

leaching and phytoavailability. Journal of

Geochemical Exploration.

Hutchinson S, Schwab A, Banks M (2003).

Biodegradation of petroleum hydrocarbons in

the rhizosphere. Phytoremediation:

transformation and control of contaminants.

Isaksson R, Balogh SJ, Farris MA (2007).

Accumulation of mercury by the aquatic plant

Lemna minor. International journal of

environmental studies, 64(2): 189-194.

Islam MS, Ueno Y, Sikder MT, Kurasaki M (2013).

Phytofiltration of arsenic and cadmium from the

water environment using Micranthemum

umbrosum (JF Gmel) SF Blake as a

hyperaccumulator. International Journal of


Kane ME, Gilman EF, Jenks MA, Sheehan TJ (1990).

Micropropagation of the aquatic plant

Cryptocoryne lucens. HortScience, 25(6): 687-

689.

Karimzadeh L, Heilmeier H, Merkel BJ (2012). Effect

of microbial siderophore DFO-B on Cd

accumulation by Thlaspi caerulescens

hyperaccumulator in the presence of zeolite.

Chemosphere, 88(6): 683-687.

Khan A, Kuek C, Chaudhry T, Khoo C, Hayes W

(2000). Role of plants, mycorrhizae and

phytochelators in heavy metal contaminated

land remediation. Chemosphere, 41(1): 197-

207.

Khellaf N, Zerdaoui M (2009). Phytoaccumulation of

zinc by the aquatic plant, Lemna gibba L.

Bioresource technology, 100(23): 6137-6140.

Kirkham M (2006). Cadmium in plants on polluted

soils: Effects of soil factors,

hyperaccumulation, and amendments.

Geoderma, 137(1): 19-32.

Kisku G, Barman S, Bhargava S (2000).

Contamination of soil and plants with

potentially toxic elements irrigated with mixed

industrial effluent and its impact on the

environment. Water, air, and soil pollution,

120(1-2): 121-137.

Knepper TP (2003). Synthetic chelating agents and

compounds exhibiting complexing properties in

the aquatic environment. TrAC Trends in

Analytical Chemistry, 22(10): 708-724.

Krämer U (2010). Metal hyperaccumulation in plants.

Annual review of plant biology, 61: 517-534.

Kumar JN, Soni H, Kumar RN, Bhatt I (2008).

Macrophytes in phytoremediation of heavy

metal contaminated water and sediments in

Pariyej Community Reserve, Gujarat, India.

Turkish Journal of Fisheries and Aquatic

Sciences, 8(2): 193-200.

Kumar N, Bauddh K, Kumar S, Dwivedi N, Singh D,

Barman S (2013). Accumulation of metals in

weed species grown on the soil contaminated

with industrial waste and their

phytoremediation potential. Ecological


Kumar SR, Arumugam T, Anandakumar C,

Balakrishnan S, Rajavel D (2013). Use of Plant

Species in Controlling Environmental

Pollution-A. Bull. Env. Pharmacol. Life Sci., 2:

52-63.

Landmeyer JE (2012). Introduction to

phytoremediation of contaminated groundwater.

Springer.

Lombi E, Zhao F, Dunham S, McGrath S (2001).

Phytoremediation of heavy metal–contaminated

soils. Journal of Environmental Quality, 30(6):

1919-1926.

Ma LQ, Komar KM, Tu C, Zhang W, Cai Y,

Kennelley ED (2001). A fern that

hyperaccumulates arsenic. Nature, 409(6820):

579-579.

Mahmood T, Malik SA, Hussain ST (2010).

Biosorption and recovery of heavy metals from

aqueous solutions by Eichhornia crassipes

(water hyacinth) ash. BioResources, 5(2): 1244-

1256.



204

Manousaki E, Kadukova J, Papadantonakis N,

Kalogerakis N (2008). Phytoextraction and

phytoexcretion of Cd by the leaves of Tamarix

smyrnensis growing on contaminated non-

saline and saline soils. Environmental research,

106(3): 326-332.

Marchand L, Mench M, Jacob D, Otte M (2010).

Metal and metalloid removal in constructed

wetlands, with emphasis on the importance of

plants and standardized measurements: a

review. Environmental pollution, 158(12):

3447-3461.

Marchiol L, Sacco P, Assolari S, Zerbi G (2004).

Reclamation of polluted soil: phytoremediation

potential of crop-related Brassica species.

Water, Air, and Soil Pollution, 158(1): 345-356.

Mbengue S, Alleman LY, Flament P (2014). Size-

distributed metallic elements in submicronic

and ultrafine atmospheric particles from urban

and industrial areas in northern France.

Atmospheric Research, 135: 35-47.

McGrath SP, Zhao FJ (2003). Phytoextraction of

metals and metalloids from contaminated soils.

Current Opinion in Biotechnology, 14(3): 277-

282.

McIntyre T (2003). Databases and protocol for plant

and microorganism selection: hydrocarbons and

metals. Phytoremediation: Transformation and

Control of Contaminants, 887-904.

Mishra VK, Tripathi B (2008). Concurrent removal

and accumulation of heavy metals by the three

aquatic macrophytes. Bioresource technology,

99(15): 7091-7097.

Mojiri A, Aziz HA, Zahed MA, Aziz SQ, Selamat

MRB (2013) Phytoremediation of Heavy

Metals from Urban Waste Leachate by

Southern Cattail (Typha domingensis). Inter J

Sci Res Environ Sci, 1.

Monferrán MV, Pignata ML, Wunderlin DA (2012).

Enhanced phytoextraction of chromium by the

aquatic macrophyte Potamogeton pusillus in

presence of copper. Environmental Pollution,

161: 15-22.

Moreno FN, Anderson CW, Stewart RB, Robinson

BH (2008). Phytofiltration of mercury-

contaminated water: volatilisation and plant-

accumulation aspects. Environmental and

Experimental Botany, 62(1): 78-85.

Mueller B, Rock S, Gowswami D, Ensley D (1999).

Phytoremediation decision tree. Prepared by–

Interstate Technology and Regulatory

Cooperation Work Group, 1-36.

Mukhopadhyay S, Maiti S (2010). Phytoremediation

of metal enriched mine waste: a review. Glob J

Environ Res, 4: 135-150.

Mulbry W, Kondrad S, Pizarro C, Kebede-Westhead

E (2008). Treatment of dairy manure effluent

using freshwater algae: algal productivity and

recovery of manure nutrients using pilot-scale

algal turf scrubbers. Bioresource technology,

99(17): 8137-8142.

Mwegoha W, Mbuya OS, Jain A, Ugochukwu NH,

Abazinge MD (2007). Use of chicken manure

extract for biostimulation and enhancement of

perchlorate rhizodegradation in soil and water

media. Bioremediation journal, 11(2): 61-70.

Nan H, Jifang Z, Dexin D, Guangyue L, Jie Y, Xin C,

Jia Y (2013). Screening of Native

Hyperaccumulators at the Huayuan River

Contaminated by Heavy Metals.

Bioremediation Journal, 17(1): 21-29.

Ndimele P, Ndimele C (2013). Comparative effects of

biostimulation and phytoremediation on crude

oil degradation and absorption by water

hyacinth (Eichhornia crassipes [Mart.] Solms).

International Journal of Environmental Studies,

70(2): 241-258.

Nzengung VA, Jeffers P (2001). Sequestration,

phytoreduction, and phytooxidation of

halogenated organic chemicals by aquatic and

terrestrial plants. International Journal of


gu J Phycoremediation: key issues for

cost-effective nutrient removal processes.

Biotechnology Advances, 22(1): 81-91.

Olguín EJ Sánchez-Galván G (2010). Aquatic

phytoremediation: Novel insights in tropical

and subtropical regions. Pure and Applied

Chemistry, 82(1): 27-38.

Oswald WJ, Gotaas HB (1957). Photosynthesis in

sewage treatment. Trans. Am. Soc. Civ. Eng,

122: 73-105.

Padmavathiamma PK, Li LY (2007).

Phytoremediation technology: hyper-

accumulation metals in plants. Water, Air, and

Soil Pollution, 184(1-4): 105-126.

Pant D, Singh P (2013). Pollution due to hazardous

glass waste. Environmental Science and

Pollution Research, 1-23.

Pilon-Smits E (2005). Phytoremediation. Annu. Rev.

Plant Biol., 56: 15-39.

Pivetz BE (2001). Ground Water Issue:

Phytoremediation of Contaminated Soil and

Ground Water at Hazardous Waste Sites, DTIC

Document.

Prasad M (2004). Phytoremediation of metals and

radionuclides in the environment: the case for

natural hyperaccumulators, metal transporters,

soil-amending chelators and transgenic plants.

Heavy metal stress inlants: from biomlecules to

ecosystems, 345-392.

Faraji


205

Rajkumar K, Sivakumar S, Senthilkumar P, Prabha D,

Subbhuraam C, Song Y (2009). Effects of

selected heavy metals (Pb, Cu, Ni, and Cd) in

the aquatic medium on the restoration potential

and accumulation in the stem cuttings of the

terrestrial plant, Talinum triangulare Linn.

Ecotoxicology, 18(7): 952-960.

Rawat I, Ranjith Kumar R, Mutanda T, Bux F (2011).

Dual role of microalgae: phycoremediation of

domestic wastewater and biomass production

for sustainable biofuels production. Applied

Energy, 88(10): 3411-3424.

Reichenauer TG, Germida JJ (2008).

Phytoremediation of organic contaminants in

soil and groundwater. ChemSusChem, 1(8‐9):

708-717.

Ridzuan MBB, Yeoh DEC, Hung YT (2010). Heavy

Metal Removal by Crops from Land

Application of Sludge. Environmental

Bioengineering, Springer: 211-232.

Romeh A (2009). Phytoremediation of Water and Soil

Contaminated with Imidacloprid Pesticide by

Plantago Major, L. International Journal of


Römkens P, Bouwman L, Japenga J, Draaisma C

(2002). Potentials and drawbacks of chelate-

enhanced phytoremediation of soils.

Environmental pollution, 116(1): 109-121.

Roongtanakiat N (2009). Vetiver phytoremediation

for heavy metal decontamination. PRVN

Technical Bulletins(2009/1).

Safonova E, Kvitko K, Iankevitch M, Surgko L, Afti

I, Reisser W (2004). Biotreatment of Industrial

Wastewater by Selected Algal‐Bacterial

Consortia. Engineering in life sciences, 4(4):

347-353.

Sahu R, Naraian R, Chandra V (2007). Accumulation

of metals in naturally grown weeds (aquatic

macrophytes) grown on an industrial effluent

channel. CLEAN–Soil, Air, Water, 35(3): 261-

265.

Sánchez-Galván G, Monroy O, Gómez J, Olguín EJ

(2008). Assessment of the hyperaccumulating

lead capacity of Salvinia minima using

bioadsorption and intracellular accumulation

factors. Water, Air, and Soil Pollution, 194(1-

4): 77-90.

Sarkar B (2002). Heavy metals in the environment,

CRC Press.

Schnoor JL, Licht LA, McCutcheon SC, Wolfe NL,

Carreira LH (1995). "Phytoremediation of

organic and nutrient contaminants.

Environmental Science & Technology, 29(7):

318A-323A.

Shao X, Cheng H, Li Q, Lin C (2013). Anthropogenic

Atmospheric Emissions of Cadmium in China.

Atmospheric Environment.

Sheoran V, Sheoran A, Poonia P (2009).

Phytomining: A review. Minerals Engineering,

22(12): 1007-1019.

Shiyab S, Chen J, Han FX, Monts DL, Matta FB, Gu

M, Su Y, Masad MA (2009). "Mercury‐induced

oxidative stress in Indian mustard (Brassica

juncea L.). Environmental toxicology, 24(5):

462-471.

Sivasubramanian V, Subramanian V, Muthukumaran

M (2010). Bioremediatio n of Chrome-Sludge

from an Electroplating Industry Using the

Micro Alg a Desmococcus olivaceus–A pilot

study. J. Algal Biomass Utln, 3: 104-128.

Stout L, Nüsslein K (2010). Biotechnological

potential of aquatic plant–microbe interactions.

Current opinion in biotechnology, 21(3): 339-

345.

Sun G, Shi W (1998). Sunflower stalks as adsorbents

for the removal of metal ions from wastewater.

Industrial & engineering chemistry research,

37(4): 1324-1328.

Sung M, Lee CY, Lee SZ (2011). Combined mild soil

washing and compost-assisted

phytoremediation in treatment of silt loams

contaminated with copper, nickel, and

chromium. Journal of hazardous materials,

190(1): 744-754.

Tu C, Ma LQ, Bondada B (2002). Arsenic

accumulation in the hyperaccumulator Chinese

brake and its utilization potential for

phytoremediation. Journal of Environmental

Quality, 31(5): 1671-1675.

Tu S, Ma LQ, Fayiga AO, Zillioux EJ (2004).

Phytoremediation of arsenic-contaminated

groundwater by the arsenic hyperaccumulating

fern Pteris vittata L. International Journal of


Türker OC, Böcük H, Yakar A (2013). The

phytoremediation ability of a polyculture

constructed wetland to treat boron from mine

effluent. Journal of hazardous materials.

Ucer A, Uyanik A, Kutbay HG (2013). Removal of

Heavy Metals Using Myriophyllum

verticillatum (Whorl-Leaf Watermilfoil) in a

Hydroponic System. Ekoloji Dergisi, 22(87).

van der Ent A, Baker A, van Balgooy M, Tjoa A

(2013). Ultramafic nickel laterites in Indonesia

(Sulawesi, Halmahera): Mining, nickel

hyperaccumulators and opportunities for

phytomining. Journal of Geochemical

Exploration.

Vangronsveld J, Herzig R, Weyens N, Boulet J,

Adriaensen K, Ruttens A, Thewys T, Vassilev



206

A, Meers E, Nehnevajova E (2009).

Phytoremediation of contaminated soils and

groundwater: lessons from the field.

Environmental Science and Pollution Research,

16(7): 765-794.

Verma V, Singh Y, Rai J (2007). Biogas production

from plant biomass used for phytoremediation

of industrial wastes. Bioresource Technology,

98(8): 1664-1669.

Vishnoi SR, Srivastava P (2007). Phytoremediation–

Green for Environmental Clean. Proceedings of

Taal2007: The 12th World Lake Conference.

Wand H, Kuschk P, Soltmann U, Stottmeister U

(2002). Enhanced removal of xenobiotics by

helophytes. Acta biotechnologica, 22(1‐2): 175-

181.

Wang KS, Huang LC, Lee HS, Chen PY, Chang SH

(2008). Phytoextraction of cadmium by

Ipomoea aquatica (water spinach) in

hydroponic solution: Effects of cadmium

speciation. Chemosphere, 72(4): 666-672.

Wang X, Dossett MP, Gordon MP, Strand SE (2004).

Fate of carbon tetrachloride during

phytoremediation with poplar under controlled

field conditions. Environmental science &

technology, 38(21): 5744-5749.

Wang Z, Calderon MM (2012). Environmental and

economic analysis of application of water

hyacinth for eutrophic water treatment coupled

with biogas production. Journal of

environmental management, 110: 246-253.

Ximénez-Embún P, Madrid-Albarrán Y, Cámara C,

Cuadrado C, Burbano C, Múzquiz M (2001).

Evaluation of Lupinus Species to Accumulate

Heavy Metals From W aste Waters.

International Journal of Phytoremediation, 3(4):

369-379.

Xing W, Wu H, Hao B, Huang W, Liu G (2013).

Bioaccumulation of Heavy Metals by

Submerged Macrophytes: Looking for

Hyperaccumulators in Eutrophic Lakes.

Environmental science & technology, 47(9):

4695-4703.

Xue S, Chen Y, Reeves RD, Baker AJ, Lin Q,

Fernando DR (2004). Manganese uptake and

accumulation by the hyperaccumulator plant

Phytolacca acinosa Roxb.(Phytolaccaceae).

Environmental Pollution, 131(3): 393-399.

Yamagata N (1979). Industrial emission of cadmium

in Japan. Environmental Health Perspectives,

28: 17.

Yang L, Wu Z, Wu J, Zhang Y, Li M, Lin Z,

Bañuelos G (2013). Simultaneous removal of

selenite and electricity production from Se-

laden wastewater by constructed wetland

coupled with microbial fuel cells. Selenium in

the Environment and Human Health, 212.

Yifru DD, Nzengung VA (2008). Organic carbon

biostimulates rapid rhizodegradation of

perchlorate. Environmental Toxicology and

Chemistry, 27(12): 2419-2426.

Yoon J, Cao X, Zhou Q, Ma QL (2006).

Accumulation of Pb, Cu, and Zn in native

plants growing on a contaminated Florida site.

Science of the Total Environment, 368(2): 456-

464.

Zavoda J, Cutright T, Szpak J, Fallon E (2001).

Uptake, selectivity, and inhibition of

hydroponic treatment of contaminants. Journal

of environmental engineering, 127(6): 502-508.

Zhang XH, Liu J, Huang HT, Chen J, Zhu YN, Wang

DQ (2007). Chromium accumulation by the

hyperaccumulator plant Leersia hexandra

Swartz. Chemosphere, 67(6): 1138-1143.

Zhang X, Zhao FJ, Huang Q, Williams PN, Sun GX,

Zhu YG (2009). Arsenic uptake and speciation

in the rootless duckweed Wolffia globosa. New

Phytologist, 182(2): 421-428.



©2014 IJSRPUB

208

Municipal Landfill Leachate Treatment Techniques: An Overview

Shuokr Qarani Aziz1, 2, *

, Hamidi Abdul Aziz2, Mohammed J.K. Bashir

3, Amin Mojiri

2

1Department of Civil Engineering, College of Engineering, University of Salahaddin–Erbil, Iraq

2School of Civil Engineering, Engineering Campus, Universiti Sains Malaysia, 14300 Nibong Tebal, Penang, Malaysia

3Faculty of Eng. and Green Tech. (FEGT), Universiti Tunku Abdul Rahman, 31900 Kampar, Perak, Malaysia

*Corresponding Author, [email protected], [email protected], H/P: 00964 750 462 5426

Abstract. Production of raw leachate from landfills regards as shortcoming for the sanitary landfills. Unprocessed landfill

leachate requires treatment prior disposal to the natural environment. In this work, leachate channeling, biological, and

physical-chemical treatment processes for treatment of formed landfill leachate were presented in details. Definite treatment

processes were resulted in high removal of pollutants such as ammonia (NH3-N), chemical oxygen demand (COD),

biochemical oxygen demand (BOD5), phenols, color etc. Practically, age of produced leachate have influence on the treatment

systems. Efficient treatment methods were explained for fresh, medium, and stabilized landfill leachates.

Keywords: Landfill, leachate, treatment, municipal solid waste, pollutants, removal efficiency

1. INTRODUCTION

Sanitary landfill is the most common municipal solid

waste (MSW) disposal method due to such advantages

as simple disposal procedure, low cost, and landscape-

restoring effect on holes from mineral workings.

However, the production of highly contaminated

landfill leachate is a chief weakness of this system

(Aziz et al., 2010). The generation of highly

contaminated leachate that can seep into the ground

and contaminate the ground water, surface water, and

soil is principally a main negative aspect associated to

municipal sanitary landfill disposal method (Bashir et

al., 2012. Additionally, leachate is liquid contain large

amounts of organic compounds measured as chemical

oxygen demand (COD), biochemical oxygen demand

(BOD5), ammoniacal nitrogen (NH3-N), halogenated

hydrocarbons suspended solid, significant

concentration of heavy metals, and inorganic salts

(Aziz, 2013; Bashir et al., 2012; Aziz et al., 2010;

Uygur and Kargi, 2004).

If not treated and disposed safely, landfill leachate

could be a major source of water contamination

because it could percolate through soil and subsoil,

causing high pollution to receiving waters. Thus, the

treatment of hazardous leachate constituents before

discharge has been made a legal requirement to

prevent pollution of water resources and to avoid both

acute and chronic toxicities (Aziz et al., 2011).

To reduce the negative impacts of discharged

leachate on the environment, several techniques of

water and wastewater treatment have been used,

including aerobic and anaerobic biological treatment,

chemical and electrochemical oxidation processes,

chemical precipitation, adsorption using various

adsorbent, reverse osmosis, coagulation–flocculation,

membrane processes, and ion exchange (Aziz et al.

2013; Bashir et al., 2012; Aziz et al., 2011; Abbas et

al 2009; Renou et al., 2008).

The current work was aimed to present appropriate

landfill leachate treatment techniques for unprocessed

landfill leachates. Giving detailed information on

leachate channeling, biological, and physical-chemical

treatment processes for raw leachates was another

goal for the present study. Furthermore, suggestion of

efficient treatment methods for different sorts of

produced landfill leachates from various sanitary

landfills was illustrated as well.

2. LANDFILL LEACHATE

2.1 Types of leachate based on landfill design

According to Yamamoto, (2002) and Matsufuji et al.

(1993), in an anaerobic landfill, solid wastes are

dumped in an excavated area of a plane field, which is

filled with water in an anaerobic condition. Typically,

anaerobic sanitary landfills are recognized by its

sandwich-shaped cover. On the other hand, semi-

aerobic landfills have a leachate collection duct. The

opening of the duct is surrounded by air, and the duct

is covered with small crushed stones. Moisture

content in solid waste is small, and oxygen is supplied

to the solid waste from the leachate collection duct.

The schematic diagram of anaerobic and semi-aerobic

(Fukuoka method) landfills is demonstrated in Figure

1 (JICA, 2005). Characteristics of fresh landfill

leachates at semi-aerobic and anaerobic landfills are

given in Table 1.

Aziz et al.


209

Fig.1: Schematic diagram of anaerobic and semi-aerobic landfill

2.2 Significance of leachate problem

Tatsi et al. (2003) and Renou et al. (2008) reported

that landfill leachate can be characterized by two main

factors: quantity (volumetric flow rate) and quality

(chemical composition). El-Fadel et al. (2002)

indicated that the quantity of leachate generation is

affected by landfill moisture distribution influencers

namely refuse age, pretreatment, permeability,

compaction, particle size and density and direct

landfill moisture contributors like rainfall, snowmelt,

groundwater intrusion, initial moisture and leachate

recirculation. Seriously, leachate production is rapid

in tropical countries such as Malaysia since the

rainfall normally exceeds the quantity that can be

evaporated during the rainy season (Lema et al.,

1988).

Usually, leachate contains a complex variety of

substances and organic compounds such as humic

substances, fatty acids, heavy metals and many other

hazardous chemicals. Regardless of the concentration

changes and show a discrepancy based on a complex

set of interconnected factors, the complexity of the

landfill leachate can be categorized on the basis of

four major groups of pollutants i.e. dissolved organic

substances, inorganic macro-components, heavy

metals and xenobiotic organic compounds

(Widziewicz et al., 2012; Schrab et al., 1993).

Bashir et al (2010; 2012) and Aziz et al. (2011)

reported that the common features of raw leachate

generated from Malaysian landfill sites are its high

strength of recalcitrant compounds (as reflected by its

chemical oxygen demand (COD) value) and high

concentrations of ammonia –nitrogen (NH3-N).

Ammonia resulting from the decomposition process of

organic nitrogen, has been recognized not only as a

major long-term noxious waste, but also as the

primary cause of acute toxicity (Ernst et al., 1994;

Baun et al., 1999). Bashir et al. (2010) stated that the

existence of high amount of NH3–N in leachate over a

long period of time is one of the most important

problems routinely faced by landfill operators. This

high quantity of unprocessed NH3–N leads to the

depletion of dissolved oxygen which is also

recognized as eutrophication. Because NH3-N is stable

under anaerobic situations, it typically accumulates in

the leachate (Ernst et al., 1994). With a concentration

of higher than 100 mg/L, untreated NH3-N is highly

toxic to aquatic organisms (Widziewicz et al., 2012;

Burton and Watson-Craik, 1998; Silva et al., 2004;

Bagchi, 1994). Unless appropriately treated, leachate

that seeps from a landfill can get into and contaminate

the underlying groundwater.

In line with the abovementioned, if the leachate

escapes to the water bodies, it is very complicated and

costly to have it controlled and cleaned up,

consequently posing potentially serious hazards to

living organisms, as well as public health in the long

term. In most cases, it is very hard to restore the

contaminated water bodies to its original state.

Recently, the hazard of groundwater pollution due to

leachate seepage has turn out to be a main

environmental concern worldwide. Therefore, an

adequate engineering plan and design of a municipal

landfill can avoid or reduce the seepage of leachate

from reaching the water bodies.

Typically, the concentration of leachate parameters

changes with the age of the leachate. The phases of

leachate are transition (0 -5 years), acid-formation (5 -

10 years), methane fermentation (15 - 20 years), and

final maturation (greater than 20 years). The age of

the landfill is one of the most important factors that

affect leachate characteristics (Kostova, 2006; Aziz,

2013) The levels of some leachate characteristics such



210

as biochemical oxygen demand (BOD5), COD, total

organic carbon (TOC), NH3-N, nitrite ( NO2-N) and

total dissolved salts (TDS) in different phases are

presented in Table 1.

According to the literature, as a landfill becomes

older, the biological decomposition of the deposited

wastes shifts from a relatively shorter initial period to

a longer decomposition period, which has two distinct

sub-phases: acidic and methanogenic. Leachates from

these distinct stages contain different constituents;

therefore, young leachates tend to be acidic due to the

presence of volatile fatty acids.

Fig. 2: Diagram of landfill leachate treatment techniques, based on Abbas et al. (2009); Renou et al. (2008)


If raw leachate is disposed without treatment, it could

become a major source of water pollution because it

can percolate through soils and sub-soils, causing high

contamination of the receiving water. The treatment of

potentially hazardous constituents of leachate prior to

discharge is a legal requirement to avoid

contamination of water resources to prevent both

acute and chronic toxicity (Ziyang et al., 2009; Oman

and Junestedt, 2008; Sanphoti et al. 2006; ARRPET,

2004).

Aziz et al.


211

To reduce the negative impact of discharged

leachate on environment, several techniques of water

and wastewater treatment have been used. The

technologies which were developed for the treatment

of landfill leachate could be classified as physical,

chemical, and biological (Abbast et al., 2009; Renou

et al., 2008). Normally, the techniques are applied as

an integrated system because it is not easy to achieve

the satisfying treatment efficiency by using only one

technology. Traditional treatment techniques

generally demand multistage process treatment. To set

up acceptable treatment process for removal of

contaminates from leachates, various physicochemical

and biological techniques and/or their different

combinations could be applied.

The implementation of the most suitable technique

for the treatment of leachate is directly governed by

the characteristics of the leachate. An overview of

leachate treatment methods is shown in Figure 2,

(Abbas et al., 2009; Renou et al., 2008)

Comparison of the above techniques for different

landfill ages with changeable success is illustrated in

Table 2. In addition, the most important advantages

and disadvantages of the different leachate treatment

methods are reviewed in the following sections. In

general, biological treatment processes are effective

for young or freshly (<5 years) produced leachate, but

are ineffective for leachate from older landfills (>10

years old). In contrast, physical–chemical methods

which are not favoured for young leachate treatment

are advised for older leachate treatment (Ghafari et al.,

2009).

3.1. Leachate Channeling

3.1.1. Combined treatment with domestic sewage

General means of landfill leachate disposal is piping

into the sewerage system for discharge into the sea or,

if possible, for combined treatment with domestic

sewage at traditional wastewater treatment plant. It

was favoured for its low operating costs and simple

maintenance (Ahn et al., 2002). This alternative has

been increasingly inquired due to the presence of

organic inhibitory compounds in leachate with low

BOD5/COD and heavy metals that might decrease

treatment efficiency and increase concentrations in the

effluent (Cecen and Aktas, 2004). A disagreement to

this treatment option is that phosphorus (brought by

sewage) and nitrogen (brought by leachate) are not

required to be added to the treatment scheme (Abbas

et al., 2009).

Cecen and Aktas (2001) investigated the combined

biological treatability of domestic wastewater and

landfill leachate in both continuous flow and semi-

continuously fed batch activated sludges, with

recycling mechanism. In addition, the researchers

added powdered activated carbon (PAC) in order to

examine the improvement in nitrification process and

organic carbon removal. The obtained results showed

that in both types of operations, NH3-N and COD

removal efficiencies decreased with an increase in the

leachate to total wastewater ratio. When the leachate

ratio increased, the positive effects of PAC on the

removal of COD and nitrification process became

more obvious.

3.1.2. Recycling

A common system used in many landfills consists of

recycling landfill leachate back through the tip

because it was one of the cheapest alternatives (Lema

et al., 1988). Bae et al. (1998) explained that the

leachate recirculation increased the moisture content

in a controlled reactor method and offered the

distribution of enzymes and nutrients between

solids/liquids and methanogens. Chugh et al., (1998)

stated that lowering COD and methane production

was observed to be important as the recycled leachate

quantity was 30% of the initial waste bed quantity.

The recirculation of leachate not only improves the

leachate characteristics, but also shortens the required

time for stabilization of leachate from several decades

to 2-3 years (Reinhart and Al-Yousisfi, 1996). High

recirculation rates of leachate could negatively

influence anaerobic degradation of solid wastes.

Recirculation of leachate could cause the inhibition of

methanogenesis as it may lead to high concentrations

of organic acids, pH less than 5, which are toxic for

the methanogens. In addition, if the amount of

recycled leachate is very high, problems such as

saturation, acidic conditions, and ponding may happen

(Abbas et al., 2009).



212

Table 1: Characteristics of raw leachate at semi-aerobic and anaerobic landfills (Aziz et al., 2010)

No. Parameter

Semi- aerobic Pulau Burung site Kulim site Standard B

Unaerated Intermittently aerated Anaerobic Discharge limit b

Range Averagea Range Averagea Range Averagea

1 Phenols (mg/L) 0.35-2.07 1.2 2.85-10.5 6.7 1-5.25 2.6 …

2 Total nitrogen (mg/L N-TN) 200-700 483 700-1800 1200 100-600 300 …

3 Ammonia-N (mg/L NH3-N) 360-730 542 1145-2150 1568 130-1039 538 …

4 Nitrate-N (mg/L NO3--N) 900-3200 2200 2900-7900 5233 400-2600 1283 …

5 Nitrite-N (mg/L NO2--N) 44-270 91 20-120 49 30-60 52 …

6 Total phosphorus (mg/L PO43--TNT) 10-43.0 21 10.0-25 17 8.0-40 19 …

7 Ortho-Phosphorus (mg/L PO4

3- mv ) 84-274 141 94-210 159 57-197 94 …

8 BOD5 (mg/L) 67-93 83 146-336 243 135-476 326 50

9 COD (mg/L) 600-1300 935 1680-4020 2345 630-2860 1892 100

10 BOD5/COD 0.051-0.12 0.096 0.036-0.186 0.124 0.088-0.35 0.205 0.5

11 pH 8.05-8.35 8.20 8.14-8.37 8.28 6.93-8.26 7.76 5.5-9

12 Electrical conductivity (ms/cm) 10.14-13.630 12.17 21.500-22.500 22.10 5.250-13.92 8.55 …

13 Turbidity (FAU) 600-3404 1546 149-211 180 490-4500 1936 …

14 Color (Pt Co) 1944-4050 3334 2310-4390 3347 1950-7475 4041 …

15 Total solids (mg/L) 5138-7404 6271 8860-11084 9925 4520-10568 6336 …

16 Suspended solids (mg/L) 906-2220 1437 374-1372 837 232-1374 707 100

17 Total iron (mg/L Fe) 2-29.5 7.9 0.9-8.8 3.4 0.6-11.4 5.3 5

18 Zinc (mg/L Zn) 0-3 0.6 0.01-2 0.5 0-1 0.2 1

19 Total coliform … … … <50 (0.77-0.85)x104 0.81x104 …

20 E-Coli … … … 0.00 (0.18-0.22)x104 0.20x104 …

a Average value of six samples b Standard B of the Environmental Quality (Sewage and Industrial Effluents) Regulations 1979, under the Environmental Quality Act of Malaysia, 1974 (MDC, 1997).

Aziz et al.


213

Table 2: Effectiveness of leachate treatment techniques versus leachate age (Abbas et al., 2009)

No. Type of treatment Leacgate age (year)

Young (<5) Medium (5-10) Old (>10)

1 Combined treatment with

domestic sewage

Good Fair poor

2 Recycling Good Fair poor

3 Aerobic processes Good Fair poor

4 Anaerobic processes Good Fair poor

5 Coagulation/flocculation Poor Fair Fair

6 Chemical precipitation Poor Fair Poor

7 Adsorption Poor Fair Good

8 Oxidation Poor Fair Fair

9 Stripping Poor Fair Fair

10 Ion exchange Good Good Good

11 Microfiltration Poor - -

12 Ultrafiltration Poor - -

13 Nanofiltration Good Good Good

14 Reverse osmosis Good Good Good

3.2. Biological treatment

The biological purification processes have been well

recognized and effectively employed for the treatment

of domestic wastewater. Biological treatments are

categorized as aerobic or anaerobic depending on

whether or not the biological processing medium

needs oxygen (O2) supply. In aerobic treatment,

organic contaminants are mostly transformed into

carbon dioxide (CO2) and sludge by using the

atmospheric O2 transferred to the wastewater. While,

in anaerobic processing, organic matter is converted

into biogas, a mixture mainly comprising of CO2 and

CH4, and biological sludge (Lema et al., 1988).

Because of its simplicity, reliability, and high-cost

effectiveness, biological purification (suspended or

attached growth) is mainly employed to treat landfill

leachate containing extreme concentrations of BOD5

(Renou et al., 2008). Biological techniques have been

recognized to be very effective for young leachate

treatment since the BOD5/COD ratio is high (> 0.5).

However, the biodegradability ratio commonly

decreases as the landfill ages, due to the presence of

pollutants that decrease biomass activity and/or are

refractory to biological processes (Lema et al., 1988).

The main fraction of mature or biologically treated

leachate is large unmanageable organic compounds

that are not easily removed through biological

treatment. Thus, to meet the allowable standards for

direct discharge of leachate into the environment, a

development of integrated treatment methods is

required, i.e. a combination of biological, physical,

and chemical, and other process steps (Tauchert et al.,

2006).

3.2.1 Aerobic treatment

Destruction of biodegradable organic matter occurs

via aeration in aerobic process. Aerobic biological

methods consist of suspended-growth and attached-

growth biomass processes. The former method has

been extensively applied for treatment of municipal

landfill leachate and wastewater (Abbas et al., 2009;

Renou et al., 2008).

i) Suspended-growth biomass process

Aerated lagoons

Aerated lagoons have commonly been viewed as a

successful and economical technique for the removal



214

of pathogens, organic and inorganic matters. Low

operation and maintenance costs have made them an

accepted option for wastewater treatment, particularly

in developing countries. Further, it requires a little

professional skill to operate the system (Maynard et

al., 1999). Maehlum (1995) studied biological

treatment of leachate using anaerobic–aerobic lagoons

and constructed wetlands and more than 70% of

nitrogen, phosphorus and ferrous removals were

obtained for diluted landfill leachate. Generally,

treatment of landfill leachate in lagoons and wetlands

need 10 to 20 days retention time (Robinson et al,

1992, Maehlum, 1995). In spite of its lower costs, this

process might not be entirely satisfactory treatment

option for leachate treatment (Zaloum and Abbott,

1997). Finally, large area requirements, aerosol and

formation are other shortcomings of this method

(Robinson et al, 1992, Maehlum, 1995).

Activated sludge process

Activated sludge process could be defined as a

suspended growth process that utilizes aerobic

microorganisms to biodegrade organic matters in

wastewater or leachate. Activated sludge technology

has been extensively used for the treatment of

wastewater and leachate. However this process has a

good ability in the removal of nutrients but it has

some weaknesses such as: 1) Excess sludge

production, 2) Sludge settleability (i.e. high SVI

values), 3) High energy requirements, 4) The need for

longer contact times, and 5) Microbial inhibition due

to high ammonium-nitrogen (NH4-N) strength

(Loukidou and Zouboulis, 2001; Lin et al., 2000;

Lema et al., 1988).

Hoilijoki et al. (2000) added plastic carrier

material to activated sludge process and examined

nitrification of anaerobically pre-treated landfill

leachate in a lab-scale at different temperatures (5-

10°C). They stated that the concentration of effluent

parameters (COD, BOD5, and NH3-N) for aerobic

post-treatment were 150-500 mg/L, < 7 mg/L , and <

133 mg/L, respectively. PAC supplemented to

activated sludge processes improved nitrification

efficiency in biological treatment of landfill leachate.

Aghamohammadi (2006) studied treatment of

semi-aerobic landfill leachate from PBLS by using

PAC augmented activated sludge process. The author

reported that the leachate characteristic had great

affect on the removal of organic matter using

activated sludge treatment. Removal efficiencies of

colour, COD, and NH3-N during the treatment of

landfill leachate for leachate from PBLS were 21%,

29%, and 60%, respectively.

SBR (Sequencing batch reactor)

SBR process varies from activated-sludge

techniques, because SBR merges all treatment units

and processes into a single basin; whereas traditional

systems rely on various tanks. Typical SBR is divided

into five time periods: fill, react, settle, draw, and idle.

SBR is used for the treatment of wastewater and

landfill leachates (Mahvi 2008; Al-Rekabi et al.,

2007). Due to low BOD5/COD ratio, high

concentration of COD, NH3-N, heavy metals, and

other compounds in landfill leachate, the capability of

SBR in leachate treatment is relatively weaker than

for municipal and industrial wastes (Uygur and Kargi,

2004). In literature, SBR was used for the treatment of

leachate with low BOD5/COD ratio of 0.09 to 0.37

(Guo et al., 2010; Spagni et al., 2008; Klimiuk and

Kulikowska, 2006).

SBR augmented powdered activated carbon (PAC-

SBR) showed higher removal efficiencies of NH3-N,

COD, colour, and TDS when compared with normal

SBR. Furthermore, PAC-SBR improved sludge

volume index (SVI) (Aziz et al. 2011; 2013)

ii) Attached-growth biomass systems

A number of attached-growth biomass processes by

using biofilm have been recently developed to

overcome the problems of activated sludge,

specifically sludge bulking (Dollerer and Wilderer,

1996). These methods offer the advantage of not

suffering from loss of active biomass. In addition,

nitrification is less affected by low temperature than in

suspended-growth methods, and by inhibition due to

high nitrogen content.

Moving-bed biofilm reactor (MBBR)

MBBR technique is based on the use of suspended

porous polymeric carriers, kept in continuous

movement in the aeration basin; whereas the active

biomass grows like a biofilm on their surfaces. The

main advantages of this process compared to normal

suspended growth methods seems to be:1) Higher

biomass concentrations, 2) Lower sensitivity to toxic

compounds, 3) Lower sludge-settling periods, and 4)

Both organic and high NH3-N removals in a single

process. Based on literature, this system could remove

60-81% of COD and 85-90% of NH3-N from landfill

leachate (Loukidou and Zouboulis, 2001; Horan et al.,

1997). Welander et al. (1998) reported that MBBR

resulted in about 90% and 20% removal of nitrogen

and COD while no inhibition of nitrification was

encountered during the treatment of high strength

ammonia leachate. Further, using adsorbent (i.e.

granular activated carbon) offers a suitable surface to

adsorb organic substance and enhanced

biodegradation. Thus, a steady-state equilibrium is

Aziz et al.


215

recognized between adsorption and biodegradation

(Horan et al., 1997).

Trickling filters

Trickling filters have been tested for lowering the

biological nitrogen from landfill leachate. Because of

low-cost filter media, biofilters remain an attractive

and interesting alternative for nitrification process

(Jokela et al., 2002). In literature, biological nitrogen

removal from municipal landfill leachate was

examined by Jokela et al. (2002) and nitrification of

leachate over 90% was achieved by using biofilters.

They concluded that nitrification in a low-cost

biofilter followed by denitrification in a landfill body

appeared to be applicable for the removal of nitrogen

in landfill leachate in colder climates (Jokela et al.,

2002). Maximum NH3-N removal of 75 % was

obtained by Martienssen and Schops (1997).

3.2.2. Anaerobic treatment

Anaerobic digestion is the oldest system used for

treatment of wastewater. It is suitable for the

treatment of high strength organic pollutants, such as

young leachate. Opposite to aerobic processes,

anaerobic digestion method saves energy and

produces very few solids; but it suffers from low

reaction rates. Further, the produced CH4 could be

used for warming the digester that generally works at

35°C (Renou et al., 2008; Sung et al., 1997).

Anaerobic digestion comprises suspended-growth

digester (anaerobic SBR and up-flow anaerobic sludge

blanket reactor, UASBR) and attached-growth

biomass system (anaerobic filter, hybrid filter, and

fluidized bed filter).

i) Suspended-growth digester

Bull et al. (1983) and Sung et al. (1997) studied the

performances of conventional anaerobic suspended

growth digester. The researchers reported that the

typical values for COD removal in anaerobic lab-scale

basin at 35°C and ambient temperature were 80-90%

and nearly 55%, respectively.

Anaerobic SBR

Anaerobic SBR is a type of suspended-growth

digester. Some studies have shown good

performances of anaerobic SBRs. These technologies

are able to obtain solid capture and organic lowering

in one reactor (Timur and Ozturk, 1999). Uygar and

Kargi (2004) used lab-scale SBR for the reduction of

nutrient from pre-treated leachate. The researchers

reported that, at the end of cycle time of 21 h,

sequential anaerobic/aerobic operations offered

removal efficiencies of 62% for COD, 31% for NH3-

N, and 19 % for PO43-

-P. Kettunen and Rintala (1995)

explained that COD removal in the anaerobic stage

was 35%; whereas removal efficiencies of COD and

BOD5 in the combined process were up to 75% and

99%, respectively.

Timur and Ozturk (1999) examined anaerobic

treatability of municipal landfill leachate by using lab-

scale anaerobic SBR at 35oC. Based on the obtained

results, about 83% of COD removed during the

treatment was converted to methane. In addition, the

average biomass yield was 0.12 g volatile suspended

solids per gram of COD removed.

Upflow-anaerobic sludge blanket reactor (UASBR)

UASBR technique is a modern anaerobic treatment

process that can have high treatment efficiency and a

short HRT (Lin et al., 2000). In addition, Garcia et al.

(1996) reported that UASBR exhibited higher

performances compared to other types of anaerobic

reactors, when they submitted to high volumetric

organic loading rates.

A pilot-scale UASBR at low temperature was used

for the treatment of municipal landfill leachate.

Removal efficiencies of COD and BOD5 at organic

loading rate of 2-4 kg/m3/d of COD were 65-75 % and

up to 95 %, respectively (Kettunen and Rintala, 1995).

At organic loading rates between 6 and 19.7 g/L/d of

COD, Kennedy and Lentz (2000) obtained COD

removal efficiency of 92%. Generally, for anaerobic

treatment with UASBRs, the process temperatures are

reported to be between 20–35 ◦C. (Akkaya et al.,

2010). However, several studies were conducted at

temperatures between 11 to 23°C (Kettunen and

Rintala, 1995; Garcia et al., 1996). Sensitivity to toxic

substances is the main disadvantages of UASBRs

(Sung et al., 1997).

ii) Attached-growth biomass systems

Anaerobic filter

Anaerobic filter method is a high rate system that

collects the advantages of other anaerobic methods

and decreases the shortcomings. The filtration process

may be down-flow or up-flow. Packed filter media

provides the mechanism for separating the solids and

the gas that are produced within the digestion process.

In an up-flow anaerobic filtration process, biomass is

retained as biofilms on the supporting material, such

as plastic rings (Nedwell and Reynolds, 1996). At

loading rates of 1.26-1.45 kg/m3/d of COD and for

different ages of landfill leachate, anaerobic filter

removed 90% of COD. Further, anaerobic filter

resulted in total biogas production ranged between

400 to 500 Lgas/kg COD destroyed and methane

content of 75% and 85% (Henry et al., 1987).



216

Hybrid filter

Hybrid bed filter consists of an up-flow sludge

blanket at the bed and an anaerobic filter on the top.

This technique acts as a gas-solid separator and

improves solid’s retention without causing

channelling or short-circuiting. A hybrid bed filter

(consisted of anaerobic filter and up-flow anaerobic

sludge blanket reactor) with a filter volume of 2.75 L

and HRT of 2.4 d at temperature of 35 oC resulted in

removal of about 37.5 to 76 % COD from landfill

leachate (Timur and Ozturk, 1997). Nedwell and

Reynolds(1996) showed steady state COD removal

efficiencies of 81-97% under methanogenic digestion

was based on organic loading rate; and effective

treatment occurred up to a volumetric COD loading

rate of 3.75 kg COD/m3·d

1. One disadvantage of

hybrid bed filter, as well as anaerobic filter, is the

additional cost of the supporting media.

Fluidized bed filter

Imai et al. (1993) applied microorganism-attached

activated carbon fluidized bed technique for the

treatment of real landfill leachate containing

refractory organics and a high concentration of NH4-

N. They reported that the microorganism-attached

activated carbon fluidized bed method removed about

60% and 70% of refractory organics and nitrogen,

respectively. A number of researches on carbon-

assisted fluidized beds have been conducted (Gulsen

and Turan, 2004; Suidan et al., 1993). Suidan et al.

(1993) stated that about 82% of COD was removed

from leachate via fluidized bed reactor at temperature

of 35 oC and volume of 7.9 L.

3.3. Physical/chemical treatment

Physico-chemical processes are used along with the

biological processes generally to enhance treatment

efficiency or to enhance biodegradability when the

biological oxidation method is disadvantaged by the

occurrence of bio-refractory materials. A number of

physical/chemical treatment techniques which

includes coagulation/flocculation, flotation, chemical

precipitation, adsorption, ammonium stripping,

chemical oxidation, ion exchange, electrochemical

oxidation, and membrane filtration are used for

removing non-biodegradable (humic and fulvic acid)

and/or unwanted compounds (such as heavy metals)

from the landfill leachate (Abbas et al., 2009; Renou

et al., 2008; Zouboulis et al., 2004)

3.3.1. Coagulation/flocculation

Coagulation–flocculation technique is considered as a

simple physical–chemical process in landfill leachate

treatment (Aziz et al. 2009; Ghafari et al., 2009). It is

an effective pretreatment process if used prior to

reverse osmosis or biological treatment or as a last

polishing treatment so as to remove or decrease non-

biodegradable organic matter in landfill leachate

(Amokrane et al., 1997). Due to its limited efficiency

for the removal of organic matter, it is not suitable for

a full treatment of landfill leachate. Duan and Gregory

(2003) reported that the removal mechanism of the

coagulation process mainly consists of charge

neutralization of negatively charged impurities by

cationic hydrolysis products followed by integration

of colloids in an amorphous hydroxide precipitate

during flocculation process.

In literature, aluminum sulfate, ferrous sulfate,

ferric chlorosulfate, and ferric chloride were generally

used as coagulants (Zouboulis et al., 2004; Amokrane

et al. 1997). Dialynas et al. (2008) investigated that

ferric chloride is more effective than alum in

eliminating organic constituent of landfill leachate,

particularly at pH values more than 9. It was sued to

the fact that ferric chloride enlarges floc size and

reduces settling time more than alum.

Ghafari et al. (2009) explained that by using alum

as coagulant, the optimum removal efficiencies of

COD, colour, turbidity, and suspended solids from

stabilized leachae were 62.8 %, 86.4%, 88.4%, and

90.1 %, respectively. Aziz et al. (2007) used

aluminum (III) sulphate (alum), ferric (III) chloride,

ferrous (III) sulphate and ferric (III) sulphate as

coagulants. The obtained results showed that ferric

chloride was better than the other coagulants with a

colour removal of 94 %.

However, coagulation/flocculation process have a

number of shortcomings, such as inefficiency in NH3-

N removal, low removal efficiency of high strength

landfill leachate, an increase on the concentration of

iron or aluminum could also be noticed in the liquid

phase, and production of huge volume of sludge

(Duan and Gregory, 2003; Tatsi et al., 2003; Silva et

al., 2004; Amokrane et al., 1997).

Al-Hamadani et al. (2011) studied the feasibility of

using psyllium husk as coagulant and coagulant aid

(with poly-aluminum chloride and aluminum sulfate)

for the treatment of semi-aerobic landfill leachate

from PBLS. When psyllium husk was used as primary

coagulant, the removal efficiencies for COD, colour

and suspended solids were 55 %, 80% and 95%,

respectively; whereas the removal efficiencies of

COD, colour, and suspended solids for psyllium husk

as coagulant aid with poly-aluminum chloride were

64, 90 and 96%, respectively. The researchers

reported that psyllium husk was more effective as

coagulant aid with poly-aluminum chloride in the

removal of COD, colour and suspended solids, as

compared to aluminum sulfate.

Aziz et al.


217

3.3.2. Flotation

Dissolved air flotation could be defined as

disconnection of suspended solids from liquids by

carrying the particles to the surface of the liquid (Al-

Shamrani et al., 2002). For many years, flotation

technique has been widely used and focused on the

reduction of ions, colloids, microorganisms,

macromolecules, and fibers (Zouboulis et al., 2003).

This technique was used extensively for the treatment

of oily wastewater.

Zouboulis et al. (2003) examined the use of

flotation in column, as a post-treatment step for the

removal of residual humic acids (non-biodegradable

compounds) from simulated landfill leachates. Under

optimum operational conditions, the researchers

obtained about 60% humic acids removal.

Lately, using dissolved air flotation process for the

treatment of landfill leachate in the presence and

absence of coagulation process was studied by Adlan

et al. (2011) and Palaniandy et al. (2010). Based on

the obtained results by Palaniandy et al. (2010), the

removal efficiencies of colour, COD, and turbidity

from leachate using dissolved air flotation process

(without coagulation process) were 36%, 33%, and

32%, respectively; whereas the removal efficiencies

with the coagulation process were 70%, 79%, and

42%, respectively.

3.3.3. Chemical precipitation

Chemical precipitation method has been extensively

employed as leachate pre-treatment because of its

simplicity, capability, and low-cost equipment used.

This method is efficient in removing NH3-N, non-

biodegradable compounds, and heavy metals from

landfill leachate (Akkaya et al., 2010; Zhang et al.,

2009; Calli et al., 2005; Ozturk et al., 2003).

Li et al. (1999) stated that the performance of a

conventional activated sludge technique could be

significantly affected by the high concentration of

NH4+-N. Struvite (magnesium ammonium phosphate

hexahydrate, MgNH4PO4.6H2O) precipitation was

firstly distinguished as a phenomenon to be controlled

because it could cause problems throughout the

operation of wastewater treatment and other processes

where high concentrations of ammonium, magnesium,

and phosphate are present.

Ozturk et al. (2003) used Struvite as precipitant for

the removal of NH3-N from anaerobically pre-treated

leachate. The removal efficiency of NH3-N and COD

were 90% and 50%, respectively. It is confirmed that

the ammonium concentration in leachate could be

considerably reduced by struvite precipitation.

However, this process requires relatively expensive

chemicals (Kochany and Lipczynska-Kochany, 2009).

3.3.4. Adsorption

Adsorption technique is recognized as the efficient

and promising elementary approach in wastewater

treatment processes (Foo and Hameed, 2009). It is

used as a stage of integrated chemical-physical-

biological method for leachate treatment, or

simultaneously with a biological process. The most

commonly used adsorbent is granular activated carbon

or PAC (Abbas et al., 2009).

PAC as adsorbent improved performance of SBR

process. Results revealed that the PAC-SBR offered

better removal efficiencies of pollutants and improved

sludge characteristics (Aziz et al. 2011; 2013)

3.3.5. Ammonium stripping

Because of its high removal efficiency, the

ammonium stripping treatment method is the most

widely used for the removal of NH3-N from landfill

leachate. High concentrations of NH4-N are

commonly found in landfill leachates, which lead to

increasing wastewater toxicity. This method usually

changes dissolved NH3-N and dissolved CH4 in

leachate to gases. High pH values must be provided in

this method so as the column works efficiently (Hao

et al., 2010; Ozturk et al., 2003; Marttinen et al.,

2002).

Sincero and Sincero (2003) explained that

stripping was completed via introducing wastewater at

the upper part of the stripping column and permitting

it to flow the air down in the column. Marttinen et al.

(2002) reported that this method at operation

condition of pH value equal to 11, temperature of

20°C, and retention time of 24 h resulted in NH3-N

removal of 89%. In a research conducted by Silva et

al. (2004), the researchers obtained NH3-N removal

efficiency of 99.5%. Calli et al. (2005) explained that

the removal efficiencies of NH3-N and COD for initial

concentration of 3260 mg/L were 94 % and <15%,

respectively.

Comparing with other treatment techniques such as

reverse osmosis and nanofiltration ammonium

stripping discovered to be more economical in terms

of operational costs. Regardless of its advantages, the

main disadvantages of this technique are: 1) The

release of NH3-N gas to the atmosphere, 2) The

scaling of CaCO3 when lime is used for pH adjustment

in the stripping tower, and 3) Adjustment of effluent

pH value prior discharge (Li et al., 1999).

3.3.6. Chemical oxidation

Oxidation and reduction process are based on transfer

of electrons. In this process, one compound gives

electrons whereas the other receives electrons. When



218

biological treatments are inefficient, this method is

used to achieve desired objectives. Chemical

oxidation process is required for the treatment of

wastewater containing non-biodegradable and/or toxic

pollutants. So, this method is an efficient technique

for the treatment of low BOD5/COD (i.e. stabilized)

landfill leachates (Derco et al., 2010; Mohajeri et al.,

2010; Rivas et al., 2003; Droste, 1997; Marco et al.,

1997).

Amokrane et al. (1997) used oxidants, such as

chlorine, potassium permanganate, ozone, and

calcium hydrochloride, for landfill leachate treatment

and found COD removal of 20–50%. Researchers

reported that the efficiency of COD reduction for

mature and biologically pretreated landfill leachates

were 60 to 75 %, respectively by using Fenton reagent

(Lopez et al., 2004; Kang and Hwang, 2000).

Fenton oxidation was used by Mohajeri et al.

(2010) for the removal of colour and COD from Pulau

Burung stabilized landfill leachate whereby colour

and COD removal were 78% and 58%, respectively

(Mohajeri et al., 2010). Additionally, Gotvajn et al.

(2009) mentioned that the removal efficiency of NH3-

N by using Fenton oxidation was 40%.

According to Lopez et al. (2004), Fenton process

seemed to be the best compromise because it is

technically simple. In spite of simplicity of Fenton

process, it requires low pH and a modification of this

parameter is necessary. Further, high demand of

electrical energy and high oxidant doses are other

disadvantages of this method that makes the process

expensive (Bashir, 2011).

3.3.7. Ion exchange

Ion exchange process could be defined as reversible

interchange of ions among the liquid and solid phases

where no significant change in the structure of the

solid is observed. This treatment process is capable of

efficiently removing the traces of metal impurities to

meet the increasingly strict discharge standards in

developed countries. However, the application of ion

exchange is not generally used for the treatment of

landfill leachate, but it received great attention in

Germany for the removal of non-biodegradable

compounds that contained humic substances (Fettig,

1999).

Researchers explained that ion exchange resins are

generally and capably used for the removal of organic

compounds and ions from water and wastewater and

as a polishing step in landfill leachate treatment. Prior

to ion exchange process, leachate should initially be

subjected to biological treatment (Bashir et al., 2010;

Kurniawan et al., 2006). All soluble metallic elements

(anionic or cationic) could be efficiently removed or

reduced by using ion exchange technique. The resin is

prepared of synthetic organic polymers or natural

zeolite. Ions such as H+, OH

-, Na

+, and Cl

- are joined

to the resin by weak electrostatic forces. These ions

are exchanged with ions in the contaminated product

that have more similarity for the resin. Resins could

be prepared to pick particular ions. The application of

ion exchange is economically limited due to high

operational cost. Another limitation is the requirement

of suitable pre-treatment system such as the removal

of suspended solids from landfill leachate. However,

ion exchange process is proper for heavy metal

removal from leachate (Kurniawan et al., 2006).

3.3.8. Electrochemical oxidation

Recently, electrochemical oxidation process has

received important consideration for wastewater

treatment because of its efficient and simple

operation. This treatment process has the ability to

destroy refractory pollutants. In this process,

refractory pollutants could be destroyed by direct

anodic oxidation, electro-chemically, or by indirect

oxidation (Atmaca, 2009; Deng and Englehrdt, 2007,

Chiang et al., 1995). Several types of electrodes have

been examined for electrochemical treatment such as

Ti, TiO2, PbO2/Ti, SnO2/Ti, Fe, aluminum, iron, and

graphite (Atmaca, 2009; Irdemez et al., 2006; Shen et

al., 2006; Chiang et al., 1995).

Atmaca (2009) used cast iron plates for both

anodes and cathodes with surface area of 22.6 cm2. At

operational conditions of 20 min treatment duration,

constant DC current of 3A, H2O2 concentration of

2000 mg/L, and at initial pH 3, removal efficiencies of

colour and COD were 90% and 72 %, respectively.

Electrochemical oxidation process was used for the

treatment of Pulau Burung semi-aerobic landfill

leachate, Malaysia. Sodium sulphate Na2SO4 (as

electrolyte) and carbon electrodes were used in the

treatment process. At optimum operational conditions

(1414 mg/L influent COD, 79.8 m A/cm2 current

density and 4 h reaction time), the removal

efficiencies of BOD5, COD, and colour were 69.8 %,

67.6 %, and 83.7 %, respectively. Further, graphite

carbon electrode showed better performance in the

removal of BOD5, COD and colour (Bashir, 2007).

The disadvantages of this method are: 1) It is more

expensive than other treatment processes because of

high energy consumption, 2) It is less widely used for

the treatment of stabilized leachate, 3) High current

density could increase corrosion rate of electrodes,

and 4) It is inefficient for the removal of inorganic

substances and NH3-N.

Aziz et al.


219

3.3.9. Membrane filtration

A membrane could be defined as a material that

creates a thin barrier capable of selectively resisting

the move of different constituents of a fluid and

therefore affecting separation of the constituents

(Visvanathan et al., 2000). Usually, a thin layer of

material with a high surface porosity and a narrow

domain of pore size affect the physical structure of the

membrane. Different membrane filtration techniques:

microfiltration, ultrafiltration, nanofiltration, and

reverse osmosis are used in landfill leachate treatment.

i) Microfiltration

Microfiltration with pore sizes of 0.05 to 10 microns

is employed to capture microbial cells, small particles,

and large colloidal. According to landfill leachate

treatment, this method is not suitable to be used alone.

It is recommended to be used as pretreatment process

with other membrane processes (i.e. ultrafiltration,

nanofiltration or reverse osmosis) or in combination

with chemical treatment processes so as to remove

suspended matters and colloids. Piatkiewicz et al.

(2001) used this method as pre-filtration stage and

obtained COD removal of 25% to 35 %.

ii) Ultrafiltration

Ultrafiltration is a selective process utilizing pressures

up to 10 bar. This technique is efficient to remove

suspended matters either by direct filtration or with

biological treatment to replace sedimentation unit. It is

strongly dependant on the kind of material

constituting the membrane. Syzdek and Ahlert (1984)

proposed that this process might prove to be useful as

a pre-treatment method for reverse osmosis. It could

be employed to eliminate the larger molecular weight

components of leachate that tend to foul reverse

osmosis membranes (Bohdziewicz et al., 2001;

Rautenbach et al., 1996). COD removal of 50% was

obtained by using ultrafiltration alone (Bohdziewicz et

al., 2001)

Lastly, Tabet et al. (2002) reported that

ultrafiltration membranes have been successfully

employed in full scale membrane bioreactor plants by

combination of bioreactors and membrane technology.

High levels for landfill leachate treatment have been

obtained by using this method.

iii) Nanofiltration

Because of its unique properties between

ultrafiltration and reverse osmosis membranes,

nanofiltration has discovered a place in the

elimination of refractory organic compounds and

heavy metals from landfill leachate (Ozturk et al.,

2003). Removals of COD and NH3-N from landfill

leachate by using this process were 60 % to 70 % and

50 %, respectively (Trebouet et al., 2001; Linde and

Jonsson, 1995). In another research conducted by

Trebouet et al. (2001), nanofiltration was employed in

combination with physical processes which offered

acceptable COD removal (70% to 80 %) from landfill

leachate. Based on the results achieved by Linde and

Jonsson (1995), removal efficiencies of sulphate salts,

chloride, and other (Pb2+

, Zn2+

, and Cd2+

) from landfill

leachate were 88-96%, 12-47 %, and > 88 %,

respectively.

On the other hand, successful use of membrane

technique needs effective control of membrane

fouling. An extensive spectrum of components could

contribute to membrane fouling in leachate

nanofiltration which includes inorganic substances,

dissolved organic, suspended particles, and colloidals

(Trebouet et al., 2001).

iv) Reverse osmosis

Membrane techniques, specifically reverse osmosis, is

a relatively new method that seems to be a more

efficient alternative than traditional methods for

mature landfill leachate treatment. This method

involves separating two solutions with various

concentrations by using a semi-permeable membrane

(Kurniawan et al., 2006; Ahn et al., 2002; Chianese et

al., 1999).

Ahn et al. (2002) stated that a landfill leachate

treatment plant in Korea was retrofitted to improve

treatment efficiency by employing integrated

membrane technique that was composed of membrane

bioreactor and reverse osmosis method. The removal

efficiencies of COD and NH3-N from young landfill

leachate were 96 % and 97 %, respectively. Other

researchers stated that the removal of COD and NH3-

N from landfill leachate was 98% (Linde et al., 1995).

Although, reverse osmosis technique was reported

as the most efficient in the removal of COD among

various physical-chemical technologies assessed

(Peters 1998), some disadvantages have been noticed

for membrane techniques. They include membrane

clogging which reduces the overall process

performance by lowering the reject concentration

while the cleaning of such membranes also reduces

their lifetime. Additionally, the production of large

quantitities of residuals which are generally useless

and required to be discharged or need further

treatment (Li et al., 2009; Wiszniowski et al., 2006).



220

4. CONCLUSIONS

Design of landfill, age of landfill leachate, climate etc.

have great effect on leachate treatment. In this work,

detailed review on various treatment techniques were

explained. Results revealed that combined treatment

with domestic sewage, recycling, and biological

systems are efficient for processing young landfill

leachate. Commonly, physical-chemical treatment

methods are competent for treating medium and

old/stabilized leachates. Further, augmenting physical

and chemical methods or biological and physical-

chemical processes enhanced treatment of stabilized

leachate.

REFERENCES

Abbas AA, Jingsong G, Ping LZ, Ya PY, Al-Rekabi

WS (2009). Review on landfill leachate

treatments. Journal of Applied Sciences

Research, 5(5): 534–545.

Adlan MN, Palaniandy P, Aziz HA (2011).

Optimization of coagulation and dissolved air

flotation (DAF) treatment of semi-aerobic

landfill leachate using response surface

methodology (RSM). Desalination, 277(1–3):

74–82.

Aghamohammadi N (2006). Semi-aerobic leachate

treatment using powdered activated carbon

augmented activated sludge process.

Unpublished, M.Sc. thesis, School of Civil

Engineering, Universiti Sains Malaysia.

Ahn S, Congeevaram S, Choung YK, Park J (2008).

Enhanced phenol removal by floating fungal

populations in a high concentration phenol–fed

membrane bioreactor. Desalination, 221: 494–

501.

Ahn WY, Kang MS, Yim SK, Choi KH (2002).

Advanced landfill leachate treatment using an

integrated membrane process. Desalination,

149: 109–l 14.

Akkaya E, Demir A, Karadag D, Varank G, Bilgili

MS, Ozkaya B (2010). Post-treatment of

anaerobically treated medium-age landfill

leachate. Environmental Progress and

Sustainable Energy, 29(1): 78–84.

Al-Hamadani YAJ, Yusoff MS, Umar M, Bashir

MJK, Adlan MN (2011). Application of

psyllium husk as coagulant and coagulant aid in

semi-aerobic landfill leachate treatment. Journal

of Hazardous Materials, 190: 582–587.

Al-Rekabi WS, Qiang H, Qiang WW (2007). Review

on sequencing batch reactors. Pakistan Journal

of Nutrition, 6(1): 11–19.

Al-Shamrani AA, James A, Xiao H (2002). Separation

of oil from water by dissolved air flotation.

Colloids and surfaces A: Physicochemical and

Engineering Aspects, 209(1): 15–26.

Amokrane A, Comel C, Veron J (1997). Landfill

leachates pretreatment by coagulation–

flocculation. Water Research, 31(11): 2775–

2782.

Asian Regional Research Programme on

Environmental Technology, ARRPET (2004).

State of the art review landfill leachate

treatment. Asian Institute of Technology,

Tongji University.

Atmaca E (2009) Treatment of landfill leachate by

using electro-Fenton method. Journal of

Hazardous Materials, 163(1): 109–114.

Aziz HA, Alias S, Adlan MN, Faridah Asaari AH,

Zahari MS (2007). Colour removal from

landfill leachate by coagulation and flocculation

processes. Bioresource Technology, 98: 218–

220.

Aziz HA, Daud Z, Adlan MN, Hung YT (2009). The

use of polyaluminium chloride for removing

colour, COD and ammonia from semi-aerobic

leachate. Int. J. Environmental Engineering,

1(1): 20–35.

Aziz SQ (2013). Produced Leachate from Erbil

Landfill Site, Iraq: Characteristics, Anticipated

Environmental Threats and Treatment. The

16th International Conference on Petroleum,

Mineral Resources and Development, Cairo,

Egypt. 10-12 Feb., 2013.

Aziz SQ, Aziz HA, Yusoff MS, Bashir MJK (2011).

Landfill leachate treatment using powdered

activated carbon augmented sequencing batch

reactor (SBR) process: Optimization by

response surface methodology. Journal of


Aziz SQ, Aziz HA, Mojiri A, Bashir MJK, Abu Amr

S (2013). Landfill Leachate Treatment Using

Sequencing Batch Reactor (SBR) Process:

Limitation of Operational Parameters and

Performance. International Journal of Scientific

Research in Knowledge (IJSRK), 1(3): 34-43.

Aziz SQ, Aziz HA, Yusoff MS, Bashir MJK, Umar M

(2010). Leachate characterization in semi-

aerobic and anaerobic sanitary landfills: A

comparative study. Journal of Environmental

Management, 91: 2608-2614.

Bae JH, Cho KW, Bum BS, Lee SJ, Yoon BH (1998).

Effects of leachate recycle and anaerobic

digester sludge recycle on the methane

production from solid waste. Water Science and

Technology, 38(2): 159–168.

Bagchi A (1994). Design Construction and

Monitoring of Landfills, John Wiley & Sons,

New York.

Aziz et al.


221

Bashir MJK (2007). Landfill leachate treatment by

electrochemical oxidation. Unpublished, M.Sc.

thesis, School of Civil Engineering, Universiti

Sains Malaysia.

Bashir MJK (2011). Removal of colour, COD and

NH3-N from landfill leachate using anionic and

cationic resin. Unpublished, PhD thesis, School

of Civil Engineering, Universiti Sains Malaysia.

Bashir MJK, Aziz HA, Aziz SQ, Abu Amr SS (2012).

An overview of electro-oxidation processes

performance in stabilized landfill leachate

treatment. Desalination and Water Treatment,

DOI:10.1080/19443994.2012.734698.

Bashir MJK, Aziz HA, Yusoff MS, Adlan MN (2010).


(RSM) for optimization of ammoniacal nitrogen

removal from semiaerobic landfill leachate

using ion exchange resin. Desalination, 254:

154–161.

Baun A, Kløft L, Bjerbg PL, Nyholm N (1999).

Toxicity testing of organic chemicals in

groundwater polluted with landfill leachate,

Environ. Toxicol. Chem. 18: 2046–2053.

Burton, S.Q., Watson-Craik, I.A. (1998) Ammonia

and nitrogen fluxes in landfill sites:

applicability to sustainable landfilling, Waste

Manage. Res. 16, 41–53.

Bohdziewicsz J, Bodzek M, Gorska J (2001).

Application of thee pressure-driven membrane

techniques to biological treatment of landfill

leachate. Process Biochemistry, 36: 641–646.

Bull PS, Evans JV, Wechsler RM, Cleland KJ (1983).

Biological technology of the treatment of

leachate from sanitary landfills. Water

Research, 17: 1473–1481.

Calli B, Metroglu B, Inanc B (2005). Landfill leachate

management in Istanbul: Application and

alternatives. Chemosphere, 59 (6): 819–829.

Cecen F, Aktas O (2001) Effect of PAC addition in

combined treatment of landfill leachate and

domestic wastewater in semi-continuously fed

batch and continuous–flow reactors. Water SA,

27(2).

Cecen F, Aktas O (2004). Aerobic co-treatment of

landfill leachate with domestic wastewater.

Environmental Engineering Science, 21: 303–

312.

Chianese A, Ranauro R, Verdone N (1999). Treatment

of landfill leachate by reverse osmosis. Water

Research, 33 (3), 647–652.

Chiang LC, Chang JE, Wen TC (1995). Indirect

oxidation effect in electrochemical oxidation

treatment of landfill leachate. Water Research,

29 (2): 671–678.

Chugh S, Clarke W, Pullammanappallil P, Rudolph V

(1998). Effect of recirculated leachate volume

on MSW degradation. Waste Management and

Research, 16: 564–573.

Deng Y, Englehardt JD (2007). Electrochemical

oxidation for landfill leachate treatment. Waste

Management, 27(3): 380–388.

Derco J, Gotvajn AZ, Zagorc-Koncan J, Almasiova B,

Kassai A (2010). Pretreatment of landfill

leachate by chemical oxidation processes.

Chemical papers, 64(2): 237–245.

Dialynas E, Mantzavinos D, Diamadopoulos E

(2008). Advanced treatment of the reverse

osmosis concentrate produced during

reclamation of municipal wastewater. Water

Research, 42 (18): 4603–4608.

Dollerer J, Wilderer PA (1996). Biological treatment

of leachates from hazardous waste landfills

using sbbr technology. Water Science and

Technology, 34 (7–8): 437–444.

Droste R (1997) Theory and practice of water and

wastewater treatment. John Wiley and Sons,

Inc., USA.

Duan J, Greogory J (2003). Coagulation by

hydrolyzing metal salts. Advances in Colloid

and Interface Science, 100–102: 475–502.

El-Fadel M, Bou-Zeid E, Chahine W, Alayli B (2002).

Temporal variation of leachate quality from

pre-sorted and baled msw with high organic and

moisture content. Waste Management, 22: 269–

282.

Ernst WR, Hennigar P, Doe K, Wade S, Julien G

(1994). Characterization of the chemical

constituents and toxicity to aquatic organisms

of a municipal landfill leachate, Water Pollut.

Res. J. Can., 29: 89–101.

Fettig J (1999) Removal of humic substances by

adsorption/ion exchange. Water Science and

Technology, 40 (9): 173–182.

Foo KY, Hameed BH (2009). An overview of landfill

leachate treatment via activated carbon

adsorption process. Journal of Hazardous

Materials, 171: 54–60.

Garcia H, Rico JL, Garcia PA (1996). Comparison of

anaerobic treatment of leachates from an urban-

solid-waste landfill at ambient temperature and

at 35°C. Bioresource Technology, 58 (3), 273–

277.

Ghafari S, Aziz HA, Isa, MH, Zinatizadeh AA (2009).


(RSM) to optimize coagulation-flocculation

treatment of leachate using poly–aluminum

chloride (PAC) and alum. Journal of Hazardous

Materials, 163: 650–656.

Gotvajn AZ, Tisler T, Zagorc–Koncan J (2009).

Comparison of different treatment strategies for

industrial landfill leachate. Journal of

Hazardous Materials, 162(2–3): 1446–1456.



222

Gulsen H, Turan M (2004). Startup of anaerobic

fluidized bed reactor for landfill leachate

treatment. Environmental Technology, 25(10):

1107–1114.

Guo JS, Abbas AA, Chen YP, Liu ZP, Chen P (2010)

Treatment of landfill leachate using a combined

stripping, Fenton, SBR, and coagulation

process. Journal of Hazardous Materials, 178:

699 –705.

Hao YJ, Chen YX, Wu WX, Zhang SG, Liu HQ, Ji M

(2010). The pathway in–situ ammonium

removal from aerated municipal solid waste

bioreactor: Nitrification/denitrification or air

stripping? Waste Management and Research, 1-

8.

Henry JG, Prasad D, Young H (1987). Removal of

organics from leachates by anaerobic filter.

Water Research, 21 (11), 1395–1399.

Hoilijoki TH, Kettunen RH, Rintala JA (2000).

Nitrification of anaerobically pretreated

municipal landfill leachate at low temperature.

Water Research, 34(5): 1435–1446.

Horan NJ, Gohar H, Hill B (1997). Application of a

granular activated carbon–biological fluidized

bed for the treament of landfill leachate

containing high concentrations of ammonia.

Water Science and Technology, 36: 369–375.

Imai A, Iwami N, Matsushige K, Inamori Y, Sudo R

(1993). Removal of refractory organics and

nitrogen from landfill leachate by the

microorganism-attached activated carbon

fluidized bed process. Water Research, 27:

143–145.

Irdemez S, Demircioglu N, Yildiz YS, Bingul Z

(2006). The effects of current density and

phosphate concentration on phosphate removal

from wastewater by elctro coagulation using

aluminum and iron plate electrodes. Separation

and Purification Technology, 52: 218–223.

Japan International Cooperation Agency (JICA)

(2005). A practical guide to landfill

management in Pacific island countries,

Volume-1: Inland-based waste disposal.

Jokela JPY, Kettunenb RH, Sormunena KM, Rintala

JA (2002). Biological nitrogen removal from

municipal landfill leachate: Low–cost

nitrification in biofilters and laboratory scale in-

situ denitrification. Water Research, 36: 4079–

4087.

Kang YW, Hwang KY (2000). Effects of reaction

conditions on the oxidation efficiency in the

Fenton process.Water Research, 34(10): 2786–

2790.

Kennedy KJ, Lentz EM (2000). Treatment of landfill

leachate using sequencing batch and continuous

flow upflow anaerobic sludge blanket (UASB)

reactors. Water Research, 34 (14), 3640–3656.

Kettunen RH, Rintala JA (1995). Sequential

anaerobic-aerobic treatment of sulphur rich

phenolic leachates. Journal of Chemical

Technology and Biotechnology, 62(2): 177–

184.

Klimiuk E, Kulikowska D (2006) The influence of

hydraulic retention time and sludge age on the

kinetics of nitrogen removal from leachate in

SBR. Polish Journal of Environmental Studuies,

15(2): 283–289.

Kochany J, Lipczynska-Kochany E (2009) Utilization

of landfill leachate parameters for pretreatment

by Fenton reaction and struvite precipitation–A

comparative study. Journal of Hazardous

Material, 166: 248–254.

Kostova I (2006). Leachate from Sanitary

LandfillseOrigin, Characteristics,

Treatment.University of Architecture, Civil

Engineering and Geodesy, “Iskar’s Summer

School”-Borovetz, 26-29 July.

Kurniawan TA, Lo WH, Chan GYS (2006). Physico–

chemical treatments for removal of recalcitrant

contaminants from landfill leachate. Journal of

Hazardous Materials, 129(1–3): 80-100.

Lema JM, Mendez R, Blazquez R (1988).

Characteristics of landfill leachates and

alternatives for their treatment: A review. Water

Air and Soil Pollution–Water, Air, & Soil

Pollution, 40(3–4): 223–250.

Li HS, Zhou SQ, Sun YB, Feng P, Li JD (2009).

Advanced treatment of landfill leachate by a

new combination process in a full–scale plant.

Journal of Hazardous Materials, 172(1): 408–

415.

Li XZ, Zhao QL, Hao XD (1999). Ammonium

removal from landfill leachate by chemical

precipitation. Waste Management, 19(6): 409–

415.

Lin CY, Chang FY, Chang CH (2000). Co–digestion

of leachate with septage using a UASB reactor.

Bioresource Technology, 73: 175–178.

Linde K, Jonsson AS (1995). Nanofiltration of salt

solution and landfill leachate. Desalination,

103: 223–232.

Linde K, Jonsson AS, Wimmerstedt R (1995).

Treatment of three types of landfill leachate

with reverse osmosis. Desalination, 101(1): 21–

31.

Lopez A, Pagano M, Volpe A, Pinto ACD (2004).

Fenton's pre–treatment of mature landfill

leachate. Chemosphere, 54(7): 1005–1010.

Loukidou MX, Zouboulis AI (2001). Comparison of

two biological treatment process using attached

growth biomass for sanitary landfill leachate

http://www.ncbi.nlm.nih.gov/pubmed?term=%22Lipczynska-Kochany%20E%22%5BAuthor%5D

http://www.springerlink.com/content/0049-6979/

http://www.springerlink.com/content/0049-6979/

Aziz et al.


223

treatment. Environmental pollution, 111(2):

273–281.

Maehlum T (1995). Treatment of landfill leachate in

on-site lagoons and constructed wetlands.

Water Science and Technology, 32: 129–135.

Mahvi AH (2008) Sequencing batch reacttor: A

promising technology in wastewater treatment,

Iran. J. Environ. Health. Sci. Eng., 5(2): 79–90.

Marco A, Esplugas S, Saum G (1997). How and why

combine chemical and biological processes for

wastewater treatment. Water Science and

Technology, 35(4): 321–327.

Martienssen M, Schops R (1997). Biological

treatment of leachate from solid waste landfill

sites-Alterations in the bacterial community

during the denitrification process. Water

Research, 31(5): 1164–1170.

Marttinen SK, Kettunen RH, Sormunen KM,

Soimasuo RM, Rintala JA (2002). Screening of

physical-chemical methods for removal of

organic material, nitrogen and toxicity from low

strength landfill leachates. Chemosphere, 46(6):

851–858.

Maynard HE, Ouki SK, Williams SC (1999). Tertiary

lagoons: A review of removal mechanisms and

performance. Water Research, 33: 1–13.

Mohajeri S, Aziz HA, Isa MH, Zahed MA, Adlan MN

(2010). Statistical optimization of process

parameters for landfill leachate treatment using

electro-Fenton technique. Journal of Hazardous

Materials, 176: 749–758.

Matsufuji Y, Hanashima M, Nagano S, Tanaka A

(1993). Generation of Greenhouse Effect Gases

from Different Landfill Type. Engineering

Geology, 34: 181-197.

Nedwell DB, Reynolds PJ (1996). Treatment of

landfill leachate by methanogenic and sulphate

reducing digestion. Water Research, 30: 21–28.

Oman CB, Junestedt C (2008). Chemical

characterization of landfill leachates–400

parameters and compounds. Waste

Management, 28: 1876–1891.

Oztruk I, Atinbas M, Koyuncu I, Arikan Q, Gomec–

Yangin C (2003). Advanced physic–chemical

treatment experiences on young municipal

landfill leachates. Waste Management, 23(5):

441–446.

Palaniandy P, Adlan MN, Aziz HA, Murshed MF

(2010). Application of dissolved air flotation

(DAF) in semi-aerobic leachate treatment.

Chemical Engineering Journal, 157: 316–322.

Peters, T.A. (1998) Purification of landfill leachate

with reverse osmosis and nanofiltration.

Desalination, 119(1–3): 289–293.

Piatkiewicz W, Biemacka E, Suchecka T (2001). A

polish study: Treating landfill leachate with

membranes. Filtration and Separation, 38: 22–

26.

Rautenbach R, Vossenkaul K, Linn T, Katz T (1996).

Wastewater treatment by membrane processes–

New development in ultrafiltration,

nanofiltration, and reverse osmosis.


Reinhart DR, Al-Yousfi AB (1996) The impact of

leachate recirculation on municipal solid waste

landfill operating characteristics. Waste

Management and Research, 14: 337–346.

Renou S, Givaudan GJ, Poulain S, Dirassouyan F,

Moulin P (2008). Landfill leachate treatment:

Review and opportunity. Journal of Hazardous

Materials, 150: 468–493.

Rivas FJ, Beltran F, Gimeno O, Acedo B, Carvalho F

(2003). Stabilized leachates: Ozone-activated

carbon treatment and kinetics. Water Research,

37(20): 4823–4834.

Robinson HD, Barr MJ, Last SD (1992). Leachate

collection, treatment and disposal. Water and

Environmental Journal, 6(4): 321–332.

Sanphoti N, Towprayoon S, Chaiprasert P,

Nopharatana A (2006). The effects of leachate

recirculation with supplemental water addition

on methane production and waste

decomposition in a simulated tropical landfill.

Journal of Environmental Management, 81: 27–

35.

Schrab GE, Brown KW, Donnelly KC (1993) Acute

and genetic toxicity of municipal landfill

leachate. Water, Air, and Soil Pollution, 69: 99–

112.

Shen ZM, Wu, D, Yang J, Yuan T, Wang WH, Jia JP

(2006). Methods to improve electrochemical

treatment effect of dye wastewater. Journal of

Hazrdous Materials, 131(1–3): 90–97.

Silva AC, Dezotti M, Sant'Anna JGL (2004).

Treatment and detoxification of a sanitary

landfill leachate. Chemosphere, 55: 207–214.

Sincero AP, Sincero GA (2003). Physical-chemical

treatment of water and wastewater. I WA

Publishing, CRC Press, New York, USA.

Spagni A, Marsili-Libelli S, Lavagnolo MC (2008).

Optimisation of sanitary landfill leachate

treatment in a sequencing batch reactor, Journal

of Water Science and Technology, 58(2): 337–

443.

Suidan MT, Schroeder A, Nath R, Krishnan E,

Brenner R (1993). Treatment of cercla

(Comprehensive environmental response,

compensation, and liability act) leachates by

carbonassisted anaerobic fluidized beds. Water

Science and Technology, 27: 273–282.

Sung MS, Chang D, Lee HY (1997). Performance

improvement of an unstable anaerobic leachate



224

treatment system in an industrial waste landfill.

Water Science and Technology, 36(12): 333–

340.

Syzdek AC, Ahlert RC (1984). Separation of landfill

leachate with polymeric ultrafiltration

membranes. Journal of Hazardous Materials, 9:

209–220.

Tabet K, Moulin P, Vilomet JD, Amberto A, Charbit

F (2002). Purification of landfill leachate with

membrane processes: Preliminary studies for an

industrial plant. Separation Science and

Technology, 37(5): 1041–1063.

Tatsi AA, Zouboulis AI, Matis KA, Samaras P (2003)

Coagulation–flocculation pretreatment of

sanitary landfill leachates. Chemosphere, 53(7):

737–744.

Tauchert E, Schneider S, Morais JLD, Peralta-Zamora

P (2006). Photochemically–assisted

electrochemical degradation of landfill leachate.

Chemosphere, 64: 1458–1463.

Timur H, Ozturk I (1999). Anaerobic sequencing

batch reactor treatment of landfill leachate.

Water Research, 33(15): 3225–3230.

Timur H, Ozturk I (1997). Anaerobic treatment of

leachate using sequencing batch reactor and

hybrid bed filter. Water Science and

Technology, 36(6–7): 501–508.

Trebouet D, Schlumpf JP, Jaouen P, Quemeneur F

(2001). Stabilized landfill leachate treatment by

combined physicochemical–nanofiltration

processes. Water Research, 35(12): 2935–2942.

Uygur A, Kargi F (2004) Biological nutrient removal

from pre-treated landfill leachate in a

sequencing batch reactor. Journal of

Environmental Management, 71: 9–14.

Visvanathan C, Aim RB, Parameshwaran K (2000)

Membrane separation bioreactors for

wastewater treatment. Environmental Science

and Technology, 30(1): 1–48.

Welander U, Henrysson T, Welander T (1998).

Biological nitrogen removal from municipal

landfill leachate in a pilot scale suspended

carrier biofilm process. Water Research, 32(5):

1564–1670.

Widziewicz K, Kalka J, Skonieczna M, Madej P

(2012). The Comet Assay for the Evaluation of

Genotoxic Potential of Landfill Leachate. The

Scientific World Journal, Volume 2012, 1-8,

Article ID 435239, doi:10.1100/2012/435239.

Wiszniowski J, Robert D, Surmacz-Gorska J, Miksch

K, Weber JV (2006). Landfill leachate

treatment methods: A review. Environmental

Chemistry Letters, 4(1): 51–61.

Yamamoto O (2002). Solid waste treatment and

disposal experiences in Japan. Proceedings of

International Symposium on Environmental

Pollution Control and Waste Management

(EPCOWM’2002), Tunis, 417-424.

Zaloum R, Abbott M (1997). Anaerobic pretreatment

improves single sequencing batch reactor

treatment of landfill leachates. Water Science

and Technology, 35 (1): 207–214.

Zhang T, Ding L, Ren H (2009). Pretreatment of

ammonium removal from landfill leachate by

chemical precipitation. Journal of Hazardous

Materials, 166(2–3): 911–915.

Ziyang L, Youcai Z, Tao Y, Yu S, Huili C, Nanwen

Z, Renhua H (2009) Natural attenuation and

characterization of contaminants composition in

landfill leachate under different disposing ages.

Science of the Total Environment, 407: 3385–

3391.

Zouboulis AI, Chai XL, Katsoyiannis IA (2004). The

application of bioflocculant for the removal of

humic acids from stabilized landfill leachates.

Journal of Environmental Management, 70: 35–

41.

Zouboulis AI, Jun W, Katsoyiannis IA (2003).

Removal of humic acids by flotation. Colloids

and Surfaces A: Physicochemical and

Engineering Aspects, 231: 181–193.



©2014 IJSRPUB

226

Application of Optimization in Wastewater Treatment

Yee Ling Lim, Yeek Chia Ho, Abbas F. M. Alkarkhi*

School of Industrial Technology, Universiti Sains Malaysia, 11800 Penang, Malaysia


Abstract. Design and analysis of experiments (DOE) helps researchers achieves the research objective with less effort, cost

and time. In wastewater treatment process, DOE is used to study the effect of input variables such as flow rate/speed,

temperature, chemical dosing and retention time on the output and then optimize the process by finding the best operating

conditions for the input variables that meet the objective for the output variables and summarize the results in terms of model

to describe the behavior of the process. First-order and second-order models are usually used to describe the behavior of most

cases.

Keywords: Optimization, Wastewater Treatment,

1. INTRODUCTION

Design and analysis of experiments (DOE) have been

widely used in planning, analyzing and running

experiments in different areas of research, such as

wastewater treatment, food analysis, material

production and medication intake helping the

researchers to achieve the research objective with less

effort, cost and time. The application of DOE in

wastewater industries has increased since the

application of experimental design and analysis of

experiments enables us to collect data effectively and

reduce the error by excluding non-significant factor/s

in the experiment and effectively improves the results

to the target range. Three basic principles should be

considered in using experimental design, namely,

replication, randomization and blocking. Replication

allows experimenter to get an estimation of

experimental error and to increase the precision.

Randomization is the allocation of the experimental

materials and the order that the individual runs or

trials of experiment are to be carried out in random

manner. Furthermore, randomization eliminates bias

and to make sure independence among the

observations. Blocking technique is used to enhance

the accuracy by reducing or eliminates variability that

may affect the response (output) such as the difference

in skills of two or more operators and weather

conditions (rainy, cloudy, sunny) (Montgomery, 2012).

Researchers should provide clear objective of the

project before using experimental design, for instance

number of input variables and levels of each variable

(the region of each selected variable). As an example,

an outcome of an experiment includes the

identification of which factors contribute to overall

wastewater treatment process, for instance the factors

involved may include chemical dosing and total

retention time, while their levels may be 100g/L or

500g/L of chemical dosing and 3 days or 7 days for

the total retention time.

Experimental design is frequently used in industry

to optimize different experiments in order to find the

best operating conditions that produce desirable

product. Moreover, the application of DOE enables a

wastewater treatment system to optimize the

effectiveness of the process in removing the impurities,

particularly to fulfill the government regulation on the

effluent water. Optimization enables processes to

achieve highest possible removal rate within the given

range of factors at the lowest cost possible. For

instance, in a conventional treatment plant, the input

factors are flow rate/speed, temperature, chemical

dosing and retention time and the response (output) is

often more than one responses such as biochemical

oxygen demand (BOD), chemical oxygen demand

(COD), pH, concentration of cadmium, concentration

of lead and others. Thus, it is not quite possible to run

and plan a proper design for the experiment for good

yield without finding the best operating conditions

which results in highest removal.

2. APPLICATION OF DOE IN WASTEWATER

RESEARCH

The applications of design and analysis of

experiments in wastewater researches can be listed

down as follows:

(a) Production of adsorbent for wastewater, factors

involved may include heating temperature, heating

time and chemical composition while the response

may include removal of heavy metal, color removal or

even COD removal.

(b) For the cultivation of bacteria for the usage in

treatment plant on the other hand may include dosage

of carbon supply, shaking speed and incubation

temperature while the response is the growth of

bacteria.

(c) Conventional batch study involves parameters

of shaking speed, temperature, chemical dosage and

pH while the response will be the removal of the

target compound.

Lim et al.


227

(d) Study of flocculating activity involving

different physico-chemical parameters such as pH,

concentration of polymer, and mixing speed.

3. FACTORIAL DESIGN

Factorial designs are designs consist of several factors

(input variables) that influence one or more responses.

The advantage of factorial designs is to study the

effect of several factors and the interaction between

different factors simultaneously whilst the

disadvantage appears when the number of factors

included in the study increases which result in

increasing of the number of experiments. General

factorial design studies experiments with more than

one factor; each has at least two levels while special

cases of factorial designs are two-level factorial

design 2k and three-level factorial design 3

k.

3.1. Two-level Factorial Design

Two-level factorial design with k factors (2k) is more

commonly used for screening purposes in order to

identify insignificant factors and to prepare for

optimization step (Ryan, 2007). Researchers usually

collect experimental data by running only a fraction of

the factorial design which is known as fractional

factorial design (Srinivasan and Viraraghavan, 2010).

The two levels are represented by (-1) to indicate that

the factor is at low level and (+1) to represents the

high level. Two-level factorial design is used to fit

first-order polynomial model.

The levels of each factor in the actual form are

known as natural variable. The levels can be

converted into -1 and 1 by using Eq. 1, this formula is

called coded form. The relationship between the

natural variable i and the coded variables iX is

given in Eq.1.

2/)(

2/)(

levelLowlevelHigh

levelLowlevelHighx i

i

(1)

The benefit of using fractional factorial design is to

reduce the number of runs (experiments) as long as

the purpose of the experiment is to choose influential

factors, for instance, assume that 5 factors involved in

a process (25), the total number of runs is 32

experimental runs for each replicate to cover all

possible combinations whilst fraction factorial design

of 2k-1

(half factorial design) will use only 16 runs

which means less effort, time and cost to identify the

influential factors.

Example 3.1

Suppose a 25-1

two-level fractional factorial designs

with 16 runs were used to study the effects of

concentration of phenolic compounds, adsorbent

dosage, contact time, temperature and shaking speed

on phenolic compound removal (y) in wastewater.

The levels of each factor in coded and actual forms

are given in Table 1. The data for this experiment in

actual levels are given in Table 2.

Table 1: The factors and levels used for screening experiment Factor Symbol Levels

Actual Coded

Concentration of phenolic compound (mg/L) x1 20 40 -1 1

Adsorbent dosage (g/100ml) x2 0.10 0.30 -1 1

Contact time (hours) x3 120 300 -1 1

Temperature (oC) x4 50 80 -1 1

Shaking speed (rpm) x5 300 400 -1 1



228

Table 2: The results of two-level fractional factorial design

Concentration of

phenolic compound

(mg/L)

Adsorbent

dosage

(g/100ml)

Contact

time

(hours)

Temperature (oC) Shaking speed

(rpm)

% Removal by

phenolic

compounds

20 0.1 2 50 400 30.29

40 0.1 2 50 300 8.16

20 0.3 2 50 300 67.70

40 0.3 2 50 400 46.36

20 0.1 5 50 300 35.04

40 0.1 5 50 400 10.16

20 0.3 5 50 400 66.45

40 0.3 5 50 300 52.08

20 0.1 2 80 300 29.69

40 0.1 2 80 400 11.11

20 0.3 2 80 400 58.52

40 0.3 2 80 300 30.06

20 0.1 5 80 400 33.40

40 0.1 5 80 300 7.17

20 0.3 5 80 300 57.82

40 0.3 5 80 400 39.83

This experiment represents one-half fractional

factorial design with 16-runs, since the number of runs

is 16 which is not enough to provide degrees of

freedom to the error term in the analysis of variance

(ANOVA). Thus, in order to analyze this experiment

we will use normal probability plot or half normal

probability plot. Minitab statistical software (version

16) was used to analyze the data obtained from this

experiment. The analysis in the form of normal

probability plot is shown in Fig. 1. It can be seen that

shaking speed and contact time did not exhibit a

significant effect on the percentage removal of

phenolic compound. Only three factors showed

significant effect, namely concentration of phenolic

compound, adsorbent dosage and temperature.

Fig. 1: Normal probability plot of effects

An ANOVA table can be built after excluding non-

significant terms and analyze the result with only

three factors, concentration of phenolic compounds,

adsorbent dosage and temperature as a 23 design with

two replicates. Table 3 shows the results of ANOVA

for phenolic compound removal.

Lim et al.


229

Table 3: ANOVA table for phenolic compound removal

Source of variation DF Sum of square Mean sum of square F-value P-value

Main effects 3 6065.58 2021.86 173.85 0.000

2-Way interaction 3 131.83 43.94 3.78 0.059

3-Way interaction 1 10.47 10.47 0.90 0.371

Residual error 8 93.04 11.63

Pure error 8 93.04 11.63

Total 15 6300.92

It can be seen that three-factor interaction and two-

factor interaction were insignificant whilst the main

effect for selected factors was significant. Two-factor

interaction will be included in the model since this

interaction is significant at p<0.07.

The results are summarized in a regression model

to describe the behavior of the process. Regression

model for this experiment in coded units is given in

Eq. 2.

42142

4121421

81.02.76-

53.060.03.04-86.1587.1049.36

xxxxx

xxxxxxxy

(2)

3.2. 3k Factorial Design

Three-level factorial design (3k) is a special case of

factorial design where all factors included in the study

have three levels. This design is used for fitting

second-order polynomial model to the experimental

runs. The levels in 3k design are denoted as low,

intermediate and high with the assumption of 0 for

low, 1 for intermediate and 2 for high, given that k

represents the number of factors included in the study.

If the factors are quantitative, we may denote the low,

intermediate and high level as 1, 0 and +1 respectively

to facilitate the calculation.

4. RESPONSE SURFACE METHODOLOGY

(RSM)

Response Surface Methodology is very well known by

RSM which consists mathematical and statistical

techniques that are useful for modeling and analysis of

problems in which the objective is to optimize the

response or responses by finding the best operating

conditions of the input variables (independent

variables) that influence the response

(Prakobvaitayakit and Nimmannit, 2003; Montgomery,

2012).

The objective of RSM is to guide the researcher to

find the levels of independent variables that optimize

the response which means maximize, minimize or

keep in a range. RSM is widely used in water and

wastewater treatment to evaluate the effect of several

independent variables and their interactions such as

speed, flow rate and chemical dosage on a response or

responses such as COD, BOD, colour, turbidity etc.

Special designs are used to carry out experiment in

RSM. These designs are capable to fit second-order

and second-order models for instance central

composite design (CCD) with Box-Behnken design

(BBD) and three-level factorial design.

4.1. Steps in RSM

The main purpose of using RSM is to optimize a

process. This purpose can be achieved by using the

following steps:

(1) Choosing the right design that serves your

objective. All information regarding the experiment

should be available to help in choosing the design by

involving all people who work in the project.

(2) Run the experiment and analyze the data

obtained then summarize the data in a model form

usually first-order or second-order model.

(3) Validate the model obtained by checking its

precision in predicting the response and whether this

model can be used for finding the optimum conditions.

(Montgomery, 2012)

To optimize the response, it is necessary to

approximate the functional relationship between the

independent variables and the response surface. For

example, suppose an engineer would like to know the

best condition for the highest removal (yield, y) of

heavy metal from its wastewater, given the

independent influencing factors are temperature (x1)

and heavy metal concentration (x2) in wastewater.

The yield for this process can be represented by the

following function:

Removal of heavy metal = f(temperature, heavy

metal concentration)



230

5. OPTIMIZATION

Optimization is a procedure to select the best target

output from a process. This selection is made from a

series of experiments testing from a range of values

and the combination of each all factors. Optimization

may be a maximizing, minimizing and to keep the

target within range depending on the objective.

5.1. Concept of optimization

Concept of optimization is to select a combination of

the levels of factors which will result in a product with

desirable properties. This means solving problems

involving the optimization of several responses

(output), which depends upon a number of factors or

sets of parameters (input). Each desirable output

requires a special setting of input factors. For instance,

in an experimental study to maximize dye adsorption

capacity of an adsorbent made from coconut husk, the

two responses involved are decoloration and COD

removal of the dye wastewater. It can be achieved by

having higher temperature and retention time and at

the same time, reducing the rotation speed and

concentration of dye. In this optimization process, an

attempt was made to fulfill the requirement of the

wastewater treatment plant discharge (maximize the

dye adsorption capacity to reduce the discharge of dye

concentration in wastewater) while at the same time

reduces the material used (chemical dosage), obtain an

energy saving condition and reduces the duration of

the runs and most importantly to save cost. .

5.2. Desirability approach for Multiple Response

As we have mentioned before, optimization of a

process begins with selecting a proper design which

suit the process and then build a response surface

model for the response/s. The process of choosing a

set of operating conditions that is able to meet the

properties of each response is a challenge task;

however, it is possible in most cases to optimize all

responses by finding a combination of the levels of

input variables that meet the criteria of each response.

In such case, desirability approach is employed to

optimize the multiple responses process. This method

enables the process to obtain the most desirable

response/s by finding the suitable sets of operating

conditions, which means finding the most suitable

combination of the input variable (parameters). Given

that the process has a number of factors, all of these

selected factors must be in the desirable range.

A desirability function D(y) derives from the

process responses range from 0 to 1 and d(y) = 0

means the solution cannot be obtained and the

properties for one or more responses should be

changed while d(y) = 1 means the solution is possible

and is located in the desire region (Derringer and

Suich, 1980). This equation was modified by

Harrington, 1965 (Harrington, 1965). Desirability

function of each factor can be written in the form of

geometric mean as given in (3).

D = [(d1 × y1) × (d2 × y2) × (d3 × y3) × (dk × yk)]1/k

(3)

with k representing the number of responses.

Upon the need of the output, the response of the

yield, y can be either maximize, minimize or at a

selected target value. Assume that L, U and T

represent the lower, upper and target value,

respectively, for the desired response y. Desirability

function for optimizing a process to the target yield is

given in (4).

Uyif

UyTifUT

yU

TyLifLT

Ly

Lyif

yDn

m

0

≤≤][

≤≤][

0

)( (4)

Whereas, for a process response that requires a maximum yield, then the desirability function would be;

Tyif

TyLifLT

Ly

Lyif

yD m

1

≤≤][

0

)( (5)

Lim et al.


231

Lastly, for a process response that requires a minimum yield, then the desirability function would be:

Uyif

UyTifUT

Uy

Tyif

yD n

0

≤≤][

1

)( (6)

with notation m and n vital in determining the

relationship of the target value and yield, respectively.

If m=n=1, then the desirability function is linear

towards the target value. However, if m<1 and n<1,

the function would be convex and if m>1 and n>1,

the function would be concave.

6. DESIGNS USED FOR OPTIMIZATION

Selecting a suitable experimental design will greatly

help in describing the real behaviour of the process

over the region of interest. Some important properties

should be provided by the design such as rotatability

which means the predicted response is a function of

distance and not a function of direction and uniform

precision. As the location of optimum solution is

unknown before running the experiment, rotatability

provides equal precision of estimation in all directions.

Designs that have these properties are CCD and BBD

while three-level factorial design is not rotatable.

CCD and BBD design required a small number of

runs to fit a response surface model to the data

compared with full factorial designs such as three-

level factorial design and higher order. The model

obtained will be either first-order or second-order

polynomial model that fits the data. The model should

be checked for normality assumption, and model

adequacy.

6.1. Model

The most commonly response surface models used to

fit the data are first-order and second-order

polynomial models.

6.1.1. First order model

In most experimental condition, the relationship

between the response or the yield and independent

factors are unknown. In order to obtain a suitable

approximation for the relationship, a linear function or

lower order polynomial is engaged. This function is

known as a first-order model. The first order model

with k variables is given in (7).

i

k

i

io xy ∑1=

+= (7)

where, i and,0 are regression coefficients, and

ix are the coded variables.

Two-level factorial (2k) is usually used since the

number of runs is smaller than other designs and

allows the assessment of linear effects of the factors

and interactions.

6.2.2. Second order model

Second-order model is used if there is a curvature in

the first order model which means that the first-order

model is insufficient, then a higher degree of

polynomial should be used to describe the real

relationship between the response and selected

independent variables. A second-order model is given

in (8).

20

1 1

k k

i i ii i ij i ji i i j

y x x x x

(8)

where, ,,,0 iii and ij are regression

coefficients, and ix are the coded variables

(Montgomery, 2012).

It is important for the second-order model to

provide good prediction throughout the region of

interest. Hence, Box and Hunter (1957) suggested that

a second-order response surface design should be

rotatable. BBD and CCD are the most commonly used

designs to fit second-order model while three-level

factorial design is less use compared to other designs.

There is other second-order designs, which are not

frequently used, for instance, Hoke designs, Box–

Draper saturated designs, uniform shell and hybrid

designs (Khuri and Mukhopadhyay, 2010).

6.2.2.1. Central Composite Design

CCD is a very efficient design for fitting a second-

order model. CCD is a rotatable design that provides

equal precision for fitted response at points (factor

level combinations) that are at equal distances from

the center of the factor space. CCD usually has axial



232

or star points (α) outside the ‘cube’. The value of α

depends on the number of variables in the design.

Furthermore, the choice of α in the CCD is usually

based on the region of interest. CCD consists of three

components (i) a complete 2k factorial design with nF

runs, (ii) 2k axial runs and (iii) n0 center runs

(Montgomery, 2012). Therefore, the total number of

runs is equal to 2k +2k+n0. The response surface plots

can be employed to study the surfaces and locate the

optimum. An example for CCD with two factors in

coded form is given below:

x1 x2

-1 -1

Factorial

Points

1 -1

-1 1

1 1

-1.41421 0

Axial

Points

1.41421 0

0 -1.41421

0 1.41421

0 0

Centre

Points

0 0

0 0

0 0

0 0

It can be seen that there are three parts of CCD, (i)

factorial points nF = 4, (ii) axial (star) points = 4 and

(iii) four or five runs at the centre. In case of axial

point equals to 1 the CCD is called face centred

design (FCCD) (Montgomery, 2012).

Example 6.1

Suppose a chemist would like to extract vegetable

oil by using supercritical CO2 from a sample prepared

without giving any pretreatment. Three factors are

thought to be influential factors, namely, pressure,

temperature and time. The levels of each factor in

actual and coded forms are given in Table 4. FCCD

was used to carry out the experiment. The data are

given in Table 5.

Table 4: Experimental conditions in actual and coded form Factor Symbol Levels

Actual Coded

Pressure (MPa) x1 27.58 48.26 -1 1

Temperature (°C) x2 40 80 -1 1

Time (min) x3 45 70 -1 1

Lim et al.


233

Table 5: Face centred composite design in natural variables Pressure (MPa) Temperature (°C) Time

(min)

Percent yield

(%)

27.58 80 45 42.5

48.26 80 45 35.0

27.58 40 70 27.1

48.26 40 70 66.7

27.58 80 70 59.5

48.26 80 70 50.1

27.58 60 57.5 38.2

48.26 60 57.5 75.9

37.92 40 57.5 44.3

37.92 80 57.5 68.9

37.92 60 45 46.8

37.92 60 70 74.4

27.58 40 45 65.4

48.26 40 45 67.6

37.92 60 57.5 67.1

37.92 60 57.5 66.9

37.92 60 57.5 68.3

37.92 60 57.5 68.7

37.92 60 57.5 68.3

19 runs were required to cover all possible

combination of factor levels. Design Expert statistical

software (version 6) was used to analyze the data

obtained from this experiment. The results of

ANOVA are given in Table 6.

Table 6: ANOVA table for vegetable oil extraction

Source of variation Sum of squares Degree of Freedom

Mean

sum square F-value

P-value

Model 3911.79 9 434.64 13.25 <0.0003

Pressure 892.95 1 892.95 27.22 <0.0006

Temperature 50.90 1 50.90 1.55 <0.2444

Time 1200.73 1 1200.73 36.60 <0.0002

Pressure×Pressure 663.93 1 663.93 20.24 <0.0015

Temperature×Temperature 296.65 1 296.65 9.04 <0.0148

Time×Time 10.30 1 10.30 0.31 <0.5888

Pressure×Temperature 13.42 1 13.42 0.41 <0.5383

Pressure×Time 785.76 1 785.76 23.95 <0.0009

Temperature×Time 119.00 1 119.00 3.63 <0.0892

Residual 119.00 1 32.80

Total 4207.03 18

The results of the ANOVA for vegetable oil

extraction showed that the main effects of pressure

and time of extraction are statistically significant.

Furthermore, the quadratic contribution for pressure

and temperature over the main effect was significant,

whilst time did not show a significant effect. In

addition, the interaction between pressure and time

was significant at P < 0.10.

A model that best describes the yield result was

built in order to optimize the process by finding the

optimal settings of pressure, temperature, and time so

as to maximize the yield for each response. A second-

order model for the yield of oil extracted from the

sample in terms of coded variables was obtained as

given in Eq. 9. Three-dimensional response surface

plot is given in Fig. 6 to illustrate the maximum yield.



234

323121

2

3

2

2

2

1321

86.391.930.1

94.142.1059.1596.1026.245.983.65

xxxxxx

xxxxxxYield

(9)

Fig. 2: Three-dimensional response surface diagram of percent yield as a function of pressure and temperature

It was found that the maximum yield in percent

can be achieved at 41.68 MPa, 74.54°C and at

duration of 62.16 min may achieved 70.07% yield.

6.2.2.2. Box behnken design

The BBD is one of the most efficient designs used to

fit a second order model. It is a three-level design

based upon the combination of two-level factorial

designs and incomplete block designs. BBD is

rotatable or nearly rotatable and spherical design

(contains no embedded factorial or fractional factorial

design), in which the treatment combinations are at

the midpoints of the edges of the process space and at

the center. The advantage of the BBD is that it does

not contain combinations for which all factors are

simultaneously at their highest or lowest levels. Hence,

these designs are useful in avoiding experiments

performed under extreme conditions, in which

unsatisfactory results are often obtained. Furthermore,

BBD does not have axial points. Thus, all design

points fall within the safe operating zone. These

designs also ensure that all factors are never set at

their high levels, simultaneously (Ragonese, Macka et

al., 2002; Govender, 2005; Ferreira, 2007).

BBD is popular to be employed in industrial

research due to its economical design and requires

only three levels for each factor (-1, 0, 1). Typically,

BBD has the number of design arrangements of k is 3,

4, 5, 6, 7, 9, 10, 11, 12, and 16 factors (Khuri and

Mukhopadhyay, 2010).

An example for BBD with three factors will

consist of 3 sets of 22 factorial points, and 3 center

points or intermediate points. The total number of runs

is 15 as given below:

x1 x2 x3

-1 -1 0

Factorial

Points

1 -1 0

-1 1 0

1 1 0

-1 0 -1

1 0 -1

-1 0 1

1 0 1

0 -1 -1

Lim et al.


235

0 1 -1

0 -1 1

0 1 1

0 0 0

Center Point/

Intermediate 0 0 0

0 0 0

Example 6.2

A biopolymer clarifies turbid water added during

coagulation-flocculation process. The performance of

clarifying water is examined by using the turbidity

reduction as response. The experimental ranges for all

independent variables were based on the preliminary

trials as depicted in Table 7.

Table 7: Experimental range for turbid water treatment

Factor Symbol Levels

Actual Coded

pH x1 3 9 -1 1

Cation concentration (mM) x2 0.1 1.0 -1 1

Biopolymer dosage (mg/L) x3 10 50 -1 1

Seventeen runs were required to cover all possible

combination of factor levels. The experiments were

designed and analyzed using the Design Expert

statistical software (version 6). The experiment was

carried out randomly to minimize the effect of

unexpected variability in the observed responses. The

experiment runs are given in Table 8. The results of

the analysis are summarized in Table 9.

Table 8: Box-Behnken design in natural variables

pH Cation concentration

(mM)

Biopolymer dosage

(mg/L)

Turbidity reduction

(%)

3 0.10 30 55.5

9 0.10 30 91.1

3 1.00 30 99.7

9 1.00 30 98.8

3 0.55 10 97.0

9 0.55 10 98.4

3 0.55 50 68.7

9 0.55 50 99.7

6 0.10 10 99.0

6 1.00 10 99.8

6 0.10 50 76.2

6 1.00 50 99.7

6 0.55 30 99.6

6 0.55 30 99.7

6 0.55 30 99.0

6 0.55 30 99.3

6 0.55 30 99.4



236

Table 9: The result of analysis of variance (ANOVA) for turbidity reduction

Source of variation

Sum of

squares

Degree

of

Freedom

Mean

sum

square F-value

P-value

Model 2685.11 9 298.35 21.25 <0.0003

pH 562.80 1 562.80 40.08 <0.0004

Cation concentration 725.81 1 725.81 51.69 <0.0002

Biopolymer dosage 311.25 1 311.25 22.17 <0.0022

pH × pH 264.44 1 264.44 18.83 <0.0034

Cation concentration × Cation

concentration 113.85 1 113.85 8.11 <0.0248

Biopolymer dosage ×


pH × Cation concentration 333.06 1 333.06 23.72 <0.0018

pH× Biopolymer dosage 219.04 1 219.04 15.60 <0.0055

Cation concentration ×


Residual 98.28 7

Total 2783.39 16

As can be seen, the analysis of variance revealed

that a second-order model adequately fitted the

experimental data. As shown in Table 6.6, the linear

effects of pH, cation concentration and biopolymer

dosage were significant. The contribution of quadratic

effect over the linear effect for pH and cation

concentration was significant, whilst the quadratic

contribution of biopolymer dosage was insignificant.

On the other hand, the interaction between pH and

cation concentration, pH and cation concentration,

cation concentration and biopolymer dosage was

significant at P<0.10. A second-order models that

describe the behavior of turbidity reduction are built

in order to optimize the process by finding the best

settings of pH, cation concentration and biopolymer

dosage that maximize turbidity reduction. The second-

order model for turbidity reduction in terms of coded

variable are given in Eq. 10,

323121

2

3

2

2

2

1321

68.540.713.9-52.0

20.5-93.7-24.6-53.939.840.99

xxxxxxx

xxxxxreductionTurbidity

(10)

Lastly, to better present the optimization setting,

three dimensional response surface plot is given in Fig.

3 to have a clear picture on the maximum turbidity

reduction.

Fig. 3: Three-dimensional response surface for turbidity reduction as a function of pH and concentration of cation

Optimization process showed that maximum

turbidity reduction can be achieved at 99% in the

region of pH around 7.45, cation concentration at 0.67

mM and biopolymer dosage at 13.53 mg/L.

REFERENCES

Box GE and JS Hunter (1957). Multi-factor

experimental designs for exploring response

Lim et al.


237

surfaces. The Annals of Mathematical Statistics,

1: 195-241.

Derringer G and R Suich (1980). Simultaneous

Optimization of Several Response Variables.

Journal of Quality Technology 214-219.

Ferreira SLC, Bruns, R.E., Ferreira, H.S., Matos, G.D.,

David, J.M., Brandao, G.C., da Silva, E.G.P.,

Portugal, L.A., dos Reis, P.S., Souza, A.S., dos

Santos, W.N.L. (2007). Box-Behnken design:

an alternative for the optimization of analytical

methods. Analytica Chimica Acta 179-186.

Govender S, Pillay, V., Chetty, D.J., Essack, S.Y.,

Dangor, C.M., Govender, T. (2005).

Optimisation and characterisation of

bioadhesive controlled release tetracycline

microspheres. Int. J. Pharm., 24-40.

Harrington ECJ (1965). The Desirability Function.

Industrial Quality Control, 10: 494-498.

Khuri AI and S Mukhopadhyay (2010). Response

surface methodology. Wiley Interdisciplinary

Reviews: Computational Statistics, 2: 128-149.

Montgomery DC (2012). Design and Analysis of

Experiments, Wiley.

Prakobvaitayakit M and U Nimmannit (2003).

Optimization of Polylactic-Co-Glycolic Acid

Nanoparticles Containing Itraconazole Using 23

Factorial Design. AAPS Pharm. Sci. Tech., 4:

Ragonese R, M Macka, et al. (2002). The use of the

Box–Behnken experimental design in the

optimisation and robustness testing of a

capillary electrophoresis method for the

analysis of ethambutol hydrochloride in a

pharmaceutical formulation. Journal of

pharmaceutical and biomedical analysis, 6: 995-

1007.

Ryan TP (2007). Modern Experimental Design, John

Wiley & Son, Inc, New Jersey.

Srinivasan A and T Viraraghavan (2010). Oil removal

from water by fungal biomass: A factorial

design analysis. Journal of Hazardous Materials,

695-702.

Wastewater Engineering: Advanced Wastewater Treatment Systems

Documents

Transcript of Wastewater Engineering: Advanced Wastewater Treatment Systems