Fluid Waste Disposal
Fluid Waste Disposal
Fluid Waste Disposal
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ENVIRONMENTAL SCIENCE, ENGINEERING
AND TECHNOLOGY SERIES
KAY W. CANTON
EDITOR
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Preface ix
Chapter 1 Treatment of Wastewater by Electrocoagulation Method
and the Effect of Low Cost Supporting Electrolytes 1
Lazare Etiégni, K. Senelwa, B. K. Balozi, K. Ofosu-Asiedu,
A. Yitambé, D. O. Oricho and B. O. Orori
Chapter 2 Application of Sulphate-Reducing Bacteria
in Biological Treatment Wastewaters 49
Dorota Wolicka
Chapter 3 Utilization of Water and Wastewater Sludge
for Production of Lightweight-Stabilized Ceramsite 83
Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
Chapter 4 Modelling and Observation of Produced
Formation Water (PFW) at Sea 113
D. Cianelli, L. Manfra, E. Zambianchi,
C. Maggi and A. M. Cicero
Chapter 5 Disposal of Sulfur Dioxide Generated in Industries
Using Eco-Friendly Biotechnological Process – A Review 137
A. Gangagni Rao and P.N. Sarma
Chapter 6 Novel Biological Nitrogen-Removal Processes:
Applications and Perspectives 155
J.L. Campos, J.R. Vázquez-Padín, M. Figueroa,
C. Fajardo, A. Mosquera-Corral and R. Méndez
Chapter 7 Application of Microbial Melanoidin-Decomposing
Activity (MDA) for Treatment of Molasses Wastewater 181
Suntud Sirianuntapiboon and Sadahiro Ohmomo
Chapter 8 Wastewaters from Olive Oil Industry:
Characterization and Treatment 199
L. Nieto Martínez, Gassan Hodaifa,M. Eugenia Martínez
and Sebastián Sánchez
viii Contents
Wastewater is any water that has been adversely affected in quality by anthropogenic
influence. It comprises liquid waste discharged by domestic residences, commercial
properties, industry, and/or agriculture and can encompass a wide range of potential
contaminants and concentrations. In the most common usage, it refers to the municipal
wastewater that contains a broad spectrum of contaminants resulting from the mixing of
wastewaters from different sources. With the dwindling available water resources in the
world coupled with high population growth, pressure is being exerted on water and
wastewater plant managers the world over to find cost-effective methods to treat a wide range
of wastewater pollutants in a diverse range of situations. This new and important book gathers
the latest research from around the globe on fluid waste disposal with a focus on such topics
as: wastewaters from the olive industry, application of sulphate-reducing bacteria in
biological treatment wastewaters, electrocoagulation treatment method, usability of boron
doped diamond electrodes in wastewater treatment and others.
Chapter 1 - Coagulation and flocculation are traditional methods of treating of polluted
water. Electrocoagulation (EC) presents a robust novel and innovative alternative in which a
sacrificial metal anode doses water electrochemically. This has the major advantage of
providing active cations required for coagulation, without necessarily increasing the salinity
of the water. Electrocoagulation is a complex process with a multitude of mechanisms
operating synergistically to remove pollutants from water. A wide variety of opinions exist in
the literature for key mechanisms and reactor configurations. A lack of a systematic approach
has resulted in a myriad of designs for electrocoagulation reactors without due consideration
of the complexity of the system. A systematic, holistic approach is required to understand
electrocoagulation and its controlling parameters (pH, temperature, conductivity, current
density). This will enable a priori prediction of the treatment of various pollutant types.
Electrocoagulation involves applying a current across electrodes in water. This results in the
dissolution of the anode (either aluminum or iron). These ions then form hydroxides which
complex with and/or absorb contaminants and precipitate out. The precipitate with the
contaminants can then be removed from the water by settling and decantation or filtration. EC
has the potential to be applied in many other areas besides the textile and semiconductor
industry. It has been successfully tested in the pulp and paper industry, as well as tea and
coffee processing. However over electrical potential within electrodes during
electrocoagulation normally causes extra voltage, which wastes energy. There have been
attempts to reduce this extra voltage which, in these days of World energy crisis, will render
x Kay W. Canton
water are brought to the surface along with the hydrocarbons. These waters include the
‗formation water‘, that lies underneath the hydrocarbon layer, and ‗additional water‘ usually
injected into the reservoirs to help force the oil to the surface. Both formation and injected
waters, named ―produced formation waters‖ (PFWs), are separated from the hydrocarbons
onboard offshore platforms and then disposed into the marine environment through ocean
diffusers. PFWs contain several contaminants and represent one of the main sources of marine
environment pollution associated with oil and gas production.
This makes the study of PFW fate of paramount importance for a proper management of
environmental resources as well as for planning and optimizing the discharge and monitoring
procedures.
In the first part of this chapter we provide a detailed description of the chemical
characteristics of PFWs and their potential toxic effects and review the mixing processes
governing their dispersion in the marine environment. In the second part of the work we
briefly review past efforts in observing and modelling PFW spreading in the ocean. Finally,
we propose a multidisciplinary approach, integrating in situ observations and numerical
modelling, to assess dispersion of PFWs in space and time. As a case study we will refer to
the results of a previous study conducted in the Northern Adriatic Sea, a sub-basin of the
Mediterranean Sea, where a number of offshore natural gas (CH4) extraction platforms are
currently active.
Chapter 5 - Sulfur dioxide (SO2) is a known pollutant and responsible for various ill
effects on living and non-living organisms. SO2 emissions can be reduced by using non-
conventional energy sources or using conventional fuels containing less sulfur. However,
under the present circumstances SO2 emissions cannot be completely avoided due to the
reasons of rapid industrialization. Various technologies are available for the removal of SO2
from flue and waste gases. Most of these technologies fall under the category of physical,
chemical or thermal. All these technologies generate secondary pollutants ending up in
disposal problems and also cost prohibitive. Biotechnology offers relatively cheaper solutions
for the conventional problems. Due to this reason, biotechnology is making in roads into the
conventional treatment processes in all the fields. Over the last decade, efforts have been
made to develop biotechnological alternatives to conventional physico- chemical processes
for the removal of SO2 from flue gases known as Biological flue gas desulphurization (BIO-
FGD).SO2 from flue gas can be absorbed in a suitable organic media. In the aqueous phase
SO2 would be converted to sulfite and some part may again be converted to sulfate due to the
presence of dissolved oxygen. Therefore, the aqueous phase will be having both sulfate and
sulfite, which can be reduced to sulfide using Sulfate Reducing Bacteria (SRB) under
anaerobic conditions. The sulfide formed in the anaerobic reactor could be converted to
elemental sulfur using Sulfur Oxidizing Bacteria (SOB) under partial microbial aerobic
conditions. The elemental sulfur can be used either as a soil conditioner or raw material for
industrial applications. Therefore, BIO-FGD process could be an environmentally benign and
economically viable alternative for the disposal of SO2 emitted from the industries especially
from power plants and refineries. The present article reviews the state of art of BIO-FGD
process.
Chapter 6 - Since the requirement for nutrient removal is becoming increasingly
stringent, a high efficiency of nitrogen removal is necessary to achieve a low total nitrogen
concentration in the effluent. Biological nitrification and denitrification processes are
generally employed to remove nitrogen from wastewater. Unfortunately, these processes are
xii Kay W. Canton
not suitable to treat wastewater with a low COD/N ratio because it involves the addition of an
external organic carbon source and, therefore, an increase of the operational costs.
Several alternative processes for nitrogen removal can be applied in order to reduce
partially (―nitrite route‖) or totally (anammox, autotrophic denitrification) the organic matter
required. Such processes suppose not only an economical way to treat these wastewaters but
they are also more environmentally friendly technologies (lower production of CO2, N2O and
sludge; lower energy consumption). Up to now, they were basically applied to the return
sludge line of municipal wastewater treatment plants (WWTPs). However, these processes
could even be implemented in the actual WWTPs in order to achieve more compact and
energy efficient systems.
Their potential advantages can make them also feasible technologies to treat polluted
ground water or to remove nitrogen compounds from recirculating aquaculture systems.
Chapter 7 - This review will discuss the melanoidin-decomposing activity (MDA) among
microorganisms. The focus will be on the potential use of the microbial-MDA to treat the
wastewater discharged from factories using molasses as the raw material (molasses
wastewater: MWW) because molasses is one of the most useful raw materials in various types
of industries, such as the fermentation and animal feed industries. However, the wastewater
discharged from factories using molasses contains a large amount of dark brown pigment,
melanoidin pigment: MP, which is poorly decomposed and/or decolorized by normal
biological treatment processes, such as the activated sludge or anaerobic treatment systems
(anaerobic pond or anaerobic contact digester), because, the microorganisms in those
wastewater treatment systems showed very poor MDA. The distribution of MDA among
microorganisms and the mechanism of decomposing activities, in particular, were reviewed.
Also, the application of the isolated strains having the MDA to treat molasses wastewater in
the wastewater treatment plant was tested.
Chapter 8 - Countries in the Mediterranean basin are among the main producers of olive
oil. The elaboration of olive-oil is typically carried out by small companies in small facilities.
The olive-oil plants produce high and variable amounts of residual waters of olives and olive-
oil washing (OMW) that has a great impact in the environment. According to the procedure
used different types of OMW with different chemical oxygen demand can be generated, the
OMW from the three phase process (COD = 150 g O2 L-1) and the OMW from olives washing
(COD = 0.8-4.5 g O2 L-1) and olive oil washing (COD = 1.1- 6 g O2 L-1) in the two-phase
process. The uncontrolled disposal of OMW is a serious environmental problem, due to its
high organic load, and because of its high content of microbial growth-inhibiting compounds,
such as phenolic compounds. The improper disposal of OMW to the environment or to
domestic wastewater treatment plants is prohibited due to its toxicity to microorganisms, and
also because of its potential threat to surface and groundwater. These waters normally are
stored in great rafts of accumulation for their evaporation during the summer. This solution
among others until the moment dose not represent a definitive solution for this problem,
especially as the administrations more and more demanding the preparation of this spill and
the constructive quality of the rafts. Today, effective technologies have been proposed such as
the chemical oxidation process using ferric chloride catalyst for the activation of H2O2 as a
treatment of OMW produced from two-phase process. In the previous works the authors have
described the experimental results on laboratory-scale. These results have been taken to pilot-
industrial scale, making the chemical oxidation in the optimum conditions of operations:
[H2O2] = 5% (w/v), using a ferric chloride catalyst with a relation of [FeCl3]/[H2O2] = 0.25
Preface xiii
(w/w), at OMW pH and environmental temperature. The final average value of COD obtained
next to 370 mg L-1 (%CODremoval = 86.2%), and the water obtained can be destined to
irrigation or disposed directly to the municipal wastewater system for their tertiary treatment.
OMW from three-phase process does not allow direct chemical and biological purification for
its content in phenolic compounds and generally used natural and forced evaporation process.
Another way of using is the application of OMW nutrients to the growth of microorganisms
such as microalgae.
Chapter 9 - Over the past few years one main focus on the research efforts at the Institute
for Sustainable Waste Management and Technology (IAE) has been on possible applications
for reactors with boron doped diamond electrodes (BDD) in the field of (waste) water
treatment. This article deals with the technical construction of the electrodes used (continuous
reactor with a different number of plate electrodes), which were produced by a spin-off of the
institute. The electrodes consist of conductible industrial diamond particles (< 250 µm),
which are mechanically implanted on a fluoride plastic substrate. These electrodes showed a
high mechanical and chemical stability in different test runs. At the institute, treatment
methods for micro pollutants (e.g. pharmaceuticals and complexing agents) were developed
with electrochemical oxidation by BDD. In this case test runs were made on laboratory scale
and technical scale treatment units and elimination rates up to 99 % were achieved. In this
project the analytic is partly provided by the ―Umweltbundesamt GmbH‖ (UBA), one of the
project partners. This agency has been a project partner in different studies about
pharmaceuticals in the ecosystem. These techniques could also be used for the waste water
treatment of alpine cabins. Pilot projects have been set up. On the basis of these results a
follow-up project was launched last October, in which an alternative treatment process for oil-
in-water emulsions and mixtures was developed by the usage of electrochemical oxidation
with BDD. A third possible application is the disinfection of drinking water from
contaminated ground and spring water. In this process oxidation agents like ozone or OH
radicals produced in situ by the BDD reactor from the treated water are used to eliminate
bacterial contaminants (for example e. coli) in the water.
Chapter 10 - Treatment of wastewater, commonly performed at municipal sewage plants,
generates sanitized water and sewage sludge. Anaerobic degradation of sewage sludge results
in the production of different gases, including the economically valuable methane, and
digested residue (biosolids) with potential value as a crop fertilizer. Traditionally, digested
sewage sludge is disposed either into water, onto or into the earth or into the air. However,
alternative exploitation of digested sewage sludge in agriculture has several advantages over
commercial fertilizers, including environmental aspects benefiting agricultural sustainability
and increased crop yield. Additionally, residue utilization is nearly always a cheaper option
than disposal costs.
Biosolids obtained from the treatment of municipal sewage sludge consist of a mixture of
organic and mineral compounds that significantly affect soil microbial communities and their
biogeochemical activities when applied as a crop fertilizer. The microorganisms influence soil
quality through nutrient cycling, decomposition of organic matter and maintenance of soil
structure, in turn, affecting agricultural and environmental quality, and subsequently, plant
and animal health. Moreover, both soil and residue normally contain considerable quantities
of microorganisms, including both beneficial and potentially human pathogenic species that
may be supported by the new conditions in the soil. Thus, soil amended with biosolids may
xiv Kay W. Canton
present a modified microbial community composition after some time and, hence, a modified
ecosystem function.
At the end of the present chapter, we discuss whether the potential risks of recycling
biosolids to agricultural cropland are acceptable for consumers, producers and scientific
expertise, in view of the resulting alterations in soil microbial diversity, activity and
accompanying functions. Furthermore, optimal ways of managing the recycling process to
achieve the most favourable balance of benefits and risks for the community are highlighted.
Chapter 11 - The purpose of the present study was to design an integrated wastewater
treatment system for a nalla (riverlet) flowing through Indian Institute of Technology Delhi
(IITD), India, besides its cost estimation and comparison with the conventional wastewater
treatment system. The design parameters for integrated aeration-cum-adsorption tank were
worked out for 240 m3 / d flow rate of the wastewater. The important parameters used for the
design included initial COD and BOD concentration in the influent, treatment time, adsorbent
dose, pH, adsorbent particle size and the desired COD and BOD in the effluent after treatment
as prescribed by Central Pollution Control Board, (CPCB) Delhi, India. All the design
parameters of this system were similar to those of conventional system except for the
replacement of aeration tank in conventional system by the aeration-cum-adsorption tank. The
concentration of COD and BOD of the treated effluent by the integrated system were well
within the permissible limits of CPCB standards (for COD it is 100 ppm and for BOD of 30
ppm) to discharge in the canal for irrigation purpose. It was worth mentioning here that the
adsorbents used in the present study were based on discarded materials which were available
free of cost. Of course, the cost of their transportation and processing should have been taken
into account.
The total cost estimated for the conventional system and the adsorption based system
would be Rs. 198,312 and Rs. 141,275 respectively (including civil work, machinery, labour,
adsorbent and miscellaneous). The cost difference for the two systems would be
approximately Rs 57,037.
This design of integrated system has resulted into saving of cost by 28 % over the
conventional system. Thus, it is a good approach for saving of conventional energy in
addition to saving the cost of treatment and can be applicable for any country for decen-
tralized sector. Moreover, it is an open ended research and we can recommend more research
by changing the adsorbents types and operating parameters to improve the model.
Chapter 12 - The objective of this chapter is to put forward an overview of
biodegradation characteristics of wastewaters by emphasizing the significance of COD
fractionation. Recalcitrant COD fractions of effluents can be used as a tool to evaluate
whether discharge standards can be met with a prescribed biological treatment. Moreover, the
appropriate type of biological treatment applicable to the wastewater under investigation can
be addressed and the performance of an existing biological treatment system can be appraised
with reference to inert COD fractions. Besides recalcitrant COD fractions of segregated
industrial effluent streams can be regarded as an essential input of a sound industrial
wastewater management strategy adopting minimization at source philosophy. Last but not
least, data on COD fractions can be used as a solid source of information for modelling
studies that define the design and performance of biological treatment systems. In this
context, COD fractionation data on a wide spectrum of activities ranging from various
industrial sectors to hotels is presented. Segregated industrial wastewater streams together
Preface xv
with domestic sewage and end-of-pipe industrial effluents are evaluated in terms of their
biodegradation characteristics.
Chapter 13 - In the present study, two types of colour removal systems were tested on
effluent samples collected from a coffee pulping factory which discharged on average 15 m3
of wastewater daily with a colour index of about 2500 OH that was too high for direct
discharge into a river in Kenya. The two colour removal systems used were: (i) electrolysis
combined with wood ash or coffee husks leachate and (ii) electrolysis combined with
phosphate rock solutions at a rate of 0.5 g/l to 4g/l. Phosphate rock is often used as
agricultural liming agent. The surface area of the electrodes was set at close to 75 m2/m3 of
effluent with a current density of 1,200 mA/m2. The experiments were laid out in a stratified
random sampling design and the data were analysed using the Statistical Package for Social
Scientists (SPSS) computer programme version 10.0. Electrolysis combined with phosphate
rock (ELPHOS) proved to be the best process in terms of power consumption (68%
reduction) compared with the 57% reduction by electrolysis combined with wood ash
(ELCAS) and the 58% reduction by electrolysis combined with coffee husks ash (ELCHAS).
Besides the 100% colour removal, ELPHOS also reduced other effluent physico-chemical
parameters such as BOD, COD, TSS and TS by 79%, 80%, 69%, and 88% respectively. The
analysis of ELPHOS treated wastewater showed that the mill could discharge an effluent that
meets local discharge standards for colour requirements. It is recommended that recycling of
the treated water by ELPHOS back to the factory for cleaning and washing purposes be
considered since the quality meets the requirement for uses of fresh water for cleaning
purposes. Furthermore, calculation of power consumption based on a scale-up batch reactor
of 15 m3 proved less expensive to treat the factory effluent than a set of 12 one 100-L reactors
similar to the one used in the field.
Chapter 14 - The major aim of this paper is to review the major problems of water
resources in the developing countries. It is based on problems related to population growth
and pollution and how these are more likely to lead to future conflicts. We know that fresh
water is only 3 % of the total global water and 78% of this is in glaciers. This makes it a
scarce and precious resource which must be sustainably managed. The paper also analyses
some of the already existing and potential conflicts based on water resources. It reviews the
potential threats to Ugandan water resources and problems which are most likely to occur as a
result of these threats. Factors hindering treatment of wastewater as a remedy to pollution in
developing countries have also been discussed. The methodology used in this paper is based
on literature review of the most current issues that affect water resources world-wide. The
review is limited to scientific facts and no political factors affecting water resources have
been included.
It has been found that although Uganda is endowed with 66km2/year of renewable water
resources, population increase, deforestation, degradation of wetlands and pollution are major
threats to its water resources. Problems associated with water quality and quantities are more
likely to result into internal conflicts which are bound to spread beyond Ugandan borders.
Chapter 15 - Water is a vital aspect of hemodialysis. During the procedure, large volumes
of water are used to prepare dialysate and to clean and reprocess machines. This paper
evaluates the technical and economical feasibility of recycling hemodialysis wastewater for
irrigation uses, such as watering gardens and landscape plantings. Water characteristics,
possible recycling methods, and the production costs of treated water are discussed in terms
of the quality of the generated wastewater. A cost-benefit analysis is also performed through
xvi Kay W. Canton
comparison of intended cost with that of seawater desalination, which is widely used in
irrigation.
Chapter 16 - Heavy metal pollution is a serious problem in many developed and
developing countries. Lead had been recognized as a particularly toxic metal and comes into
water bodies mainly from metallurgical, battery, metal plating, mining and alloy industries. In
order to minimize the impacts of this metal on human health, animals and the environment,
lead-contaminated water and wastewater need to be treated before discharge to water bodies.
This chapter concerns an investigation of potential usage of corn-processing wastewater
as a new alternative low-cost substrate to produce biosorbent and evaluate this biosorbent to
remove Pb(II) ions from aqueous solutions. For this aim, Rhizopus oligosporus cultivated on
corn-processing wastewater and dried biomass of these fungi was used as an adsorbent. The
adsorption experiments were conducted in a batch process and the effects of contact time (1-
48 hours), initial pH (2-7), initial metal ion concentration (20-100 mg L-1) and adsorbent
dosage (0.5-5 g L-1) on the adsorption were investigated. Pb (II) ion concentrations before and
after adsorption were measured using Inductively Coupled Plasma-Mass Spectrometry.
Maximum adsorption capacity was achieved at pH 5.0. The isothermal data of dried fungal
biomass could be described well by the Langmuir equation and monolayer capacity had a
mean value of 59.88 mg g-1. The pseudo-second order reaction model provided the best
description of the data with a correlation coefficient 0.99 for different initial metal
concentrations. This result indicates that chemical sorption might be the basic mechanism for
this adsorption process and Fourier Transform Infrared Spectroscopy analyses showed that
amide I and hydroxyl groups play an important role in binding Pb (II).
Because of the high activation capacity of adsorbent and low cost of process dried R.
oligosporus biomass presents a good potential as an alternative material for removal of Pb (II)
ions from the aqueous solutions.
Chapter 17 - The main objective of this research work is to determine the presence of
di(2-ethylhexyl) phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate, in different
water samples (drinking waters, effluents and surface waters). Different analytical methods
were studied in order to know the best methodology for the quantification of these
compounds. Solid-liquid and liquid-liquid extraction were investigated and finally the liquid-
liquid extraction and analysis by gas chromatography followed by mass spectroscopy was
chosen because of offering the highest recovery rate. In the whole of this research study, the
control of background pollution by reagents and material was extremely important. The
problem of background pollution is more serious in the trace analysis of phthalates and
adipates as a consequence of their presence in almost all equipment and reagents used in the
laboratory.
Respect to the control of the selected plasticizers in the different water samples, bis (2-
ethylhexyl) phthalate and bis (2-ethylhexyl) adipate were detected in drinking water, effluents
and surface waters. On the other hand, diisodecyl phthalate was not detected in any sample.
In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 1-48 © 2010 Nova Science Publishers, Inc.
Chapter 1
TREATMENT OF WASTEWATER BY
ELECTROCOAGULATION METHOD AND THE EFFECT
OF LOW COST SUPPORTING ELECTROLYTES
ABSTRACT
Coagulation and flocculation are traditional methods of treating of polluted water.
Electrocoagulation (EC) presents a robust novel and innovative alternative in which a
sacrificial metal anode doses water electrochemically. This has the major advantage of
providing active cations required for coagulation, without necessarily increasing the
salinity of the water. Electrocoagulation is a complex process with a multitude of
mechanisms operating synergistically to remove pollutants from water. A wide variety of
opinions exist in the literature for key mechanisms and reactor configurations. A lack of a
systematic approach has resulted in a myriad of designs for electrocoagulation reactors
without due consideration of the complexity of the system. A systematic, holistic
approach is required to understand electrocoagulation and its controlling parameters (pH,
temperature, conductivity, current density). This will enable a priori prediction of the
treatment of various pollutant types. Electrocoagulation involves applying a current
across electrodes in water. This results in the dissolution of the anode (either aluminum
or iron). These ions then form hydroxides which complex with and/or absorb
contaminants and precipitate out. The precipitate with the contaminants can then be
*
Corresponding author: E-mail: lazetiegni@amatala.org
2 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
removed from the water by settling and decantation or filtration. EC has the potential to
be applied in many other areas besides the textile and semiconductor industry. It has been
successfully tested in the pulp and paper industry, as well as tea and coffee processing.
However over electrical potential within electrodes during electrocoagulation normally
causes extra voltage, which wastes energy. There have been attempts to reduce this extra
voltage which, in these days of World energy crisis, will render the electrocoagulation
process uneconomical. The inclusion of supporting electrolyte such as NaCl achieves
this. One of the methods pioneered by researchers at Moi University in Kenya is the use
of wood ash leachate as supporting electrolyte which in some cases could reduce energy
consumption by as much as 80%. Other supporting electrolytes tested are ash from
bagasse and from coffee husks. These supporting electrolytes are relatively inexpensive,
but they all generally result in large amount of coagulated sludge. Other supporting
electrolytes such phosphate rock are less effective than wood ash, but they yield almost
50% less sludge after electrocoagulation. Most of the supporting electrolytes have an
added advantage of reducing other wastewater pollution parameters such as BOD, COD,
TSS, TS, turbidity, pH and color. Because of the inherent benefits of these low cost
supporting electrolytes, electro-chemical methods could be a credible alternative to more
traditional wastewater treatment approaches.
INTRODUCTION
With the dwindling availability of water resources in the World coupled with high
population growth, pressure is being exerted on water and wastewater plant managers the
world over to find cost-effective methods to treat a wide range of wastewater pollutants in a
diverse range of situations. Traditionally coagulation, flocculation and lagooning have been
used as chemical and biological processes with varying degrees of success to treat polluted
waters. However a more cost-effective and proven method to clean an ever widening range of
water pollutants, on-site, and with minimum additives, is required for sustainable water and
wastewater management. Electrocoagulation treatment of water seems to fit this description.
Colloidal dispersions in water or wastewater often referred to as sols consist of discrete
particles held in suspension by their extreme small size (1-200 nm), state of hydration
(chemical combination with water), and surface electric charge. The chemistry of coagulation
and flocculation is primarily based on the electrical properties of the particles. Like charges
repel each other while opposite charges attract. Particles finer than 0.1 µm (1x10-7 m) in water
or wastewater remain continuously in motion due to electrostatic charges (often negative)
which cause them to repel each other.
There are two types of colloids - hydrophilic and hydrophobic. Hydrophilic are readily
dispersed in water and their stability depends on the affinity for water rather than the slight
negative charge they possess. Hydrophobic colloids on the other hand have no affinity for
water and their stability depends on the charge they possess, usually positive. The
electrostatic repulsion between the colloidal particles leads to a stable sol. The surface or
primary charge of colloidal particles comes from charged groups within the particles or the
adsorption of charged particles. The sign and magnitude of the surface charge depends on the
character of colloids, the pH (the lower the pH the more positive the charge becomes), the
ionic strength and the characteristics of the water or wastewater.
The surface of the colloid has a certain δ-potential (zeta potential) which is the magnitude
of the charge at the surface of shear. The δ-potential is derived from the diffused double-layer
Treatment of Wastewater by Electrocoagulation Method… 3
theory applied to hydrophobic colloids (Figure 1), and can be estimated using
Smoluchowski‘s (1872-1917) electrokinetic mobility equation:
δε
μ= --------------------------------- (1) (1)
ε
Zeta potential can also be calculated using the following relationship for electrostatic
force
4πqd
δ= --------------------------- (2) (2)
D
q = charge per unit area
d = thickness of the layer surrounding the shear surface through which the charge is
effective
π = pi (= 3.142857)
D = dielectric constant of the liquid
+
+
+
+
+ + +
+
+
Stern layer +
+
+ + +
+
+ Surface shear
+
+
+ + +
+ Bulk of
Particle
+ solution
+ + n
+ +
+
+
+ +
+ + +
+ Electric
+ +
+ potential
surrounding
+
+ particle
+ Zeta
potential Diffusion
+ layer of
counterions
Fixed layer
of ions
The diffused double layer (Figure 1) consists of two parts: an inner region, also
referred to as Stern layer, which includes ions bound relatively strongly to the surface
(including specifically adsorbed ions) and an outer, diffuse or movable region in which
the ion distribution is determined by a balance of electrostatic forces and random
thermal motion. The potential in this region, therefore, decays as the distance from the
surface increases until, at sufficient distance, it reaches the bulk solution value,
conventionally taken to be zero. The repulsive force of the charged double layer
scatters particles thus preventing agglomeration. Particles with high zeta potential have
a very stable sol.
The zeta potential is the overall charge a particle acquires in a specific medium. In
other words, it is a measure of the magnitude of electrical charge surrounding the
colloidal particles. The magnitude of the zeta potential gives an indication of the
potential stability of the colloidal system. Zeta potential can be equated to the amount
of repulsive force which keeps the particles in suspension. If the zeta potential is large,
then more coagulants will be needed to destabilize colloidal particles. If all the
particles have a large negative or positive zeta potential they will repel each other and
there is dispersion stability. When particles have low zeta potential values, there is no
force to prevent the particles coming together and there is dispersion instability. A
dividing line between stable and unsable aqueous dispersions is generally taken at
either +30 or -30mV.
1. pH
In aqueous media, the pH of a sample is one of the most important factors that affect its
zeta potential. A zeta potential value on its own without defining the solution conditions is
virtually meaningless. A zeta potential versus pH curve will be higher or positive at low pH
and lower or negative at high pH. There may be a point where the plot passes through zero
zeta potential. This point is called the isoelectric point and is very important from a practical
consideration. It is normally the point where the colloidal system is least stable.
2. Conductivity
The thickness of the double layer (κ-1) depends upon the concentration of ions in
solution and can be calculated from the ionic strength of the medium. The higher the ionic
strength, the more compressed the double layer becomes. The valence of the ions will also
influence double layer thickness. A trivalent ion such as Al3+ will compress the double layer
to a greater extent in comparison to a monovalent ion such as Na+. Inorganic ions can interact
with charged surfaces in one of two distinct ways (i) non-specific ion adsorption where they
have no effect on the isoelectric point and (ii) specific ion adsorption, which will lead to a
Treatment of Wastewater by Electrocoagulation Method… 5
change in the value of the isoelectric point. The specific adsorption of ions onto a particle
surface, even at low concentrations, can have a dramatic effect on the zeta potential of the
particle dispersion. In some cases, specific ion adsorption can lead to charge reversal of the
surface.
The effect of the concentration of a formulation component on the zeta potential can give
information to assist in formulating a product to give maximum stability. The influence of
known contaminants on the zeta potential of a sample can be a powerful tool in formulating
the product to resist flocculation for example.
COAGULATION
Schulze, in 1882, first showed that colloidal systems could be destabilized by the addition
of ions having a charge opposite to that of the colloid (Benefield et al., 1982). Coagulation in
water or wastewater chemistry is a process in which a chemical referred to as a coagulant is
added to destabilize dispersed colloidal particles so that they agglomerate. Coagulation
experiments using natural products such as Moringa oleifera have also been tried with
varying degrees of success (Kasser et al., 1990; Ogutveren et al., 1994; Ndabigengesere et al.,
1995; Mohammed, 2001; Bhuptawat and Chaudhari, 2003). The objectives of coagulation are
to (i) destabilize suspended and colloidal particles to enhance their removal through
sedimentation and filtration and (ii) to precipitate dissolved maters i.e. PO43-, color, natural
organic matter (NOM). Coagulation process may require several reaction steps: (i) hydrolysis
of multivalent metal ions; (ii) adsorption of hydrolysis species at the solid-solution interface
for the destabilization of colloidal particle (reduction of zeta potential); (iii) aggregation of
destabilized particles by interparticle bridging; (iv) aggregation of destabilized particles by
particle transport and van der Waals‘ forces; (v) ―aging‖ of flocs formed in the process; and
(vi) precipitation of metal hydroxides (Stumm and O‘Melia, 1968).
ELECTROCOAGULATION
Electrocoagulation is a process that applies a current across electrodes through a liquid,
using a variety of anode and cathode geometries, including plates, balls fluidized bed spheres,
wire mesh, plates (either aluminum or iron), rods, and tubes. This results in the dissolution of
the anode (Equation 3 & 12). These ions then form hydroxides which complex with and/or
absorb contaminants and precipitate from water or wastewater. They are subsequently
removed by surface complexation and electrostatic attraction according to the following
equations:
6 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
In alkaline medium,
In alkaline medium
The cation hydrolyses in water to form a hydroxide. The following equations (20 to 23)
are an illustration of this phenomenon in the case of aluminum:
ELECTROCOAGULATION MECHANISMS
The electrocoagulation overall mechanism is a combination of mechanisms that operate
concurrently or in series but synergistically. The main mechanism may vary throughout the
dynamic process as the reaction progresses, and will almost certainly shift with changes in
operating and environmental parameters and pollutant types. Highly charged cations
destabilize any colloidal particles by the formation of polyvalent polyhydroxide complexes.
These complexes have high adsorption properties, forming aggregates with pollutants. The
pollutants presumably act as a ligand to bind with iron or aluminum ions resulting in the
formation of amorphous polymeric complexes (hydroxo-complexes). These compounds with
a large specific surface area are very active and able to coagulate and adsorb pollutants soon
after their in situ generation (Rajeshwar and Ibanez 1997; Scott, 2001). Besides the
generation of polyvalent cations described above, electrocoagulation includes also the
production of electrolysis gases that are hydrogen and oxygen (Equation 5, 6, 9, 10, 14, 15 &
19).
Evolution of hydrogen gas aids in mixing and flocculation. Once the floc is generated, the
electrolytic gas binds to and creates a buoyant force on the floc leading to its flotation and
ultimately to the removal of the pollutant as a floc - foam layer at the liquid surface
(Equation 11). Other flocs that are heavier settle at the bottom of the reactor.
There are many ways in which species can interact in solution:
Electrocoagulation process has been around for some time. The process was proposed
before the turn of the last century with Vik et al. (1984) describing a treatment plant in
London built in 1889 (for the treatment of sewage by mixing with seawater and
electrolyzing). In 1909, Harries (cited in Vik et al., 1984) in the United States, received a
patent for wastewater treatment by electrolysis with sacrificial aluminum and iron anodes.
Matteson and Dobson (1995) described a device of the 1940‘s, the ―Electronic Coagulator‖
8 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
which electrochemically dissolved aluminum (from the anode) into solution, reacting this
with the hydroxyl ion (from the cathode) to form aluminum hydroxide. The hydroxide
flocculates and coagulates the suspended solids purifying the water. A similar process was
used in Britain in 1956 for which iron electrodes were used to treat river water (Matteson and
Dobson, 1995).
Because of its capability to remove several types of water pollutants, the recent thirty
years have seen an explosion of journal article reports on electrocoagulation methods
probably due to new and more stringent environmental regulations on a wide range of water
and wastewater pollutants. This has further translated into a number of electrocoagulation
devices, designed to purify water or wastewater, being put on the market.
Emulsion breaking, resulting from the oxygen and hydrogen ions that bond into the
water receptor sites creating a water insoluble complex that separate water from
pollutants.
Seeding, resulting from the anode reduction of metal ions that become new centers
for larger, stable, insoluble complexes that precipitate as complex metal ions.
Bleaching by the oxygen ions produced in the reaction chamber oxidizing pollutants
such as dyes, cyanides, biohazards, chlorolignins from pulp and paper mill effluent.
DC power supply
Flocs
Anode
Cathode
H2
Coagulation
& Flocculation
Sediments
Figure 2. Electrocoagulation process interactions (Hydrogen discharge at the cathode generates gas
micro-bubbles that cause the floatation of flocs and the increase of pH).
Treatment of Wastewater by Electrocoagulation Method… 9
Electron flooding of the water that eliminates the polar effect of the water complex,
allowing colloidal materials to precipitate. The increase of electrons creates an
osmotic pressure that ruptures bacteria, cysts, and viruses.
Oxidation reduction reactions that are forced to their natural end point within the
reactor which speeds up the natural process.
Electrocoagulation induced pH swings toward neutral although this will not always
be the case and will depend on the type of electrolyte used.
TYPE OF ELECTRODES
Electrode material can subtancially affect the performance of an electrocoagulation
reactor. The heart of EC is the dimensionally stable oxygen evolution anode which is usually
expensive. The anode material determines the cation introduced into solution. Several
researchers have studied the choice of electrode material with a variety of theories as to the
preference of a particular material. The most common electrodes were aluminum or iron
plates as described by Vik et al. (1984) and Novikova and Shkorbatova (1982). Do and Chen
(1994) have compared the performance of iron and aluminum electrodes for removing color
from dye-containing solutions. Their conclusion was that the optimal electrocoagulation
conditions varied with the choice of iron or aluminum electrodes, which in turn was
determined by initial pollutant concentration and pollutant type.
STIRRING RATE
Bazrafshan et al. (2008), while comparing chromium removal efficiency with iron and
aluminum electrodes, showed that removal efficiency of chromium with aluminum electrodes
was lower than chromium removal efficiency with iron electrodes. Metal consumption
equally was much lower with aluminum than with iron electrodes. Conversely, power
consumption was lower with aluminum than with iron electrodes for the same concentration
of pollutant. However, as the chromium concentration in the solution increased to 500.0 mg/L
, the consumption of the electrode reduced, but efficient chromium removal occurred due to
the large amount of flocs formation that helped sweep away chromium. For example, iron
electrode consumption for the initial concentration of 5.0 mg/l and voltage of 40 V was 9.01 g
while for an initial concentration of 500.0 mg/L it was 7.70 g (Bazrafshan et al., 2008). The
highest efficiency of chromium removal (for both iron and aluminum electrodes) was
measured in acidic medium (pH = 3) for an initial chromium concentration of 500.0 mg/L and
at lower concentrations, the removal efficiency was almost complete at all pH values. At high
chromium concentration, however, the complete removal would have required longer time i.e.
higher power consumption.
Some researchers have investigated the relationship between ―size‖ of the cation
introduced and removal efficiency of organic waste (Baklan and Kolesnikova, 1996;
Vlyssides et al., 1997). The size of the cation produced (10-30μm for Fe3+ compared to 0.05-1
μm for Al3+) was suggested to contribute to the higher efficiency of iron electrodes. Their
conclusion was based on a single experiment, however, using chemical absorption of oxygen
10 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
as the only measure. Hulser et al. (1996) observed that electrocoagulation is strongly
enhanced at aluminum surfaces in comparison to steel. This is attributed to a higher efficiency
due to the in situ formation of dispersed aluminum-hydroxide complexes through hydrolysis
of the aluminate ion, which does not occur with steel electrodes.
Tsai et al. (1997) employed Fe and Al anodes to simultaneously utilize
electrocoagulation, responsible for removal of high molecules, and oxidation during treatment
of a raw leachate. Iron anodes provided better COD removal at low applied voltages than did
aluminum (Englehardt et al., 2006). As a general statement the efficiency of aluminum or
iron electrodes will depend on the specific type of pollutant and also on the different set of
operating parameters (Kobya et al., 2003).
ELECTRODE PASSIVATION
One of the greatest operational issues with electrocoagulation is electrode passivation.
The passivation of electrodes is of concern for the longevity of the process. Passivation of
aluminum electrodes has been widely reported in the literature (Nikolaev et al., 1982;
Osipenko and Pogorelyi, 1977). The latter also observed that during electrocoagulation with
iron electrodes, deposits of calcium carbonate and magnesium hydroxide were formed at the
cathode and an oxide layer was formed at the anode. Nikolaev et al. (1982) investigated
various methods of preventing electrode passivation and suggested the following options for
its control:
According to these researchers, the most efficient and reliable method of electrode
maintenance was to periodically mechanically clean the electrodes or wash the electrodes
with 8% sulfuric acid between runs in batch which for large-scale, continuous processes
is a challenging issue. Corrosion promoters such as Cl - ions have been found to induce
thinning of passive layer, enhance dissolution and promote depassivation (spontaneous
depassivation).
Other types of electrodes with a wide range of materials have been tested for
electrocoagulation process. These materials include: Graphite, Platinum oxide, Iridium
oxide, lead oxide, tin oxide, boron doped diamond (BDD). Graphite electrodes are
deemed to be cheap but unstable and for most part ineffective (Barisa et al., 2009). They
become easily fouled during the electrocoagulation process and this reduces their
effectiveness. Platinum and Iridium oxide electrodes are too expensive and ineffective.
Electrodes made of lead oxide (PbO 2) and tin oxide (SnO 2) are easy to manufacture but
they are highly unstable. Boron doped diamonds are materials suitable for use as anodes
in the electrocoagulation of organic compounds. Due to their very high resistance to
deactivation via fouling and extreme electrochemical stability they show no significant
corrosion even under high current densities. They have good chemical, mechanical and
Treatment of Wastewater by Electrocoagulation Method… 11
COLOR
Color is found mostly in surface waters, although some groundwater inside deep
wells may also contain color that is noticeable (APHA-AWWA, 1992; AWWA, 1999).
Many domestic and industrial wastewaters are rarely colorless and the color levels
depend on the industrial process and the age of the wastewater i.e. the travel time in the
collection and treatment system (Kim et al., 2005). The pulping and bleaching of wood
for example generally produce large amounts of wastewaters that contain lignin
derivatives and other dissolved wood by-products. Lignin derivatives which are usually
brownish in color remain resistant to biological degradation during wastewater treatment.
The brownish color of a pulp and paper mill effluent is mainly attributed to products of
lignin polymerization formed during pulping and bleaching operations. These
chromophoric groups are mainly quinonic types with conjugated double bonds
originating from pulping processes (Luner et al., 1970). When disposed of into natural
watercourses, they add color which persists for great distance. Additionally, colored
effluents from pulp and paper mills for example result in reduced photosynthetic activity,
increased long term BOD, increased water treatment cost for users downstream, and
increased toxicity (Springer et al., 1995).
Several studies have been carried out to determine the effectiveness of EC in color
removal. In general, the findings indicate that EC is more cost effective than normal or
conventional coagulation. Moreover, other wastewater pollution parameters are reduced
(Orori et al., 2005; Kashefialasl et al., 2006, Oricho et al., 2008). Electrocoagulation
combined with wood ash or bagasse ash has also been applied on tea factory effluent. In
one study by Maghanga (2008) on tea factory effluent, the treated effluent COD, BOD
and electrical conductivity were reduced by 96.6%, 42.4%, and 20.9% respectively.
Supporting electrolytes from wood ash, phosphate rock and bagasse ash further reduced
power consumption by between 64% and 16%, confirming the effectiveness of this
process.
Treatment of Wastewater by Electrocoagulation Method… 13
EC produces effluent with less total dissolved solids (TDS) content compared to chemical
treatments. If this water is reused, the low TDS level contributes to a lower water recovery
cost.
EC requires simple equipment and is easy to operate with sufficient operational latitude
to handle most problems encountered during its running.
Wastewater treated by EC can give palatable, clear, colorless and odorless water.
Sludge formed by EC tends to be readily settable and easy to de-water, because it is
composed of mainly metallic oxides/hydroxides.
Flocs formed by EC are similar to chemical floc, except that EC floc tends to be much
larger, contains less bound water, is acid-resistant and more stable, and therefore, can be
separated faster by filtration.
The EC process can remove the smallest colloidal particles, because the applied electric
field sets them in faster motion, thereby facilitating their agglomeration and subsequent
coagulation.
The EC process often avoids uses of chemicals and so there may be no problem of
neutralizing excess chemicals and no possibility of secondary pollution caused by chemical
substances added at high concentration as when chemical coagulation of wastewater is used
alone.
The gas bubbles produced during electrolysis can carry the pollutant to the top of the
solution where it can be more easily concentrated, collected and removed.
The electrolytic processes in the EC cell are controlled electrically and with no moving
parts, thus requiring less maintenance.
process as a whole. The use of new materials, different electrode types and arrangements
(Pretorius et al., 1991; Mameri et al., 1998) and more sophisticated reactor operational
strategies (such as periodic polarity reversal of the electrodes mentioned above) have led to
significant reductions in the impact of passivation. The issue, however, is still seen as a
serious potential limitation for applications where a low-cost, low maintenance water
treatment facility is required.
GEOMETRY
Geometry of the reactor affects operational parameters including bubble path, flotation
effectiveness, floc formation, fluid flow regime and mixing/settling characteristics. From the
literature, the most common approach involves plate electrodes (aluminum or iron) and
continuous operation. Water is dosed with dissolved metal ions as it passes through the
electrocoagulation cell. A downstream unit is often required to separate pollutant and water.
SCALE-UP ISSUES
One of the cornerstones of chemical engineering is to establish key scale-up parameters
to define the relationships between laboratory and full-scale equipment.
The surface area to volume ratio (S/V) is a significant scale-up parameter.
Electrode area influences current density, position and rate of cation dosing, as well as
bubble production and bubble path length. Mameri et al. (1998) reported that as the S/V ratio
increases the optimal current density decreases.
However, the S/V ratio was not widely reported. Some of the values reported are listed in
Table 1 below:
The values reported here seem empirical with no specific criteria for their choice. A more
rigorous and consistent approach is clearly required to establish a set of design characteristics
for Electrocoagulation reactors. The prime differentiator between pollutant removal by
settling or flotation would seem to be the current density employed in the reactor. A low
current produces a low bubble density, leading to a low upward momentum flux—conditions
that encourage sedimentation over flotation (Holt et al., 2002). As the current is increased, so
does the bubble density resulting in a greater upwards momentum flux and thus more likely
removal by flotation.
Other researchers such as Zolotukhin (1989) scaled up an electrocoagulation-flotation
system from laboratory to industrial scale. The following dimensionless scale-up parameters
have been chosen to ensure correct sizing and proportioning of the reactors:
Horizontal Flow
Vertical flow
Electrodes during EC can be set up as parallel vertical or horizontal sheets as can be seen
in Figure 3.
The turbulence generated by the gases at the anode and cathode can be used in both types
of flow. However, vertical flow allows for more improved separation by electroflotation as
compared with horizontal flow.
EFFECT OF PH ON ELECTROCOAGULATION
Optimal pH reported for electrocoagulation reactions varies significantly. These
discrepancies probably derive from the complex and variability of wastewater composition,
and the different operating conditions used in the electrocoagulation studies. It has been
established that pH has a considerable effect on the efficiency of the electrocoagulation
process (Springer et al., 1995, Chen et al., 2000, Li et al. 2001). The wastewater pH
determines the speciation of metal ions and influences the state of other species in solution
and the solubility of products formed. The pH of the medium also changes during
electrocoagulation process, as observed by other investigators. This change depends on the
type of electrode material and initial pH and alkalinity. In a study by Bazrafshan et al. (2008)
on the removal of Chromium VI from synthetic chromium solutions by electrocoagulation
Treatment of Wastewater by Electrocoagulation Method… 17
using aluminum electrode, it was observed that there was an increased in the solution pH for
an initial pH of less than 7. The increase was ascribed to hydrogen evolution at the cathodes
contrary to Chen et al. (2000) assertion that the pH increase is due to the release of CO2 from
wastewater owing to H2 bubble disturbance.
At low pH, wastewater is over saturated with CO2 which can be released during H2
evolution, causing a pH increase. In addition, if the initial pH is acidic, reactions would shift
towards a pH increase (Bazrafshan et al. 2008). During the same experiment, in alkaline
medium (pH > 8), the final pH did not vary considerably but a slight drop was recorded. This
result concurs with previous published works and suggests that electrocoagulation can act as a
pH buffer (Gao et al.,2005). In the same study of chromium removal by electrocoagulation
carried out over a wide range of Cr concentrations, it was also observed that the influent pH
did not significantly affect the removal efficiencies of Cr VI. This means that for practical
applications, pH adjustment before treatment is not required.
In another study by Springer et al. (1995) on the effect of pH on the color removal
reaction by electrocoagulation, it was found that higher pH slowed the electrocoagulation
reaction, thereby increasing power consumption. In a separate study on color removal from a
pulp and paper mill effluent, Orori (2003) found that decreasing the original effluent pH led
to a significant reduction in power consumption during electrocoagulation combined with
wood ash leachate. Lowering pH from 12.0 to pH 10.0 significantly (P 0.05) reduced power
consumption by between 20 to 21% during electrochemical removal of a paper mill effluent
color. It was postulated that decreasing the original effluent pH increased ionisation of
wastewater, which increased the rate of iron (II) ions production at the anode and hydrogen at
the cathode. Consequently, decreasing pH led to increased production of positively charged
iron (II) ions, which attracted the negatively colored flocci (Springer et al., 1995). Thus
increased production rate of these ions led to an increase in the rate of color removal at lower
pH than at higher pH. Therefore lower pH facilitated color removal and lowered electrical
power consumption. Li et al. (2001) reported that COD removal was at least 20% higher at
pH 4.0 than at pH 8.0 after a 4-hour electrolysis. Vlyssides et al. (2003) found that pH was
the most significant operational parameter in electrolyzing leachate, compared to Cl-
concentration, temperature, applied voltage, SO42- concentration and leachate input rate.
Lower pH favored COD removal and saved energy consumption within the range pH 5.5 –
7.5. The disagreement in these investigations suggests further work, perhaps in terms of the
mechanisms by which pH affects COD removal in leachate electrolysis.
Theoretically, it can be stated that acidic conditions decrease the concentrations of CO32-
and HCO3- , both well-known scavengers of OH radical generated at the anodes (Li et al.,
2001), while alkaline conditions promote the Cl-→Cl2→ClO-→Cl- redox cycle. Therefore, low
pH may enhance direct oxidation, while high pH may enhance indirect oxidation (Wang et
al., 2001). Thus, solution pH influences the overall efficiency and effectiveness of
electrocoagulation. An optimal pH seems to exist for a given pollutant, with optimal pH
values ranging from 6.5 to 7.5 (Holt et al.,2002).
Kashefialasl et al. (2006) showed that the maximum efficiency of color removal during
the treatment of dye solution containing colored index acid yellow 36 by electrocoagulation
using iron electrodes was observed at pH range 7–9 as expected considering the nature of the
reaction between ferrous and hydroxide ions. When the pH of solution was lower than 6,
Fe(OH)3 was in soluble form (Fe+3) and when it was higher than 9, Fe(OH)3 was in soluble
form {Fe(OH)4-} and because Fe(OH)3 played a major role in removing color, when pH of
18 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
solution was 8, color removal was the highest. The dye solution with different initial
concentrations in the range of 20-60 mg/l was treated by EC at optimized current density and
time.
In contrast, other investigators have found that pH variation does not considerably alter
COD removal in leachate electrolysis. Chiang et al. (1995a) have reported that the pH effect
on chlorine/hypochlorite production efficiency was insignificant over the range pH 4-10
during an electrolysis experiment in saline water conducted to help elucidate the mechanism
of electrolyzing leachate. Cossu et al. (1998) found that a pseudo-first-order rate constant for
COD reduction in real leachate increased only slightly at pH 3, compared with pH 8.3. Also,
Wang et al. (2001) have reported that at pH 8.9 and 10, COD removal was approximately 4%
higher than at pH 7.5, not a very significant effect.
High current densities are desirable for separation processes involving flotation cells or
large settling tanks, while small current densities are appropriate for electro-coagulators that
are integrated with conventional sand and coal filters. A systematic analysis will be required
to define and refine the relationship between current density and desired separation effects.
Current density (current per unit area of electrode) in an electrochemical process indicates
gross reaction rate. For example under weaker oxidative conditions, leachate may darken and
Treatment of Wastewater by Electrocoagulation Method… 19
brown precipitates may form at the anode (Cossu et al., 1998; Li et al., 2001). Increasing
current density improves COD and NH3-N treatment efficiencies at the same charge loading.
Bazrafshan et al. (2008) showed that increasing electrocoagulation voltage increased the
removal efficiency of Chromium, which was also helped by higher pHs as can be seen in
Tables 1 and 2. Chiang et al. (1995b) reported that during electrolytic treatment of leachate,
COD removal at 25 mA/cm2 was approximately 50% higher than that observed at 6.25
mA/cm2, for the same charge loading (1.178 x 105 Coulombs/L). This is probably due to the
fact that increasing current density during electrolysis enhances chlorine generation, which
may have been responsible for subsequent removal of pollutants (Costaz et al., 1983; Chiang
et al., 1995a).
Li et al., (2001) have shown that the effect of current density on treatment was not
evident between 30 and 120 mA/cm2 at a low Cl- concentration (1650 mg/L), but became
noticeable when Cl- concentration reached the 5000 mg/L level.
Figure 4. Effect of current density on the efficiency of color removal from a solution with concentration
of the dye = 50 ppm (Source: Kashefialasl, et al., 2006 with permission).
20 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
This result corroborates the importance of indirect oxidation during the electrolytic
treatment of leachate. In addition, Moraes et al. (2005) reported that color removal from
leachate strongly depended upon current density. Color removal efficiency at 116 mA/cm2
was five times higher than that at 13 mA/cm2, after 180 minutes of electrochemical treatment.
In the treatment of dye solution containing colored index acid yellow 36 by
electrocoagulation using iron electrodes Kashefialasl et al. (2006) showed that as current
density increased so did color removal from the dye solution up to a certain maximum as
shown in Figure 4.
During electrocoagulation, electrical current not only determines the coagulant dosage
rate but also the bubble production rate and size and the floc growth, which can influence the
treatment efficiency by electrocoagulation (Letterman et al., 1999; Holt et al., 2002). This is
ascribed to the fact that at higher voltage the amount of anode material oxidized increases,
resulting in a greater amount of precipitate for the removal of pollutants. In addition, it has
been demonstrated that bubble density increases and their size decreases with increasing
current density resulting in a greater upwards flux and a faster removal of pollutants and
sludge flotation (Khosla et al., 1991). As the current decreased, the time needed to achieve
similar efficiencies increases. This expected behavior is explained by the fact that the
treatment efficiency is mainly affected by charge loading (Q = It), as reported by Chen et al.
(2000). However, the cost of the process is determined by the consumption of the sacrificial
electrode and the electrical energy. It has also been established that for a given time, the
removal efficiency increased significantly with increase of current density. The highest
electrical potential normally produces the quickest treatment.
SPP1
45
Power Consumption (MWh))
SPP2
40
SPP3
35
SPP4
30 SPP5
25
20
15
10
0
15 20 25 30 35 40 45
o
Temperature ( C)
Figure 5. Effect of temperature on power consumption by ELCAS at five sampling points along a pulp
and paper Mill effluent treatment system (Source: Orori, 2003 with permission).
Treatment of Wastewater by Electrocoagulation Method… 21
The results obtained at different electrical potentials showed that initial concentration of
chromium may have an effect on the efficiency of its removal and for higher concentration of
chromium, higher electrical potential or more reaction time is needed. On the other hand, if
the initial concentration increases, the time required should increase too. It is clear from
Tables 1 & 2 that at higher concentrations, longer time is needed for removal of chromium,
but higher initial concentrations of chromium were reduced significantly in relatively less
time compared to lower concentrations. The time taken for its reduction thus increases with
the increase in concentration. This can be explained by the theory of dilute solution. In dilute
solution, formation of the diffusion layer at the vicinity of the electrode slows the reaction
rate, but in concentrated solution the diffusion layer has no effect on the rate of diffusion or
migration of metal ions to the electrode surface (Chaudhary et al., 2003). Chromium removal
with respect to time by electrocoagulation process at different pH levels is shown in Tables 1
& 2.
Figure 6. Effect of electrode gap on the removal of (A) Cu, (B) Pb and (C) Cd Current density=36
Am2, Electrolysis time =10 min, Conductivity = 900 mS/cm.
22 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
EFFECT OF TURBIDITY
In order to study the effect of turbidity (10, 50 and 200 NTU) on removal efficiency of
cadmium a set of experiments was performed with different initial concentrations of cadmium
(5, 50 and 500 mg l-1). The results obtained at optimum condition (pH=10, time reaction = 60
min and voltage = 40 V) showed that the removal efficiency for various concentrations of
cadmium was fairly unchanged and hence electrocoagulation process can be applied
efficiently for cadmium removal in the presence of turbidity (Mahvi and Bazrafshan, 2007).
EFFECT OF TEMPERATURE
Raising temperature during electrocoagulation increases the rate of reaction (Shenz et al.,
2006). Springer et al. (1995) showed that the time required for color removal reaction through
electrocoagulation to reach 0.2 Absorbance Units (AU) was cut approximately by half by
increasing temperatures from 23oC to 80oC (12 min vs 7 min). Orori (2003) studied the effect
of temperature on color removal by electrocoagulation combined with wood ash leachate
(ELCAS) and the results are shown in Figure 5. It was found that at 40oC color removal
consumed less than 50% the electric power used at 20oC by ELCAS treatment. This was
attributed to fast movement of electrons at higher temperatures compared to low
temperatures.
such as electrocoagulation in aqueous and non-aqueous solutions (Lund et al., 1991; Fry,
1996). While a large number of experiments have been performed with electrodes under
conditions where no SE was deliberately added, it is increasingly common practice to operate
an EC in the presence of a certain amount of ions such as chloride or ammonium or salts such
as NaCl, Na2SO4 and NaNO3 (Lopes et al., 2004; Orori et al., 2005; Shenz et al., 2006;
Englehardt et al., 2006; Uğurlu, 2006; Oricho et al., 2008; Yildiz et al. 2008). Some of the
electrolytes used in past experiments are shown in Table 4 and their respective effects on
effluent color removal.
Hu et al., (2003) carried out an experiment on defluoridation by EC and studied the effect
of coexisting anions. The results showed that the type of dominant anion had a direct impact
on the EC defluoridation reaction. Defluoridation efficiency was nearly 100% and most of the
fluoride removal reaction occurred on the surface of the anode in the solution without the co-
existing anions, due to the electro-condensation effect. In the solutions with co-existing
anions, most of the defluoridation took place in the bulk solution. The residual fluoride
concentration was a function of the total mass of Al3+ liberated. It was found that sulfate ions
inhibited the localized corrosion of aluminum electrodes, leading to lower defluoridation
process because of lower current efficiency. However the presence of chloride or nitrate ions
prevented the inhibition of sulfate ions, and the chloride ions were more efficient. Different
corrosion types occurred in different anion-containing solutions and the form of corrosion
affected the kinetic over-potential of the EC reaction.
When the concentration of NaCl salt or any other supporting electrolyte in solution
increases, solution conductivity increases. Consequently, with respect to the solution voltage
if any SE is added:
where:
the necessary voltage for access to a certain current density will reduce, and the consumed
electrical energy will be decreased (Kashefialasl et al., 2006). Excess SE affects the current in
the bulk of the solution, which is maintained mostly by the ions of the SE, and migration
effects on charged substrates can be neglected. The SE can also have some affects on the
double layer reducing the Zeta potential of the substrate ions and helping their agglomeration
or coagulation.
Orori et al. (2005) showed that when the volume of wood ash leachate increased during
color removal from a pulp and paper mill effluent, the power consumption reduced
considerably by almost 80%. Similar results were also obtained by Etiégni et al. (2007) and
Oricho et al. (2008).
24 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
Type of Electrolyte Applied voltage (v) Conductivity (μS/cm) Color removal (%)
*NaCl 2.9 19.13 83
*BaCl2 5.1 9.67 77
*KCl 2.7 18.75 80
*NaBr 4.2 11.77 80
*KI 5.3 9.18 79
*Na2SO3 3.8 13.8 76
*Na2CO3 4.2 15.8 76
**Wood ash leachate 23 4823.12 100
***Phosphate rock 23 1150-1730 100
****Bagasse ash 24 310 100
leachate
*****NH4NO3 3.9 - 100
*****Na2SO4 3.5 - 94
**Alum 23 3456.23 100
**Ca(OH)2 23 3358.34 100
*Source: Kashefialasl et al., 2006: (Current density =127.8A/m2, Time of electrolyses =6min)
**Source: Orori , 2003 (Current density= 250 A/m2, Time of electrolysis= 150 s)
***Source: Etiégni et al., 2007: (Current density = 222.2 A/m2, Time of electrolysis = 55 s)
**** Source: Maghanga, 2008: (Current density = 55 A/m2, Time of electrolysis = 4 min)
***** Source: Lopes et al., 2004: (Current density = 2 mA/cm2, Time of electrolysis = 70-96 hrs)
20
Power (Whr)
15
10
0
0 1000 2000 3000 4000 5000
Electrolyte Dosage (g/m3 )
Source: Etiégni et al., 2009
It appears that leachates from wood ash contain a wide range of ions or supporting
electrolytes that may be helping or assisting the electrocoagulation reaction (Figure 7). Chou
et al. (2009) studied the effect of NaCl concentration on the removal efficiency of indium
Treatment of Wastewater by Electrocoagulation Method… 25
(III). They showed that there was an increase removal efficiency up to 83% when NaCl (used
as supporting electrolyte) concentration was 8 g/l. The concentration of supporting electrolyte
was adjusted to the desired levels by adding a suitable amount of NaCl to the synthetic
wastewater. Increasing the concentration of the supporting electrolyte from 0 to 200ppm led
to an increase in indium (III) ion removal efficiency, whereas with the concentration of the
supporting electrolyte increasing, the specific energy consumption decreased by almost 80%.
When the concentration of the supporting electrolyte increased, the solution ohmic resistance
decreased, so the current required to reach the optimum applied voltage diminished,
decreasing the consumed energy (Chou et al. (2009).
Although some SEs are available commercially, they can be extracted from material
otherwise considered as waste. Several research papers have been recently published on the
use of leachate from ash emanating from wood, coffee husk or bagasse as supporting
electrolyte (Orori et al., 2005; Etiégni et al., 2007, Oricho et al., 2008).
Several studies have been carried out on the operating cost of electrochemically treated
wastewater (Bayramoglu et al., 2004; Can et al., 2006; Bayramoglu et al., 2007). In a study
by Bayramoglu et al. (2004) for the treatment of textile wastewater by EC using aluminum
and iron electrode materials, the effect of wastewater characteristics and operational variables
on the technical performances of COD and turbidity removal efficiencies as well as on the EC
operating cost were determined.. Only direct costs such as material (electrodes and chemical
reagents) and energy costs were considered for the calculation of the operating cost. Other
cost items such as labor, maintenance and solid/liquid separation costs, depreciation of fixed
investment such as rectifier and electro-coagulators were not taken into account. This
simplified cost equation was used to evaluate the effect of various process variables on the
operating cost:
where Cenergy and Celectrode, were consumption of energy and electrode material per kg of COD
removed, which are normally obtained experimentally. Unit prices, a and b, determined for a
specific market are as follows: a= electrical energy price and b= electrode material price for
aluminum or for iron. Using equation 6, Bayramoglu et al. (2004) found that for iron
electrode, the operating cost decreased initially with pH until pH = 5, where it remained
constant up to pH=7, beyond which it increased (Figure 8). For aluminum electrodes, the EC
cost increased with initial pH.
26 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
Initial pH
Source: Bayramoglu et al. (2004) with permission
When cells are set-up in an electrocoagulation process, one can choose from different
modes or connections depending on the required voltage, the expected output and the overall
efficiency of the EC system. Kobya et al. (2007) studied the effect of wastewater pH, current
density and operating parameters for two sacrificial electrode materials, Fe and Al, and three
electrode connection modes - namely monopolar-parallel (MP-P), monopolar-serial (MP-S)
and bipolar-serial (BP-S) on the EC operating cost. The highest consumption of electrode
material occurred with bipolar series mode (BP-S); approximately 0.27 kgm−3 for Fe
electrode and between 0.18–0.23 kgm−3 for Al electrode. Monopolar parallel (MP-P) mode
showed the lowest electrode consumption for both electrode Fe and Al materials: 0.12 kgm−3
for Al electrode and 0.16 kgm−3 for Fe electrode (Kobya et al., 2007). When the consumption
of energy was compared for the three modes, as seen in Figure 10, only a minor change was
observed with pH for all of the systems using Fe or Al electrodes.
MP-S and BP-S modes exhibited high consumptions of energy because of the serial
connection that required higher potential. When MP-P mode was used, it consumed the
lowest energy or approximately 0.63 kWhm−3 and 0.7 kWhm−3 for Fe and Al electrode
respectively. The effect of the initial pH on amount of sludge production is depicted in
Figure 12. Sludge amounts vary from 0.65 to 1.0 kgm−3 for Fe electrode and from 0.9 to 1.3
kgm−3 for Al electrode (Kobya et al., 2007). In general, more sludge was produced with BP-S
mode than with MP-P mode because of high electrode material consumed leading to high
power consumption and higher electro-coagulant produced in situ. MP-P mode for both
Treatment of Wastewater by Electrocoagulation Method… 27
electrode materials was therefore economically more feasible owing to its low electrical
energy consumptions and amount of sludge produced (Figure 12).
Figure 9. Different types of electrode connection modes: a-Monopolar parallel (MP-P), b-Monopolar
Serial (MP-S), c-Bipolar parallel (BP-P) modes. (Source : Kobya et al., 2007, with permission).
67% with MP-P mode for a current density of 50Am−2. However, for aluminum electrode, the
effect of the current density was more pronounced on COD removal, especially for MP-P
mode and that lower current densities were more favourable. For example a 30Am−2 was
preferred with MP-S mode.
EFFECT OF POLYELECTROLYTE
AND SE ON THE EC OPERATING COST
As a general rule, EC operating cost has been found to reduce with the addition of
polyelectrolyte up to an optimum concentration beyond which it usually rises, although this
will also depend on the type of polyelectrolyte. Can et al. (2006) showed alum and
polyaluminum chloride (PAC) increased operating EC operating cost when their
concentration increased (Figure 13). However, Orori et al. (2005) found that increasing the
concentration of wood ash leachate reduced power consumption and reduced operating cost,
although the cost of electrode replacement and sludge removal was not included in the overall
operating cost calculations.
Treatment of Wastewater by Electrocoagulation Method… 29
Conductivity, μS/cm
Wood ash is the residue powder left after the combustion of wood. The main producers of
wood ash are wood industries, power plants, homesteads especially in Third World countries.
Large amount of this residue are produced every day. Typically 6-10 percent of burned wood
results in ash. Wood ash is commonly disposed of in landfills or agricultural lands, but with
rising disposal costs ecologically friendly alternatives are becoming more attractive. These
alternatives will be based on the ash composition. It has been demonstrated that wood ash
composition is a function of the wood combustion temperature as can be seen in Table 5
(Etiégni and Campbell, 1991).
Wood combustion produces a highly alkaline ash that can be used to neutralize acidic
effluent. As can be seen in Table 5 below Ca, K, Mg, Mn, Fe and Na are important elements
found in wood ash. Misra et al., (1993) analyzed samples of wood ash using Inductively
Coupled Plasma Emission Spectrometer (ICPES) and X-ray diffraction (XRD) to identify the
minerals present in wood ash. A list of the compounds identified in ash is shown below in
Table 6. The low temperature ash at 600oC showed strong peaks corresponding to calcium
carbonate. Pine and aspen ash contained relatively higher amounts of potassium compared to
poplar ash and showed strong peaks corresponding to K2Ca(CO3)2. Pine ash contained
calcium manganese oxide, Aspen ash had sulfates of calcium and potassium, and poplar ash,
silicates of K, Mg, and Ca. At higher temperatures (1000oC) where most industrial wood-fired
boilers operate, with the dissociation of carbonates, XRD patterns showed predominant
presence of calcium and magnesium oxides. In addition, pine ash which contains more
Treatment of Wastewater by Electrocoagulation Method… 31
manganese showed the presence of calcium manganese oxide and manganese oxide.
Similarly, poplar, being richer in sodium, displayed weak peaks corresponding to sodium
calcium silicate. It appears that when the ash is left standing in air, calcium oxide reacts
with atmospheric water vapor to form calcium hydroxide. However calcium hydroxide is
unstable at temperatures over 600oC. Table 6 also indicates that small amounts of potassium
may be present as K2SO4 as the peaks corresponding to this compound become distinct at
higher temperatures. Low temperature ash produced from the wood waste appears to contain
predominantly calcium carbonate while at high temperatures the content changes to
predominantly calcium oxide. What this Table shows is the close relationship of ash
composition with combustion temperature.
Many of these elements, when in solution, will behave as counter-ions. Wood ash
leachate added to wastewater does the following:
One of the most important factors that need to be considered when using wood ash
leachate as supporting electrolyte is the time required to allow leaching to take place and the
ash to water ratio for leaching.
In an experiment conducted on wood ash leaching, Etiégni and Campbell (1991) found
that the total dissolved solids (TDS), K, Na, and Mg concentration increased linearly as the
ash to water ratio increased (Figure 14). However, the percentage of ash dissolving did not
change significantly, as approximately 10% of the ash dissolved at 50 g/L and 9% at 390 g/L.
32 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
(EC) of the leachates followed the same trends as the TDS (Etiégni and Campbell, 1991). The
reaction order of TDS leaching was found to be equal to n = 0.08 indicating that TDS was
essentially independent of pH.
Table 6. XRD analysis of wood ash showing relative intensity of the strongest peaks (%).
Compound Pines Aspen Poplar White oak White Oak Bark Douglas Fir
Bark
600oC 1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC 600oC 1300oC
CaCO3 100 100 100 100 100 100
K2Ca(CO3)2 86 21 11
Ca(OH)2 24 34 40 * 3 8
MgO 8 100 26 4 49 4
CaO 19 100 16 100 100 100 100
Ca4Mn3O10 21
Ca2MnO4 75
K2Ca2(SO4)3 12
Mg6MnO8 12
K2SO4 11 9 2 *
K2MgSi3O8 46
CaSiO3 *
Na2CaSiO4 *
Ca2SiO4 23 17
22 11
These findings have major implication on the use of wood ash leachate. They show that
one can expect to extract most of the useful ion species within one hour. This should therefore
constitute a critical design parameter, especially if the EC reactor must be a continuous one.
One such reactor was recently patented and is shown in Figure 18 below. The most important
fixture of this system is the ash tray and the mixing tank with three compartments. Baffles A1
and A2 prevent vortexes created by vigorous mixing in compartment 1. They are eliminated
and allow laminar flow into compartment 2 and 3, thereby permitting ash particles to settle at
the bottom of the mixing tank. The resulting mixture- ash leachate and wastewater- flows into
the electrolytic tank where electrolysis takes place. Three variables are important here if the
electrocoagulation system is continuous. The 1st one is the time required to allow the ash to
leach which will be at least one (1) hour, the second will be the ash to water ratio and the
third will be the detention time needed to permit maximum electrolysis before treatment. All
these variables must be determined experimentally for each type of wastewater. If we assume
plug flow:
C = Co e-t/tR (26)
where:
C = concentration of effluent
Co = concentration of influent
tR = Hydraulic detention time in the electrolytic tank.
34 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
4
Ash=30 g/l
3
2 Ash=20 g/l
Ash=10 g/l
1
Ash=5 g/l
0
0 5 10 15 20 25 30
Time (hr)
Figure 16. Leaching of Na at different ash concentrations.
12
Potassium concentration (g/L)
10 Ash=40 g/l
Ash=30 g/l
8
6
Ash=20 g/l
4
Ash=10 g/l
2
0 Ash=5 g/l
0 5 10 15 20 25 30
Time (hr)
Figure 17. Leaching of K at different ash concentrations.
The hydraulic detention time would have been determined experimentally. The
volumetric flow rate of the factory effluent will then enable one to size up the electrolytic
tank:
Ash tank
Wastewater
+ - A A
Electrodes 2 C2 1 C1
C3
Mixing Tank
Electrolytic tank
Leached ash
Leached ash
draw off pipe
Settled sludge
Sludge
Sludge settling
Tank
Treated wastewater
Figure 18. Process description of ELCAS for color removal from wastewater.
Because of the inherent requirement of leaching of wood ash, most reactors have been
designed as batch. Other materials have also been investigated as potential sources of
supporting electrolyte. They include: rock phosphate, ash from bagasse, coffee husk (Etiégni
et al., 2007; Etiégni et al., 2009). Rock phosphate composition is shown in Table 7. The
advantage with these materials is that they are produced on site, usually by the same factory
that really needs them, at almost no cost. They have been successfully used to treat industrial
wastewaters as shown in the picture below (Figure 19). The electrode configuration is
depicted in Figure 20 below. This assembly with an electrode surface to volume ratio of 75
m2/m3 of effluent at a current density of 200 mA/m2 was successfully applied to completely
decolorize wastewater from a coffee factory effluent. They can quickly neutralize acidic
effluent and once they have been used, they can be safely applied on land as soil amendment.
Rock phosphate produces less sludge than wood ash, but it is less effective for the
reduction of power consumption (Etiégni et al., 2007). The only draw-back with wood ash
leachate is that its tend to yield large amount of sludge (Orori, 2003; Barisa et al., 2009). The
existence of these low cost supporting electrolytes is another tool at the disposal of
Treatment of Wastewater by Electrocoagulation Method… 37
wastewater treatment professionals. In terms of performance, the use of SE can only help the
overall treatment of wastewater and more importantly reduce the overall power consumption.
The kinetic rate equation for representing the leaching of TDS or metallic ions from
wood ash can be described by the following mth order reaction kinetics:
dC
= −kCm ------------------------------------
(28) (26)
dt
where C represents the TDS or the metallic ion concentration, m is the order of reaction, k is
the reaction rate constant, and t is the time. For a first-order reaction, the above Eq. (28)
becomes:
The slope of the plot of ln Ct/C0 versus time gives the value of the rate constant k1, in t−1.
Here, C0 is the initial concentration in milligrams per liter, Ct is the concentration value in g
per liter at time t, and t is the time. The leaching reaction of wood ash can be described by a
first order reaction where m = 1 and k = 0.007 in the case of potassium shown in Figure 21.
38 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
Figure 20. Batch reactor with electrode configuration for color removal experiment.
-1
Table 8. Mean electrodes mass loss year by electrochemical color removal methods.
Sampling -1
Electrodes mass Loss year
point
ELCAS ELCCA ELCAL ELCON
Tones % Tones % Tones % Tones %
SP1 1.63 0.024 5.52 0.079 2.43 0.034 9.57 0.134
SP2 2.87 0.041 9.67 0.138 4.83 0.068 14.96 0.209
SP3 3.68 0.053 12.7 0.181 6.25 0.089 16.93 0.237
SP4 4.66 0.067 15.39 0.220 8.12 0.115 18.53 0.259
SP5 5.98 0.086 20.38 0.291 10.83 0.153 22.36 0.312
0.350 ELCAS
ELCCA
Percent electrodes mass loss per year (%)
0.300 ELCAL
ELCON
0.250
0.200
0.150
0.100
0.050
0.000
SP1 SP2 SP3 SP4 SP5
Sampling points
Figure 23. Electrodes mass loss comparison for all electrochemical color removal methods at all
sampling points
Figure 23 shows the percent electrodes mass loss per year for all electrochemical color
removal methods at these various sampling points. Percentage electrodes mass loss (<0.5%)
was not statistically significant compared with the initial electrodes mass for all electrochemical
Treatment of Wastewater by Electrocoagulation Method… 41
color removal methods used. Electrodes mass loss increased significantly from SP1 to SP5 for
all the EC methods as time for current flow increased. This was attributable to increasing
original effluent color intensity from SP1 to SP5, which required longer time for color removal.
Faraday‘s law can be used to relate the mass (m) of electrolytically generated aluminum
going into solution to the operating current (I) and the run time (t). In this relationship, M is
the atomic weight of aluminum or iron, z is the number of electrons transferred in the anodic
dissolution (here z = 3 or z= 2), while F is Faraday‘s constant (96486 C mol-1).
ItM
m= ………………………………………………………
(30) (27)
zF
Using this equation, the amount of coagulant or soluble metal delivered to the solution
may be calculated.
SLUDGE PRODUCTION
One of the biggest draw-backs of wood ash leachate as a supporting electrolyte during
electro-coagualtion is its tendency to produce large amount of sludge. Orori et al. (2005)
showed that ELCAS produced the highest sludge quantity compared to ELCAL and ELCCA,
probably because of the contribution from wood ash leachate as can be seen in Tables 8 and
9. This has also been confirmed by other researchers (Barisa et al., 2009).
There have been several applications of electrocoagulation experiments with wood ash
leachate as supporting electrolyte. Orori et al. (2005) reported successful total color removal
from a pulp and paper mill effluent using ELCAS process with close to 80% reduction of
electric power consumption. This process has also been applied to remove color from Tea and
coffee factory effluent.
where CSE is the cost of supporting electrolyte which is almost negligible. It may include only
some labor and transportation cost to the point of utilization. If we assume that the cost of
42 Lazare Etiégni, K. Senelwa, B. K. Balozi et al.
electric power and that of electrode replacement will be cut by a least 50%, the operating cost
of EC will be substantially reduced, if one was to use wood ash leachate.
CONCLUSION
Numerous research papers have shown that electrocoagulation method can be a more
effective alternative treatment process for pollution abatement. However, certain factors such
as electrode replacement and power cost are likely to make the method less affordable,
although new sets of electrodes such as polypyrrole (PPy) and boron doped diamond (BDD)
electrodes are becoming available on the market. BDD are expensive but more resistant and
require higher current density. Fortunately, supporting electrolytes such as wood ash
leachates, which are inexpensive and can make the process more affordable are available. A
logical, systematic approach to a fundamental understanding of electrocoagulation with wood
ash leachate as SE is clearly required. The design phase can then proceed on solid scientific
and engineering knowledge. A large number of key mechanisms are dependent on a few
operating parameters. The authors of this chapter have had quite a wealth of experience
working with this SE. A trade-off between the competing factors such as sludge production,
electrical conductivity and reduction of power consumption must be evaluated with respect to
other optimum operating conditions.
Treatment of Wastewater by Electrocoagulation Method… 43
LIST OF ABBREVIATION
EC = Electrocoagulation
SE = Supporting electrolyte
ELCAS = Electrocoagulation combined with wood ash leachate
ELPHOS = Electrocoagulation combined with rock phosphate leachate
ELCON = Electrocoagulation alone
ELCAL = Electrocoagulation combined with alum
ELCCA = Electrocoagulation combined with calcium oxide/hydroxide
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 49-81 © 2010 Nova Science Publishers, Inc.
Chapter 2
Dorota Wolicka*
Institute of Geochemistry, Mineralogy and Petrology, Faculty of Geology,
University of Warsaw, Żwirki i Wigury 93, 02-089 Warsaw, Poland,
1. INTRODUCTION
1.1. Trends in Environmental Biotechnology
Human activity is strictly linked with the production of waste i.e. materials and
substances that are undesired and cannot be used further. On the one hand these substances
are natural to the environment, eliminated from further technological process by their
uselessness (e.g. mining waste), or represent new products i.e. anthropogenic waste, being the
by-product of industrial and agricultural activities. A separate group comprises municipal
waste that is not linked with production but results from human dwelling.
Utilization actions aiming at neutralizing and/or removal of waste are focused on
substances that due to their existing or potential chemical activity may negatively influence
the biosphere. Non-active substances represent alien elements in the natural environment, but
due to their passive character, their utilization is concentrated on non-conflicting storage.
Active pollutants influencing the natural environment penetrate it as gaseous emanations,
fluids (sewage and effluents) and solids.
Restriction of emission and removal of hazardous gaseous emanations should be
conducted in places where they are formed. Imperfection of the applied technology or its lack
results in atmospheric pollution. This problem can be of local (around industrial plants, e.g.
chemical works, food processing plants, around farmsteads and stock farms), country or
global range (emission of CO2, nitrogen compounds, gases hazardous to the ozone layer).
Control of such pollutants beyond their source areas is difficult or even impossible.
*
Corresponding author: email: d.wolicka@uw.edu.pl.
50 Dorota Wolicka
Control of liquid pollutants such as sewage or effluents or solid waste is always focused on
their chemical transformation in order to obtain end-products that are neutral to the environment.
These methods are applied to sewage, whereas in the case of solid waste they may be
used in the soil-water environment with essential and stable water supply, which as a solvent
mobilizes solid compounds susceptible to leaching forming effluents within the dump sites or
in their foreland, and facilitates the growth of microorganisms at the boundary between the
solid and liquid phases.
The main aim of each sewage treatment method is protection of the natural environment
against unfavourable influence caused by introduction of such wastes. For many years
attention was drawn on disturbance of the oxygen balance caused by presence of organic and
ammonium compounds. Due to this fact, pollution treatment methods were dominated by
methods ensuring distinct reduction of BOD5, COD, nitrification of ammonium compounds,
and effective utilization of active sludge. Next, focus was drawn on eliminating inorganic
compounds of phosphorus and nitrogen, degradation of non-biodegradable or poorly
biodegradable compounds. The primary aim became, however, decreasing treatment costs,
what is economically justified.
These aims can be realized using many methods and on every stage of liquid waste
utilization. New equipment or technology may be introduced, or those previously applied can
be modified and optimized. It should be remembered that preventing environmental pollution
does not begin at the stage of sewage treatment, but much earlier, and requires wide-range
activities. The most correct attitude is preventing hazards at their source, particularly in the
case of industrial waste. In the first place activities should be undertaken to diminish the
volume and harmfulness of pollutants that are by-products of industrial processes or steps
should be taken to work out a technology treating several types of sewage and/or waste in one
process. The scale of this problem can be illustrated by the dairy industry, where the volume
of sewage flowing out of a dairy plant reaches several cubic meters per 24 h and comprises
0.5 to 3 times the volume of processed milk. Similar quantities of sewage are produced by oil
refineries, where 1 ton of processed oil results in about 10 to 18 cubic meters of refinery-
petrochemical waste. Analogous results have been observed in the case of solid waste
produced during technological processes, where the volume of waste distinctly exceeds the
final product. For example, the industry of phosphorus fertilizers produces 5 t of waste
(phosphogypsum) from 1 t of phosphorites. Such waste not only poses serious hazard to the
biological equilibrium in the environment, but also distinctly increases the cost of the
technological process, which should in this case also include utilization of liquid and solid waste.
Treatment of different types of sewage requires application of many physical, chemical
and mechanical methods. They include e.g. retaining of suspensions from sewage on screens,
sieves and in settling tanks, neutralization of acidic or alkaline sewage, adsorption of sewage
components on relevant adsorbents, coagulation of non-subsiding suspensions, or extraction
of sewage components by relevant solvents. Although commonly applied, these methods are
not environment friendly, do not entirely solve the problem of waste neutralization, and often
change only the physical-chemical composition or form of the waste. In such cases,
anthropogenic waste is produced, which after many years of active production in the
industrial plant may even become anthropogenic deposits. They are often the source of many
rare elements e.g. elements from the lanthanoid or actinoid series (rare earths), which occur in
considerable quantities in the waste, but cannot be recovered due to lack of efficient process.
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 51
Thus it seems crucial to search for pro-ecological methods focused on the utilization of
industrial sewage and waste that are hazardous to the natural environment. To such methods
belong biological methods that are used in a wide range in the case of some types of organic
sewage. All these methods use microorganisms to remove organic and some inorganic
compounds from sewage. They decrease the volume of pollutants in the sewage and retain
correct parameters determined by norms for sewage water introduced into surface waters.
Selection of the appropriate method of sewage treatment depends on the type of sewage, its
composition, volume, as well as degree of pollution of the water reservoir, to which the
refined sewage will be introduced. These methods can be variously modified, and in some
cases multistage sewage treatment is carried out with application of different methods.
Biological methods may be applied only to the treatment of sewage, in which the
concentration of toxic compounds does not hamper the incubation of microorganisms. Due to
this fact, continuous cultures are applied in biological treatment plants, which allow obtaining
the maximal incubation speed of microorganisms through continuous supply of fresh medium
and removal of metabolism products. In a classical continuous culture (chemostat), the time
of sewage flow should correspond to the microorganism growth speed, which is determined
experimentally by change of flow speed from the moment when the biomass stable level is
attained.
Recently, search is focused on methods that would allow simultaneous biodegradation of
several wastes. Concurrent biodegradation of two industrial wastes seems an interesting issue
from the economy of the process. Costs linked with simultaneous biodegradation of two
hazardous industrial wastes are always lower than for each of them separately.
Biotransformation of phosphogypsum in the environment of organic sewage as a liquid state
to its dissolution may be an interesting example. This trend in environmental biotechnology
results in a large number of reports focused on treatment of high-sulphur sewage using
sulphidogenesis methods (Lens et al., 1998) or biotransformation of solid waste in organic
sewage environments. There are also single publications on the application of sulphidogenesis
on the treatment of sewage after enrichment with phosphogypsum formed as a by-product in
many industrial branches (Deswaef et al., 1996; Kaufman et al., 1996; Wolicka et al., 2005;
Wolicka & Kowalski, 2005; Wolicka & Kowalski 2006a; Wolicka & Kowalski, 2006b;
Wolicka, 2008b; Wolicka & Borkowski, 2008; Wolicka & Borkowski, 2009).
Concluding, the most aggressive in the natural environment are liquid substances and
they pose the most serious hazard to the natural environment.
Nowadays, sulphate reducing bacteria (SRB) are becoming more frequently applied in
the biodegradation of anthropogenic waste. These bacteria in course of anaerobic respiration
produce hydrogen sulphide, which can bind heavy metals in poorly soluble and non-toxic
sulphides of metals. This is one of the many advantages of sulphidogenesis application in
environmental biotechnology. Additionally, due to the toxic activity of hydrogen sulphide,
SRB may eliminate various microorganisms from the environment, including pathogenic
forms, what causes their domination in a sulphate rich environment. The bacteria may be
utilized during the biodegradation of two industrial wastes, of which one may be solid waste
as the sulphate source, and the second – liquid waste as the carbon source. Application of
such process allows simultaneous biodegradation of two arduous industrial wastes, thus
shortening the biodegradation time. Moreover, post-culture deposits generated in this process
i.e. carbonates and/or calcium phosphates can potentially be utilized in agriculture. An ideal
example is the biotransformation of phosphogypsum.
It should be remembered, however, that anaerobic methods are not devoid of
disadvantages, including:
Other metals that are present in the alloy e.g. Cu, Zn, Ni or Cr may also take part in the
reaction.
Sulphate reducing bacteria (SRB) are heterotrophs and absolute anaerobes. They utilize
sulphates, as well as other partly oxidized sulphur compounds (sulphites, tiosulphites and
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 53
tetrationates), and elemental sulphur as the final electron acceptor in the respiration processes
(Postgate, 1984; Gibson, 1990). Electron donors for this microorganism group are organic
compounds such as e.g. alcohols, carboxylates, phenols, aliphatic and aromatic hydrocarbons,
amino acids and some carbohydrates.
Diverse SRB physiology influences their distribution in the natural environment as well
as in anthropogenic environs e.g. polluted by crude oil and oil products (Wolicka&
Borkowski, 2007a; Wolicka, 2008a). The presence of SRB has been noted in aquatic and
terrestrial environments (Hao et al., 1996). They occur in soils, deposits of fresh water and
marine reservoirs, in silts at the moutha of river deltas etc. (Chi Ming So & Young 1999),
thermal springs and in geothermal regions, in crude oil, refining and petrochemical waste,
natural gas intakes and on corroding steel (Hao et al., 1996). They may be found in all types
of bioreactors purifying sewage in anaerobic conditions, from which they can be isolated
(Przytocka-Jusiak et al., 1997; Baena et al., 1998, 1999, 2000; Hernandez et al., 2000). The
most characteristic environments of SRB occurrence are marine deposits (Bak & Widdel,
1986; Szewzyk & Pfennig, 1987; Lovley et al., 1995; Aeckersberg et al., 1998; Caldwell et
al., 1999; Kniemeyer et al., 2003), in which sulphate concentration reaches averagely 28mM
(Wit, 1992), as well as oil fields and crude oil reservoirs (Voordouw et al., 1996; Mueller &
Nielsen, 1996; Jenneman & Gevertz, 1999; Magot et al., 2000; Wolicka, 2008a). Their
presence has also been noted in environments polluted by crude oil and oil products, dairy
work sewage, whey, refining-petrochemical waste and distillery decoctions (Wolicka &
Kowalski, 2005; Wolicka, 2006; Wolicka, 2008b; Wolicka & Borkowski, 2009).
Additionally, SRB always accompany crude oil and were for a long time considered as
microorganisms indicative of oil deposits (Postgate, 1984). Suggestions pointing to the
presence of SRB in brines from oil fields come from 1926. They are the first attempt to
explain the permanent presence of sulphides in crude oil reservoirs (Jenneman & Gevertz,
1999). SRB are the most commonly isolated group from oil fields and their groundwater
(Rueter et al., 1994; Mueller & Nielsen, 1996; Aeckersberg et al., 1998; Wilknes et al., 2000;
Magot et al., 2000; Rozanowa et al., 2001).
Although SRB are considered as absolute anaerobics, their presence has also been noted
at the boundary between the oxygenated and anaerobic zone in sediments, or even within the
oxygenated zones. The presence of SRB is often recognized in the environment due to the
presence of the characteristic odour of hydrogen sulphide as well as due to black colouring being
the effect of precipitation of poorly soluble metal sulphides (Postgate, 1984; Gibson, 1990).
The preferred carbon sources for SRB are low-molecule organic compounds such as
organic acids e.g. lactic, formic, pyruvic, and malic acids; volatile organic acids e.g. acetic acid,
54 Dorota Wolicka
and alcohols e.g. ethanol, propanol, methanol, and butanol. Some SRB species are known to
utilize amino acids as the sole carbon source: Desulfovibrio aminophilus (Baena et al., 1998),
Desulfobacterium vacuolatum (Rees et al., 1997), and Desulfovibrio mexicanus (Hernandez-
Eugenio et al., 2000). Some species, e.g. Desulfotomaculum antarcticus may utilize glucose as
the sole carbon source, but this is a rare case in SRB (Fauque et al., 1991). Rather common in
turn is the utilization of aromatic and aliphatic compounds (Widdel & Bak, 1992).
All organic compounds that represent the optimal carbon source for SRB are the products of
fermentation formed during anaerobic biodegradation of carbohydrates, proteins and lipids (Fauque
et al., 1991; Hao et al., 1996). This results from the fact that SRB do not produce hydrolytic enzymes
and become involved in the anaerobic biodegradation of organic matter in the last stage. The only
exception is the archeon Archeoglobus fulgidus, which can produce hydrolytic enzymes.
Many papers devoted to the anaerobic biodegradation of crude oil contain information of
the utilization of SRB in this process, as well as in the biodegradation of oil products both in
refinery-petrochemical waste and in bioremediation of soils polluted by crude oil and oil
products. Biodegradation of n-hexane, n-octane and n-decane by various SRB has been
described by Gieg & Suflita (2002). Aeckersberg et al. (1998) described two mesophilous
strains Hxd3 and Pnd3 utilizing n-alkanes C12-C20 and C14-C17, whereas So & Young (1999)
described a mesophilous strain AK-01 utilizing n-alkanes C13-C18. Tribe TD3 was able to
grow on media with n-decane and n-alkanes C6-C16 (Reuter et al., 1994).
Utilization of aromatic compounds such as benzene, toluene and xylene in SRB cultures
has been noted by Edwards & Garbic-Galic (1992), Beller et al. (1992), Edwards et al. (1992)
and Ball & Reinhard (1996). Benzene, toluene, ethylbenzene and xylene (orto-, meta-, para-
xylene) were utilized by thermophilous sulphidogenic consortia ALK-1 and LLNL-1
described by Chen & Taylor (1997). Benzene decomposition by thermophilous SRB was
noted by Lovley et al. (1995), Przytocka-Jusiak et al. (1997) and Caldwell et al. (1999).
reaction ∆G0’
(kJ/mol)
3 lactate → 2 propionate + acetate + HCO3- + H+ -165
2 lactate + SO42- + H+ → 2 acetate + 2CO2 + HS- + 2H2O -189
4 propionate + 3SO42- → 4 acetate + 4HCO3- + 3HS- + H+ -151
acetate + SO42- → 2HCO3- + HS- -60
acetate + 4S + 3H2O → 4H+ + HCO3- + 4HS- + CO2 -24
4H2 + SO42- + CO2 → 3H2O + HS- + HCO3- -152
H2S + SO42- + H+ → HS-+ 4H2O -172
Besides easily accessible carbon sources and the presence of oxidized sulphur
compounds, many factors influence the life and growth of SRB. These physical and chemical
factors include: concentration of dissolved oxygen, temperature, pH, Eh, and presence of
accompanying microflora. Studies by Hao et al. (1996) have shown that concentration above
1.0 mg O2/l leading to the increase of the redox potential, in effect inhibits SRB activity. On
the other hand, some tribes such as e.g. Desulfovibrio desulfuricans, D. vulgaris, D.
56 Dorota Wolicka
bacteria to obtain energy indispensable for growth, because they do not reduce sulphates. A
representative of this group is Desulfovibrio sulfodismutans, which has the ability to reduce
each of these compounds according to the following reactions (Hao et al., 1996):
In anaerobic conditions elemental sulphur may also represent the final electron acceptor.
This process is known as sulphur respiration and is not as common as sulphate reduction. A
representative of this group is Desulfuromonas acetoxidans that utilizes acetate (rarely
ethanol, propanol) as the carbon source and electron donor, and conducts its complete
oxidation to CO2 according to the following reaction (Bothe & Trebst, 1981).
acceptor – elemental sulphur. The ability to reduce elemental sulphur has been noted also in
other microorganisms, among which occur:
Sulfospirillum – utilizing most frequently H2 as the electron donor.
Desulfurella – thermophilous bacteria utilizing acetate as the electron donor.
Campylobacter – not able to reduce sulphates but reducing sulphur, sulphites,
tiosulphates, nitrates as well as fumaric acid, utilizing acetate as the electron donor.
Pyrodictium – a thermophilous archeon able to utilize diatomic hydrogen as the
electron donor, and elemental sulphur as electron acceptor.
Particular SRB species differ in several features, including: cell shape, mobility,
occurrence, preferred electron donors, complete or incomplete oxidation of organic
compounds, content of GC pairs in DNA, formation of spores, presence of desulphoviridine,
cytochromes, and optimal growth temperature (Gibson, 1990).
Based on rRNA analysis, SRB have been subdivided into four groups:
Till 1984, SRB were considered to be dominated in the natural environment by other
microorganisms. Therefore, SRB were commonly isolated on media containing an easily
accessible carbon source, e.g. lactate, pyruvic acid or ethanol, and sodium sulphate, well
soluble in water, as the electron acceptor (Postgate, 1984). However, the last twenty years
have brought many reports on the fact that SRB are microorganisms that have the ability to
biodegrade a wide spectrum of organic compounds. In due course, SRB began to be isolated
on media containing compounds that are the main pollutants in the environment. Thus, active
SRB communities are isolated from environments polluted by substances that are subject to
treatment. For example, anaerobic SRB communities can be isolated from refinery-
petrochemical sewage, soils polluted by oil products, or other organic sewage, and are then
used in the anaerobic utilization process of these wastes.
Very soluble sulphates such as sodium sulphate were considered the optimal electron
acceptors in the sulphidogenic process. There are also reports indicating the fact that poorly
soluble sulphates such as bassanite (CaSO4 x 0,5H2O), gypsum (CaSO4 x 2H2O), anglesite
(PbSO4), or barite (BaSO4) may also become an easily accessible electron acceptor for SRB
(Karnachuk et al., 2002).
During multiplication of anaerobic microorganism communities, typically two methods
of SRB selection are applied: the ―microcosms‖ method and multiplication on agar medium
(Figure 1). Reproduction of selected microorganism communities should take place in strictly
anaerobic conditions. Therefore, liquid media are often supplemented with compounds
decreasing the redox potential, e.g. cysteine, sodium tioglicolate, or sodium sulphide, as were
as indicators of oxygen level such as resaurine.
58 Dorota Wolicka
what is very difficult to achieve and obliges the application of e.g. neutral gases to decrease
the redox potential of the culture. A different issue is the multiple passaging of the culture,
which is indispensable in the process but at the same time causes oxygenation of the medium.
SRB isolation from agar media is facilitated by the black colouring of the cultures on media
containing iron caused by formation of iron sulphides. Another difficulty comes from the fact
that SRB grow in strict physiological and spatial relationships with other groups of bacteria,
e.g. they form consortia with methanogenic archaea. Decomposition products formed by one
group within the consortium become the substrate for other bacteria groups taking part in the
anaerobic decomposition of organic matter. During SRB identification it often turned out that
strains considered pure, in fact comprised a microorganism community that was not possible
to sub-divide, or after isolation each of the components was not capable of independent growth.
produced mainly by food industry, e.g. dairy plants, sugar plants, fruit and vegetable
processing plants, meat processing plants, breeding plants, and slaughter houses.
Dairy sewage contains high volume of organic compounds and BOD. The main
constituents of this sewage are saccharides, fats and proteins that compose milk. This sewage
undergoes biodegradation easily, after which the environment becomes acidified and
hydrogen sulphide, toxic for most microorganisms, is produced. The sewage cannot be
introduced into the receiver without treatment. The volume of sewage produced by a single
dairy plant reaches several thousands of cubic meters per 24 h. The composition of dairy
sewage resembles highly diluted whole milk, with dissolved lactose and protein (casein) in an
identical proportion as in milk. Based on the analysis of milk composition used in the
processing plant and the production profile, the sewage composition can be assumed. The
volume of sewage formed in dairy plants comprises from 0.5 to 3 times the volume of
processed milk. Typically, 1 g BOD5 in sewage is considered to correspond to 9 g BOD5 in
milk. The BOD5 of sewage reaches 450–5800 mg O2/l (averagely about 1800 mg O2/l).
The total volume of dairy sewage comprises strongly polluted production sewage formed
during flushing of products or washing of vessels, equipment, facilities, as well as slightly
polluted cooling waters (comprising 60–90% of sewage volume). Sewage may also contain
low quantities of flavouring matter, gelling and clarification agents, etc., as well as
disinfectants, cleaning and degreasing agents, and agents removing limescale from utensils.
Sewage reaction varies between 7.0 and 8.8, with the exception of acidic sewage from
production of casein and selected cheeses, which varies between 5.5 and 6.5. Due to the
content of lactose, dairy sewage is easily biodegraded, and its pH falls down to very low
values. The composition of sewage from particular dairy plants may vary considerably
depending on the type of manufactured products (Tab. 2, 3, 4). The table 2 presents the
typical composition of dairy sewage from a plant producing cream, cottage cheese and kefir.
The ballast of dairy sewage is more variable when it encompasses buttermilk, whey and
skimmed milk. The application of some sewage constituents e.g. whey to feed breeding
animals, or buttermilk to produce beverages, may distinctly influence this ballast. Whey
contains about 72% lactose, 10% protein, 0.5% fat, as well as mineral compounds and
vitamins. In the dairy industry it is formed as a by-product during cheese or casein
manufacture. One volume of produced cheese results in almost 10 volumes of whey, which is
a crucial problem in cheese manufacture (Kutera & Talik, 1996).
Table 2. The typical composition of sewage from plant producing milk products.
Casein and lactose are not optimal carbon sources for SRB. Instead, products of their
hydrolysis may become electron donors for SRB, as these bacteria do not produce hydrolytic
enzymes and therefore they take part in the process of organic matter biodegradation at the
level of volatile fatty acids (Mizuno et al., 1998). Due to this fact, utilization of proteins and
disaccharides is a rare case in SRB.
Determinations Content
pH 9.6
Oxidization [mg O2 / 290
dm3]
COD [mg O2 / dm3] 1200
BOD5 [mg O2 / dm3] 515
ether ekstract [mg / dm3] 160
Suspensions [mg / dm3] 480
Determinations Share
Lactose 4.7%
Fats ~3.2%
Proteins ~3%
Mineral salts 0.8%
Water to 100%
Sewage polluted by organic compounds that are not easily biodegraded includes sewage
produced in e.g. oil refinery and petrochemical plants, pulp and paper industry, textile
industry, leather industry, as well as plant and animal utilization plants.
Refinery-petrochemical sewage contains from about 0.3 to 2% of organic compounds
from crude oil, such as hydrocarbons, alcohols, aldehydes, esters, alkali, acids and their salts
from the refining-petrochemical industry, as well as oily pollutants (with a various degree of
emulsification) and tars.
In petrochemical plants sewage is formed in the process of raw material cleaning, as well
as during the manufacture of many chemical products, half-products and usable products
from crude oil and its fractions and from earth gas (most commonly distillation and
rectification processes). Sewage composition depends on the type of production in a given
petrochemical plant. The main pollutants in the sewage are hydrocarbons, alcohols,
aldehydes, phenols, esters, alkali, acids and theirs salts. Other common constituents of
petrochemical sewage include phenol, ethylene, propylene, butadiene, acetone, glycol,
plastics, synthetic rubber, epoxy resins, and surface active agents. Petrochemical sewage is
commonly treated jointly with refinery sewage.
Petroleum processing includes thermal processing comprising distillation, rectification
and cracking. In modern refinery plants processes of further fuel refinement are carried out,
including petrol reforming and hydrorafinery (desulphurization) of diesel oil. The first phase
of crude oil processing comprises removal of gas in gas separators, mechanical pollutants by
62 Dorota Wolicka
water flushing and sedimentation in settler tanks, whereas emulsions are broken in
electromagnetic field, by heating or addition of deemulsifiers. After these initial stages, crude
oil is distilled by separation of hydrocarbons into fractions characterized by similar boiling
temperatures. At normal atmospheric pressure are produced: petrol, heavy petrol, naphtha,
diesel oil and oil fuel. Distillation of fuel oil at lower pressure leads to the production of
diesel oil, light, medium and heavy petrol (paraffin oils).
Hydrocarbon fractions obtained during distillation are refined in order to obtain purified
products. Modern oil refinery methods are based on the adsorption of pollutants or application
of selective dissolvents that dissolve the pollutants not dissolving the oil. The most commonly
applied dissolvents are phenol and furfuryl alcohol.
The content of sewage produced in oil refineries depends on the quality of oil and degree
of its processing, and varies from 10 to 18 cubic meters per tonne of processed crude oil.
Sewage is produced during:
The colour of sewage varies from milk white to dark brown and it has a specific odour
(Table 5). The main pollutants in sewage from refining plants include naphtha, sulphides,
thiols, phenols, fatty and naphthene acids, naphthene sulphoacids, mineral oils, aliphatic and
aromatic hydrocarbons (often chlorinated), aldehydes and alcohols formed during oxidation
of hydrocarbons, and suspensions. Naphtha occurs in sewage in forms of emulsions or surface
films.
Table 5. Characteristics of sewage formed during naphtha refinery.
Inorganic sewage that may be biodegraded during sulphidogenesis results from e.g.
production of artificial fertilizers, mines, manufacture of various inorganic chemical
compounds, metal ore processing plants, and electroplanting plants. Depending on the type of
processing plant, it contains various types of pollutants, including biogenic elements, acids,
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 63
alkali, salts and heavy metals. Toxic inorganic sewage may contain: heavy metals, inorganic
acids, alkali, cyanides, etc. as the main pollutants. Such sewage is formed in metal ore and
sulphur mines, in heat power plants, smelters, electroplanting plants, mineral acid works, and
in explosive factories.
The presence of SRB has been noted in practically all reactors treating various types of
organic sewage, e.g. dairy sewage (Baena, 1998, 1999, 2000; Hernandez, 2000), however
they still play a small role due to the low concentration of sulphates ca. 258 g/m3 (Talik,
Kutera, 1997). The source of sulphates for SRB may be not only easily soluble compounds
such as Na2SO4, but also insoluble mineral phases such as hanebachite (CaSO4), jarosite
(KFe3[OH]6 (SO4)2), anglesite (PbSO4) or barite (BaSO4) (Karnachuk et al., 2003). A good
source of sulphate ions are solid anthropogenic wastes such as phosphogypsum formed
during the production of phosphoric acid.
Addition of sulphates to primarily non-sulphur sewage results in high-sulphur sewage
similar to sewage produced during the manufacture of molasses (2.9 g SO4/L), citric acid
from sugar cane (2.5–4.5 g SO4/L) or wood industry (1–2 g SO4/L) (Colleran et al., 1995).
High concentration of sulphates in sewage hampers treatment using methanogenesis; in turn
they can be purified with application of sulphidogenesis. In reactors treating high-sulphur
sewage, SRB are solely responsible for the final stages of biodegradation of organic
pollutants (Colleran et al., 1995).
Biodegradation of two different industrial wastes including sewage and solid waste seems
both an interesting and indispensable procedure for economic reasons. Biotransformation of
phosphogypsum in various industrial sewages such as refinery-industrial sewage, dairy
sewage or distillery decoctions has already been described (Wolicka & Kowalski, 2006a;
2006b; Wolicka, 2008b). Products of phosphogypsum biotransformation and biodegradation
of organic compounds in stationary cultures were carbonates and/or phosphates. The obtained
results indicate the possibility of obtaining secondary post-culture deposits that can later be
applied e.g. as fertilizers in farming.
Due to the lack of hydrolytic enzymes in most SRB, application of a two-step process of
the treatment of organic sewage with phosphogypsum seems a good solution (Kaufman et al.,
1996; Deswaef et al., 1996) (Figure 2). In the first phase, the biological reactor is colonized
by an assemblage of acidogenic and sulphate reducing bacteria. Acidogenic bacteria are
capable of producing acetate from organic compounds that are not easily accessible to SRB.
In turn, the produced acetate can be a good energy substrate for the second group of bacteria.
In the second stage, the reactor is colonized mainly by SRB that use simple organic
compounds flowing from the first reactor. In both reactors waste gypsum is the source of
sulphates. This method of sulphidogenesis allows for effective phosphogypsum
biodegradation in the organic sewage setting, which can at first contain organic compounds
that are not easily accessible for SRB (Deswaef et al., 1996). The process described herein
can be conducted in bioreactors basing on the flooded deposit with biofilm. A crucial problem
in these methods is the possibility of biofilm overgrowth on grains of filling material, a
process that can be additionally accelerated by sparingly soluble gypsum. The possible
solution may be application of reactors in which the circulating elements are the filling
64 Dorota Wolicka
material grains with biofilms. As a result, excessive biofilm is removed and good contact of
microorganisms with the solution of sewage with phosphogypsum is secured. Due to
continuous circulation, the accumulation of non-reduced calcium sulphate in the biofilm is
simultaneously hampered.
Figure 2. Treatment of organic waste water with phosphogypsum in two stages process.
Reactors with biofilm on solid material (filler), such as plastic, ceramic and glass
particles, sand grains, expanded clay pellets, etc. This type pf bioreactor is
characterized by the fact that biomass is retained in the reactors and sewage can flow
faster than it is assumed from the growth speed of microorganisms.
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 65
Heavy metals are metallic elements with density over 4.5 g/cm3 and atomic mass
exceeding 50. This group comprises such metals as: arsenic, manganese, zinc, chrome, iron,
cadmium, lead, nickel, copper, mercury, cobalt, and molybdenum. Their characteristic feature
is the ability of very long persistence in the natural environment, what is linked with sparing
solubility of some heavy metals chemical compounds and also with large ability to
accumulate in living organisms. Heavy metals can be subdivided into four groups:
1. elements with very high potential hazard to environment, e.g.: cadmium, mercury,
chrome, silver, zinc, gold, antimony, tin, thallium;
2. elements with high potential hazard to environment, e.g.: molybdenum, manganese,
iron;
3. elements with medium potential hazard to environment, such as vanadium, nickel,
cobalt and wolfram;
4. elements with low potential hazard to environment, e.g. zirconium, tantalum,
lanthanum, niobium.
66 Dorota Wolicka
Natural processes that influence the mobility of heavy metals in the environment include:
Due to these processes the primary occurrence of heavy metals is transformed. Migration
of heavy metals as well as their uncontrolled appearance in the environment is greatly
influenced by anthropogenic processes. Industrial activity, e.g. exploitation of metal ores,
transport of ore to processing plant, transformation, final management and utilization of used
products are the main anthropogenic sources of heavy metals. The resulting solid and gaseous
wastes as well as sewage are accumulated in water, soil and atmosphere.
High content of heavy metals can be found in sewage deposits and industrial sewage.
Industrial waste containing heavy metals derive mainly from: smelters, electroplating plants,
tanning, fertilizer, pesticide, dyeworks, textile, and electrochemical industries, from plants
producing batteries, accumulators, catalysts, etc. Copper in sewage is derived from e.g.
metallurgy, dyeworks, and textile industries, and is also emitted during production of
pesticides and fertilizers. Electroplating and paper industries, refineries, and steelworks
supply high content of nickel to the environment, whereas production of batteries and paints,
textile and plastic industries, polymer stabilizer industry, printing and graphic plants deliver
high contents of zinc. Heavy metals are also introduced to water with industrial sewage and
wastes, with effluents from fields or smelter dumps.
Products of sewage treatment are sewage deposits in which the heavy metal concentration
distinctly exceeds that in sewage. The harmful influence of heavy metals on living organisms
and the natural environment is undisputable. Therefore actions should be undertaken to
eliminate or/and restrict their emission, as well as minimize their negative influence.
Industrial plants in which sewage with heavy metals are produced should have installations
for their preliminary treatment before sewage is sent to the receiver. Treatment of isolated
sewage streams is more effective that treatment of mixed sewage.
Sewage containing heavy metals are commonly treated using: chemical (neutralization,
reduction and/or oxidation, precipitation), physical and chemical (sorption, extraction, ion
exchange) and electrochemical methods. The correct method depends on the type of sewage,
their content, phase and concentration of particular particles and required treatment degree.
Recently, biological methods are also more frequently applied (Figure 3); these methods
allow recovery of heavy metals and cause the transformation of toxic cations of heavy metals
into sparingly soluble sulphides, a desired effect particularly in solid waste management
(Ekstrom et al., 2008; Gibert et al., 2002).
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 67
Figure 3. Biological methods to the remove heavy metals from waste water.
Anaerobic organic sewage treatment is based mainly on the activity of several groups of
bacteria: fermentation, acetogens, methanogens, sulphate reducing bacteria and denitrifying
bacteria. Some role is played also by bacteria reducing iron and oxygenated forms of other
metals, but the content of these bacteria is very low.
The most well known and commonly applied method of anaerobic sewage treatment uses
the activity of the first three mentioned microorganism groups. Each microorganism group
partly oxygenates an organic compound to the relevant end products, which are next
assimilated by the next link of the food chain till complete oxygenation (according to the
scheme in Figure 4).
This system of biological sewage treatment in the biodegradation of organic pollutants
begins with the activity of fermentation bacteria, which are responsible for hydrolysis and
fermentation of particular organic compounds representing the main pollutant in the sewage.
Many microorganisms display their ability to fermentation, in course of which various
organic compounds are formed becoming an easily accessible carbon source for SRB.
The product of acetic acid fermentation, carried out by acetogens, is acetic acid. This
reaction is catalyzed by enzymes produced mainly by acetate bacteria, but takes place also
during other biochemical transformations.
Various acetate bacteria species display smaller or larger abilities of further oxidation of
acetic acid according to the following formula:
The ability of lactic acid fermentation occurs in bacteria of the proper lactic acid
fermentation. These bacteria oxidize simple carbohydrates and disaccharides to lactic acid (as
the main product) and to various by-products. Lactic acid fermentation was observed e.g. in
Micrococcus, Escherichia and Microbacterim. This fermentation can be sub-divided into two
types: homofermentation – where lactic acid is the main product and there is a low content of
by-products; and heterofermentation – where more by-products are formed in relation to
lactic acid.
Lactic acid fermentation is a typical case of metabiosis i.e. growth of one group of
microorganisms after another. In the first stage heterofermentation bacteria develop that
acidify the environment and create favourable conditions for the development of bacteria
from proper lactic acid fermentation (homofermentation). Ethanol is also one of the by-
products of homofermentation, resulting from the anaerobic decomposition of carbohydrates
into ethylic alcohol and carbon dioxide.
The products of butyric acid fermentation may be acetic acid, succinic acid and ethanol.
This fermentation is carried out by Clostridium (C. acetobutylicum, C. butylicum), during
which butanol is formed instead of butyric acid, and acetone instead of acetic acid. At first
this fermentation is identical with butyric acid fermentation. However, when the produced
acids decrease the reaction to about 4, fermentation changes course and neutral compounds
such as butanol and acetone are formed instead of acids. There are also bacteria species which
reduce acetone to isopropyl alcohol. It is worth noting that all fermentation products are an
easily accessible carbon source for SRB.
The last stage of biological sewage treatment is the activity of methanogens. Using
hydrogen, carbon dioxide and acetate produced by acetogens, they produce methane. It is
estimated that about 1/3 methane produced is the product of CO2 reduction, and the rest – of
acetate decarboxylation.
The two main advantages of this treatment method is the formation of low quantities of
excessive sludge, difficult to utilize, and the possibility of using the formed biogas as fuel.
The energetic value of biogas increases with the content of methane, but it should be
remembered that the content of methane in biogas is inversely proportional to the
concentration of SO42- in the treated sewage. Sulphates that are present in variable content in
organic sewage do not directly influence the activity of methanogens, but they favour the
selection of SRB, which according to reaction stoichiometry transform SO42- to H2S. H2S has
toxic influence on methanogens, and decreases the quality of biogas.
SRB have been noted in all types of bioreactors purifying sewage in anaerobic
conditions. In such conditions they compete with various bacteria groups for the available
organic compounds at almost all decomposition stages except hydrolysis, because most
microorganisms capable of dissimilative sulphate reduction do not develop hydrolytic
enzymes. The only exception is Archeoglobus fulgidus. In reactors treating high-sulphur
sewage e.g. from pulp and paper, textile, pharmaceutical, metallurgic, paint and varnish, and
plastics industries SRB are solely responsible for the final stages of organic pollutants
decomposition (Colleran et al., 1995).
Some SRB species such as Desulfuromonas acetoxidans may form syntrophic
relationships with photosynthesizing bacteria belonging to green sulphur bacteria
(Chlorobiaceae). Under sun light the phototrophs assimilate carbon dioxide and oxidize
hydrogen sulphide, which is the electron donor, to elemental sulphur that is released from the
cell. In the following stage Desulfuromonas reduces it to hydrogen sulphide, with
70 Dorota Wolicka
simultaneous oxidation of acetate and production of carbon dioxide. In some cases syntrophic
relationships have been noted in other SRB species, e.g. Desulfovibrio. The common
development of these bacteria is an example of a syntrophic relationship, in which substrates
are distributed in two directions The product of the activity of one bacteria group becomes the
substrate for the next group (Figure 5).
As mentioned above, SRB co-occur in all types of bioreactors applied in anaerobic
treatment of sewage along with methanogens. The growth and increase of SRB activity
depends on winning the rivalry with other bacteria groups on each stage of biodegradation.
Competition of various microorganism groups in the presence of sulphates may take place in
several different stages of the biodegradation process (Colleran et al., 1995):
The process of anaerobic sewage treatment using sulphidogenesis is linked with the
necessity to remove hydrogen sulphide produced by SRB. This requirement seems the most
important disadvantage of the method; therefore many reports regarding this topic have been
published (Kobayashi et al., 1983; Khanna et al., 1996; Lee & Kim, 1998; Tichy et al., 1998).
So far, the often proposed solution is chemical (abiotic) oxidation of hydrogen sulphide to
elemental sulphur; this process is, however, very expensive due to the application of catalysts
and energy demanding aeration. Therefore, solutions with application of microorganisms
capable of replacing chemical removal of hydrogen sulphide are sought for. It seems that the
most appropriate microorganisms are photosynthesizing and chemolithoautotrophic bacteria.
Both groups comprise microorganisms capable to oxidation of hydrogen sulphide, although
the metabolism linked with this process is entirely different.
The most important problem during application of microorganisms in removing hydrogen
sulphide is the selection of tribes and physical-chemical conditions producing elemental
sulphur in course of oxidation; the sulphur would then be emitted from the reactor. It is thus
indispensable that the final effect of sulphur transformation in anaerobic high-sulphur sewage
treatment using sulphidogenesis and additional refinery stages would be elemental sulphur
being a potential source in many industrial branches. It poses much less hazard to the natural
environment during stacking than other solid waste containing sulphur, e.g. waste gypsum.
72 Dorota Wolicka
Photosynthesizing sulphur bacteria. This is the largest and probably the most
appropriate group with regard to application in hydrogen sulphide removal that encompasses
microorganisms capable of oxidizing hydrogen sulphide to elemental sulphur (Gemerden,
1986; Eraso & Kaplan, 2001). It comprises a diverse group of microorganisms classified into
many taxonomic units which share the ability to carry out processes depending on presence of
light. Typically, four sub-groups are distinguished: green sulphur bacteria, purple sulphur
bacteria, purple non-sulphur bacteria and the Chloroflexus group comprising non-sulphur
green bacteria. Due to their ability to cumulate elemental sulphur the most appropriate for
application in biotechnology of hydrogen sulphide oxidation are the first two groups.
Green sulphur bacteria conduct photosynthesizing processes in which the electron
sources include reduced sulphur compounds such as sulphides and tiosulphate. These bacteria
contain high amounts of bacteriochlorophyll c, d, e, as well as much lower quantities of
bacteriochlorophyll a (Brock & Madigan, 2006; Baneras et al., 1999). They are assigned to
the order Chlorobiales, which comprises such species as Chlorobium limicola, Chlorobium
limicola thiosulfatophilum, and Pelodictyon sp. They occur in aqueous environs, particularly
in thermally stratified lakes, but can also be found in saline conditions or in highly exceeded
temperatures. Many species of green bacteria are easily isolated from the natural environment
and multiplied in laboratory conditions, what allows potential application in biotechnology.
The most important feature of the group with regard to this fact is the ability to cumulate
elemental sulphur outside the cells, what distinguishes them from another group of
photosynthesising bacteria – purple sulphur bacteria of potential significance in the process.
Microorganisms belonging to purple sulphur bacteria contain high amounts of
bacteriochlorophyll a and have the ability to cumulate sulphur, but only within the cell, with
the exception of the family Ectotiorhodospiraceae that cumulates sulphur outsides the cell
(Eraso & Kaplan, 2001). The basic source of electrons in the photosynthesizing processes
includes hydrogen and hydrogen sulphide, and the stored sulphur may potentially be oxidized
to sulphate. The bacteria belong to the order Chromatiales, and the best known species is
Chromatium okeni.
producing sulphur such as Thiobacillus thioparus, a large number of bacteria species oxidize
hydrogen sulphide directly to sulphates. Secondly, these bacteria often require lower pH in
the environment, and thirdly, the effluent after refinement using sulphidogenesis may contain
organic compounds distinctly hampering the activity of chemolithoautotrophic sulphur
bacteria. Due to this fact, the present application of these bacteria is restricted to attempts of
metal bioleaching and in refinement of mineral sewage containing e.g. tiosulphates.
Attempts to apply green and purple sulphur bacteria in the removal of hydrogen sulphide
are justified in all processes that generate the formation of sewage strongly contaminated by
sulphides, hydrogen sulphide and tiosulphates. Such sewage is formed during gas
desulphurization, in crude oil processing plants, chemical plants, and mostly during treatment
of high-sulphur sewage using aerobic methods. Solutions proposed hitherto are based on the
utilization of photosynthesizing bacteria populations in photoreactors, in which are ensured
conditions favourable for their growth. The most important and practically the only necessity
is the assurance of a light source, what is linked with some cost, but due to the possibility of
using day-light, the cost is distinctly lower than in the case of a chemical process. Lee & Kim
(1998) proposed the application of the so-called optical-fibre bioreactor, to which light is
brought by optical fibres. The bacteria Chlorobium limicola thiosulfatophilum were used in
the bioreactor with successful results. Henshaw et al. (1998) also applied Chlorobium
limicola in a system similar to a chemostat, in a suspended-growth continuous stirred tank
reactor, which was illuminated by a lamp emitting infrared light. The process was highly
effective, with 90% conversion of sulphides in the solution to elemental sulphur. Henshaw et
al. (1999) tested the influence of the material used to construct the bioreactor (mainly various
types of plastics permitting infrared radiation) and showed the lack of significant influence of
the applied material on the growth of Chlorobium limicola. It is worth noting that a rather low
light intensity was applied in this experiment, from 3.4 ∙ 10-3 to 4.7 ∙ 10-3 W m-2.
Application of photosynthesizing bacteria is linked with the important issue of organic
compounds presence in sewage flowing to the reactor. Due to anaerobic conditions SRB may
grow in the reactor; at this stage of refinement this is highly undesirable, influencing the
effective reduction of hydrogen sulphide pollution. Theoretically, the presence of organic
compounds may be unfavourable to the metabolism of photosynthesizing bacteria, although it
seems that they are partly capable to utilize simple organic compounds. This problem is
crucial during attempts of applying photosynthesizing bacteria to remove hydrogen sulphide
from sewage earlier treated during sulphidogenesis. Such sewage may contain also indistinct
amounts of organic matter and may be also the source of SRB. Borkowski & Wolicka (2007a)
applied an assemblage of photosynthesizing bacteria isolated from the natural environment on
a flooded deposit and indicated that yeast extract flowing with synthetic medium may
distinctly inhibit the efficiency of sulphide oxidation. In turn, in Borkowski & Wolicka
(2007b) the synthetic medium flowing onto the flooded deposit colonized by
photosynthesizing bacteria was replaced by filtered influent from a sulphidogenic bioreactor,
in which phosphogypsum with distillery decoctions was treated. In this case, 60%
effectiveness of sulphide content reduction was obtained (from 163 mg L-1 to 70 mg L-1) with
partial conversion to elemental sulphur.
74 Dorota Wolicka
During simultaneous treatment of organic sewage and solid gypsum waste such as
phosphogypsum using anaerobic treatment by sulphidogenesis, it is possible to refine the
sewage already after the main process (Figure 6).
The process should comprise the following stages:
1. The effluent from the anaerobic reactor after sulphidogenesis contains sulphides and
hydrogen sulphide; after primary filtering it is transported to a bioreactor with
photosynthesizing bacteria. The effluent can contain also organic compounds and
SRB, the number of which should distinctly decrease after filtration. The effect is
almost complete reduction of the remaining organic pollutants and elemental sulphur.
2. The solid waste after phosphogypsum may contain high amounts of calcite and
slightly less phosphates and due to this fact may be applied in agrotechnical
activities. However, if the phosphogypsum contained a high content of heavy metals
or if these metals flew with sewage, after utilization using sulphidogenesis the
deposit is enriched in sulphides of heavy metals. Such deposit, beside the fact that it
is very toxic, is also a significant source of metal that can be recovered during
bioleaching by natural or selected communities of chemolithoautotrophic sulphur
bacteria representing e.g. Thiobacillus and Acidithiobacillus.
4. CONCLUSION
At present it is obvious that natural environment protection is a basic and essential task.
Taking into account the great volume of sewage produced in anthropogenic processes by
industrial plants, it is important to find solutions aimed at decreasing the influence of toxic
and dangerous sewage on the natural environment. One of the basic methods of solving the
water quality issue is rational and professional approach to problems linked with the
treatment of various types of organic sewage.
It is commonly known that biological methods using microorganisms can be applied in
the biodegradation of hazardous petrochemical waste. Many groups of anaerobic bacteria are
able to biodegrade various types of organic waste to non-toxic compounds or even to
inorganic compounds after complete mineralization. Biological methods do not require
introduction of chemical compounds to the environment that would negative influence the
biocenosis. Most of all, biological methods take place naturally in the environment, and
Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 75
anthropogenic influence in this case is focused only on stimulating their range and intensity.
Final products of microbiological decomposition include carbon dioxide and water.
Simultaneous biodegradation of several industrial wastes, e.g. organic sewage and solid
waste seems an interesting and optimal solution from the point of economy. Costs linked with
simultaneous biodegradation of two arduous industrial wastes are almost always much lower
than separate treatment processes.
Application of SRB in biological treatment of sewage is becoming popular and is entirely
justified.
ACKNOWLEDGMENTS
I would like to thanks Dr Andrzej Borkowski for his help in preparing of this manuscript.
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Application of Sulphate-Reducing Bacteria in Biological Treatment Wastewaters 81
Chapter 3
ABSTRACT
Disposal of wastewater treatment sludge (WWTS) and drinking-water treatment
sludge (DWTS) is one of the most important environmental issues nowadays. Traditional
options for WWTS and DWTS management, such as landfilling, incineration, etc., are no
longer acceptable because they can cause many environmental problems. Conversion of
WWTS and DWTS into useful resources or materials is of great interest and must be
intensely investigated. To attain this goal, WWTS and DWTS were used as components
for making ceramsite. Part I: SiO2 and Al2O3 were the major acidic oxides in WWTS and
DWTS, so their effect on characteristics of ceramsite was investigated. Results show that
WWTS and DWTS can be utilized as resources for producing ceramsite with optimal
contents of SiO2 and Al2O3 ranging from 14–26% and 22.5–45%, respectively. Bloating
and crystallization in ceramsite above 900 ℃ are caused by the oxidation and
volatilization of inorganic substances. Higher strength ceramsite with less Na-Ca
feldspars and amorphous silica and more densified surfaces can be obtained at
18%≤Al2O3≤26% and 30%≤SiO2≤45%. Part II: Fe2O3 and CaO were the major basic
oxides, so their effect on characteristics of ceramsite was also investigated. The optimal
contents of Fe2O3 and CaO are in the range of 5%–8% and 2.75%–7%, respectively.
Higher strength ceramsite with more complex crystalline phases and fewer pores can be
obtained at 6%≤Fe2O3≤8%. Lower strength ceramsite with more pores and amorphous
phases can be obtained at 5%≤CaO≤7%, which implies that excessive Ca2+ exceeds the
needed ions for producing electrical neutrality of silicate networks. Part III: To
84 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
1. INTRODUCTION
WWTS is a mixture of biosolid generated in the treatment of the organic substances of
municipal sewage along with disease-causing pathogens and inorganic substances such as
sand and metal oxides [1, 2]. It is therefore of great significance to find a proper way to
dispose it to avoid secondary pollution. The generally adopted sludge disposal is landfilling,
but the option takes up valuable space and may generate methane that contributes to the
greenhouse effect [3]. Another commonly used method is thermal treatment [3-6] involving
incineration, gasification, and pyrolysis, which can reduce the leachability of heavy metals in
the obtained materials with dramatically decreased volume of sludge [5]. This method
proposes an alternative waste management technology for sludge disposal, but some of the
final products still have to be deposited in landfills. DWTS mainly consists of organic and
inorganic compounds in solid, liquid and gaseous forms, with variable physical, chemical,
and biological characteristics [7, 8]. DWTS are frequently chemically-treated and dewatered
before disposal in the developed countries [7, 8]. Typical practices of disposal are landfill,
recycling, reutilization [7-9], and coagulants-regeneration. Although various alternative
options for disposal, regeneration and beneficial reuse of DWTS have mainly been explored
in the past decade, the search for cost effective and eco-friendly disposal options has become
an urgent priority due to tighter environmental regulations and declining public acceptance of
other solutions in the developed countries. In the developing countries, the untreated DWTS
are normally discharged into the municipal sewers which directly flow into surface water [7].
Although, this type of disposal may seem inexpensive, it is usually ecologically unfavorable.
However, the traditional options for WWTS and DWTS management are no longer
acceptable because they can cause many environmental problems such as atmospheric
contamination, soil contamination, water contamination, etc. Alternative options need to be
explored in order to solve their handling in a more environmentally sound manner.
Today, some wastes with high contents of water, inorganic/organic matter, and pathogen
agents have been postulated as a precursor of materials that can be successfully employed in
several environmental applications [10-12]. Converting waste into useful products can
alleviate the problems of disposal and offer a new reserve for depleting resources [10-20]. For
example, many inorganic/organic wastes, such as steelworks slag, mining residues, slag,
paper sludge, fly ash, and even municipal solid waste have been widely tested as the
alternatives to produce clay, shale, fuel, glass ceramics, adsorbent, activated carbon, cement,
Utilization of Water and Wastewater Sludge for Production of Lightweight … 85
etc. [14-31]. So, the conversion of WWTS and DWTS into useful resources or materials is of
great interest and must be intensely investigated. Recent research has been carried out on the
reuse of WWTS for production of innovative aggregates or ceramsite, to be both used as
construction materials or filter media [32-34]. Much research has successfully examined the
feasibility of using clay and WWTS to produce the ceramsite [34, 35]. Valuable information
about the chemical speciation of heavy metals in ceramsite and their potential environmental
risks is also obtained. Results show that heavy metals are properly stabilized in ceramsite and
cannot be easily released into the environment again to cause secondary pollution [36, 37].
The resulting ceramsite is made with WWTS and clay. To avoid more consumption of clay
and protect the earth‘s surface environment, searching for other materials to replace clay in
ceramsite production needs to be encouraged to achieve the sustainable development of
natural resources. DWTS is one of the best substitutes for clay because its major components
are similar to those of clay. From the viewpoint of natural resource recovery and
conservation, utilization of DWTS as a substitute for clay for production of ceramsite is of
great interest and significance.
In our studies, WWTS and DWTS, the latter being often rich in amorphous Fe or Al
oxides because of Fe or Al salts used for coagulation of source water to remove turbidity and
taste and to speed sedimentation [7-9], have successfully been sintered at temperatures about
1000 ℃ to produce ceramsite. (The obtained parameters for production of ceramsite are
shown as follows: DWTS/WWTS=45/55, water glass/(DWTS+WWTS)=20%, sintering
temperature=1000 ℃, sintering time=35min), which has the potential of cost reduction of
sludge treatment (ceramsite can be sold as a commercial product) [38]. The aim of sludge-
ceramsite production is to form new and useful, less soluble, and more geo-chemically stable
products. The significance of utilization of DWTS and WWTS as components for sludge-
ceramsite production are the waste recycling and re-utilization, as well as the reduction of
waste volume and the total destruction of pathogenic agents and organic pollutants. This kind
of conversion can revolutionize handling of such sludge types allowing them to be reused as
low-cost raw materials, rather than as waste requiring costly disposal.
The main concerns in this chapter are (I) whether the impact of the specific and important
constituents, such as acidic oxides (SiO2 and Al2O3) and basic oxides (Fe2O3 and CaO), on
characteristics of ceramsite is significant; (II) whether it is safe to use ceramsite made from
sludge containing heavy metals, especially when heavy metals are not separated but stabilized
in the product; and (III) whether physical and chemical changes in the environment over time
will result in an unacceptable increase in heavy metal leaching. By considering the above
issues, the specific objectives are established and shown as follows:
Part I: The major acidic oxides, such as SiO2 and Al2O3 (Al2O3 is considered as acidic
oxide), in DWTS and WWTS may strongly affect the bloating behavior and crystal formation
of the ceramsite during the heat treatment process. To validate this hypothesis, the present
study has been conducted (i) to utilize DWTS and WWTS for production of ceramsite; (ii) to
investigate the effect of SiO2 and Al2O3 on the physical characteristics (bulk density, particle
density, water adsorption, and porosity) of ceramsite, (iii) to characterize the ceramsite within
the optimal content ranges of SiO2 and Al2O3 by thermal analysis, morphological structures
analysis, XRD (X-ray diffraction) and compressive strength, and (iv) to analyze the sintering
mechanisms, as well as to establish effective parameters for evaluation.
86 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
Part II: It was found that the characteristics of sintered ceramsite might be strongly
influenced by the basic oxides, such as Fe2O3 and CaO, which was of great importance for the
sintering process to gain more crystallization phases or liquid phases [21, 38]. The required
sintering temperature for production of ceramsite might be lowered by higher content of the
basic oxides in the raw materials. The formation of pores and thermal bloating might be also
influenced by the basic oxides, which could activate the carbon phase to decompose to release
water and gases. The study in this part attempts to address this question by investigating the
following: (i) to investigate the effect of Fe2O3 and CaO (basic oxides) on the characteristics
of ceramsite; (ii) to characterize the ceramsite by thermal analysis (DT-TGA), XRD,
morphological structures analyses, and compressive strength measurements and (iii) to
analyze the sintering mechanisms, which are useful in optimizing the sintering processes for
production of ceramsite.
Part III: Heavy metals leachability, binding capacities, and binding strengths in ceramsite
may be governed by the major parameters such as sintering temperature, and in conjunction
with the external conditions (pH and oxidative condition) [14, 36, 37]. The aim of this part is
to establish effective parameters for evaluation of heavy metals stabilized in ceramsite and to
analyze the binding mechanisms. Thus, laboratory experiments for evaluating short-term and
long-term durability of the heavy metal compounds solidified in ceramsite and theoretical
analysis are thoroughly evaluated. Specific objectives of this study are to determine the effect
of sintering temperature (from 900 ℃to 1100 ℃at 25 ℃intervals), pH (from 1 to 12 at 1 unit
intervals), and H2O2 concentration (0.1, 0.5, 1, 1.5, 3, and 5 mol L-1) on the stabilization of
heavy metals (Cd, Cr, Cu, and Pb) in ceramsite, and to demonstrate the leaching behaviors of
heavy metals in it. Results will allow the fundamental understanding and quantitative
description of the stabilization of heavy metals in sludge-ceramsite. Furthermore, the
evaluation of the toxicity of so stabilized/solidified heavy metals in the ceramsite is helpful to
optimize the ceramsite-making processes, reduce its toxicity, and meet environmental safety
requirements.
The WWTS used in this study was obtained from the Wen–chang Wastewater Treatment
Plant, Harbin, China. The dewatering of WWTS was conducted by using a belt filter press,
and cationic polymeric flocculants were used for the flocculation and dewatering of the
activated sludge. The sludge cake generated from the activated sludge process is
approximately 1.6×105 kg d–1 in wet weight with 24% solids, which is directly landfilled.
DWTS were collected from the chemical coagulation/flocculation unit of the 3rd drinking-
water treatment plant in Harbin, China. The 3rd plant uses a conventional process with
aluminum sulfate [Al2(SO4)3] as the primary coagulant and a small amount of activated silicic
acid with no pH adjustment. Treatment includes mechanical lattice flocculation basins,
settling basins, and fast filters. Sludge and backwash water are not discharged simultaneously.
The DWTS and WWTS were treated by air–dry method and were ground into sizes below
100 μm that were sufficiently fine to be mixed homogeneously. DWTS, WWTS, and water
Utilization of Water and Wastewater Sludge for Production of Lightweight … 87
component analyses
SiO2 Al2O3 Fe2O3 CaO MgO P2O5 K2O Others Carbonaceous Matter
16.28 6.35 5.15 4.10 1.67 1.65 1.12 <0.89 <62.90
element analyses
Zn Fe Mn Si Cu Ca Mg O Cr
0.13 3.81 0.14 8.26 0.03 3.18 1.02 48.83 0.03
P Na K Al Ni C S N Others
0.84 0.34 0.95 3.36 0.10 27.41 0.04 1.03 <0.50
component analyses
SiO2 Al2O3 Fe2O3 TiO2 K2O MgO CaO Others Carbonaceous Matter
64.89 24.95 2.50 1.20 1.20 0.50 0.50 <0.5 <3.60
elements analyse
Zn Fe Mn Si Cu Ca Mg O Cr
0.06 1.83 0.08 32.62 0.02 0.37 0.31 44.56 0.02
P Na K Al Ti C Ba Zr Others
0. 11 0.26 0.98 15.81 0.65 1.88 0.03 0.01 <0.40
88 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
component analyses
SiO2 Al2O3 Fe2O3 CaO MgO K2O Na2O Others Carbonaceous Matter
40.61 27.36 6.99 2.62 1.89 1.28 1.05 <1.0 <17.20
element analyses
Zn Fe Mn Si Zr Ca Mg Sr Cr
0.02 4.47 0.57 20.55 0.02 1.95 1.10 0.02 0.01
Cl Na K Al Rb C Ti O Others
0.11 0.74 1.03 14.16 0.01 6.52 0.16 48.06 <0.50
NO Cd Cr Cu Pb
Minimum 1 100 100 50
Medium 25 500 250 500
Maximum 50 1000 500 1000
In leaching tests, the reference sludge samples were made with heavy metals by adding
metal solutions [Cd(NO3)2, K2CrO4, Pb(NO3)2, and CuSO4 were of the analytical grade] into
the dried WWTS, mixing and allowing them to react for 30 days. The contents of Cd, Cr, Cu
and Pb were designed according to the basic data obtained through analysis of activated
sludge at different places in China as shown in Table 4. The synthetic metal solution was
prepared by dissolving 0.05 g L–1 of Cd2+, 0.1 g L–1 of Cr6+, 0.1 g L–1 of Pb2+ and 0.5 g L–1 of
Cu2+ in deionized water. The simulated content of heavy metals was prepared by adding the
tested heavy metal compounds into WWTS. The content of heavy metals added to the WWTS
are shown in Table 5. The ceramsite prepared for leaching was made with this WWTS.
2.2. Methods
The WWTS containing heavy metals and DWTS was treated by air-dry method and
was ground into sizes below 100 μm that are sufficiently fine to be homogeneously
mixed. The ceramsite for determination of the stabilization of Cd, Cr, Cu and Pb was
Utilization of Water and Wastewater Sludge for Production of Lightweight … 89
made of DWTS, WWTS, and sodium silicate. The raw materials were mixed and
pelletized to particle sizes of 5–8 mm and left at a room-temperature of about 20 ℃ for
a few days (about 3 days) and then the samples were dried at 110 ℃ in a DHG-9070A
blast roaster (China) for 24h. The heating of samples started at 20 ℃, heated at a rate of
8 ℃/min in a SX 2 -10-12 muffle furnace (China), and the samples were soaked at 200 ℃,
600 ℃, and 800 ℃ for a duration of 10min and at 1000 ℃ for a duration of 35min, and
then these samples were naturally cooled until they reached room temperature .
Bulk density which includes all voids and spaces in the volume, particle density
which is also the apparent specific gravity of the aggregates includes all intraparticle
voids, water absorption determined from the weight differences between the sintered
and water saturated samples (immersed in water for 1h), and porosity ((1–bulk
density/particle density) ×100%) were analyzed [34]. To achieve statistical soundness,
at least three replicates were carried out for each sample. Thermal behaviors of samples
were examined by thermodifferential and thermogravimetric analyses (DT–TG) using a
ZRY–2P simultaneous DT–TG analyzer (China) while the samples were heated at a rate
of 8 ℃/min from 20 ℃to 1080 ℃ in air. Samples weighed from 4 to 10 mg in mass, and
they were put into a Pt–Rh crucible with 20 taps. All curves were evaluated using the
TA–instruments software. The second derivative differential thermal curve was used
for determination of peak temperature. Scanning electron microscope (SEM) and
energy dispersive spectrum (EDS) analyses were conducted by using S–570 scanning
electron microscope and TN–5502 X–ray energy dispersive spectrum (Japan).
Compressive strength of ceramsite was analyzed by using an INSTRON 5569 automatic
material testing machine (USA) while the sintered ceramsite with diameter of 6–8 mm
was placed vertically on the platform of the press and was pressed at a crosshead speed
of 0.5 mm min –1 until it was crushed.
Toxicity of ceramsite samples was determined by using a revised method derived
from toxicity characteristic leaching procedure (TCLP), a standard method used to
determine waste leaching toxicity, updated on the basis of hazardous waste extraction
procedure (EP) by USEPA [36]. By using this method, the leaching test was conducted
with the solution prepared at a liquid–solid ratio of 1L/200g, and stirred at 110rpm for
24h or 30d. To achieve statistical soundness, at least three experiments were carried out
with each sample. The supernatant was analyzed using Perkin–elmer Optima 5300DV
Inductively Coupled Plasma Atomic Emission Spectrometer (ICP–AES, U.S.A). Total
contents of heavy metals in WWTS or sintered ceramsite were extracted by acid
digestion (using HNO 3 /HClO4 /HF) according to USEPA SW3050 and were examined
by ICP–AES. Ceramsite were ground in a small agate mortar and XRD patterns of
powder ceramsite were recorded on a D/max–γ β X–ray diffractometer with 50mA and
40Kv, Cu Kα radiation (Japan). XRD analyses were conducted by using an XRD
pattern database (International Centre for Diffraction Data, ICDD) and the samples
were scanned for 2ζ ranging from 10 to 90°. Major components of raw materials and
ceramsite containing heavy metals were analyzed by using Philips PW 4400 XR
spectrometer (X–ray fluorescence–XRF, Netherlands).
90 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
Figure 1A shows that at a SiO2 content in the raw materials ≤ 30%, the bulk density
gradually increases from 510 kg/m3 to 783 kg/m3, while the water adsorption and porosity
gradually decrease from 34.9% to 6.9% and from 50.6% to 36.6%, respectively; at a SiO2
content is greater than 30%, the bulk density gradually decreases from 728 kg/m3 to 600 kg/m3,
and the water adsorption and porosity gradually increase from 12.4% to 25.2% and from 40.9%
to 43.2%, respectively. The maximum particle density (1260kg/m3) is obtained when the
content of SiO2 in the raw materials is 27.2%. During the sintering process, the Si4+ is solidified
in tectosilicate with network tetrahedron of SiO44- (Si–O–Si). The raw materials with proper
contents of SiO2 can enhance the formation of liquid phase at 1000 ℃, which encloses the solid
particles and packs the pores in the solid particles. The pores of ceramsite decrease and the
binding forces between the solid particles is improved through the reaction described above that
also called capillary action (the interaction between contacting surfaces of a liquid and a solid
that distorts the liquid surface from a planar shape). Therefore, the variation of SiO2 contents
can either reduce or enhance the liquid phase in the ceramsite bodies and make the ceramsite
bodies denser or more expanded at 1000 ℃. It is concluded from Figure 1A that 22.5%-45.0%
can be considered as the optimal range of SiO2 contents for production of ceramsite.
During the sintering process, Al2O3 can be considered as a skeleton material in the
sintered ceramsite. Al2O3 can react with other components to form silicate mineral groups
(such as anorthite-CaO·Al2O3·2SiO2) with relative lower eutectic points around 1000 ℃ and
the reaction effectively lower the sintering point of the materials and enhance the formation of
liquid phase [39, 40]. Hence, proper quantities of Al2O3 are beneficial to the liquid-phase-
sintering at 1000 ℃and improve the characteristics of ceramsite. As shown in Figure 1B, the
maximum bulk density (723.9kg/m3) and minimum porosity (42.6%) are obtained when the
content of Al2O3 is 15.8%; the maximum particle density (1310kg/m3) and minimum water
adsorption (13.6%) are obtained when the content of Al2O3 is 18%. Higher water absorption
and lower density of ceramsite (5%≤Al2O3<14% and 26%<Al2O3≤30%) is caused by the
more intraparticle voids, which reduce the strength of ceramsite and increase the cracking and
bulging of the ceramsite [32]. It is therefore concluded that 14%-26% can be considered as
the optimal range of Al2O3 contents in the mixtures for production of ceramsite.
It should be noted from the above results that the optimal contents ranges of SiO2 and
Al2O3 are quite wide in this study. The reasons for selecting these quite wide ranges of SiO2
and Al2O3 contents are as follows: (1) the eutectic point of raw materials is not dramatically
influenced by variation of SiO2 and Al2O3 in the optimal contents ranges; (2) both SiO2 and
Al2O3 are acidic oxides in raw materials and the Si4+ in the network tetrahedron can be
substituted by Al3+ (to some extent, the effect of acidic oxides on the characteristics of
ceramsite is similar to each other); and (3) the physical characteristics of ceramsite in these
ranges are more desirable. The following tests (thermal analysis, X-ray diffraction (XRD),
morphological structures analyses, and compressive strength measurements) were conducted
with ceramsite within each optimal contents range of the tested oxides (SiO2 and Al2O3).
Utilization of Water and Wastewater Sludge for Production of Lightweight … 91
1300
50
A
1200
45
1100 40
Density (kg m )
Percentage (%)
-3
1000 35
Bulk density Particle density
900 Water adsorption Porosity 30
800 25
700 20
600 15
10
500
5
400
15 20 25 30 35 40 45 50 55 60
SiO2 content (%)
1400 55
B
1300 50
1200
45
Density (kg m )
-3
1100
Percentage (%)
40
1000 Bulk density Particle density
Water adsorption Porosity 35
900
30
800
700 25
600 20
500 15
400 10
4 6 8 10 12 14 16 18 20 22 24 26 28 30
Al2O3 content (%)
Figure 1. Effect of SiO2 and Al2O3 contents on the physical characteristics of ceramsite (■ bulk density,
particle density, ▲ water adsorption, × porosity).
The rates of weight loss of the raw materials subjected to sintering were examined at
different burning temperatures with the use of a ZRY-2P simultaneous DT-TG analyzer.
These tests were conducted in dry air atmosphere to observe the weight loss undergone by the
ceramsite samples during sintering. The peaks-valleys on the differential thermal (DT) curves
presented the endothermic/exothermic changes during the heating process. TG and DT plots
of the samples are closely interrelated.
Figure 2A demonstrates that there is little difference in DT-TG curves for mixtures with
SiO2 contents of 22.5%, 30%, and 45%. The temperature ranges for significant weight loss
for mixtures with SiO2 contents of 22.5%, 30%, and 45% are 235.5-521.4 ℃, 233-520 ℃, and
233.8-522.1 ℃, respectively. The high values of weight loss in these temperature ranges
indicate the release of the structural water and a significant amount of organic materials. The
maximum slope of weight loss for mixtures with SiO2 contents of 22.5%, 30%, and 45% are
demonstrated at 285.5 ℃, 290.6 ℃, and 297.2 ℃. For the 3 mixtures, there are three reaction
92 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
stages in 320-355 ℃, 475-600 ℃, and above 900 ℃ as distinguished by the three distinct
endothermic changes over the temperature range. The endothermic reactions observed for the
3 samples above 900 ℃ have been correlated with the formation of a lattice that means a
transition from amorphous to a crystalline phase, which in this case implies the crystallization
of silicate minerals.
8
A
7
TG (mg)
3
0
DTA (μV)
-5
-10
SiO2=22.5%
SiO2=30%
-15 SiO2=45%
-20
-25
0 200 400 600 800 1000
Temperature (℃)
9
B
8
7
TG (mg)
6
5
4
3
2
10
5
0
DTA (μV)
-5 Al2O3=14%
-10 Al2O3=18%
-15 Al2O3=26%
-20
-25
-30
0 200 400 600 800 1000
Temperature (℃)
Figure 2. DT and TG analyses for (A) ceramsite with SiO2 contents of 22.5%, 30%, and 45%, and (B)
ceramsite with Al2O3 contents of 14%, 18%, and 26%.
Utilization of Water and Wastewater Sludge for Production of Lightweight … 93
KQ
K KEA
Q
Q SiO2=45%
A K
K KE Q Q QK
Q Q
K EKA
SiO2=30%
A K K KE Q
Q QK
S Q Q Q
Q
Relative intensity
EKA K
KES SiO2=22.5%
KAE K S QK
Q Q
Q
KQ
K
EKA Q
KES
A Al2O3=26%
K
S K QQ Q Q Q K
EKA
KA K K KE
Q Al2O3=18%
Q Q
Q K
EKA
KA K K
K Al2O3=14%
Q Q
Q K
20 30 40 50 60 70
2theta
Figure 3. Effect of SiO2 and Al2O3 content on XRD patterns of ceramsite. Band labeling: A, albite-
anorthite; E, enstatite; K, kyanite; Q, quartz; S, sillimanite.
As shown in Figure 2B, there are two exothermic peaks for the mixtures with Al2O3
contents of 14%: the peak at 332.6 ℃is due to the removal of absorbed water and the burning
of carbon on the surface of ceramsite, and that at 447.7 ℃ corresponds to removal of
structural water and carbon in ceramsite bodies. The trend of DT curve for the mixtures with
Al2O3 contents of 18% is similar to that of the mixtures with Al2O3 contents of 26%. The
continuous endothermic changes for the 3 mixtures begin to occur at 450 ℃, which indicate
the endothermic reactions are caused by the gradual transformation of crystalline phases
(silicate mineral groups) with lower weight loss as shown in TG curves. These results suggest
that the exothermic reaction below 450 ℃ may give ceramsite products with reduced
crystallinity and the endothermic reactions above 450 ℃ are slowly conducted that may give
the crystals with enhanced stability and strength.
However, the composition of raw materials in the optimal ranges of SiO2 and Al2O3 does
not dramatically vary, resulting in a predictable characteristic of thermokinetics. The
thermokinetics in the actual sintering process can not be properly represented in the DT–TG
test, because the shape of ceramsite is spherical and the release of substances may firstly occur
on ceramsite surface and then occur in ceramsite interior. Because the combustible organic
matters and some of the inorganic matters have to completely release, the DT–TG test still
reveals the intrinsic thermokinetics for ceramsite sintering. Results of DT–TG test reinforce the
basis of sintering temperature profile adopted in this study (as demonstrated in section 2.2).
94 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
In this study, XRD analyses were applied to obtain mineral compositions of powder
ceramsite specimens by using an XRD pattern database (International Centre for Diffraction
Data, ICDD). The samples were scanned for 2ζ ranging from 10 to 90°. Using X-ray
diffraction, it can be seen from Figure 3 that the changes in the degree of crystallinity of the
ceramsite, the formation of new phases, and transformation of some phases.
Kyanite (Al2SiO5), quartz (SiO2), and Na-Ca feldspars [albite-Na(AlSi3O8), anorthite-
Ca(Al2Si2O8)] are the main crystalline phases of the ceramsite with SiO2 contents of 22.5%
and a small amount of sillimanite [(Al1.98Fe0.2) SiO5] and enstatite [Mg2(Si2O6)] are also noted
in the patterns. The formation of Na-Ca feldspars, sillimanite, and enstatite in ceramsite is
caused by the presence of flux such as Fe2O3, CaO, and MgO in raw materials and the solid-
liquid-phase reactions during the sintering process. Kyanite, quartz, and Na-Ca feldspars are
the main crystalline phases of the ceramsite with SiO2 contents of 30%. The major
mineralogical compositions found in ceramsite with SiO2 contents of 45% are quartz, kyanite,
and Na-Ca feldspars. It should be noted that the broad intensity is attributed to the amorphous
silica (between 15° and 30°) and other amorphous phases [39]. The degree of crystallinity of
the ceramsite increases as the SiO2 contents increase from 22.5% to 45%.
Ceramsite with Al2O3 contents of 14% and 18% consists of both amorphous and
crystalline material with a little amorphous material mostly due to silica and the crystalline
peaks attributable to kyanite and quartz with small amounts of Na-Ca feldspars and enstatite.
Quartz, kyanite, and Na-Ca feldspars with small amounts of sillimanite and enstatite are the
identified crystalline phases of the ceramsite with Al2O3 contents of 26%. The main
crystalline phases of the 3 ceramsite are similar to each other. During the sintering process,
the Al3+ ion may substitute for the Si4+ in a network tetrahedron, contributing to the stability
of the network. Therefore, Al2O3 can enter the silica network as AlO44- tetrahedra to replace
some of the SiO44- groups; however, since the necessary +4 for the tetrahedra is replaced by
the +3 valence of Al, alkali cations must supply the necessary other electrons to produce
electrical neutrality (41). Therefore, the presence of Al2O3 in ceramsite results in the
formation of numerous crystalline phases (multi-peak pattern) from which aluminosilicates
with iron, calcium, magnesium, and sodium have a high probability to exist. The results
suggest that Al2O3 plays a significant role in the formation process of crystals, but the major
crystalline phases do not dramatically change under the variation of Al2O3 contents.
To understand more about the surface morphology and crystalline phases of ceramsite
bodies, the samples were gilt with Au and SEM analyses were conducted. The
microstructures of the samples were observed using a SEM (S-570 scanning electron
microscope) and the crystalline structures are shown in Figure 4.
The SEM observations show that the porous structures become more compact due to the
increasing SiO2 contents. The observations clearly show the particulate nature of the crystals
in the ceramsite with SiO2 contents of 22.5%. Sintered crystals bonding is evident by
cohesive necks growing at the particle contact points (Figure 4A1). As seen in ceramsite with
SiO2 contents of 30%, there are abundant small pores with thin boundary and big crack with
no boundary present on the surface, which are caused by release of gases (Figure 4A2). The
Utilization of Water and Wastewater Sludge for Production of Lightweight … 95
water absorption of ceramsite with SiO2 contents of 30% remains relatively low despite the
more pores, because the gas produced can not create sufficient voids in the ceramsite bodies.
Ceramsite with SiO2 contents of 45% clearly shows the advance in densification and neck
growth between the particles. However, in Figure 4A3 the ceramsite samples show clear neck
growth between the particles but that the particle size is much greater than that of the other
two ceramsite, suggesting the occurrence of melted neck growth. The slight expansion of the
ceramsite with SiO2 contents of 45% that occurs when sintered at 1000 ℃ is clearly
associated with the formation of a significant volume of approximately slit-shaped pores.
It can be seen from Figure 4B1 that the surface of the ceramsite with Al2O3 contents of
14% is rough and with a few pores. The micrograph of ceramsite with Al2O3 contents of 18%
clearly shows that some pores (6.0μm<pore size<10.0μm) are irregularly distributed in the
microstructure and the pores become bigger by comparing with the others (Figure 4B2).
Melting phenomena are also observed on the crystalline surface of ceramsite. The release of
gases and melting of raw materials allow the big pores to form and the water absorption of
ceramsite with Al2O3 contents of 18% remains relatively low despite the bigger pores,
because the impervious skin layer of the pellets restricts water ingress. The microstructure of
ceramsite with Al2O3 contents of 18% is also related to the lower viscosity of the liquid phase
produced at the temperature of 1000 ℃and the consequent improvement on the densification
during the natural cooling process. The irregular porous structures with a lot of small pores
and agglomerated crystallizations are observed and the microstructure of ceramsite with
Al2O3 contents of 26% confirms the formation of several indistinct crystalline phases, as
shown in Figure 4B3. The above results indicate that more densified and lower porous
ceramsite can be obtained as Al2O3 contents≥18%.
Figure 4. Scanning photomicrographs for ceramsite with SiO2 contents of (A1) 22.5%, (A2) 30%, and
(A3) 45%, and ceramsite with Al2O3 contents of (B1) 14%, (B2) 18%, and (B3) 26%.
96 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
18
A
16 SiO2 content=22.5%
14
12
10
8
6
4
2
0
-2
0.00 0.05 0.10 0.15 0.20 0.25 0.30 0.35 0.40
Compression deformation (mm)
Figure 5. Compressive strength analyses for (A) ceramsite with SiO2 contents of 22.5%, 30%, and 45%,
and (B) ceramsite with Al2O3 contents of 14%, 18%, and 26%.
The compressive strength is the maximum compressing force a sinter can withstand
before it breaks [32, 33, 42]. The sintered ceramsite with diameter of 6-8 mm was placed
vertically on the platform of the press (an INSTRON 5569 automatic material testing machine
made by INSTRON, USA) and was pressed at a crosshead speed of 0.5 mm min-1 until it was
crushed. The compressive strength (N mm-2) of the sintered ceramsite was the compressing
force (N) divided by the pressed area (mm2). The compressive strength results were the
average values of three tests for each composition. The Chinese National Bureau of Standards
(CNBS) requires a minimum compressive strength value of 7.50 MPa (1 MPa=1 N mm-2) for
the application of sintered lightweight ceramsite (bulk density< 900 kg m-3) in civil
engineering.
In this study, results in Figure 5A show that the maximum compressive strength of
ceramsite increases from 13.87 to 14.55MPa (above CNBS value) as the SiO2 contents
increase. The densified and melting of ceramsite surface are probably the main reasons for the
increase in the compressive strength of the ceramsite with higher SiO2 contents (SiO2
Utilization of Water and Wastewater Sludge for Production of Lightweight … 97
3.2. Part II: Effect of Basic Oxides (Fe2O3 and Cao) on the Characteristics of
Ceramsite
As shown in Figure 6A, bulk density gradually increases from 497 kg/m3 to 780 kg/m3 as
the contents of Fe2O3 increases from 3% to 16%; the maximum particle density
(1815.8kg/m3), the minimum water adsorption (8.4%), and the maximum porosity (58.6%)
are obtained when the content of Fe2O3 is 10%. The liquid phases can be apparently increased
by increasing the contents of Fe2O3 in the raw materials, which enhance the formation of FeO
with higher viscosity and lower melting point ( 3Fe2O3 + CO >820℃ 2Fe3O4 + CO2;
Fe3O4 + CO >820℃ 3FeO + CO2 ) [45]. But, it should be noted that a sinter with higher viscosity
is not good for the release of gases and the formation of pores. 5%-8% is therefore selected as
the optimal content range of Fe2O3 for production of ceramsite.
Ceramsite has higher bulk density and particle density as the CaO contents are in range of
2.75%-7%, indicating the presence of less intraparticle voids in ceramsite bodies, hence
obtaining a lower water adsorption and porosity (as shown in Figure 6B). Higher water
adsorption and porosity of ceramsite with lower density may affect the compressive strength
and chemical stability of the lightweight ceramsite [12]. 2.75%-7% is selected as the optimal
content range of CaO for production of ceramsite.
98 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
60
1800 A
54
1650
48
1500
Density (kg m )
Percentage (%)
-3
42
1350
1200 Bulk density Particle density 36
Water adsorption Porosity 30
1050
900 24
750 18
600 12
450 6
3 4 5 6 7 8 9 10 11 12 13 14 15 16
Fe2O3 content (%)
1400 60
1300 B 55
1200 50
Percentage (%)
Density (kg m )
-3
1100 45
1000 40
900 Bulk density Particle density 35
800 Water adsorption Porosity
30
700 25
600
20
500
15
400
2 3 4 5 6 7 8 9 10
CaO content (%)
Figure 6. Effect of oxides contents on the physical characteristics of ceramsite (A-Fe2O3 and B-CaO).
(■ bulk density, □ particle density, ▲ water adsorption, × porosity).
The following tests (thermal analysis, XRD, morphological structures analyses, and
compressive strength measurements) were conducted with ceramsite within each optimal
contents range of the tested oxides (6%≤Fe2O3≤8% and 2.75%≤CaO≤7%).
Two exothermic peaks of the mixtures with Fe2O3 contents of 5% are detected at 337.5 ℃
and 473.2 ℃. Two exothermic peaks of the mixtures with Fe2O3 contents of 6% and one
exothermic peak of the mixtures with Fe2O3 contents of 8% are detected at 336.7 ℃, 462.5 ℃,
and 336.7 ℃, respectively. The intensity of exothermic peaks between 400-500 ℃ decreases
as Fe2O3 contents increases and the peak disappears when Fe2O3 contents reach 8%. For each
test, the detected exothermic peaks with significant weight loss below 500 ℃ are caused by
the release of structural water and mixed gases (CO2, CO, SO2, etc.) (Figure 7A).
Endothermic changes are observed and little substances are vaporized from the 3 mixtures
above 900 ℃, which indicate the endothermic reaction is caused by the transformation of
crystalline phases. Increase of Fe2O3 contents in the mixtures will profoundly influence the
thermal properties and lead to more gas produced under the reaction of Fe2O3 at temperatures
>820 ℃[45].
Utilization of Water and Wastewater Sludge for Production of Lightweight … 99
9 A
8
TG (mg)
7
6
5
4
DTA (μV)
-5
-10
-15
Fe2O3=5%
-20 Fe2O3=6%
-25 Fe2O3=8%
-30
-35
0 200 400 600 800 1000
Temperature (℃)
8 B
7
TG (mg)
6
5
4
3
6
0
DTA (μV)
-6
CaO=2.75%
-12
CaO=5%
-18 CaO=7%
-24
-30
0 200 400 600 800 1000
Temperature (℃)
Figure 7. Effect of oxide contents on the thermal properties of ceramsite (A-Fe2O3 and B-CaO).
As shown in Figure 7B, it is interesting that there are two endothermic valleys at 280.6 ℃
and 279 ℃ for the mixtures with CaO contents of 5% and 7%, respectively. These
endothermic valleys may partly attribute to the removal of water from hydrated products,
which is likely to include most of calcium silicate hydrate (C–S–H) formed by the reaction of
CaO, water, and silicate. From the results, it is clear that hydrated products formed at higher
CaO contents are one of the first decomposed phases during the heating. There is little
difference in DT-TG curves for the 3 mixtures above 300 ℃. Weak weight loss for 3 mixtures
are detected above 500 ℃ in TG analyses, which indicate that there is little volatilization of
substances and the endothermic changes are caused by the oxidation of mostly inorganic
substances and the transformation of crystalline phases [35].
The main crystalline phases of the 3 ceramsite are similar to each other and Quartz
(SiO2), kyanite (Al2SiO5), and Na-Ca feldspars [albite- Na(AlSi3O8), anorthite- Ca(Al2Si2O8)]
are the main crystalline phases of the 3 ceramsite (Figure 8). The formation of these
100 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
It is observed that the ceramsite with Fe2O3 contents of 5% presents rough and densified
surface and has some irregularly distributed approximately slit-shaped pores. It can be
explained that the pores form as the residual glassy phase viscosity falls to a level when gas-
forming inorganic decomposition reactions can produce the observed pores [44]. Ceramsite
with Fe2O3 contents of 6% and 8% present rough surfaces with a few small pores. These are
believed to be a result of the softening of the glassy phase present in the ceramsite, along with
a simultaneous incomplete release of gas at 1000 ℃. Comparing the images in Figure 9 (A1-
A3) reveals that the higher the Fe2O3 contents, the smaller and fewer the pores. The SEM
observations for the 3 ceramsite are in general agreement with the trends shown by the
physical data (Figure 6A). The above results indicate that ceramsite has better crystallization
and melting of the bodies as 6%≤Fe2O3≤8%.
As shown in Figure 9B1, ceramsite with CaO contents of 2.75% presents rough and
densified surface and has a few irregularly distributed pores. Crystals with many-sided
morphology and a few small pores are observed on the surface of ceramsite with CaO
contents of 5%. For ceramsite with CaO contents of 2.75% and 5%, melting phenomena are
observed on both of the crystalline surfaces. SEM image of ceramsite with CaO contents of
7% suggests that the majority of the coarse and fine minerals are located in close contact with
the silicates matrix. The relative higher alkaline earth oxide content (CaO), present in the
ceramsite will act as a fluxing agent during the sintering process [46, 47]. Besides better
porosity, ceramsite surfaces are glossier when densification is more completed at lower CaO
contents. This might partly explain the differences in compressive strength for the 3 samples
because of ceramsite with denser surface generally have better strength. The differences in the
microstructures (Figure 9B1, 9B2, and 9B3) are also suggested to be the reasons for the
differences in their water absorption, because the size and quantity of pores are the chief
determinants of water absorption.
Utilization of Water and Wastewater Sludge for Production of Lightweight … 101
KQ
FKA
K AE K EKSF
K
Fe2O3=8%
S Q Q Q Q
K EKA
AE K K EK Q Fe2O3=6%
Relative intensity
Q Q
K EKA
AE Q
Q Fe2O3=5%
K K K Q
Q
KQ
EKA
KAE Q
S K
KS Q CaO=7%
Q Q
EKA
K KS
E K Q CaO=5%
Q Q Q
CaO=2.75%
Q Q Q Q
2theta
Figure 8. Effect of oxide contents on the XRD patterns of ceramsite. Band labeling: A, albite-anorthite;
E, enstatite; F, ferrosilite magnesian [(Fe,Mg)SiO3]; K, kyanite; Q, quartz; S, sillimanite
[(Al1.98Fe0.2)SiO5].
Figure 9. Effect of oxide contents on the morphological structures of ceramsite (A-Fe2O3 and B-CaO).
102 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
16
14
Fe2O3 content=5%
12
10
Fe2O3 content=6%
8 Fe2O3 content=8%
Figure 10. Effect of oxides contents on the compressive strength of ceramsite (A-Fe2O3 and B-CaO).
The compressive strength of ceramsite with Fe2O3 contents of 5%, 6%, and 8% are
14.97MPa, 15.14MPa, and 15.67MPa, respectively (Figure 10A). Ceramsite with Fe2O3
contents of 5% has irregular-shaped elongated pores that decrease the compressive strength.
Because of its porous structure, which is caused the foaming reactions and the tiny pores
produced during sintering, the resultant specimens have worse compressive strength property.
In contrast, ceramsite with a few pores formed at higher Fe2O3 contents lead to relatively
higher compressive strength. The difference in crystalline phases may also explain the higher
gain in compressive strength for ceramsite with 8% Fe2O3 as compared to ceramsite with 5%
and 6% Fe2O3. The Fe3+ may act as Al3+, replacing the Si4+ parent ions of silicates to be
enclosed in the framework of silicates and lower the body strength.
It can be seen from Figure 10B that the compressive strength of ceramsite with CaO
contents of 2.75%, 5%, and 7% are 15.13MPa, 14.26MPa, and 13.13MPa, respectively. Ca2+
may act as metal ion ionically bonded in the interstices of the silicate network to produce
electrical neutrality that was broken due to the substitution of Si4+ by Al3+ [39]. Increasing the
content of CaO usually makes the crystalline particles easier to form at a given temperature
but increases its chemical reactivity in the silicate network. The compressive strength
decreases as the content of CaO increase from 2.75% to 7%, which implies that excessive
CaO exceeds the needed ions for producing electrical neutrality and lowers the body strength.
Another reason may be that the lower strength is thought to be due to the lower hardness of
the resulting crystalline phases [45].
Utilization of Water and Wastewater Sludge for Production of Lightweight … 103
To indicate the actual resistance provided by the structures of ceramsite for leaching of
heavy metals, the effect of sintering temperature on the solidification of heavy metals was
investigated. It can be seen from Figure 11 that the leaching contents of Cd, Cr, Cu, and Pb of
the 3 specimens decrease at the 24th hour or on the 30th day as sintering temperature increases.
The leaching contents of Cd and Cu of the 3 specimens at the 24th hour or on the 30th day are
all near 0μg/g and almost do not change above 975 ℃. The increase in sintering temperature
above 1000 ℃ has a slight influence on the leaching contents of Cr and Pb. It should be also
noted that Cd and Cu are the elements most sensitive to sintering temperature (≥1000 ℃), and
the leaching contents of the 4 heavy metals on the 30th day are higher than those at the 24th
hour. It is demonstrated that the heat treatment of ceramsite is advantageous to improve the
thermodynamic stability of heavy metals but that, to some extent, thermal treatment at
temperature below 1000 ℃reduces the binding capacity for heavy metals.
Above results indicate that after subjecting the sintered ceramsite to leaching conditions
that sintering at≥1000 ℃exhibits good binding ability for Cd, Cr, Cu, and Pb in ceramsite and
can be considered as a proper means for solidifying/stabilizing heavy metals. At temperatures
below 1000 ℃, densification and crystals growth are the main processes occurred in ceramsite
and heavy metals can not be completely solidified in ceramsite with the expanded bodies and
104 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
the semideveloped crystalline phases. This implies that only weak physical bonds are formed
between these heavy metals and the compounds in the ceramsite, and thus heavy metals are
easy to be leached, which may pose a detrimental impact on the environment. More densified
and lower porous ceramsite can be obtained at temperatures ≥1000 ℃and the high temperatures
are beneficial to the formation of liquid phases and improve the strength characteristics of
ceramsite [48], which inevitably increase the solidification efficiencies of heavy metals in
ceramsite. Furthermore, at higher temperatures, the heavy metals in the raw materials can be
entrapped inside the crystalline structures through the chemical and physical transformation.
It is generally recognized that the leaching of heavy metals obtained by such leaching
procedures is always operationally affected by many experimental factors especially the
mineral compositions of the samples [48]. In this study, XRD analyses were applied to obtain
mineral compositions of powder ceramsite specimens by using an XRD pattern database
(International Centre for Diffraction Data, ICDD). Samples sintered at 1000 ℃ with different
added contents of heavy metals (minimum, medium, and maximum) were selected and the
samples were scanned for 2ζ ranging from 10 to 90°. The analyses are also helpful to test the
forms of heavy metals present in ceramsite in order to judge the actual improvement in metal
binding resulting from the thermal treatment.
It can be seen from Figure 12A that the main crystalline phases of the 3 kinds of
ceramsite with different heavy metals contents are similar. The results indicate that the effect
of the contents of heavy metals on the formation of main crystalline phases in ceramsite is
minor. Kyanite (Al2SiO5), quartz (SiO2), and Na-Ca feldspars [albite-Na(AlSi3O8), anorthite-
Ca(Al2Si2O8)] are the main crystalline phases of the 3 ceramsite specimens and a small
amount of sillimanite [(Al1.98Fe0.2) SiO5] and enstatite [Mg2(Si2O6)] are also noted in the
patterns. The formation of Na-Ca feldspars, sillimanite, and enstatite in ceramsite is caused by
the presence of flux such as Fe2O3, CaO, and MgO in the raw materials and the solid-liquid-
phase reactions during the sintering process. Better crystallization produces better strength
and chemical durability [49], which implies that the solidifying efficiencies of heavy metals
in ceramsite may be partly influenced by the formation and types of crystals. It is thus evident
that the fraction of metals available for leaching is greatly reduced as a result of the
crystallization and chemical fixation within the aluminosilicates or silicates frameworks
acting during thermal treatment, as reported by the previous studies [48, 50]. The evidence of
chemical immobilization of metals is gained from the mineralogical characterization, in that
the heavy metals compounds in the 3 specimens can be identified in ceramsite by XRD
analyses (as shown in Figure 12B).
Figure 12B shows that the 4 heavy metals in ceramsite sintered at 1000 ℃ are in steady
forms and the main compounds are crocoite (PbCrO4), chrome oxide (Cr2O3), cadmium
silicate (CdSiO3), and copper oxide (CuO). The results indicate that the interactions between
the chromate and insoluble Pb compounds (e.g., lead oxides) result in the formation of
crocoite (PbCrO4), revealing the low leachability of Pb. Thus, the solidifying efficiency due
to the formation of PbCrO4 via consumption of the CrO42- present in the raw materials is
expected to be significant and the sintering process enhances Pb2+ to form insoluble PbCrO4
with CrO42- [51]. It is also observed that some Cr6+ is deoxidized to Cr3+ reflecting by the
identification of Cr2O3 in XRD patterns. The formation of CuO provides direct evidence that
Cu can be trapped inside the newly formed silicates or aluminosilicates minerals and it is
reasonable to think that Cu is not located in the framework, but is most likely buried inside
the newly formed minerals.
Utilization of Water and Wastewater Sludge for Production of Lightweight … 105
KQ
(A) -1 -1 -1
maximum--50mg kg Cd;1000mg kg Cr;500mg kg Cu;1000mg kg Pb
-1
-1 -1 -1 -1
medium-- 25mg kg Cd;500mg kg Cr;250mg kg Cu;500mg kg Pb
-1 -1 -1 -1
minimum-- 1mg kg Cd;100mg kg Cr;100mg kg Cu;50mg kg Pb
EKA
AE
Relative intensity
K K KES
K Q Q
Q Minimum
EKA
AE
K K K KES
Q Q Q Q Medium
EKA
K
KES
K K Q
AE S Q
Q
Q Maximum
-1 -1 -1 -1
(B) maximum--50mg kg Cd;1000mg kg Cr;500mg kg Cu;1000mg kg Pb
10 20 30
medium-- 40kg-1Cd;500mg
25mg 50 kg-1Cr;250mg
60 kg-1Cu;500mg
70 kg80Pb
-1 90
-1 -1 -1 -1
minimum-- 1mg kg Cd;100mg kg Cr;100mg kg Cu;50mg kg Pb
2theta (deg)
Minimum
Relative intensity
Medium
Maximum
30 35 40 45 50 55 60 65 70
2theta (deg)
Figure 12. XRD analyses of the main crystalline phases (A) (Band labeling: A, albite-anorthite; E,
enstatite; K, kyanite; Q, quartz; S, sillimanite) and the forms of heavy metals (B) (Band labeling: ■,
CuO; ▼, PbCrO4; ▲, CdSiO3; ●, Cr2O3) in ceramsite sintered at 1000 ℃.
The forms of Cr, Cu, and Pb (i.e. Cr2O3, CuO, and PbCrO4) particularly suggest that
more incorporation of the compounds into the aluminosilicates or silicates matrix is occurred
after the heat-induced transformation. For Cr, Cu, and Pb, this transformation of the chemical
state of the incorporated metals is clearly supported by the leaching results and XRD studies.
Majority of Cd is uniformly distributed in ceramsite bodies to form CdSiO3 indicating that Cd
has entered the liquid-solid-phases and stabilized in the crystalline structures of ceramsite
during the sintering process. The transformation of these heavy metals to crystalline state is
advantageous for the long-term stability of the metals and the crystalline solids are likely to
have the improved capacity to bind heavy metals [50]. Moreover, it is possible that the heat-
induced transformation of crystallization causes most of the heavy metals ions to transfer
from the surface to the interior of the particles. It is thus concluded that the fraction of heavy
106 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
metals available for leaching is greatly reduced and the solidifying efficiencies are strongly
enhanced by the crystallization and chemical incorporation within the aluminosilicates or
silicates frameworks during the thermal treatment.
The pH-dependent and the following H2O2-dependent leaching tests of heavy metals are
conducted with ceramsite sintered at 1000 ℃ to investigate the effect of external conditions
on the stabilization of heavy metals. It can be seen from Figure 13 that the maximum leaching
contents of the 4 heavy metals is achieved at the 24th hour or on the 30th day when pH is 1 and
the leaching contents almost do not change as pH>1. It is interesting to note that the
cumulative amounts of Cr leached from ceramsite (1000mg kg-1) on the 30th day are the
maximum and about fourfold as much as those at the 24th hour. Although the above
phenomenon is present, it should be noted that Cr cannot be easily released from ceramsite
during the pH-dependent leaching test. Most of Cr6+ may be deoxidized to Cr3+ (Cr2O3) in the
reducing environment formed locally during the sintering process and both Cr6+ and Cr3+ can
be stabilized in the structures of crystalline networks [41].
The pH-dependent leaching tests indicate that the relationship between heavy metals
mobility and pH is complicated and the solubility of heavy metals in ceramsite are
dramatically influenced by lower pH such as pH=1, also reported in previous studies [36]. In
the leaching tests, the pores in ceramsite are slowly and sufficiently filled by the leachant so
that the pores water pH will progressively decrease especially when pH is 1 and then the
heavy metals in this region will start to leach. A fraction of the leached metals will diffuse
toward the leachant and these initial released metals likely originate from the outer surface
pores of ceramsite in contact with the leachant. However, the results from the present work
indicate that the residual amounts of heavy metals in the sintered ceramsite appeared to be
efficiently immobilized within the silicates or aluminosilicates matrix, as demonstrated by the
availability values as expressed in terms of the leaching contents of each heavy metal. This
implies that stronger chemical bonds are formed between these heavy metals and the
components in ceramsite, making heavy metals difficult to be leached, in other words, the
toxic metals present in the ceramsite pose no harmful impact on the environment.
It can be seen from Figure 14 that the leaching contents of Cd, Cr, Cu, and Pb at the 24th
hour or on the 30th day increase as the H2O2 concentration increases. This phenomenon
indicates that heavy metals can be significantly solidified in ceramsite for a long period of
time due to the heavy metals in the pores surface present in stable forms and will remain
steadily even in the oxidative environment. Furthermore, the leached space of the pores
hinders the initiation of more contact oxidation, which decreases the available space for water
ingress and improves the solidification efficiencies of Cd, Cu, and Pb in ceramsite. The
leaching contents of heavy metals slowly increase at H2O2 concentration≥1.5mol L-1, which
can be attributed to the reason that the solubility of heavy metals and oxidative ability of
H2O2 have already reached an equilibrium state in the leachant.
Utilization of Water and Wastewater Sludge for Production of Lightweight … 107
-1 -1 1.8
1.4 1mg kg Cd 24h 100mg kg Cr 24h
-1 -1
25mg kg Cd 24h 500mg kg Cr 24h
-1 -1
1.2 50mg kg Cd 24h 1000mg kg Cr 24h 1.5
-1 -1
1mg kg Cd 30d 100mg kg Cr 30d
1.0 -1
25mg kg Cd 30d
-1
500mg kg Cr 30d 1.2
-1 -1
50mg kg Cd 30d 1000mg kg Cr 30d
Cr (ug g )
-1
0.8
Cd (ug g )
-1
0.9
0.6
0.6
0.4
0.2 0.3
0.0 0.0
16 -1
100mg kg Cu 24h -1 0.5
-1
50mg kg Pb 24h
14 250mg kg Cu 24h -1
-1
500mg kg Pb 24h
500mg kg Cu 24h -1
12 -1
1000mg kg Pb 24h 0.4
100mg kg Cu 30d -1
-1
50mg kg Pb 30d
Pb (ug g )
Cu (ug g )
250mg kg Cu 30d
-1
-1
-1
10 -1
500mg kg Pb 30d
500mg kg Cu 30d -1 0.3
1000mg kg Pb 30d
8
6 0.2
4
0.1
2
0.0
0
1 2 3 4 5 6 7 8 9 10 11 12 1 2 3 4 5 6 7 8 9 10 11 12
pH pH
Figure 13. Effect of pH on the leaching of heavy metals.
10 -1 -1
1mg kg Cd 24h 100mg kg Cr 24h
-1 -1 2.8
9 25mg kg Cd 24h 500mg kg Cr 24h
-1 -1
8 50mg kg Cd 24h 1000mg kg Cr 24h 2.4
-1 -1
1mg kg Cd 30d 100mg kg Cr 30d
7 -1 -1
25mg kg Cd 30d 500mg kg Cr 30d 2.0
-1
6 -1
50mg kg Cd 30d 1000mg kg Cr 30d
Cr(ug g )
Cd(ug g )
-1
-1
5 1.6
4 1.2
3
0.8
2
1 0.4
0 0.0
0.8 -1 -1
100mg kg Cu 24h 50mg kg Pb 24h 0.24
-1 -1
0.7 250mg kg Cu 24h 500mg kg Pb 24h
-1 -1
500mg kg Cu 24h 1000mg kg Pb 24h 0.20
0.6 -1 -1
50mg kg Pb 30d
100mg kg Cu 30d
-1 -1
250mg kg Cu 30d 500mg kg Pb 30d 0.16
0.5
Pb(ug g )
Cu(ug g )
-1
-1
-1 -1
500mg kg Cu 30d 1000mg kg Pb 30d
0.4 0.12
0.3
0.08
0.2
0.04
0.1
0.0 0.00
Figure 14. Effect of
0 H2O12 on the
2 leaching
3 4 of heavy
5 0 metals.
1 2 3 4 5
H2O2 concentration(mol/L) H2O2 concentration(mol/L)
108 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
The reasons for the effective solidification of heavy metals in oxidative condition can be
attributed to many phenomena and are presented as follows: During the sintering process, the
Al3+ ion may substitute for the Si4+ in a network tetrahedron, contributing to the stability of
the network. Therefore, Al2O3 can enter the silica network as AlO44- tetrahedra to replace
some of the SiO44- groups; however, since the necessary +4 for the tetrahedra is replaced by
the +3 valence of Al, cations must supply the necessary other electrons to produce electrical
neutrality [41]. These complex chemical reactions allow for the incorporation of a large
number of cations including Cd, Cr, Cu, Pb, and others into the structure [48, 52]. Therefore,
these results indicate that, even subjecting the ceramsite to rigorous leaching conditions
(H2O2 concentration=5mol L-1), ceramsite structures systems still exhibit good binding ability
for heavy metals and cannot be easily leached from ceramsite to the leachant, which can be
considered chemically durable. It can be concluded from the results that most of the heavy
metals are incorporated inside the crystalline structures of ceramsite and these new recycled
materials are environmentally sustainable.
It seems that higher contents of Al2O3 may enhance the (Cu and Pb) substitution of parent
ions (Al3+ or Ca2+) in ceramsite and may restrain the replacement of parent ions by Cd and Cr.
This phenomenon indicates that heavy metals can be significantly solidified in ceramsite for a
long period of time because the heavy metals with stable forms in the porous surface can not
be easily washed out. It should be noted that Cu leached from ceramsite with lower Cu
contents on the 30th day is higher than those with higher Cu contents, which may be attributed
to the special reaction between Cu and the components in ceramsite. Because of the
enhancement of diffusion for the ions of Cu in the sintering processes, cation-exchange
reactions between the ions of Cu and the crystals in ceramsite are easily to occur and,
consequently, cation exchange is partly responsible for the immobilizations of such heavy
metal ions (Cu). The higher Cu contents in ceramsite may enhance their substitution of parent
ions (Al3+ and Ca2+) and may therefore enable the leaching content of Cu to decrease [23].
4. CONCLUSION
It can be concluded from the results in this chapter that DWTS can be used as a substitute
for clay to be mixed with WWTS and water glass for producing ceramsite with optimal
content of SiO2 and Al2O3 ranging from 14–26% and 22.5–45%, respectively. Thermal
properties are profoundly influenced by the variation of Fe2O3 contents in the mixtures.
Ceramsite with higher Fe2O3 contents have more complex crystalline phases and fewer pores,
which accordingly contribute to relatively higher compressive strength. As the content of
CaO increases, pores, Na-Ca feldspars, and amorphous phases increase while compressive
strength decreases, which implies that excessive Ca2+ exceeds the needed ions for producing
electrical neutrality of silicate networks. This study also demonstrates the feasibility of
transforming heavy metals into stable forms in ceramsite and the successful reduction of
heavy metals mobility after the heat treatment. When the sintering processes are conducted
with the ceramsite containing heavy metals, a decrease in the solubility of the heavy metals
can be achieved due to the changes in their structural location and chemical forms. Sintering
temperature and the changes in composition of leachant solution (acidic, neutral, alkaline and
oxidative) have significant influence on the leaching and solidification behaviors of the heavy
Utilization of Water and Wastewater Sludge for Production of Lightweight … 109
metals in ceramsite. Heavy metals stabilized in ceramsite are in steady forms and cannot be
easily released into the environment again to cause secondary pollution. The results in this
chapter present a new step towards the safe utilization of sludge-ceramsite. The findings of
this study can revolutionize the handling of such kinds of sludge in the future for their reuse
as low-cost materials, rather than their becoming waste requiring costly disposal.
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112 Zou Jinlong, Yu Xiujuan, Dai Ying and Xu Guoren
Chapter 4
ABSTRACT
Through the last decades, the ever increasing energetic demands have been
accomplished by exploiting new natural reservoirs, including offshore oil and gas deposits
located in marine coastal areas. During the extraction and production phases, large amounts
of water are brought to the surface along with the hydrocarbons. These waters include the
‗formation water‘, that lies underneath the hydrocarbon layer, and ‗additional water‘ usually
injected into the reservoirs to help force the oil to the surface. Both formation and injected
waters, named ―produced formation waters‖ (PFWs), are separated from the hydrocarbons
onboard offshore platforms and then disposed into the marine environment through ocean
diffusers. PFWs contain several contaminants and represent one of the main sources of
marine environment pollution associated with oil and gas production.
This makes the study of PFW fate of paramount importance for a proper
management of environmental resources as well as for planning and optimizing the
discharge and monitoring procedures.
In the first part of this chapter we provide a detailed description of the chemical
characteristics of PFWs and their potential toxic effects and review the mixing processes
governing their dispersion in the marine environment. In the second part of the work we
briefly review past efforts in observing and modelling PFW spreading in the ocean.
Finally, we propose a multidisciplinary approach, integrating in situ observations and
numerical modelling, to assess dispersion of PFWs in space and time. As a case study we
will refer to the results of a previous study conducted in the Northern Adriatic Sea, a sub-
basin of the Mediterranean Sea, where a number of offshore natural gas (CH4) extraction
platforms are currently active.
114 D. Cianelli, L. Manfra, E. Zambianchi et al.
INTRODUCTION
The worldwide exploitation of oil and gas reserves located in marine coastal areas has
drastically increased during the last decades due to the gradual depletion of the on-land
deposits. A by-product resulting from both oil and gas extraction is the Produced Formation
Water (PFW), which to date represents the largest volume waste from the production phase of
the offshore oil and gas facilities (e.g. Ray and Engelhardt, 1992; Berry, 2005).
The disposal of the PFWs is carried out through the reinjection into the reservoir, the
discharge into the ocean or the transport onshore. Here we focus on the disposal of the PFWs
into the marine environment by means of surfacing or submerged outfalls.
The complex chemical composition of the PFWs (ranging from heavy metals to soluble
hydrocarbons) as well as its physical state (solution, particulate-matter, suspensions, etc.)
determines the potential impact of the PFWs on the receiving environment (OGP, 2005). In
the last decade the release of PFW into the sea has received increasing attention due to both
the potential long term effects (and associated ecological risks) of some of its chemical
compounds and the high water volumes discharged (Rye et al.,1996).
In order to determine the actual impact of the PFWs on the marine environment, the
chemical and physical processes driving their distribution into the ambient fluid have to be
investigated. In particular, an exhaustive assessment of the effect of PFWs on the
environment at discharge locations may be derived from the site-specific chemical
characteristics of the effluent and environmental factors (meteo-oceanographic conditions).
After the release into the sea, the PFW undergoes several different processes such as the
dilution into the ambient fluid (which may occur more or less rapidly depending on local
oceanographic conditions), the evaporation and biodegradation (that change the concentration
and composition of inorganic constituents) (Strømgren et al., 1995 and references therein),
the volatilization towards the atmosphere and the settling at the bottom.
In particular, a crucial process determining the dilution of PFW and the reduction of its
concentration in the sea water is the mixing of the effluent with the ambient fluid (e.g.
Baumgartner et al., 1994). Such a dispersion process occurs in two phases: a rapid initial
mixing phase (near-field) taking place immediately after the release within the first tens of
meters from the discharge point, and the subsequent passive dispersal phase (or far-field) that
evolves at larger distance over time scales of hours or days.
Nowadays several field measurement and dispersion modelling studies investigating the
fate of the PFW and predicting the near and far field dilution rates with high accuracy are
available (e.g. Neff, 2002). The approach integrating field and laboratory observations with
numerical modelling presently represents the most suitable tool to assess the effects of PFW
discharges and the potential risks for the local marine ecosystems. Such studies can be
considered as a baseline to improve monitoring programs and to optimize different sampling,
monitoring and assessment techniques as well as the industry discharge practices (Durell et
al., 2006; Cianelli et al., 2008).
In this paper we firstly describe origin, chemical characteristics, potential toxic effects
and methods of disposal of the PFWs. In the first part we also provide a description of the
dispersion processes that a marine discharge may experiment and the possible effect of
environmental factors (water column stratification and ambient current) on the effluent.
Modelling and Observation of Produced Formation Water (PFW) at Sea 115
When PFW is discharged into the sea, its chemical compounds undergo a number of
dispersion, removal and degradation processes (Burns et al., 1999). In particular, weathering
is a combination of physical, chemical and biological processes that influence the fate of
PFW discharged into the sea and consequently the distribution and the effects of compounds
present in it. The most important sub-processes of weathering are dilution (discussed in the
next section), evaporation, chemical reactions, adsorption, sedimentation and biodegradation.
Evaporation of PFW-contained contaminants depends on vapour tension of compounds
and some environmental conditions (e.g. current velocity, water depth, turbulence, wind
velocity). The compounds with lower molecular weight and higher vapour tension tend to
evaporate into the air rapidly. Some aromatic hydrocarbons (like, e.g., BTEX) may be
dissolved in seawater but easily volatilize into the air. The compounds of PFW are also
exposed to different chemical reactions (e.g. oxidation, hydrolysis and precipitation). Some
PFW elements (e.g. trace metals) exist at lower oxidation states until not in contact with
oxygen; they undergo oxidation when PFW is brought to the surface and then precipitate as
hydroxides. As an example consider the transformation of dissolved iron (Fe2+) to Fe3+ and to
hydroxides or the oxidation of hydrogen sulphide to elemental sulphur that precipitates
(OGP, 2005).
Modelling and Observation of Produced Formation Water (PFW) at Sea 117
Some substances tend to adsorb on suspended solids both within PFW and in the water
column; in this way, they become potentially available to filter feeders and/or transported into
the sediments as a result of flocculation and complexation processes; the adsorption depends
on factors such as the polarity and solubility of chemical substances, the particle size
(preference for fine fraction <60µm) and organic carbon content in particulates and sediments
(Marchetti, 2000).
Sedimentation yields accumulation of these compounds in benthic organisms and may
result in inserting them into the trophic chain. PFW compounds can be biodegraded from the
microbial community both in seawater and in the sediment; aerobic processes take place in
the water column while in the sediment aerobic or anaerobic degradation can occur.
The biodegradation is conditioned by factors such as the chemical structure of
compounds, the availability of micro-organisms and the environmental conditions (e.g.,
oxygen and nutrient concentrations); substances with low molecular weight are readily
broken down while multi-ring aromatic hydrocarbons are relatively stable and remain longer
in the plume of PFW (OGP, 2005).
After the separation from hydrocarbons, PFWs are discharged directly into the sea from
offshore platforms by means of surface or submerged outfalls, usually mounted in a vertical
position. The effluent, also indicated as plume, is released at a certain depth and undergoes
physical, biological and chemical processes acting over increasing temporal and length scales
while moving away from the release point. Among these processes, the most important ones
together with their typical scales are reported in Table I.
The mixing in the proximity of the diffuser is influenced by buoyancy and momentum
fluxes, which are generated by the release of PFWs from the diffuser itself and by their
interaction with the current field (self-induced turbulence).
The buoyancy is originateds by the density difference between the effluent and the
ambient fluid, while the momentum is associated with the velocity at which the effluent is
ejected. The interaction with the medium is intense and determines a fast mixing of the two
fluids. This process persists until the turbulent kinetic energy generated by these processes is
dissipated. The region where this initial mixing occurs is named ―near field‖ or ―initial
mixing zone‖ or alternatively ―initial dilution zone‖; the outer limit of this area is generally
118 D. Cianelli, L. Manfra, E. Zambianchi et al.
set where the plume meets a boundary layer (e.g., the sea surface, the bottom of the water
column or an equilibrium layer where the effluent density equals that of the environment).
By the end of the initial mixing, the effluent is stabilised and passively transported by the
ocean currents in an area known as the ―far field‖, where the mixing is entirely due to the
environmental turbulence and takes place more slowly than in the near field.
At distances greater than that of the far field, PFWs may be subject to chemical and
biological decay processes determining their long term fate (Roberts, 1990).
ve va (1)
Sa =
ve
where ve is the effluent flux volume discharged, while va is the ambient fluid flux volume
(Baumgartner et al., 1994). In the vicinity of the diffuser port, Sa will be close to 1 as va will
be very small. Accurately measuring va is not straightforward; for this reason, very often what
is measured is the concentration of pollutant in a selected position at a given distance from the
discharge.
The mass balance equation thus gives:
where cp is the average concentration of the effluent cross section, ce is the initial
concentration in the effluent and ca that in the ambient fluid.
Reorganising equation #2, Sa becomes (Glenn, 1997):
ce ca
Sa (3)
cp ca
If the ambient concentration is null (ca = 0), the equation #.3 simplifies as (Baumgartner
et al., 1994):
Sa ce (4)
cp
This equation shows that, if the ambient concentration is null, the dilution factor defined
in equation (1) and based on flux volumes also describes the dilution in terms of concentrations.
Modelling and Observation of Produced Formation Water (PFW) at Sea 119
The mass or volume dilution factor Sa is entirely defined by the density of the plume and,
in principle, is independent on the contribution given by each individual pollutant.
In recent years increasing emphasis has been devoted to the understanding of the
concentration of pollutants in the ambient fluid at the end of the initial dilution phase. If both
the effluent and the receiving fluids carry pollutant substances, the dilution reduces until in a
limit it becomes zero when the concentration of pollutants in the environment equals or
surpasses that of the effluent. On these grounds, for some purposes the ambient concentration
cannot be taken as null, but the actual amount of substances in both fluids must be accounted
for. In this case, the effective dilution factor (Saei) has to be used to measure the dilution of
each pollutant in the plume (Baumgartner et al., 1994):
cei cai
Saei (5)
cpi cai
where the suffix ―i‖ refers to the i-th substance and is used to underline that the effective
dilution must be specified for each single substance. If cai = 0, equation (5) will be simplified
to equation (4).
Therefore, Saei measures the dilution of each individual substance rather than of the
transport fluid. It differs from the average dilution (equation (4)) when the ambient fluid
already contains a given amount of the pollutant (background pollution) (Frick et al., 2002).
The background pollution acts as the minimum concentration that the plume can assume,
and consequently sets an upper limit to the effective dilution.
Effluent dynamics
The typical concentration profile of an effluent discharged in a fluid at rest is Gaussian
(see, e.g., Csanady, 1980), i.e. highly diluted at the margins and concentrated in the centre
(Figure 1). Such scheme is commonly used to represent the dynamics of a plume (Roberts,
1990; Skatun, 1996; MacIntyre et al., 1995; Glenn, 1997) and shows the formation of a
minimum dilution zone in the central part where the highest concentrations are reported.
Plume
concentration
profile
Port diffuser
Figure 1. The dynamics of an effluent discharged in a fluid at rest from an horizontal diffuser. The
figure shows the Gaussian profile of the concentration of the plume with high dilution at the margins
and low dilution in the central part.
120 D. Cianelli, L. Manfra, E. Zambianchi et al.
Generally speaking, two kinds of plumes are considered (Frick et al., 2002): jet plumes
and buoyant plumes. The jet plume (also referred to as pure jet) is a jet current with high
momentum and no buoyancy, with a density similar to that of the ambient fluid at the
discharge location. On the contrary, the buoyant plume (or pure plume) has null momentum
and negative (i.e., the density of the plume is higher than that of the medium at the discharge
location) or positive (i.e., the density of the plume is higher than that of the medium at the
discharge location) buoyancy. The remaining types of plumes (defined as buoyant jets) have
intermediate characteristics and may have at the same time momentum and buoyancy. Most
of underwater discharges belong to this last category.
Figure 2 shows a positively buoyant plume released by an horizontal diffuser located at a
given depth of the water column. The effluent, lighter than the ambient fluid, is vertically
curved and accelerated by its buoyancy. The region inside which the buoyancy and
momentum of the discharge itself along with the effect of the local ambient current field
results in rapid turbulent mixing of the plume, represents the near field.
CURRENT
depth
CENTERLINE
Plume
Diffuser
Horizontal distance
Port elevation
Figure 2. The dynamics of a positively buoyant plume. The figure shows an effluent discharged from a
diffuser under stratified water column conditions.
The entrainment process tends to decrease the density difference between the effluent and
the medium. If the ambient fluid has constant density, the density of the diluted effluent will
asymptotically reach the environmental density without equalling it, thus continuing its rise
(or sink in the case of negative buoyancy), possibly reaching the surface (or the bottom).
In stratified conditions, the effluent can reach a depth of the same density, called trap
level or neutral buoyancy level. In the upper part of the trap level the ambient density is lower
than that of the plume, whereas below it the density is higher; as a consequence, the effluent
is trapped between these two layers.
In general, when the effluent reaches the trap level it still has upward momentum
bringing it in a region where the density is lower than that of the plume. For this reason, the
buoyancy becomes negative, slowing down the plume until its vertical velocity is inverted
and the plume becomes trapped. This second kind of trap level is different from that described
above; during the rise and fall, the entrainment process still works determining changes in the
Modelling and Observation of Produced Formation Water (PFW) at Sea 121
average density of the plume. In an idealised case, the effluent oscillates around a variable
trap level creating an oscillatory motion at the Brunt-Väisälä frequency (or buoyancy
frequency).
Once the initial kinetic energy is dissipated, the plume reaches a given depth and
undergoes further transformations due to the ambient stratification and to the difference
in density between the inner and the outer part of the plume. When neutral buoyancy is
reached in the trap level, the density in the centre of the plume will be the same of the
medium at the same depth. As mentioned above, the water column must be stably
stratified (density increasing with the depth) to prevent any further vertical motion of the
plume.
U0 (6)
Fr
a e
gD
r
where:
D = port diameter
ρa = environmental density at discharge location
ρe = effluent density
ρr = reference density, usually equal to ρe
g = gravity acceleration
U0 is the velocity at which the plume is discharged through the port, and it is equal to:
4q (7)
U0
D2
where q is the flux volume at the port. If the diffuser is multi-port, then q=Q/n° ports, Q
being the total flux volume.
The quantity under the square root in eq. (6) must be positive. If ρa < ρe the absolute
density difference is used for the square root, and only afterwards the minus sign is included
in the equation.
Fr is an adimensional number that allows the classification of the plume at the discharge
location, differentiating between a jet plume and a buoyant plume; this parameter also
measures the relative importance of buoyancy and momentum during the initial mixing. The
numerator of eq. (6) is associated with the momentum of the plume, while the denominator
represents the buoyancy.
A low absolute Fr (around unity) indicates that the mixing process is dominated by
buoyancy (buoyant plume), while higher values (usually greater than 10) are typical of
122 D. Cianelli, L. Manfra, E. Zambianchi et al.
Ua3 (8)
FRo
a e Q
g
a L
where Ua is the velocity of ambient current at port dept, while L is port length.
The degree of initial dilution of an effluent is a function of environmental and technical
factors. For example, a crucial parameter is the type of diffuser used to discharge the effluent
(Roberts, 1990). The diffuser can be made of one or more ports, and even the diameter of
each port can affect the velocity of the effluent and consequently its classification as jet or
buoyant plume. The orientation of the diffuser can influence the plume trajectory, and in
presence of a multi-port discharge also the distance between the ports can play a role. Another
property affecting the entrainment and the dilution is the ambient velocity. Intense currents
can increase the shear between the plume and the medium, thus enhancing turbulence, which
in turn favours the entrainment of marine water within the effluent, and consequently its
dilution.
The stratification of the water column is another parameter influencing plume dynamics
and dilution. In highly stable conditions, the plume can get trapped under the surface and the
movements along the water column might be so reduced to prevent the dilution in the near
field. The higher the rise (or sink) of the plume, the higher the chances for the effluent to mix
with the medium. Under weak stratification (winter), an increase in the rise (or sink) depth
promotes the dilution. In presence of strong stratification (summer) the situation is reversed
(Petrenko et al., 1998).
In the North Sea an important national monitoring programme was conducted by Norway
to evaluate the fate of PFW in the marine environment and its effects on marine organisms.
For these objectives, surveys were carried out from 1999 to 2003 to track PFW hydrocarbons
in the environment and analyse the levels of these compounds in fishes living in Norwegian
waters. Hydrocarbon levels in the marine environment were monitored using in situ natural
and artificial sampling ―systems‖ such as mussels and semi permeable membrane devices
(SPMD) located at different distances from the PFW discharge source. The results showed
that even though hydrocarbons were recorded as far as 10 km from the discharge point,
significant concentrations in terms of their possible biological effects were measured within a
distance of 500 metres from the discharge. Some data on the potential impact of PFW
originating from North Sea platforms are reported in the scientific literature (e.g. Somerville
et al., 1987; Brendehaugh et al., 1992; Neff et al., 1992; Stagg et al., 1995; Stromgren et al.,
1995; Frost et al., 1998; Henderson et al., 1999; Holdway, 2002). These studies show that the
toxicity of PFW to marine organisms is low and would likely have acute effects only within
the immediate mixing zone around production platforms. Effects of PFW may include altered
benthic communities dominated by short-lived opportunistic polychaetes up to 100 m from
offshore platforms.
In the Netherlands the usefulness of the Chemical Hazard and Risk Management
(CHARM) model to regulate use and discharge of chemicals employed in offshore
exploration and production phases has been studied: trace metals and PAH concentrations
have been analysed in mussels along a gradient from a PFW discharge. The results show that
no increase was observed in mussel tissue in metal concentrations while a significant increase
was measured for naphthalene (OGP, 2005).
Bioaccumulation studies have been conducted in the Gulf of Mexico to evaluate PFW
contaminant concentrations in edible tissue of fish and invertebrates collected near platforms
(USEPA, 1993). The concentrations measured in this work resulted not harmful to fish,
molluscs or to human health. However, Rabalais et al., (1992) observed bioaccumulation of
hydrocarbons as far as 1000 metres from PFW diffusers and an ecological impact of PFW
discharge in terms of the decrease of some benthic species near offshore platforms.
Burns et al. (1999) used chemical tracers (benzene and toluene) to describe the
distribution of PFW discharged from an offshore platform located in Northwest Shelf of
Australia; they evaluated hydrocarbon bioaccumulation in bivalves in the vicinity of
discharge point in connection with growth rates in natural marine bacterial and phytoplankton
assemblages. The conclusion of chemical and biological analyses, associated to modelling
studies, was that dissipation and degradation processes were fast and the area of potential
biological impact extended as far as 900 m from the discharge.
In a series of experiments described by Raimondi and Schmitt (1992), Osenberg et al.
(1992), Krause (1993) and Krause et al. (1992) various invertebrates were exposed to PFW
discharged from an outfall in the Santa Barbara Channel near Carpinteria (Southern
California). Patterns of sub-lethal effects (in terms of reduced survivorship of larvae,
settlement, ability to metamorphose, viability, ability of sperm to fertilize eggs and
reproductive success) observed on different marine organisms (abalones, mussels and sea
urchins) were inversely correlated with the distance from the diffuser.
In the Adriatic Sea (Mediterranean basin) a monitoring programme is presently carried
out in order to evaluate potential impact of PFW discharges. The monitoring plan includes
chemical-physical analyses on seawater samples, chemical analyses on particulate matter
124 D. Cianelli, L. Manfra, E. Zambianchi et al.
within the water column, sediment and biota (mussels). Besides, biological investigations
(ecotoxicological bioassays, biomarker analyses, study of benthic animal community and fish
assemblage analysis) can fruitfully complement those analyses. Chemical composition and
effects of PFW originated from Adriatic gas installations have been studied by several authors
(Cicero et al., 2003; Mariani et al., 2004; Faraponova et al., 2007; Gorbi et al., 2007; Manfra
et al. 2007; Fattorini et al., 2008; Gorbi et al., 2008). The chemical analyses show high
concentrations of zinc and cadmium in sediment near the PFW discharge point. These
contaminants are present in PFWs and they may be derived from corrosion or chipping of
galvanized structures on the platform or in the oil/water separator system. Arsenic also
showed higher quantities in biota near offshore platforms but chemical speciation allowed to
exclude the anthropogenic impact connected with exploitation activities and revealed a
natural regional gradient of arsenic levels (as arsenobetaine and arsenocholine) in mussel
tissues. Different marine organisms (bacteria, algae, crustaceans, sea urchins and fish)
exposed to PFW showed toxic responses but no significant toxic effect was observed when
organisms were exposed to sea waters and to sediments collected along the PFW plume. The
reason for this lies probably in the fact that these platforms drain only small volumes of PFW
into the sea and the dilution process is rapid in the near field. Toxicity tests on PFW allowed
to define a range of sensitivities for different test-organisms. Toxic effects of PFW on
organisms may be due to absorption of water-soluble components through their surface
epithelia and/or to oral ingestion and digestion of particulates. Besides, studies on benthic
community show that densities of some benthic organisms (e.g. species ecologically linked to
M. galloprovincialis) increase near offshore platforms (Trabucco et al., 2006).
Model aims: as a first natural step, the investigator has to define detailed model aims
taking into account the physical processes involved in the system to be modelled. It
is worth underlining that a model may target scientific as well as operational goals.
Model features: starting from the model goals, a list of the required model
characteristics in terms of input-output flexibility has to be formulated.
Modelling and Observation of Produced Formation Water (PFW) at Sea 125
After selecting the model, a scale analysis is necessary to determine the relevant scales of
the problem and the complexity of the model in order to adequately reproduce the studied
system. A rigorous procedure to test a numerical model ensures that it is appropriate to
simulate the functioning of our natural system. The following tests have to be carried out:
A broad range of numerical models varying in complexity, accuracy and other features
has been conceived for pollutant transport problems. It is worth underlining that accurate
model implementation, calibration and testing are imperative to ensure the reliability of
model results.
The models may be classified into different groups according to the problem description
adopted; transport processes can be described using the Eulerian or, equivalently, the
Lagrangian approach. The difference lies in the expected output: the Eulerian approach will
result in pollutant concentration maps, whereas the Lagrangian one will yield trajectories of
pollutant particles.
Following a general classification, the models simulating the dynamics of an effluent
discharged into the fluid can also be divided into two main groups: empirical and theoretical
models (MacIntyre et al., 1995; Glenn, 1997).
The most frequently applied theoretical models are: the UM3 model (Three-dimensional
Update Merge), the DKHW model (Davis, Kannberg, Hirst model for Windows) and the
JETLAG (Lagrangian Jet) model (Baumgartner et al., 1994; Frick et al., 2002). Some of the
most widespread empirical models are the CORMIX (Cornell Mixing Zone Expert System)
and the RSB (Roberts-Snyder-Baumgartner) models (Baumgartner et al., 1994; Glenn, 1997;
Frick et al., 2002).
The better understanding of the physical processes involved in the pollutants transport
problem has yielded a growing development of these advanced numerical models which
accurately reproduce the dispersion processes and may provide an estimate of pollutants
126 D. Cianelli, L. Manfra, E. Zambianchi et al.
concentration generated by the effluent discharges. In particular, in the case of PFW releases,
numerical modelling allows to simulate the dispersion process taking into account both the
discharge and receiving environment conditions.
For such reasons since the 1990s several transport models simulating the initial mixing
process as well as the effect of the ambient currents and turbulence in the far- field zone on an
effluent discharged into the sea have been successfully developed. Here we present a brief
review of some of the most representative modelling studies on the dispersion of PFW into
marine environment.
Some of the previously cited numerical models have been successfully applied to
evaluate the fate of PFWs discharged in coastal areas (e.g. Washburn et al, 1999; Berry, 2005;
Cianelli et al, 2008). Washburn et al. (1999) used the RSB model to perform a field and
modelling study around a diffuser located in California; they demonstrated that a crucial
factor controlling the exposure of organisms to PFWs around the discharge point is the depth
of the effluent in the water column. Berry (2005) developed an analysis of potential
environmental effects associated with PFW discharge adopting an integrated modelling
approach. In particular, the CORMIX model was applied to describe the dispersion of PFW
released at Sable Island Bank (Canada); the results of this work suggested that the potential
risks for the environment are low due to the rapid dilution of the wastewater plume. In the
following section we will summarize the results of a case study (Cianelli et al., 2008) on the
dispersion of PFW discharged in the Adriatic Sea (Mediterranean Sea), where the initial
mixing has been simulated by means of the UM3 model.
At present several modelling studies using various approaches have also been
implemented and applied to PFWs discharged from platforms located in the main oil and gas
extraction areas.
In the North Sea, where the discharge of PFW from oil and gas production reached an
annual volume of almost 400 million m3/year in 2003 (e.g. Durell et al., 2006), monitoring
programmes have been conducted since the mid-1990s. These local and regional field studies
have been used to optimize the monitoring plan as well as to implement and validate the
numerical models simulating the dispersion and fate of the PFW chemical compounds
released into the marine environment.
The CHARM (Chemical Hazard Assessment and Risk Management) model (Stagg et al.,
1996) was developed to predict the potential risks due to the chemicals released offshore and
was validated with field measurements of concentration of selected PFW compounds.
The DREAM (Dose related Risk and Effect Assessment Model) model was applied in the
Norwegian sector of the North Sea to estimate the dispersion of PAH (polycyclic aromatic
hydrocarbons) (Durell et al., 2006) and to predict the ecological risks associated with PFW
discharges (Neff et al., 2006). A comparison with field measurements showed that the
DREAM model results complement the in situ and laboratory data and that the numerical
approach represents an useful tool for PFW discharges and impact assessment (Durell et al.,
2006; Neff et al., 2006).
The dispersion and dilution of PFW discharged from 95 oil platforms operating in the
North Sea have also been simulated by means of the PROVANN (Produced Water in
Norwegian) model (Rye et al., 1998). PROVANN is a three-dimensional model simulating
the transport, dilution and degradation of chemical compounds released into the marine
environment from one or more simultaneous discharge points (Reed et al., 1996). The
Modelling and Observation of Produced Formation Water (PFW) at Sea 127
numerical predicted concentrations were compared with the field data; the model provided
useful results in terms of potential exposures to marine biota (Rye et al., 1998).
Estimates of the PFW concentration in the Gulf of Mexico, the North Sea and the Bass
Strait (Australia) were computed by means of the OOC (Offshore Operators Committee) Mud
and Produced Water Discharge model (Brandsma and Smith, 1996). In a more recent work
Smith et al. (2004) validated the OOC model using field data on mud and PFW discharges
from platforms located respectively in California and in the Gulf of Mexico. In both studies,
the model predicted plume depth and trajectory were in good agreement with field
observations for a wide range of discharge and receiving environment conditions. In
particular, in the near field zone the simulated PFW concentrations matched very accurately
the measured data (Smith et al., 2004).
Independently on the numerical approach followed, all the previously described works
demonstrate that, at present, modelling PFW dispersion both in near- and far- field zones may
play a crucial role in a ―prevention first― policy and represents an important first step in the
design of a decision-making action.
In order to study the effects of PFW discharge into the sea it is also very important to
consider:
a) which matrix has to be investigated (e.g. water column, sediment and biota);
b) what sampling pattern and frequency have to be chosen;
c) what parameters have to be measured.
The most useful matrices in a PFW monitoring plan are: the water column since it
receives the PFWs and because its physical-chemical characteristics influence the PFW
128 D. Cianelli, L. Manfra, E. Zambianchi et al.
dispersion; the sediment, which is a conservative matrix and a vehicle of transport for
contaminants; finally, the biota providing information about ecological effects of discharges.
Following the OSPAR guidelines, in a monitoring plan several sampling stations have to
be scheduled. These stations need to be located taking into account the predicted extension of
the area of influence as derived by discharge and local environmental conditions.
During a monitoring plan several different parameters have to be measured (Maggi et al.,
2006):
The comparison of these parameters with the background level can provide an indication
on the potential levels of concern. For example, if the concentration of some substances in
PFW is found to be significantly higher than the seawater background levels, then the
bioavailability and ecotoxicity of these substances on selected species may have to be further
considered. The toxicity studies are used as a complement to chemical measures for
quantifying the potential toxic effects of PFW, together with the bioavailability of chemical
substances and the possible synergies (see proceedings in Ray and Engelhardt, 1992 and in
Reed and Johnsen, 1996; Manfra et al., 2007; Mariani et al., 2004). The in situ physical and
chemical data may be also utilized in mathematical models to assess the time and spatial
extent of PFW discharge effects at sea. For example, some chemical compounds of PFW
(salts, nutrients, isotopes, ions) may be used as tracers. Their analysis in PFW and
environmental samples (seawater and sediment) permits to track the PFW into the sea. A
substance may be conveniently used as a tracer if it has some proprieties: conservative,
representative of PFW, easy to analyse and to monitor. Burns et al. (1999) used benzene and
toluene as tracers of PFW discharged into the North Sea. These compounds are volatile and
so their concentrations in seawater often are very low. For this reason, they may be good
tracers only when offshore platforms discharge high volumes of PFW, which permits to find
detectable concentrations of BTEX into the sea. Cianelli et al. (2008) successfully used a
chemical additive of PFW (DEG diethylene glycol) as tracer of PFW in the Adriatic Sea.
Field measurements are often unpractical and expensive thus an exhaustive assessment of
the effect of PFW discharge on marine environment needs to integrate in a monitoring plan
field observations with results of numerical models of effluent dispersion. These models
allow to perform realistic predictions of temporal and spatial extent of the PFW plume in
selected regions as well as to easily evaluate several different discharge scenarios (dispersion
under different environmental conditions or discharge characteristics).
The application of multidisciplinary monitoring programmes allows a more complete
assessment of the PFW fate and effects in the marine environment. Moreover, integrating
field observations with a modelling approach may provide public administration and decision
makers with useful directions to protect the marine ecosystem.
Modelling and Observation of Produced Formation Water (PFW) at Sea 129
The historical current data, the density profiles of the receiving water column, the
measured DEG concentrations and the outfall pipe geometrical features represented the initial
conditions of the numerical model.
The near-field dispersion of the PFWs discharged from the three gas platforms was
simulated by means of the UM3 (Three-dimensional Updated Merge) model (Baumgartner et
al., 1994; Frick et al., 2002). UM3 is a three-dimensional Lagrangian theoretical model that
quantifies the entrainment by applying both the Taylor entrainment (or shear) and the
projected-area-entrainment (PAE) (or forced) hypotheses (e.g. Winiarski and Frick, 1976)
(Figure 3). The Taylor entrainment is due to the shear between the effluent discharged and the
receiving water, while the PAE entrainment is the rate at which mass is incorporated into the
plume in presence of ambient current (Baumgartner et al., 1994).
UM3 also computes the equations for conservation of mass, momentum and energy at
each time step along the plume trajectory. The numerical output parameters (Table III)
allowed to evaluate the space and time extent of the PFW plume.
The results demonstrated a good agreement between the chemical analyses results and the
numerical model outputs; both data sets highlighted that the PFW plume spread within a
limited layer 2 to 4 meters thick, centered at the source depth (Figure 4).
In low current conditions the initial dilution phase occurred within 15 m or less from the
outfall.
The results showed that also in the Adriatic Sea the PFW dispersion was mainly
modulated by the seasonal stratification for the modelled current speeds. During the summer
the trap level depth of the plume was deeper than in the winter season because of the high
level of stability at the diffuser depth, while the low stratification occurring in winter
conditions supported the vertical rise of the plume within the water column.
v v
a b
Figure 3. Entrainment hypotheses applied by the UM3 model: a) shear entrainment and b) forced
entrainment for a cylindrical fluid element.
Modelling and Observation of Produced Formation Water (PFW) at Sea 131
10
cln summer
bnd summer
11 cln winter
bnd winter
12
13
0 2 4 6 8 10
plume horizontal distance from diffuser (m)
Figure 4. The vertical section of the plume vs the horizontal distance from the diffuser in summer (grey
lines-closed circles) and winter (black lines-open circles) conditions. The solid lines represent the
centerline of the plume while the dashed lines portray the plume boundaries (adapted from Cianelli et
al., 2008).
Length
of Trap Froude Effluent Ambient N Dilution Initial
Platform initial level number density density Frequency factor mixing
mixing depth Sa time
zone
(m) (m) ( t) ( t) (rad s-1) (s)
Summer
DEG
1 15.0 Surface 0.16 23.95 25.75 0.0758 175 90
2 2.3 9.8 0. 66 26.73 24.80 0.3380 55 70
3 14.8 11.5 0.15 24.80 25.02 0.3326 74 400
Winter PFW
1 10.0 Surface 0.09 23.95 29.00 0.0919 170 60
2 6.8 7.8 0.06 26.73 27.78 0.1539 135 185
3 6.7 10.8 0.03 24.80 26.50 0.2323 200 150
The modelled Froude number values (Table III) indicated that the water column stability
at the platform locations influenced the PFWs dilution more than the local current field. In
high stability conditions (summer) the simulations showed low values of the dilution factor Sa
due to the limited extent of the initial mixing zone of the plume (Table III), whereas during
132 D. Cianelli, L. Manfra, E. Zambianchi et al.
winter the weak stratification sustained the dilution of the plume over a wide zone of the
water column. In agreement with other works on PFW dispersion (e.g., Neff, 2002), the
multidisciplinary approach applied in this case study in the Adriatic Sea also indicated rapid
initial dispersion times which cause negligible or non toxic effects on marine organisms
(Manfra et al., 2007).
The integrated numerical-chemical approach applied by Cianelli et al. (2008) allowed to
assess the dispersion processes influencing the potential effects induced on marine
ecosystems by PFW discharges and provided suggestions to optimize the monitoring protocol
presently adopted in the Adriatic Sea. In particular, on the basis of the results of this case
study the monitoring plan in the Adriatic Sea was improved increasing the number and
resolution of sampling stations in the horizontal and vertical around the discharge locations.
CONCLUSIONS
The PFWs currently represent the largest waste stream derived from oil and gas offshore
industry. The disposal of PFW in the marine environment presents the potential for a negative
environmental impact. For this reason the regulatory environmental authorities of the
countries involved in the extraction and production activities have established to continuously
monitor the PFW discharges from offshore oil and gas platforms.
In the last very few decades the comparison between laboratory and field observations
and with numerical modelling results showed that acute effects may occur only within few
hundred meters from the platform locations due to the rapid mixing and dilution of the PFW
plumes. Consequently even the organisms entrained in the plume may be exposed in the
worst cases to concentrations of pollutants decreasing on time scales of tens of minutes or few
hours.
In order to assess the fate and effect of the PFW in the marine environment the most
promising monitoring plans have to integrate field observations and numerical modelling;
such an approach may enhance our ability to understand and manage the potential effects of
the extraction and production activities. The integrated monitoring approach is nowadays
essential to support decisions in the assessment of the risk caused by the PFW discharges on
marine ecosystems and should be available and commonly used by consultants,
administrations and decision makers. An example is provided by the US EPA (Environmental
Protection Agency) and by the Norwegian environmental authorities that usually include the
dispersion modelling tools in the regulation of the offshore PFW discharges (e.g. Ray and
Engelhardt, 1992).
In the context of a cost-effective management of the marine environment, the disposal of
PFWs is an important issue. The systematic application of integrated monitoring programmes
may be very fruitful, as it can provide adequate indications to dispose the PFW properly, so as
to help protect the marine environment while imposing at the same time the least economic
burden to the oil and gas industry.
Modelling and Observation of Produced Formation Water (PFW) at Sea 133
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Modelling and Observation of Produced Formation Water (PFW) at Sea 135
Chapter 5
ABSTRACT
Sulfur dioxide (SO2) is a known pollutant and responsible for various ill effects on
living and non-living organisms. SO2 emissions can be reduced by using non-
conventional energy sources or using conventional fuels containing less sulfur. However,
under the present circumstances SO2 emissions cannot be completely avoided due to the
reasons of rapid industrialization. Various technologies are available for the removal of
SO2 from flue and waste gases. Most of these technologies fall under the category of
physical, chemical or thermal. All these technologies generate secondary pollutants
ending up in disposal problems and also cost prohibitive. Biotechnology offers relatively
cheaper solutions for the conventional problems. Due to this reason, biotechnology is
making in roads into the conventional treatment processes in all the fields. Over the last
decade, efforts have been made to develop biotechnological alternatives to conventional
physico- chemical processes for the removal of SO2 from flue gases. The only method
available at present is Biological flue gas desulphurization (BIO-FGD).SO2 from flue gas
can be absorbed in a suitable organic media. In the aqueous phase SO2 would be
converted to sulfite and some part may again be converted to sulfate due to the presence
of dissolved oxygen. Therefore, the aqueous phase will be having both sulfate and sulfite,
which can be reduced to sulfide using Sulfate Reducing Bacteria (SRB) under anaerobic
conditions. The sulfide formed in the anaerobic reactor could be converted to elemental
sulfur using Sulfur Oxidizing Bacteria (SOB) under partial microbial aerobic conditions.
The elemental sulfur can be used either as a soil conditioner or raw material for industrial
applications. Therefore, BIO-FGD process could be an environmentally benign and
economically viable alternative for the disposal of SO2 emitted from the industries
*
Corresponding author: Telephone: +91-40-27191664, Fax: +91-40-27193159, E-mail: gangagnirao@yahoo.com
138 A. Gangagni Rao and P.N. Sarma
especially from power plants and refineries. The present article reviews the state of art of
BIO-FGD process.
.
INTRODUCTION
Sulfur dioxide (SO2) represents the main fraction of anthropogenic sulfur emissions
worldwide. Anthropogenic SO2 emission is mainly caused by combustion of sulfur containing
fossil fuels such as coal and oil. Thermal based electrical power generating plants account for
nearly 70% of all SO2 emissions. The flue gas contains SO2 in the range of 2000-4000 ppmv
depending on the sulfur content of the coal or fuel that is being used. Another major source of
SO2 is industrial combustion processes, metallurgical operations, roasting and sintering, coke
oven plants, processing of titanium dioxide, pulp production and the thermal treatment of
municipal and industrial wastes. Some non-combustion processes like production of sulfuric
acid, treatment of metallic surfaces and oil refining processes (AIR Trends, 1995) also
contribute to SO2 emissions.
SO2 is a known pollutant and responsible for various ill effects for living and non-living
organisms. (Koren,1991). It has a number of unwanted environmental effects like acid rain
and formation of acidic aerosols. Therefore the need for SO2 removal from flue gases is
evident and acknowledged by many countries (AIR Trends, 1995). Presently ambient air
quality standards specify that SO2 concentration should not exceed 80, 60 and 15 μg/m3 in the
industrial, residential and sensitive areas respectively. Many methods are available for
controlling SO2 pollution in the industrial scenario. One of those methods is discharging into
the atmosphere via stacks. Options for reduction of sulfur emissions at source include (Johns
Eow, 2002; Rosenberg et al., 1975) measures of energy management, increasing the
proportion of non-combustion renewable energy sources of the total energy supply, fuel
switching (e.g. from high to low sulfur coals and or liquid fuels, or from coal to gas), fuel
desulfurization and advanced combustion technologies (e.g. coal gasification combined with
gas desulfurization). Another category of processes which aim at removing already formed
SO2 is referred to as flue-gas desulfurization (FGD) (Pfeiffer, 1975). FGD is probably most
widely used technique for control SO2 emissions from industries (DeNevers, 2000). State-of-
art technologies for flue gas treatment processes are all based on the removal of SO2 by wet,
dry or semi dry (also referred to as wet and dry absorption processes) and catalytic chemical
processes.
Biotechnology is being considered as an emerging technology in environmental
protection, as it involves the use of microorganisms, which are more suitable for pollution
control due to their versatility and minimize high chemical and catalyst costs to a certain
extent. Apart from that microbiological processes operate around ambient temperature and
atmospheric pressure. This eliminates the need of power utilization for heat and pressure and
brings the energy costs down to the minimum. Disposal of by products of conventional
physicochemical systems is another important drawback that may be overcome in a biological
process. Over the last decade, efforts have been made to develop a biotechnological
alternatives to conventional physico- chemical processes for the removal of SO2 from flue gas
Disposal of Sulfur Dioxide Generated in Industries Using Eco-Friendly … 139
(Buisman and Prins, 1994). These biological methods for SO2 removal from flue gas are
either direct or indirect. In direct biological methods, the feasibility of utilization of SO2 as an
electron acceptor from flue gases by Desulfovibrio desulfuricans is studied (Lee and Sublette,
1991). Desulfotomaculum orientis grows on hydrogen (H2), carbon dioxide (CO2), and SO2
with production of hydrogen sulfide (H2S), while D. desulfuricans transforms SO2 to H2S
using minerals and pretreated sludge as carbon source (Deshmane et al., 1993). Selvaraj et al.
(1997a) carried out studies on various immobilized cell bioreactors to maximize the
productivity of the bioreactor for SO2 reduction for effective flue gas desulfurization. The
direct processes for SO2 removal through biological systems like biofilters suffer lot of
limitations, such as the presence of high concentrations of CO2 (10–15%) in the flue gas,
which can be inhibitory to the growth of the microorganisms in the biofilm. Further, direct
biological processes are slow and it is hard to maintain a consistent supply of emissions at
constant flow rate and at appropriate concentrations of the flue gas constituents in the
bioreactor system. This problem can be solved by absorbing SOx selectively in a suitable
solvent and the sulfate/sulfite-rich scrubbed solution can then be reduced to elemental sulfur
using biotechnological approach. (Buisman et al., 1990a; Buisman and Prince, 1994; Cork et
al., 1986; Janssen et al., 1995). In indirect biological processes, studies are carried out by
using Thiobacillus ferrooxidans, for desulfurization of waste gas containing SO2 in a two step
process. In this process, SO2 is scrubbed with ferric sulfate solution and the resultant ferrous
sulfate solution is treated by Thiobacillus ferrooxidans bacteria and some are with sorbent
regeneration process. Lee and Sublette (1991) proposed coupling the reduction of SO2 to H2S
by mixed culture of sulfate- reducing bacteria containing D. desulfuricans to a Claus process
as a means of by-product recovery from a dry generable scrubbing process of desulfurization
of flue gas. Ligyphilip and Deshusses (2003) reported that complete treatment of SO2 from
flue gases could be possible in a two-stage process consisting of a biotrickling filter followed
by biological post-treatment unit. Buisman and Prins (1994) have proposed a new
biotechnological process for flue gas desulfurization which consists of an alkaline SO2
absorption step followed by two biological steps. The first stage of the process is a chemical
one in which flue gases containing SO2 are absorbed in a suitable solvent in the form of
sulfate/sulfite/bisulfite and then biologically reduced to H2S. In the first biological step,
sulfate is reduced to H2S, and in the second biological step H2S is converted into sulfur by
colorless sulfur bacteria (Buisman et al., 1990a; 1991; Jassen et al., 1995). This process is
cheaper than the presently used physicochemical processes and can remove up to 98% of
SO2. Moreover, instead of gypsum or waste sorbent, this process produces sulfur that can be
potentially be reused in the industry. It overcomes the drawbacks involved in direct
biofiltration of SO2 and also includes the benefit of low-cost biological approach. In these
biological processes the biological entities involved are sulfate reducing bacteria
(Desulfovibrio, Desulfotomaculum, etc.), photosynthetic bacteria (genera of the families
Chlorobiaceae and Chromatiaceae), and autotrophic Thiobacillus sp.
A.G.Rao et al (2007) carried out work on microbial conversion of SO2 in flue gas to
sulfide using bulk drug industry wastewater as an organic source by mixed cultures of sulfate
reducing bacteria. In this study mixed cultures of sulfate reducing bacteria (SRB) are isolated
from anaerobic cultures and are enriched with SRB media. Studies on batch and continuous
reactors for the removal of SO2 with bulk drug industry wastewater as an organic source
using isolated mixed cultures of SRB have revealed that isolation and enrichment
methodology adopted in the present study are apt to suppress the undesirable growth of
140 A. Gangagni Rao and P.N. Sarma
anaerobic bacteria other than SRB. Studies on anaerobic reactors showed that the process is
sustainable at COD/S ratio of 2.2 and above with optimum sulfur loading rate (SLR) of 5.46
kg-S/m3/day, organic loading rate (OLR) of 12.63 kg COD/m3/day and at hydraulic
residence time (HRT) of 8 hours. Free sulfide (FS) concentration in the range of 300 – 390
mg-FS/l is found to be inhibitory to mixed cultures of SRB used in the present studies.
Bio -FGD occurs mainly in three steps
(i) Conversion of SO2 to sulfite/ sulfate
SO2 is absorbed either in water or in aqueous slurries of limestone and converted into
sulfite and sulphate in a chemical scrubber (Janssen et al., 1995).
HSO3-+1/2O2 Æ SO42 - + H+
The presence of oxygen in the flue gas results in the oxidation of part of the sulfite
into sulfate.
(ii) Biological conversion of sulfate/ sulfite into sulfide
Sulfite and sulfate thus formed is reduced to sulfide (Widdel, 1988) by Sulfate
Reducing Bacteria (SRB).
HS-+1/2 O Æ Sº + OH-.
HSO3-+1/2O2 Æ SO42 - + H+
SO2 is absorbed either in water or in aqueous slurries of limestone and converted into
sulfite and sulphate in a chemical scrubber (Janssen et al., 1995) which can be further
converted into H2S and then into elemental sulfur by SRB and SOB respectively.
Disposal of Sulfur Dioxide Generated in Industries Using Eco-Friendly … 141
Bandyopadhyay and Biswas (2006) studied the scrubbing of SO2 (initial concentration
ranging between 400 and 1780 ppm) in a tapered bubble column scrubber using water and
dilute sodium alkali. The results indicated that maximum water scrubbing efficiency of SO2
achieved in the tapered bubble column is 58.81 % / (W/m3) at a QL/QG ratio of 8.30 m3/1000
actual cubic meter (ACM). On the other hand, maximum water scrubbing efficiency of SO2 of
34.38 % /(W/m3) is reported at a QL/QG ratio of 10.17 m3/1000 ACM in a standard single
stage bubble column where as almost 100% SO2 removal (i.e., zero penetration) is achieved
in the scrubber developed in alkali scrubbing at an optimum QL/QG ratio of 3.0 m3/1000
ACM.
Biological method involves passing of the flue gas through a biotrickling filter with
Desulfovibrio desulfuricans. Under optimum conditions, this organism is shown to reduce
SO2 to H2S within a contact time of 1-2 s. H2S thus produced is further converted to sulfate.
A drawback in this approach is that D. desulfuricans is a strict anaerobe, and maintaining
anaerobic conditions in biotrickling filters treating flue gases containing on an average 2-8%
residual oxygen remains a challenge (Chou and Lin, 2000).
Ligyphilip and Deshusses (2003) reported that complete treatment of SO2 from flue gases
in a two-stage process consisting of a biotrickling filter followed by biological post-treatment
unit is investigated. The biotrickling filter could remove 100% of influent SO2 from simulated
flue gas at an empty bed residence time of 6 s for a concentration range of 300-1000 ppmv. All
the absorbed SO2 is recovered in the biotrickling filter liquid effluent as sulfite (a product of
chemical reaction of SO2) and sulfate (product of biological oxidation of sulfite). Biotrickling
filter liquid effluent is further processed biologically in a single post-treatment unit consisting of
a combined anaerobic and microaerophilic reactor for simultaneous reduction of sulfate and
sulfite to sulfide and oxidation of sulfide to elemental sulfur. The post-treatment unit is used to
treat effectively the biotrickling filter effluent and produce elemental sulfur. The sulfur
production efficiency of the reactor is about 80% of the SO2 treated.
bed (EGSB) reactor, inoculated with acclimated sulfidogenic granular sludge, operated at 33
o
C, fed with acetic acid as COD source and sulfate as electron acceptor, has a sulfate
conversion efficiency of 80–90% at a high sulfate loading rate of 10.4 g SO42− S/l.d. The
limitations of the EGSB technology with respect to the sulfate conversion rate appeared to be
related to the biomass wash-out and deterioration of granules occurring at superficial upflow
velocities above 10 m/h. Increasing the recirculation flow caused a drop in the sulfate
reduction rate and efficiency (Dries et al., 1998). The gas lift reactor design using H2 as
electron donor achieved so far the highest sulfate reduction rates, up to 30 g SO42− l−1 day−1
(van Houten et al., 1994, 1995b). Yet, this system is limited by mass transfer resistance of the
gaseous substrate H2/CO2 to the biofilm and by the H2 lost when competition with methanogens
and acetogens takes place. In a study carried out by Jan sipma et al (2007), it is observed that
sulfate reduction rates are limited by the amount of CO supplied and its conversion efficiency
(about 85%) at higher CO loads likely resulting from low biomass retention.
and with acetogenic bacteria (AB) for intermediate substrates such as short-chain fatty acids
(VFA) and alcohols is important. It will determine to what extent sulfide and methane, the
end products of the anaerobic mineralization processes, will be produced. The following
factors affect the biological reduction of sulfate and sulfite.
EFFECT OF TEMPERATURE
The sulphate reducing bacteria can be classified into mesophiles (growth
temperature<40°C), moderate thermophiles (growth temperature: 40-60°C) and extreme
thermophiles (growth temperature>60°C) based on their optimum growth temperature.
CONCENTRATION OF O2
SRB are strict anaerobic bacteria that are often in anoxic conditions in their natural
biotopes. One of the most excessive biotopes where sulfate reduction coexists with anoxic
conditions is provided in cyanobacterial mats. SRB are able to deal with temporary exposures
to elevated oxygen concentrations up to 1.5mM (Sigalevich et al., 2000).
functional groups, to denature proteins, and to compete with essential cations. Utgikar et al.
(2001) reported that the insoluble metal sulphide formed is not toxic to SRB by itself but it
blocks the access to substrate and the nutrients that are essential for bacteria by forming a
precipitate coating the SRB.
ELECTRON DONORS
Waste gas scrubbing waters originating from flue gas desulphurization units present a
special problem, since they do not contain organic compounds to support the SRB. In order to
biologically treat these waste streams, an external carbon and energy source has to be
supplied. The choice for the appropriate electron donor is based on the suitability for the
sulfate reduction process, and the availability in large enough quantities at a cheaper or no
cost. In any potential application of SRB bioreactors, the selection of electron donor source is
expected to have an impact on the economics of the process and will be a site-specific design
criterion.
Cork et al; (Cork et al., 1983; Cork, 1982a; Cork and Garunas, 1982b; Cork and Ma,
1982c) proposed a microbial process for the removal of H2S from a gas stream based on the
photosynthetic bacterium Chlorobium thiosulfatophilum as an alternative to the Claus or
Stetford process.
In nature sulfide can be oxidized biologically in three different ways (Kuenen, 1975):
2 HS- + O2 Æ 2 S0 + 2 OH-
sulfur particles is obviously detrimental to the efficiency of the system because it leads to an
increased consumption of the required electron donor and increased sulfide levels in the
anaerobic reactor, which may cause inhibition of the biomass (Koster et al., 1986).
Recognized groups of SOB are mainly phototrophic bacteria and chemotrophic bacteria.
Phototrophic bacteria (green or purple in colour) use light as energy source to reduce CO2 to
carbohydrates. Reduced sulfur compounds are used as electron donors for this reduction. The
reduction takes place under anaerobic conditions. Phototrophic C. limicola is an ideal
bacterium in these biological removal processes due to its ability to grow under anaerobic
conditions using only inorganic substrates and a light source and its efficient extracellular
production of elemental sulfur from H2S. The chemotrophic sulfur bacteria (colourless
bacteria) obtain energy from the chemical aerobic oxidation of reduced sulfur compounds i.e.
spontaneous reaction of H2S or elemental sulfur with dissolved oxygen at the water surface.
SOB can also be classified into lithotrophic bacteria, which use inorganic substances as
source for hydrogen and organotrophic bacteria, which use organic substances as source for
hydrogen. The process for sulfide removal is based on aerobic oxidation by the group of
colorless sulfur bacteria (Kuenen and Beudeker, 1982). To this group belong the organisms
Disposal of Sulfur Dioxide Generated in Industries Using Eco-Friendly … 147
with widely different types of physiology and morphology. Genera belonging to the group of
colorless sulfur bacteria are: Thiobacillus, Thiomicrospira, Thermothrix, Thiothrix, Thiospira
Pseudomonas, Thioovulum, Sulfolobus, Beggiatoa, and Thioploca. In Thiobacillus sp.,
Thiobacillus ferrooxidans, Thiobacillus thiooxidans, Thiobacillus novellas, Thiobacillus
thioparus, Thiobacillus denitrificans etc. are major sulfide reducing bacteria.
Visser et al. (1997a) isolated the dominant autotrophic sulfide oxidising strain present in
elemental sulfur producing bioreactors, which is found to be a new Thiobacillus species,
designated as Thiobacillus sp. W5. This organism is then used as a model organism for
studying sulfur production by thiobacilli in wastewater treatment reactors (Visser et al.,
1997b). Interestingly, the end product of sulfide oxidation at a given sulfide loading rate by
Thiobacillus sp. W5 is almost exclusively elemental sulfur. A very closely related bacterium,
Thiobacillus neapolitanus, converted only 50% of the sulfide to elemental sulfur while the
other 50% is completely oxidised to sulfate. Comparison of the metabolic properties of
Thiobacillus sp. W5 with those of Thiobacillus neapolitanus revealed that Thiobacillus sp.
W5 has a competitive advantage over Thiobacillus neapolitanus in bioreactor environments
because its sulfide oxidising capacity is double that of Thiobacillus neapolitanus.
Interestingly, the maximum specific oxygen uptake rates of the two organisms are very
similar. No other significant biochemical differences are observed between the two
organisms. This means that the limited sulfide oxidation rate of Thiobacillus neapolitanus
gives this species a competitive disadvantage, as it can oxidise only 50% of the incoming
sulfide to elemental sulfur. Thus, these bioreactor environments select species such as
Thiobacillus sp. W5.
2 HS- + 4 O2 Æ 2 SO42- + 2 H+
Previous studies by Janssen et al. (1995) have demonstrated that the molar oxygen to
sulphide ratio for sulphur production would be around 0.6 –1.0. However, in practical
situation such as a real wastewater treatment plant, it is difficult to maintain a narrow
sulphide/oxygen ratio. On the other hand maintaining optimum redox potential could control
sulphide oxidation more precisely (Janssen et al., 1998; Khanal and Huang, 2003). At reactor
dissolved oxygen (DO) concentrations higher than 0.1 mg l-1, sulfate is the main product of
148 A. Gangagni Rao and P.N. Sarma
sulfide oxidation. At DO less than 0.1 mg l-1, sulfur is the major end product of the sulfide
oxidation.
Effect of pH
Effect of Temperature
SUMMARY
Biotechnological methods for the removal of SO2 gas from flue gas streams is a
promising alternative when compared to the other available physical and chemical methods
especially flue gas streams having low SO2 concentrations. This process is also applicable for
the removal of SO2 in tail gases emitted from processes treating high concentrations of SO2.
The biotechnological methods involve exploitation of microbial metabolism to convert the
toxic SO2 compound into inoccus elemental sulfur. The BIO-FGD method is a three step
process; 1. Absorption of SO2 gas and conversion of the gas into sulfites and sulfates. 2.
Anaerobic treatment of sulfite and sulfate containing solution by suitable microbial consortia
for conversion into sulfides 3. Aerobic oxidation of sulfides to convert into elemental sulfur
by microorganisms.
Conversion of SO2 into sulfites and sulfates is a physical and chemical process which is
not elaborated in the present review. Extensive discussions are done for anaerobic treatment
of sulfites and sulfates by sulfate reducing bacteria. Metabolism of SRB requires organic
carbon as carbon source and sulfates and sulfites as energy sources. Proper selection of
suitable organic sources and the ratio of COD/SO4 2- play a major role in efficient removal of
sulfates and inhibition of methanogenesis. The anaerobic sulfate reducing bacteria are
sensitive to variation in environmental conditions such as pH, temperature, presence of O2
and presence of metal ions. Controlling of all the parameters and optimization is crucial for
the efficient operation of the sulfate reducing reactor.
In the biotechnological process, dissolved sulfide (HS-) is converted to elemental sulphur
by the aerobic metabolism of SOB. The insoluble sulfur can easily be removed from the water
stream. Conversion of sulphide to elemental sulphur is a sensitive step and is dependent on
dissolved oxygen concentration. Temperature, sulphide loading rate and pH are the other
critical parameters in aerobic sulphide oxidation process. Optimization of parameters in all
stages of BIO-FGD process is crucial for efficient removal of SO2 gas from flue gas streams.
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154 A. Gangagni Rao and P.N. Sarma
Chapter 6
ABSTRACT
Since the requirement for nutrient removal is becoming increasingly stringent, a high
efficiency of nitrogen removal is necessary to achieve a low total nitrogen concentration
in the effluent. Biological nitrification and denitrification processes are generally
employed to remove nitrogen from wastewater. Unfortunately, these processes are not
suitable to treat wastewater with a low COD/N ratio because it involves the addition of an
external organic carbon source and, therefore, an increase of the operational costs.
Several alternative processes for nitrogen removal can be applied in order to reduce
partially (―nitrite route‖) or totally (anammox, autotrophic denitrification) the organic
matter required. Such processes suppose not only an economical way to treat these
wastewaters but they are also more environmentally friendly technologies (lower
production of CO2, N2O and sludge; lower energy consumption). Up to now, they were
basically applied to the return sludge line of municipal wastewater treatment plants
(WWTPs). However, these processes could even be implemented in the actual WWTPs
in order to achieve more compact and energy efficient systems.
Their potential advantages can make them also feasible technologies to treat polluted
ground water or to remove nitrogen compounds from recirculating aquaculture systems.
1. INTRODUCTION
The main reserve of nitrogen is located in the atmosphere, where it is found as N2, but
this molecule can not be directly used by most living beings. Both bacteria and blue-green
algae can carry out nitrogen fixation converting N2 into ammonia and nitrate. Once nitrogen
has entered the biosphere via biological fixation, it is subject to a series of conversion steps,
from plant protein to animal protein, and finally ends up in dead organic matter. When this
organic matter is mineralised in the soil, most of the inorganic nitrogen compounds produced
(ammonium, nitrite, nitrate) are taken up again by plant or microbial biomass for the production of
protein. The nitrogen cycle also includes electrochemical reactions (electrical storms)
although their importance is small when compared to biological processes (Gijzen, 2001).
Human activity is changing the rate of some processes involved in the natural nitrogen
cycle which causes accumulation of nitrogenous compounds in water and, therefore, its
associated pollution. The most harmful anthropogenic processes are: a) Mining and the use of
fossil fuels since inactive nitrogenous compounds are introduced again into the active cycle
and b) fixation of nitrogen from the atmosphere by chemical processes (fertilizer industries)
and by intensive cultivation of nitrogen fixation plants.
The problems generated by the nitrogenous compounds in wastewaters will depend on
their oxidation state:
Ammonia can have toxic effects on aquatic life; these effects may be either acute
(i.e., fish mortality) or chronic (impacts on reproduction, tumours, etc.). The presence
of ammonia also causes algal blooms (eutrophization) which influence the water
system in two ways. First, they hamper the penetration of sunlight, causing death of
underwater grasses. Secondly, the decomposition of dead algae causes depletion of
oxygen, which is normally essential to most organisms living in water.
Nitrite can bind to iron on hemoglobin reducing transfer of oxygen to cell tissues; the
result is suffocation accompanied by a bluish tinge of the skin. Nitrite and nitrate in
drinking water have been medically linked to methemoglobinemia, a sometimes fatal
blood disorder affecting infants (―blue baby syndrome‖). On the other hand, the
disposal of wastewater containing nitrite and/or nitrate can generate both NO and
N2O by an incomplete denitrification. These compounds contribute to the destruction
of the ozone layer.
Nitrate pollution impedes the production of drinking water. During chlorination of
drinking water, carcinogenic nitrosamines may be formed by the interaction of nitrite
with compounds containing organic nitrogen.
To avoid these problems, legislation has imposed maximum disposal limits for
nitrogenous compounds. These limits can be achieved by physico-chemical or biological
treatments, the last one being the most used for economical reasons.
assimilation, denitrification and anammox, which are carried out by different microorganisms
(Figure 1).
The COD/N ratio of wastewater will determine which of these biological processes is the
most suitable to remove nitrogen:
2.1. Nitrification-Denitrification
During denitrification both nitrate and nitrite are reduced to nitrogen gas under anoxic
conditions, organic matter being used as electron donor (Equation 3). This process is carried
out by denitrifying bacteria.
on
icati NH+4+
o nif mpo
sitio
n NH4
Ammonium
m Deco
Nit
Am Faecal s
is i
rifi
ol
An
matter
dr
Fixation
Hy
ca
amm
ti on
N org Urine
ox
n
Animal Urea tio
protein ila N2
s im
As N2
Atmospheric
Anim
on Nitrogen NO2-
s i ti
po ion
at
al
m
o Fix
n
ec
Chemical
feed
industry
tio
D
a
ing
fic
N org
io
tri
at
Vegetal
ni
fic
De
protein
tri
Ni
Assim NO3-
il ation
Predenitrification N2 O2
Wastewater NH4+
COD Denitrification Nitrification Effluent
NH4+ COD
NO3-
Postdenitrification O2 N2
Wastewater NO3-
Nitrification Denitrification Effluent
COD
NH4+
COD
Integration of both nitrification and denitrification also allows reducing the amount of
chemicals needed to control pH during the treatment since alkalinity generated during
denitrification compensates for the pH decrease due to nitrification.
Nitrification occurs under aerobic conditions while anoxic conditions are necessary for
denitrification. This implies the use of two different tanks which can be provided in two
possible configurations (Figure 2): a) Predenitrifying configuration—wastewater is fed into
the denitrifying reactor and later nitrification is carried out. A stream from the aerobic tank
containing nitrate and/or nitrite is recirculated to the first unit to carry out denitrification.
Therefore, nitrogen removal efficiency depends on the recycling ratio; b) Postdenitrification
configuration—wastewater is fed into the nitrifying unit and its effluent enters into the
denitrifying reactor. This configuration is very simple, easy to control and no recycling is
needed. Nevertheless, organic matter is oxidized in the aerobic unit and an external carbon
source must be added for denitrification which increases the operational costs. This
configuration is only used when the COD/N ratio of wastewater is low.
In the predenitrification configuration, the organic matter coming from the denitrification
unit causes the proliferation of heterotrophic bacteria in the aerobic unit. These
microorganisms compete with nitrifying bacteria for oxygen and its concentration must be
maintained at levels around 1–2 mg O2/L to avoid the failure of the nitrification process. On
the other hand, the concentration of nitrifiers in the aerobic tank is low due to their slow
growth rate and, therefore, the required volume of the aerobic unit is high. The use of carrier
material in this unit would increase the concentration of nitrifiers and, then, decrease the
required volume [Pegasus system (Tanaka et al., 1996)].
(T)
O2 S [4]
max
KO2 O2 K S (1 NH3 )
S
KINH3
where μmax is the maximum growth rate (d-1), O2 the dissolved oxygen concentration (mg
O2/L), KO2 the oxygen affinity constant (mg O2/L), S the substrate concentration (NH4+ for
ammonia-oxidizers and NO2- for nitrite-oxidizers) (mg N/L), KS the substrate affinity
constant (mg N/L), NH3 the free ammonia concentration which inhibits both ammonia- and
nitrite oxidizers (mg N/L) and KINH3 the free ammonia inhibition constant (mg N/L).
1 mol Nitrate
(NO3-)
40%Organic matter
Nitrifiers
Heterotrophs
Aerobic-Nitrification
Anoxic-denitrification
25% O2
75% O2
Partial nitrification-denitrification
Nitrification-denitrification
Advantages;
4.6 g O2/g NH4+-N oxidized 25% Reduction of O2 demand
7.5 g DQO/g NO3--N reduced 40% Reduction of required organic matter
40% Reduction of biomass generated
Table 1 shows the values of the different kinetic parameters. Since the oxygen affinity
constant of nitrite-oxidizers is higher than that of ammonia-oxidizers, a decrease of the
dissolved oxygen in the reactor would exert a higher effect on the former one. Therefore,
nitrite generation would be favoured at low dissolved oxygen concentrations (Figure 4a).
However, this operational strategy implies also the decrease of the ammonia-oxidizers
activity and a control system for oxygen is required when the influent characteristics are not
constant.
It is also known that ammonia oxidation is inhibited by higher free ammonia
concentrations than nitrite oxidation. Therefore, partial nitrification could be achieved by
maintaining free ammonia levels in the reactor which only causes the inhibition of nitrite-
oxidizers (Figure 4b). Albeit, the maintaining of a certain concentration of free ammonia in
the system means that the effluent does not fulfil the disposal requirements. Another
disadvantage of this strategy is the possible adaptation of nitrite-oxidizers to free ammonia
and the restoration of their activity.
Since the activation energy of ammonia oxidation is higher than that of nitrite oxidation,
an increase of temperature will cause a higher effect on the first step. In practise, the ammonia
oxidation rate is higher than the nitrite oxidation rate at temperatures higher than 30 ºC
(Figure 4c). From an economic point of view, this strategy would be only feasible when
effluents have already such temperature, for example effluents of anaerobic digesters.
These bacteria belong to the phylum Planctomycetes and their yield coefficient is low
(Y= 0.038 g VSS/g NH4+-N). This low sludge production reduces the management costs but
makes the start-up of anammox reactors quite long. For this reason, systems with high
biomass retention capacity must be used. Another characteristic of these microorganisms is
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 161
the decrease of their activity in the presence of oxygen, nitrite or organic matter (Dapena-
Mora et al., 2004; 2007).
1
A
ammonia
0.6
0.4
nitrite
0.2 oxidizers
0
0 1 2 3 4
O2 [mg/L]
1.2
ammonia
B
Growth rate (d -1)
1
oxidizers
0.8
0.6
0.4 nitrite
0.2 oxidizers
0
0 0.2 0.4 0.6 0.8 1
NH3 [mg N/L]
3
C
2.5
Growth rate (d -1)
ammonia
oxidizers
2
1.5
1
nitrite
0.5 oxidizers
0
10 15 20 25 30 35 40
T [ºC]
Figure 4. Possible strategies to carry out partial nitrification by changing dissolved oxygen
concentration (A), free ammonia concentration (B) and temperature (C).
To apply the anammox process, the effluent should contain suitable concentrations of
both ammonia and nitrite. Ammonia is generally present in wastewater but nitrite is not and
has therefore to be generated by the oxidation of 50% of the ammonia. During partial
nitrification, organic matter is also oxidized which prevents its possible negative effects on
the anammox reactor. The combination of anammox and partial nitrification to treat
wastewaters with high nitrogen content and without organic matter gives some advantages
compared to the conventional nitrification-denitrification process (Figure 5): 1) The oxygen
162 J.L. Campos, J.R. Vázquez-Padín, M. Figueroa et al.
requirements are 60% lower; 2) No organic matter must be added; and 3) Sludge production
is 85% lower (Fux and Siegrist, 2004).
17 g CODbiomass
1.5 g CODbiomass
Under limiting oxygen conditions (lower than 0.5% of air saturation) a mixed culture of
both ammonia-oxidizers and anammox bacteria can be obtained. This culture converts
ammonia directly into nitrogen gas with nitrite as intermediate product. Nitrifiers consume
oxygen and generate both nitrite and an anoxic environment for anammox microorganisms.
Then, ammonia can be removed in a single unit under autotrophic conditions. Different
acronyms were used to define this process: OLAND (Oxygen-Limited Aerobic Nitrification
and Denitrification) (Windey et al., 2005), aerobic deammonification (Wett, 2006) and
CANON (Completely Autotrophic Nitrogen removal Over Nitrite) (Sliekers et al., 2002;
2003). The two former names are based on the idea that the own ammonia-oxidizers carried
out the denitrification process. However, nowadays, it was demonstrated that anammox
bacteria are responsible of the denitrification process, the last acronym being the most
suitable to define the process.
Two possible strategies to start-up a CANON system are possible: 1) to inoculate an
anammox reactor with nitrifying biomass and to supply air to maintain microaerobic
conditions or 2) to operate a nitrifying reactor under oxygen limited conditions to obtain the
desired ammonia to nitrite molar ratio inside the system and then to inoculate anammox
biomass (Pynaert et al., 2004; Gong et al., 2007). The second strategy seems to be more
suitable because an important decrease of the anammox activity is observed when the first
strategy is applied (Sliekers et al., 2002; 2003; Liu et al., 2008). Moreover, only a few amount
of anammox biomass is necessary to start-up the CANON process with this second strategy
(Vázquez-Padín et al., 2009a; 2009b).
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 163
Sulfur compounds can be present in wastewaters together with carbon and nitrogen
compounds and the interactions between the biological cycles of the three elements can be
used to remove them (Figure 6).
The biological interaction between sulfur and nitrogen cycles is given by autotrophic
denitrification which consists on the reduction of nitrogen oxides (NO3- and/or NO2-) into
nitrogen gas by using reduced sulfur compounds as electron donors (S2O3-2, Sº and/or H2S)
(Equations 6, 7 and 8). The end product is sulfate which is less harmful than nitrate,
especially when the effluent is disposed in a marine environment.
These bacteria can also use oxygen as electron acceptor to oxidize sulfur compounds. The
end product is sulfate under high dissolved oxygen levels while, at low oxygen concentrations
( 0.1 mg O2/L), only a partial oxidation into S° occurs (Equations 9 and 10). If the aim of the
treatment is to remove nitrate, the presence of oxygen must be avoided since the
microorganims will preferentially use it as electron acceptor.
Bioaugmentation of nitrifying bacteria in the flocculent sludge could reduce the required
volume of the aerobic tank. This may be achieved by promoting nitrification of the sludge
digester effluent with biomass coming from the return sludge stream. Only part of this stream
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 165
should be applied to the effluent in order to maintain its temperature around 20-30 °C. This
process has been successfully applied at industrial scale and was named BABE (Salem et al.,
2002a; Berends et al., 2005). Another advantage of its application is the increase of the
denitrification capacity of the WWTP since part of the volume of the aerobic tank is not
required and can be operated under anoxic conditions.
The BABE process is operated with denitrification in order to control pH. In this aspect,
the sludge supplied to the process allows minimizing the amount of external organic matter
added. Under anoxic conditions, the electrons needed to reduce nitrate are given by
endogenous respiration of sludge. The BABE process can be carried out in a system with one
or two units. The system with one only unit is operated in cycles. Firstly, the reject water and
sludge are fed to the system to nitrify under aerobic conditions. During a second stage,
denitrification occurs and sludge partially settles. At the end of this stage, the liquid fraction
and the non settled sludge are fed to the main stream of the WWTP. The two reactors
configuration consists of an anoxic reactor following by an aerobic tank (Figure 8). In the
anoxic tank, the supernatant of the sludge digester is mixed with the sludge which also acts as
carbon source and even external organic matter could be added if necessary. A recirculation
between both units is maintained in order to supply nitrate to the anoxic tank.
The BABE process with a single unit has been tested at industrial scale in the WWTP of
Garmerwolde (The Netherlands) with a capacity of 300,000 inhabitants-equivalent (Salem et
al., 2004). The implementation of this process allowed improving ammonia concentration in
the effluent from 13.3 to 5.2 mg NH4+-N/L. In order to upgrade the WWTP of Walcheren to
fulfil disposal requirements ( 10 mg N/L) (140,000 inhabitants-equivalent, The Netherlands)
two alternatives were evaluated: An increase of both anoxic and aerobic tanks and the
implementation of the BABE technology. The second option allowed reducing 50% of the
required area and supposed saving costs of 115,000 Euros/year (Salem et al., 2002b) (Table 2).
This technology has also been implemented in the WWTP of Hertogenbosch (350,000
inhabitants-equivalent, The Netherlands).
The SHARON process was developed in 1997 and the first plant at full scale was built
the same year. Nowadays, full-scale sludge liquor treatment with partial
nitrification/denitrification in SHARON reactors has already been introduced in 7 WWTPs
(van Loosdrecht and Salem, 2006) (Table 3). Economic balances demonstrated the cost
savings using the combined SHARON-denitrification processes to treat reject water in
comparison to physicochemical processes or conventional nitrification-denitrification
processes (Table 4). The costs distribution of this treatment are: 47% installation costs, 15%
energy costs, 4% maintenance, 8% working costs, 18% methanol and 7% management of
sludge produced.
Table 4. Estimation of costs for nitrogen removal in the sludge line for a WWTP of
500,000 inhabitants-equivalent (van Kempen et al., 2001).
3.3. SHARON-Anammox
Secondary
Sludge return settler
Anammox
reactor
Water line
Sludge line
Sludge
digester
Dehyrated
sludge Dehydration
system
In the one-unit system, partial nitrification and anammox processes are simultaneously
carried out under microaerobic conditions. For this propose systems with flocculent
ammonia-oxidizing and anammox granular biomasses can be used. Another option is the
utilization of granular biomass where ammonia-oxidizing bacteria grow in the outer layers
consuming oxygen and generating nitrite and, therefore, suitable conditions are promoted in
the inner layers to develop the anammox process (Figure 11). Under a practical point of view,
systems of one unit are preferred because higher removal rates can be achieved (smaller
reactors) and their N2O emissions are low although systems with two units are more flexible
and stable against influent fluctuations (Kampschreur et al., 2008).
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 169
PARTIAL
+
NITRIFICATION NH4
O2 ANÓXICA
AEROBIC
NO2- ANOXIC
NO2-
N2 + NO3-
NH4+
ANAMMOX
Figure 11. Simultaneous partial nitrification and anammox processes in granular systems.
Up to now, there are 4 anammox plants operating at full scale (Abma et al., 2007), three
of them in The Netherlands and one in Japan (Table 5). All of them have reached their design
capacity treating wastewater from different origins which indicates the wide applicability of
the process. It is important to point out that the length of the first reactor (72 m3) start-up was
3 years while the fourth plant was started up in 2 months (Van der Star et al., 2007). This fact
was mainly due to a higher knowledge about anammox process and a greater availability of
inoculum.
Table 5. Full-scale anammox plants around the world (Abma et al., 2007).
Design Load
capacity achieved
Waterboard
Hollandse Delta,
IWL,
The Netherlands
Waterstromen,
The Netherlands
Potato
(1 unit) processing 1200 7001 6 months
Mie prefecture,
Japan
Under the denomination of deammonification (DEMON), there are other two full scale
plants in Austria (500 m3, 300 kg N/d) and Switzerland (400 m3, 250 kg N/d) treating the
supernatant of sludge digesters (Wett, 2006; 2007). The start up length of the first plant was
2.5 years while the plant located in Switzerland, inoculated with sludge from the first one, got
its design load in only 50 days.
Economic estimations were calculated with data obtained from these full scale plants and
compared to those obtained when the nitrification-denitrification process is used. The results
showed important economic and environmental benefits (Table 6).
Nitrification- Partial
denitrification nitrification-
Anammox
Energy (kWh/kg N) 2.8 1
Methanol (kg/kg N) 3 0
Sludge production (kg VSS/kg 0.5-1.0 0.1
N)
CO2 emissions (kg/kg N) 4.7 0.7
Total costs1 (Euros/kg N) 3-5 1-2
1
Capital and operational costs included.
The application of this process to reject water would suppose important saving costs
(Siegrist et al., 2008). Most of the municipal WWTPs designed only for organic matter
removal were equipped with primary settlers with a hydraulic retention time (HRT) of 2–3 h
to reduce the organic matter applied to the biological reactor. Albeit the requirement of
nitrogen removal caused that the HRT of primary settlers was reduced to less than 1 h in
order to ensure the availability of organic matter during denitrification. This fact implied the
decrease of biogas generated during sludge anaerobic digestion. If the partial nitrification-
anammox process is applied to treat reject water, the denitrification capacity of the WWTP
could be decreased without negative effects on the overall nitrogen removal efficiency. This
would allow increasing again the HRT of the primary settler to enhance production of biogas
and significantly decrease the requirement of oxygen in the aerobic tank to remove organic
matter. On basis of a nitrogen removal efficiency of 75%, the treatment of reject water would
allow decreasing a 25% the denitrifying capacity. Then, this process would need 25% less
organic matter which can be separated in the primary settler by adding flocculant increasing
the biogas production in 25% (Figure 12).
Taking into account that oxygen requirements for organic matter removal and
nitrification suppose 70-80% of the total energetic costs of the plant, this new configuration
could achieve a reduction up to 50% of the energy consumed (Table 7).
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 171
Figure 12. N and COD balances of a municipal WWTP with two possible configurations of the primary
settler: a) HRT: 0.5-1 h; b) HRT: 2 h, addition of flocculant and treatment of reject water with the
partial nitrification-anammox process (Adapted from Siegrist et al., 2008).
Supernatant
Nitrification
Nitrification space
Limiting Limiting
factor BABE
process
Denitrification
Aeration
capacity
Limiting
factor Denitrification
space
Acetate
Counter-ion SHARON
Organic NH4+
matter
HCO3- SHARON
ANAMMOX
Figure 13. Diagram of selection of the best technology to treat reject water.
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 173
NO3-
Activated sludge
reactor Settler
Influent Sludge
Anaerobic tramp Effluent
Anoxic
This technology is based on a two units configuration. The first unit is an anaerobic
reactor with three compartments (2 anaerobic + 1 anoxic) containing flocculant sludge. The
second unit is a conventional activated sludge system with a settler. In the first unit, anaerobic
digestion of organic matter and sulfate reduction into sulfide are carried out. During
autotrophic denitrification, sulfide is again oxidized into sulfate with the nitrate coming from
the effluent recirculated.
Secondary
settler
Thickening tank
Thickening tank Secondary sludge
Primary sludge
Water line
Thermal hydrolysis
reactor Dehydration Sludge line
system
Biogas
Dehydrated
sludge
Sludge digester
Crystallization CANON
unit system
P recovery
4. APPLICATION PERSPECTIVES
4.1. Municipal WWTPs Improvement
Excess sludge treatment and disposal of conventional WWTPs supposes between 50 and
60% of operational costs. For this reason, recently, a great effort in the development of new
technologies to reduce sludge production was done (Kroiss, 2004).
When a sludge digester is already present in the WWTP, the implementation of a sludge
disintegration unit (for example, thermal hydrolysis) previous to the anaerobic digester is the
best option to maximize the recovery of energy from the sludge (Figure 15).
This treatment would allow an increase in methane production and would decrease the
HRT of the sludge digester. Ammonia concentration in the digester supernatant would
increase, the application of a CANON system would be even more profitable from an
economical point of view. This system could be operated to obtain an effluent with a
stoichiometric NH4+/PO4-3 ratio to obtain struvite (Equation 11). The phosphorus recovered
can be used as fertilizer.
This proposed scheme would allow: a) Reducing the size of the WWTP; b) Obtaining a
positive net balance of energy in the WWTP; c) Reducing both sludge generation and CO2
emissions.
In the last years, nitrate levels in ground waters exceeding the European Regulation (11.3
mg NO3--N/L) were observed. The conventional method to remove nitrate is ionic exchange
although the application at full scale of reverse osmosis also gave good results. Nevertheless,
both processes generate a residual stream which needs a postreatment. An alternative to these
technologies is the denitrification. In the case of heterotrophic denitrification, organic matter
(ethanol or methanol) must be added as electron donor that leads to a secondary
contamination. This can be avoided if nitrate removal is done by autotrophic bacteria using
elemental sulfur since it is not a toxic compound and it is insoluble in water. This process will
generate sulfate and is recommended to apply to ground water with low endogenous sulfate
levels to avoid sulfate concentrations higher than 400 mg SO4-2/L.
The application of autotrophic denitrification to ground water has been limited by the low
biomass retention. Therefore, recent works are focused on combining this process with
membrane (McAdam y Judd, 2006) or biofilm technologies (Soares, 2002) to achieve a
complete retention of the biomass. The configurations proposed are the following (Figure 17):
(a) Bioreactor with extractive membrane: In this configuration, nitrate is extracted from
water by molecular diffusion through the membrane to a stream containing both
denitrifying biomass and electron donor (Figure 17a).
Biogas
FeCl3 + Flocculant
(SS and P removal)
Influent
Effluent
Primary
settler
Figure 16. Municipal wastewater treatment using a CANON system to remove nitrogen.
(b) Bioreactor with filtration membrane: Denitrifying biomass is mixed with polluted
ground water and electron donor. In this case, the membrane is used to separate
biomass from treated water by application of pressure (Figure 17b).
(c) Biofilm reactor: Elemental sulfur particles could be used as both electron donor and
support of autotrophic denitrifying biomass. A column filled with elemental sulfur
granules and operated in an upflow mode could be a system very simple, stable and
easy to maintain (Figure 17c).
176 J.L. Campos, J.R. Vázquez-Padín, M. Figueroa et al.
Recirculation
A Treated Denitrifying bacteria
Water with nitrate
water Sulfur
Membrane
Denitrifying
biomass
Residual
water Water with Membrane
nitrate
Denitrifying Membrane
biomass
Residual
water Membrane
C Water Denitrifying
with sulfate bacteria
Water
with nitrate
Figure 17. Systems to remove nitrate from ground water: a) bioreactor with extractive membrane, b)
bioreactor with filtration membrane and c) biofilm reactor.
Factors such as limitations of water quality, land costs, disposal requirements and
environmental impact are driving the aquaculture sector to more intensive practices. The use
of recirculating systems allows reducing water used and disposed during aquaculture
activities. Besides, it has another advantages: a) Saving of pumping costs; b) Control of pH
and temperature which optimize fish production; c) Presence of pathogens is minimized
which reduces mortality during the broodstock stage.
Since ammonia is toxic for fish at concentrations higher than 1.5 mg NH4+-N/L, this
compound must be removed by a nitrifying biofilter to avoid its accumulation in the system.
Ammonia is oxidized into nitrate which is less toxic for fishes, its recommended limit being
around 50 mg NO3--N/L. However its effect depends on the specie and growth stage and,
therefore, its removal is advisable. The use of denitrifying biofilter with elemental sulfur
would be the most suitable option to maintain nitrate concentration as low as possible (Figure 18).
The sulfate generated during the autotrophic denitrification would cause neither
environmental nor toxicity problems when marine species are cultured (Vidal et al., 2002).
Novel Biological Nitrogen-Removal Processes: Applications and Perspectives 177
5. CONCLUSIONS
Implantation of bioaugmentation, partial nitrification or anammox processes in the
reject water stream of WWTPs supposes an economical and feasible alternative to
improve effluent quality in terms of nitrogen content.
The most suitable technology must be chosen depending on the WWTP operational
conditions.
Perspectives of advanced nitrogen removal processes application are very promising
in fields such as wastewater, drinking water or aquaculture systems.
UV unit
Effluent
NH4+
Solid Organic matter
Recycling water wastes
Denitrifying
biofilter
Nitrifying
SO4-2 NO3-
biofilter
ACKNOWLEDGMENTS
This work was funded by the Spanish Government (TOGRANSYS project CTQ2008-
06792-C02-01/PPQ and NOVEDAR_Consolider project CSD2007-00055).
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 179-200 © 2010 Nova Science Publishers, Inc.
Chapter 7
ABSTRACT
This review will discuss the melanoidin-decomposing activity (MDA) among
microorganisms. The focus will be on the potential use of the microbial-MDA to treat the
wastewater discharged from factories using molasses as the raw material (molasses
wastewater: MWW) because molasses is one of the most useful raw materials in various
types of industries, such as the fermentation and animal feed industries. However, the
wastewater discharged from factories using molasses contains a large amount of dark
brown pigment, melanoidin pigment: MP, which is poorly decomposed and/or
decolorized by normal biological treatment processes, such as the activated sludge or
anaerobic treatment systems (anaerobic pond or anaerobic contact digester), because, the
microorganisms in those wastewater treatment systems showed very poor MDA. The
distribution of MDA among microorganisms and the mechanism of decomposing
activities, in particular, were reviewed. Also, the application of the isolated strains having
the MDA to treat molasses wastewater in the wastewater treatment plant was tested.
*
Corresponding author: E-mail: suntud.sir@kmutt.ac.th
182 Suntud Sirianuntapiboon and Sadahiro Ohmomo
INTRODUCTION
A large body of scientific research has been conducted on the conversion of renewable
biomass to useful materials for political and practical concerns over the state of the
environment (Underfolker and Hickely, 1954). In particular, bio-fuel from renewable biomass
is expected to help alleviate the ongoing energy crisis. Molasses, a by-product from sugar
cane or sugar beat in sugar industries, is one of the largest sources of biomass and a very
important material worldwide.
Molasses consists of about 50% sugar (reducing sugar) or related substrates, about 10%
non-sugar organic substances, and about 10% minerals (Underfolker and Hickely, 1954). Due
to these components, molasses was widely used in various fermentation industries, such as
alcohol fermentation, amino acid fermentation, antibiotics fermentation, and baker‘s yeast
fermentation, (Chang and Yang, 1973; Chaung and Lai, 1978), as a low-cost and readily
available raw material when diluted with water. Molasses from sugar cane is mainly produced
in tropical areas of the world, especially in south-east Asia (Philippines, Indonesia, Thailand,
etc).
Wastewater from the fermentation processes using molasses is densely colored by
molasses pigment, called melanoidin pigment (MP), and contains a large amount of organic
matters, which leads to high biological oxygen demand (BOD5) and chemical oxygen demand
(COD) values (Sirianuntapiboon et al., 1988a; Antonia et al, 2000). Therefore, the wastewater
can be treated by normal biological treatment processes, such as the activated sludge system,
aerated lagoon, or anaerobic pond, to remove the organic matter. However, MP is poorly
decomposed and still remains in the wastewater after treatment by above processes. No
suitable method for the treatment of large amounts of this type of wastewater has been
developed yet, so this is a problem that still needs to be solved. For example, this problem has
led to an increase in the production cost of the ethyl alcohol fermentation process from
molasses, because about 75% of potential energy in ethyl alcohol is wasted due to the popular
treatment process as concentration and combustion of the wastewater (Chaung and Lai, 1978;
Chang and Yang, 1973). Therefore, the development of a low-cost and simple wastewater
treatment system that utilizes microbes to decompose and decolorize MP is urgently needed.
In this paper, the MP-decomposing activities (MDA) in microorganisms are reviewed
with a focus on the distribution of MDA in microorganisms, the mechanisms of their
activities, and the application process for the treatment of wastewater from the factories using
molasses.
100,000 Dalton) and the condensed-MP precipitates under acidic conditions of pH less than 3.
The color-density of MP solution under acidic conditions (pH 5.0) is weakened by about 20%
in comparison with the alkaline condition and shows a maximum absorption at a wavelength
of 475 nm (Ohmomo et al, 1985a; Sirianuntapiboon et al, 1988a). The colored substances in
some foods, such as Shoyu and Miso (Japanese seasoning), as well as molasses, are typical
MPs (Ohmomo et al., 1985a).
Two kinds of MP solutions were used in screening the microorganisms for MDA and MDA
determination as natural-MP (NMP) and synthetic-MP (SMP) solutions. The NMP solution was
fundamentally prepared from the MWW (Ohmomo et al., 1987a; Sirianuntapiboon et al.,
1988a). Two kinds of MWW as stillage from an alcohol factory (U-MWW) and treated-MWW
especially from an anaerobic pond (An-MWW) could be used for preparation of the NMP
solution. However, the molecular weight distribution of NMP solution was not always
184 Suntud Sirianuntapiboon and Sadahiro Ohmomo
equalized, because of the conditions of the sugar making process, fermentation of molasses, and
treatment process and condition of MWW (Sirianuntapiboon and Chairattanawan, 1998). Due
to these uncertainties, synthetic-MP (SMP) was more widely used for the screening of
microbial-MDA. SMP was synthesized by heating the solution containing 1 mol/L glucose, 1
mol/L glycine and 0.5 mol/L Na2CO3 at 1210C for 3 hr (Sirianuntapiboon, et al., 1988a;
Ohmomo et al., 1985a). After heating, the solution was adjusted to a pH of 7.0 with 1.0 mol/L
NaOH solution and ultra-filtrated using membrane filters of molecular weight cut-offs between
1,000 Dalton and 10,000 Dalton. The fractions having molecular weights from 1,000 to 10,000
Dalton was harvested and freeze-dried to make a SMP powder. The solution of SMP is prepared
as giving an optical density of 3.5 (OD=3.5) at a wavelength 475 nm in 0.1 mol/L acetate buffer
(pH 5.0) before being used in the experiments.
The medium containing SMP or NMP was used for screening the microorganisms having
MDA. Fungal and bacterial strains were isolated by using media containing MP. The tested-
microorganisms were inoculated on the surface of an agar medium suitable for growth (the
media contained MP) and were cultured to make a colony. If the tested-microorganism had
MDA, a clear zone around the colony was formed (Sirianuntapiboon, et al., 1988a).
Furthermore, the microorganisms forming the clear zone around the colony was cultured in a
liquid medium suitable for the growth and the medium color intensities before and after
cultivation were compared in order to calculate the decolorization yield (Sirianuntapiboon, et
al., 1988a). If the tested-strain produced organic acids and reduced medium pH, the color
density of the culture filtrate was reduced due to the acidic-pH. Therefore, the color intensity
of culture filtrate should be measured after dilution with 0.1 mol/L acetate buffer (pH 5.0) to
prevent the error of the color intensity reduction in acidic-pH condition (Sirianuntapiboon, et
al., 1988a).
Figure 1. Outline for the formation of MP: Maillard‘s reaction (Monica et al., 2004)
Biological Removal of Melanoidin Pigments 185
(Watanabe, et al., 1982; Ohmomo, et al., 1985a; Raghukumar and Rivonkar, 2001; Miyata, et
al., 1998; Fitzgibbon, 1998). But, the strains of brown-rod fungi having the ability to
decompose cellulose never showed the MDA (Aoshima et al., 1985). Most of the fungi
having MDA grew well on shaking cultures using a medium containing glucose, sucrose, or
maltose and showed very strong MDA (Watanabe, et al., 1982). However, the MDA was
weak on the cultures using xylose or arabinose as the carbon source, while the growth rate
was high. Additionally, nitrogen sources were also affected to growth and MDA. Organic
nitrogen sources, such as peptone and casamino acid were the best nitrogen source in
obtaining a high growth rate and strong MDA. Ammonium salt was also good for growth, but
it gave only half level of MDA of that with peptone. Nitrate salts gave poor growth and weak
MDA. The highest decolorization yield (75-80%) of Coriolus versicolor Ps4a was obtained in
a shaking culture using a MP-medium containing 5% glucose and 0.5% peptone at 30oC for
4-6 days, and the MDA was the decomposition of MP by decreasing the molecular weight of
the MP (Ohmomo et al., 1985b). A research group at Kobe University also screened for MDA
among white-rod fungi and detected a strong MDA from Coriolus pubescens, Hirschioporus
fuscoviolaceus, Polyporellus brumalis, etc., and at the same time, they found strong browning
activity in some unidentified fungi (Tamaki et al., 1985). Moreover, Trametes versicolor
(Benito et al., 1997), Phanerochaetes chrysosporium (Fahy et al., 1997; Kumer et al., 1998;
Kumer et al, 1997; Fitgibbon et al., 1998; Dahiya, 2001a), Coriolus hirsutus (Miyata et al.,
1998), Flavodon flavus (Raghkumar et al., 2004) and others have since been reported as
having MDA.
For the screening of thermophilic fungi, the stain of Ascomycetes, mainly belonging to
the genus Aspergillus, strain G-2-6 was isolated as showing the strongest MDA and identified
as Aspergillus fumigatus. The strain gave the maximum decolorization yield of 75% on a
shaking culture using a medium containing glycerin and peptone as the carbon and nitrogen
sources, respectively, at 45oC for 3 days, and the MP-removal mechanism was the
decomposition of MP (Ohmomo et al., 1987a). At the same time, strain Y-2-32 was also
isolated as showing the strongest MDA and identified as Aspergillus oryzae. However, the
MDA of this strain was not the MP-decomposition mechanism, but MP was only adsorbed
onto the surface of mycelia (Ohmomo et al., 1988). The ability to adsorb MP among living
mycelia and dead mycelia was almost the same and was recovered after washing of mycelia
with buffer solution. The reuse of mycelia after washing was possible (Ohmomo et al.,
1988b). Aspergillus niger 180 was also selected (Miranda et al., 1996) and the MDA of
Aspergillus niger when combined with Penicillium decumbens and Penicillium lignorum was
also confirmed (Antonia et al., 2003).
Additionally, Rhizoctonia sp. D90 (=Mycelia sterilia D90), which belongs to class
Deuteromycetes, was screened as having strong MDA (Sirianuntapiboon et al., 1988a,
Sirianuntapiboon et al., 1988b; Sirianuntapiboon et al., 1995). This strain gave the maximum
decolorization and COD removal yields of more than 90% and 80%, respectively
(Sirianuntapiboon et al., 1988a; Sirianuntapiboon et al., 1988b; Sirianuntapiboon, 1995). The
MDA of Geotrichum candidum (Kim and shoda, 1999), Oscillatoria boryana (Kalavathi et
al., 2001) and Paecilomyces canadensis (Terasawa et al., 2000) were also detected.
Biological Removal of Melanoidin Pigments 187
Both aerobic bacteria and anaerobic bacteria strains having MDA were isolated
(Ohmomo et al, 1987b; Ohmomo et al, 1988a; Sirianuntapiboon et al, 2004b), and both
strains were applied to the conventional wastewater treatment systems (in the laboratory
scale), activated sludge system and anaerobic treatment system, respectively (Kumar et al,
1997; Mohana et al, 2007; Ghosh et al, 2002). Lactic acid bacteria strains were isolated in
order to find a strain of anaerobic bacteria with MDA that could be applied in an anaerobic
treatment system (Kumar et al., 1997; Kumar et al., 1998; Mohana et al, 2007; Ohmomo et al,
1987b). Among them, a hetero-fermentative strain W-NS identified as Lactobacillus hilgardii
showed the strongest MDA and the maximum decolorization yield was about 30% for sugar
cane molasses, about 40% for beat molasses, and about 65% for synthetic glycine-glucose-
MP under the presence of 1% glucose at 35-40oC (Ohmomo et al., 1987b). Furthermore,
acetogenic bacteria BP103, an aerobic bacteria, was isolated from Thailand showed a strong
MDA. This strain decolorized 75-80% of molasses wastewater under the presence of 3%
glucose and 0.5% peptone at 30oC (Sirianuntapiboon et al., 2004a). The MDA was also
detected in Bacillus smithii, which decolorized about 36% of molasses wastewater at 55oC
within 20 days (Nakajima-Kambe et al., 1999). The MDA were also detected in Pseudomonas
fluorescens (Jagroop et al., 2001; Dahiya et al., 2001), Pseudomonas putida (Ghosh et al.,
2002) and some strains belonging to the genus Methanothrix and Methanosarcina (Boopathy
and Tilche, 1991; Boopathy, 1992). In addition, the mixed culture of Streptomyces
warraensis and Basidiomycetes fungi was also tested for decolorization of MWW (Terasawa
et al., 2000).
Our research group tried to isolate the yeast strain having MDA from several sources of
fruit and soil in Thailand. Citeromyces sp. WR-43-6 was first isolated as the MDA strain in
the yeast group. This strain gave a maximum decolorization yield of more than 70% and at
the same time, BOD5 and COD were removed by more than 76% and 98%, respectively
(Sirianuntapiboon et al., 2004b).
cells as a macromolecule and its intracellular accumulation in the cytoplasm and around the
cell membrane as a MP complex, which was then gradually degraded by an intracellular
enzyme. However, the main-MP decolorization or removal mechanisms of each MP-
decolorization strain was different. The details of MDA and MP-removal mechanisms are
described below:
3.1. Microbial-Decomposition of MP
Few papers related to the mechanism of the enzyme system in MDA have been
published, despite the numerous publications on the decolorization and decomposition of MP
from molasses. As a first step in resolving the mechanisms in microbial MDA, sorbose
oxidase was partially purified from the mycelia of white-rod-fungi, Coriolus sp. No. 20
(Watanabe et al., 1982). It was thought that the mechanism of MDA by this enzyme system
would act by oxidizing sorbose and release activated-oxygen (oxygen radical) to decompose
MP. However, no clear evidence for this system was found. Two enzymes, P-III and P-IV,
related to the decolorization of MP were partially purified from the mycelia of Coriolus
versicolor Ps4a (Ohmomo et al., 1985a; Ohmomo et al, 1985b). Enzyme P-III with a
molecular weight of about 50,000 Dalton gave the decolorization yield of about 11% under
aerobic conditions and in the presence of glucose. However, enzyme P-IV with a molecular
weight of about 45,000 Dalton gave the decolorization yield of about 13% under anoxic
conditions and without glucose. The maximum decolorization yield of each enzyme was low;
however, the multiplicative effect (Figure 2) for decolorization with both enzymes was
observed with the decolorization yield of about 40%, which was higher than that calculated
sum of decolorization yields of both enzymes of 24%. Decolorization activity of enzyme PIII
and PIV was 11% and 13%, respectively.
Furthermore, lactic acid and various amino acids were detected as the reaction products
of these enzymes (Ohmomo et al., 1985c). Conversely, the relation of manganese-dependent
peroxidase to the decomposition of MP was suggested in Coriolus hirsutus (Miyata et al.,
1998, 2000) and Flavodon flavus (Raghkumar and Rivonkar, 2001, Raghukumar et al, 2004).
For the determination of molecular weight distribution of MP decomposed by
microorganisms, it was found the molecular weight of MP of the treated SMP or NMP
solutions were shifted to smaller molecular weight fractions than that of the initial solutions,
as shown in Figure 3, and this shift was also detected in the case of fungi and bacteria
(Ohmomo et al., 1988; Sirianuntapiboon et al., 1988b).
In addition, the production of hydrogen peroxide and activated oxygen by photo-synthetic
cyanobacteria closely participated in the decomposition of MP by Oscillatoria boryana BDU
92181 (Kalavathi et al., 2001). Further, the induction at low level MP (10 g/liter) and the
inhibition at high level MP (20 g/liter) for the production of peroxidase was reported in
Geotricum candidum (Lee et al., 2000).
Biological Removal of Melanoidin Pigments 189
Figure 2. Multiplicative effect between enzyme P-III and P-IV for the MP Decolorization (Ohmomo et
al., 1985a).
Chromatograms of MP solution were obtained by using gel filtration on a Sephadex G-50 column.
Symbols: , initial MP solution; , solution treated by Citeromyces sp. WR-43-6.
Figure 3. Molecular weight distribution in MP solution before and after decolorization by Citeromyces
sp. WR-43-6 (Sirianuntapiboon et al., 2004b).
3.2. Microbial-Adsorption of MP
The adsorption of MP onto the cell surface was suggested in Aspergillus oryzae Y-2-32,
which strongly decolorized MWW. The decolorization of this strain was due to the adsorption
of MP onto the cell surface, and its yield depended on the amounts of cell mass. However, the
MP adsorption ability disappeared when washed with 0.1% Tween 80 solution or 0.1% SDS
190 Suntud Sirianuntapiboon and Sadahiro Ohmomo
solution. This means that there is a relation between the cell surface components, such as
muco-polysaccharides, and MP adsorption (Ohmomo et al.,a 1988b). The adsorption of MP
onto the cell surface should be the first step of the MDA, and this strain has no next step, such
as incorporation of MP into the cell and/or decomposition of MP.
However, the electron microscopic observation for the adsorption and incorporation of
MP onto the cell (cell membrane and cytoplasm) of Rhizoctonia sp. D90 (= Mycelia sterilia
D90) was reported (Sirianuntapiboon et al., 1995). Tremetes versicolor gave a strong
decolorization yield, maximum 80%, and about 10% of the yield was due to the adsorption
onto the cell surface (Benito et al., 1997). These reports could suggest that the MDA displays
a two-step reaction of adsorption of MP onto the cell surface and incorporation of MP into the
cell as shown in Figure 4 and Figure 5.
a b
a
a: Cross-section of 7-day-old mycelium, grown in synthetic melanoidin medium, showing electron-
dense materials distributed in the cytoplasm.
b: Cross-section of 7-day-old mycelium, grown in potato dextrose medium, showing well-defined cell
organelles such bas the cell wall (cw) and the cell membrane (cm).
Figure 4. Electron Micrographs of Rhizoctonia Sp. D-90 in SMP medium and potato dextrose medium
(Sirianuntapiboon et al., 1995).
b
a
a: Cross-section of mycelium that had been grown in potato dextrose medium for 7 days (100,000 x
magnification), showing the clear cytoplasm and cell membrane (cm).
b: Cross-section of mycelium that had been grown in potato dextrose medium for 7 days and then in
SMP-medium for another 4 days (100,000 x magnification), showing electron-dense materials
distributed in b
the cytoplasm.
Figure 5. Electron micrographs of Rhizoctonia sp. D-90 collected at various stages of cultivation
(Sirianuntapiboon et al., 1995).
Biological Removal of Melanoidin Pigments 191
Continuous decolorization was carried out in a bubbling column (diameter 26 mm x length 400 mm) at
30 0C with aeration. The column contained 100 ml of waste water and 25 g of immobilized
mycelia (corresponding to 0.8 g dry mycelial weight). → shows the start of continuous
decolorization at a dilution rate of 0.022 hr-1.
Many studies have been conducted on the decolorization process of MWW using fungal
mycelia. For example, the mycelia of Coriolus versicolor Ps4a maintained a decolorization
yield of about 75% for the continuous process under a dissolved oxygen (DO) concentration
of 1 mg/L, dilution rate of 0.03 hr-1, and with the addition of 0.5% glucose and 0.05%
peptone. This mycelia, immobilized in Ca-alginate gel, maintained the decolorization yield of
about 65% for the continuous decolorization process under the dilution rate of 0.022 hr-1 for
16 days operation and removed about 53% of the COD value and 46% of the total carbon
concentration, as shown in Figure 6 (Ohmomo et al., 1985b).
Furthermore, the continuous decolorization of molasses wastewater by mycelia of Coriolus
sp. No. 20 (Sirianuntapiboon and chairattanawan, 1998), the decolorization of MWW by the
mycelia of Phanerochaetes chrysosporium immobilized into Ca-alginate gel (Fahy et al., 1997),
the continuous decolorization of MWW by a mixture of immobilized-Coriolus versicolor IFO
30340 and Paecilomyces canadensis NC-1 strains (Terasawa et al., 2000), the decolorization of
MWW by the mycelia of Coriolus hirsutus IFO 4917l in conjunction with the activated sludge
process (Miyata et al., 2000) and the decolorization of MP from MWW by the immobilized
mycelia of Flavodon flavus (Raghkumar et al., 2001) were also reported.
In addition, under batch-type conditions, Mycelia sterilia D90 gave a maximum
decolorization yield of about 80% and COD removal yield of about 70% in a three time
replacement reaction for a continuous 24 day operation (Sirianuntapiboon et al., 1988b).
Aspergillus fumigatus G-2-6 immobilized into a Ca-alginate gel maintained a decolorization
yield of about 60% in the continuous replacement reaction for 18 days of operation and more
than 70% in the continuous decolorization process under the dilution rate of 0.014 hr-1 with
removal yields of about 51% (BOD5) and 56% (COD) (Ohmomo et al., 1987a). Aspergillus
niger 180 maintained a decolorization yield of about 40% and a simultaneous COD removal
192 Suntud Sirianuntapiboon and Sadahiro Ohmomo
yield of about 70% in the continuously fed batch system with MWW (Miranda et al., 1996).
The decolorization of wastewater from alcohol fermentation using beat molasses was tested
by using a process that combined the mycelia of Penicillium decumbens with the mycelia of
Penicillium lignorum and Aspergillus niger, which resulted in a maximum decolorization
yield of about 70% and a simultaneous COD removal yield of about 50% (Antonia et al.,
2003). Also, Fujila et al (2000) also reported that polyurethane foam-immobilized white root
fungi could be applied into the bioreactor for treatment of the molasses wastewater
Lactobacillus hilgardii W-NS immobilized into Ca-alginate gel gave about a 35%
decolorization yield on the 7 times replacement decolorization process (for 30 days
cultivation) and about 30% decolorization yield for continuous process under the dilution rate
of 0.02 hr-1 and no aeration for 16 days, as shown in Figure 7 (Ohmomo et al., 1987b).
Acetogenic bacteria BP 103 gave the decolorization yield of about 70% on the 6 times
continuous replacement reaction for 30 days operation; however, it maintained only a 30%
decolorization yield on the continuously fed batch reaction for 30 days operation
(Sirianuntapiboon et al., 2004). Pseudomonas fluorescens adsorbed onto cellulose and coated
by collagen gave about a 94% decolorization yield for the continuous replacement
decolorization process (for 4 days operation) (Jagroop et al., 2001). A two-step column
bioreactor using two bacteria strains of Pseudomonas putida U and Acetomonas sp. Ema was
applied for the decolorization of molasses wastewater (Ghosh et al., 2002). The results
showed that the COD and color intensity of MWW were reduced by 44.4% and 60.0%,
respectively, in the first step by Pseudomonas putida U. Then the COD of the effluent of the
first reactor was reduced by 44.4% in the secondary step with Acetomonas sp. Ema.
The continuous feeding system for decolorization of MWW by Citeromyces sp. WR-43-6
was tested and obtained a stable MP, COD and BOD5 removal efficiencies of about 50 - 60%,
99% and 89%, respectively, as shown in Figure 8. (Sirianuntapiboon et al., 2004b).
5. DISCUSSION
MP is recognized as a material difficult to decompose (MDD) because it is hardly
removed (decolorized or adsorbed) by normal biological treatment processes, such as the
activated sludge system, aerated lagoons, anaerobic ponds, etc (Boopathy, 1992; Boopathy
and Tilche, 1991; Sirianuntapiboon et al, 1988b). This review has outlined the potential
application for microbiological removal of MP from MWW. Actually, the MDA has been
detected among various groups of microorganisms, in spite of the classification of MP as an
MDD. Although the removal or decolorization of MP by microbial processes is still only
done on a laboratory-scale level, it is evident from this review that the microbial-MDA can be
a useful process to treat, at least to decolorize, MWW. Among the microbial-MDA, the
Biological Removal of Melanoidin Pigments 193
potential ability of lactic acid bacteria (LAB) should be useful for the development of the
treatment process because LAB is a facultative bacteria and the treatment process requires no-
aeration. Immobilized LAB cells should be particularly advantageous because the treatment
process is very simple and it is expected to be a low cost operation. And organic acids were
generated was the raw material for the metanogenic bacteria group in the methane
fermentation step of anaerobic treatment process. However, the MDA of LAB has only been
detected on Lactobacillus hilgardii WN-S (Ohmomo et al., 1987c) and the decolorization
yield was not very high when compared to some fugal strains, such as Coriolus versicolor
Ps4a (Ohmomo et al., 1985b) and Aspergillus fumigatus G-2-6 (Ohmomo et al., 1987a). The
screening of LAB having higher decolorization yields is expected.
The fungal-MDA mentioned above and the yeast cells have a significantly higher
decolorization yield. However, these abilities are unsuitable for treating MWW because these
microorganisms are aerobic microorganisms and require oxygen for growth and MDA.
Nevertheless, microbial-MDA has great potential for decolorizing MP. It is hoped that
the MDA, such as those discussed in this review, will be exploited to their maximum
potential in the near future. Also, the review article was consisted mostly our research group
activity, we believed that the use of microbiological process to treat MWW was more suitable
according to low cost, reusable of the end product and intermediate. It is therefore,
recommended that further research regarding the MDA mechanism (both biological
adsorption and degradation mechanisms) be conducted to further advance the understanding
of biological MP removal. Also, the observation of some operation parameters of activated
sludge systems (both aerobic and anaerobic process) such as dilution rate, organic loading,
MP loading, sludge age and so on was necessary for the application of the potential-isolated
strain in the conventional wastewater treatment processes.
Continuous decolorization was carried out in a bubbling column (diameter 26 mm x length 400 mm) at
37 °C. The column contained 100 ml of waste water and 60 g of immobilized cells mycelia
(corresponding to 13.2 mg dry mycelial weight). Continuous decolorization was started by the
feeding of wastewater after decolorization for 4 days (showed ↑and↓). The feeding rate was
adjusted to 2.0 ml/hr (dilution rate of 0.02 hr-1).
Symbols: , wastewater adjusted to pH 5.0 by NaOH; , wastewater adjusted to pH 7.3 by Ca(OH)2.
Decolorization was carried out in an μ-carrier magnetic stirrer containing 10 ml of cells (4 x 109
cells/ml) and 1 liter of waste water added 2.0% of glucose, 0.1% of NaNO3 and 0.1% of KH2PO4
(pH 6.0) at 30℃ at an impeller speed of 150 rpm. From after 8 days culture, 100 ml of fresh
medium was added into the system everyday. ↓ shows the time for feeding 10% of fresh medium.
Symbols: , decolorization yield; , reducing sugar; , medium pH; ▲, dry cell weight.
6. CONCLUSION
The microbial-MP removal process is the most suitable way to treat wastewater from
factories that use molasses as a raw material. However, it is still only carried out on a
laboratory scale. Many types of microorganisms, such as fungi (class Basidiomycetes and
Ascomycetes and Dueteromycetes) and bacteria (Lactic acid bacteria and Acetogenic
bacteria) were found to have MDA. The mechanisms of microbial-MDA might be the
adsorption of MP on to the cell (cell membrane and cytoplasm) and/or the adsorbed or
adsorbed MP was degraded by both intracellular and extracellular enzymes. Some of the
microbes showed MP adsorption ability as the main activity, but others showed both MP-
adsorption and MP-adsorption activities. MDA was both induced and inhibited by the level of
MP. Also, the aerobic, facultative, and anaerobic conditions of the microbial-MP removal
processes were investigated. Rhizoctonia sp D-90 and Coriolus sp. No.20 were applied to the
conventional wastewater treatment processes to treat MWW under aerobic conditions. Lactic
acid bacteria was also introduced the anaerobic wastewater treatment process to decolorize
MWW. Both Acetogenic bacteria BP 103 and Cyteromyces sp. WR-43-6 showed MDA with
MWW under both aerobic and facultative conditions. Suspended growth (batch type and
continuous type) and attached growth (Bio-film reactor) systems were tested for the
microbial-MDA process for the treatment of molasses wastewater. However, all the
experiments were still in the laboratory scale.
Biological Removal of Melanoidin Pigments 195
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 197-217 © 2010 Nova Science Publishers, Inc.
Chapter 8
ABSTRACT
Countries in the Mediterranean basin are among the main producers of olive oil. The
elaboration of olive-oil is typically carried out by small companies in small facilities. The
olive-oil plants produce high and variable amounts of residual waters of olives and olive-
oil washing (OMW) that has a great impact in the environment. According to the
procedure used different types of OMW with different chemical oxygen demand can be
generated, the OMW from the three phase process (COD = 150 g O2 L-1) and the OMW
from olives washing (COD = 0.8-4.5 g O2 L-1) and olive oil washing (COD = 1.1- 6 g O2
L-1) in the two-phase process. The uncontrolled disposal of OMW is a serious
environmental problem, due to its high organic load, and because of its high content of
microbial growth-inhibiting compounds, such as phenolic compounds. The improper
disposal of OMW to the environment or to domestic wastewater treatment plants is
prohibited due to its toxicity to microorganisms, and also because of its potential threat to
surface and groundwater. These waters normally are stored in great rafts of accumulation
for their evaporation during the summer. This solution among others until the moment
dose not represent a definitive solution for this problem, especially as the administrations
more and more demanding the preparation of this spill and the constructive quality of the
rafts. Today, effective technologies have been proposed such as the chemical oxidation
process using ferric chloride catalyst for the activation of H2O2 as a treatment of OMW
produced from two-phase process. In the previous works the authors have described the
*
Corresponding author. E-mail: ghodaifa@quim.ucm.es (G. Hodaifa), Tel.: +34 913 944 115; Fax: +34 913 944 114.
200 L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al.
1. INTRODUCTION
Olive oil extraction produces vast amounts of liquid and solid wastes. The elimination of
olive mill wastewater (OMW) is one of the main environmental problems related to the olive
oil industry in Mediterranean countries, where Spain and Italy are the greatest producers.
The OMW was genrated during a few months of the year (November-February). This
liquid waste comes from the vegetable water of the fruit and the water used in the different
steps of oil production and contains olive pulp, mucilage, pectin, oil, etc. suspended in a
relatively stable emulsion.
Olive oil is obtained by the traditional method of discontinuous pressing or by the
continuous centrifugation of a mixture of milled olives and hot water. In both systems three
phases are produced: (i) olive oil; (ii) solid by-product (olive pomace); and (iii) aqueous
liquor, which represent 20, 30 and 50%, respectively, of the total weight of processed olives.
The disposal of highly pollutant olive by products, especially the aqueous liquor, is an
important environmental problem which needs to be solved.
The aqueous liquor comes from the vegetation water and the soft tissues of the fruits. The
mixture of this by-product with the water used in the different stages of oil elaboration
constitutes olive mill wastewater (alpechin in Spanish). The quantity of OMW produced in
the process ranges from 0.5 to 2 L kg-1 of olive, depending on the oil extraction system.
In the main olive-oil-producing countries the implementation of systems based on the
olive-pomace centrifugation has become more widespread. These include three- and two-
phase centrifugation systems. Effluents from two-phase systems are composed essentially by
olive oil and olive-mill wastewater. The olive-oil and the wastewater must be separated.
Continuous three phase extraction systems are still widely used in olive oil mills,
especially in Italy, where in most cases they have not yet been replaced by more recent two-
phases systems, which involve a reduction of OMW volumes but an increased concentration
in organic matter [1]. Three phases extraction systems involve the addition of large amounts
of water (up to 50 L/100 kg olive paste), resulting in the worldwide production of more than
30 millions m3 per year of OMW [2]. This represents a great environmental problem, since
this by-product is characterized by a high organic load; among the different organic
substances found in OMW, including sugars, tannins, phenolic compounds, polyalcohols,
pectins and lipids [3]. The toxicity, the antimicrobial activity and the consequent difficult
biological degradation of OMW are mainly due to the phenolic fraction [4]. The partition
coefficients (oil/water) of most olive phenols, ranging from 6 10-4 to 1.5 are in fact in
favour of the water phase: the olive fruit is very rich in phenolic compounds, but only 2% of
Wastewaters from Olive Oil Industry: Characterization and Treatment 201
the total phenolic content of the olive fruit passes in the oil phase, while the remaining
amount is lost in the OMW (approximately 53%) and in the pomace (approximately 45%) [5].
On the other hand, the phenolic compounds, which are very abundant in the OMW and are
the major responsible of their polluting load, are characterized by a strong antioxidant activity
[6].
The production of olive oil generates large volumes of wastes that vary in composition
depending on which of the three production systems is used. The traditional method, which is
based on the combined use of a crusher and hydraulic press, does not require the addition of
water and yields a very high-quality olive oil. However, the system presents significant
disadvantages to bulk production such as elevated labour requirements and a discontinuous
process. This system has almost disappeared in many production areas. The three phase
‗continuous system‘, on the other hand, has many advantages, such as low labour costs and
continuous production. However, it has disadvantages in the amount of wastewater produced.
In some countries this system has almost disappeared due to the introduction of the two-phase
decanter system. The two-phase decanter system is the newest system and is able to operate
without water and thus dramatically reduces processing costs and the amount of wastewater
produced. However, the semi-solid cake produced by this method has a high moisture content
(55– 60%). One of the main disadvantages of the waste from 2- and 3-phase decanter
systems is the presence of polyphenolic compounds in both the cake and the vegetation water.
The presence of polyphenols limits the use of cake in animal food, since it can cause
digestion problems in cattle [7]. Polyphenols also have been identified as being responsible
for damage to soil when used for irrigation since they inhibit the growth of soil microflora
[8]. The semi-solid waste from the two-phase decanter system cannot be purified by
traditional methods so alternative solutions such as incineration and composting have been
adopted for its treatment [9].
To solve the problems associated with OMW, different elimination methods have been
proposed based on evaporation ponds, thermal concentration and physical-chemical and
biological treatments, as well as its direct application to agricultural soils as an organic
fertilizer. However, the most frequently used methods nowadays are the direct application to
agricultural soils and storage in evaporation ponds, which produces a sludge.
The disposal of OMW is becoming a critical problem in the Mediterranean countries.
Traditionally the olive oil production sector was made up of a large number of small mills
widespread throughout the production area. The volume of OMW produced in each mill was
very small and its disposal was very widespread. These effluents rarely reached the water
courses and their negative effects were only noticed in those places close to the mills. During
the 1950s the industrialization of the olive oil production sector started, with the
concentration of small producers in co-operatives and the creation of big factories with high
milling capacities. Larger mills meant a greater local concentration and volume of OMW,
which was discharged into the rivers without treatment.
For this reason in 1981 the Spanish Government prohibited the discharge of OMW into
the rivers and subsidized the construction of ponds for its storage during the milling period
and the evaporation of its water during the warm Andalusian summer.
Chemical oxidation based on Fenton‘s reagent (hydrogen peroxide in the presence of
ferric salts) has been used to decompose organic and inorganic compounds in the laboratory,
and also real effluents from several sources such as the textile and chemical industry, or
202 L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al.
refineries [10]. The process is based on the formation of various oxidizing agents which
degrade pollution in wastewaters, but the nature of these species is under discussion [11,12].
Assays by Fenton at the end of the 19th century demonstrated that hydrogen peroxide and
ferrous-salt solutions could oxidize tartaric and malic acids as well as other organic
compounds. Later, Haber and Weiss [13] suggested that OH was formed through the
reaction (1). These radicals could react via the oxidation of Fe2 to Fe3 (unproductive
reaction) or via the attack on organic matter:
At pH < 3.0, the reaction is autocatalytic, since the Fe3+ decomposes H2O2 to O2 and H2O
through a chain reaction:
Today, it is thought that other species of Fe(IV) or Fe(V) (such as FeO2 and ferrule
complexes), are the actual active agents of the process. In the presence of peroxide, the Fe2+
concentration is low compared to the Fe3+concentration, since the reaction (4) is slower than
reaction (6). Both radicals, HO and HO2 , react with organic matter, but HO2 is less
reactive. The speed constant for the reaction of ferrous ion with H2O2 is high, and Fe(II) is
oxidized to Fe(III) after a few seconds or minutes in excess H2O2. Therefore waste
destruction through Fenton‘s reagent is thought to be a process catalysed only by Fe(III)–
H2O2 and Fenton‘s reaction with an excess of H2O2 is essentially a process of Fe3+/H2O2. For
Wastewaters from Olive Oil Industry: Characterization and Treatment 203
this reason these kinds of reactions also occur with transition metal ions such as Fe(III) or
Cu(II), and they are known as Fenton-type reactions:
As a general rule, the degree and total mineralization speed are independent of the initial
oxidation state of the Fe. However, the initial efficiency and speed of demineralization are
higher when starting from Fe(II). On the contrary, Fe(III) salts produce an stationary Fe(II)
concentration. In this case a pH lower than 2.8 must be used.
Fenton‘s process has been effective for the degradation of aliphatic compounds and
aromatic chlorates, PCBs, nitro aromatics, azoic colorants, chlorobenzene, phenols, chlorate
phenols, chlorates, octachloro-p-dioxin, and formaldehyde. Only a few compounds cannot be
attacked by this reagent, such as acetone, acetic acid, oxalic acid, paraffins, and
organochlorinated compounds. This reagent is a good oxidizer for herbicides and other soil
pollutants such as hexadecane or Dieldrin. It is used to decompose dry-cleaning solvents and
to decolour wastewater containing different kinds of colorants and other industrial wastes,
reducing its COD. Fenton‘s reaction has been successfully applied to reduce the COD of
municipal and underground water and also for the treatment of lixiviates. It is highly useful
for the pre-treatment of non-biodegradable compounds.
The advantages of this method are the following: Fe2+ is an abundant non-toxic
substance, and hydrogen peroxide is easy to handle and environmentally friendly. Unlike
other oxidative techniques, chlorinated compounds do not result and there are no mass-
transfer limitations because the system is homogeneous. The reactor design for the
technological application is very simple. However, it requires continuous and stoichiometric
addition of Fe2+ and H2O2 as well as a high Fe concentration. It is important to take into
account that excessive Fe2+ amounts may cause the proper conditions for HO to be trapped,
according to the above-mentioned equations.
At pH > 5.0 particulate Fe3+ is generated, although this produces sludges demanding later
management, and, at the end of the process, water is usually alkalinized by means of
simultaneous addition of flocculants to remove waste iron.
The molar stoichiometric H2O2/substrate reaction should theoretically oscillate between 2
and 10 if a reagent is used for the destruction of soluble compounds. In practice, however,
this relation may reach values of up to 1000, since the destructible compound of many
environmental samples is usually accompanied by other compounds that can be attacked by
HO . The peroxide/Fe/compound relation may be maintained by intermittently adding
oxidizer or be fixed at the beginning of the reaction.
In the laboratory, the metal aggregate is traditionally made in the form of pure ferrous
salts, but high prices of these salts hamper their use at the industrial level. Instead of salts,
Fe2(NH4)2SO4 containing 20% active iron is used. Other iron compounds have been used,
including some solids such as goethite to remove trichloroethylene [14]. In such cases total
mineralization is not reached, but some intermediaries resistant to the treatment are created
(carboxylic acids), that react slowly with HO , and there is a predomination of the
unproductive reaction where ferrous ion is converted to ferric ion. More toxic products than
the initial ones, such as quinones, can sometimes form, and these must be thoroughly
controlled.
204 L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al.
Rivas et al., [15] have recently studied the oxidation of p-hydroxybenzoic acid (PHB)
with Fenton‘s reagent, a pollutant usually found in effluents generated by the food industry.
An optimal molar relation Fe/H2O2/pHB of around 5.10-3/2.65/1.10-2 was established. The
addition of tert-butyl alcohol, an HO trap, had very little influence on the process, and
therefore it is interpreted that other radicals were present, as mentioned above. The formation
of phenol, catechol, hydroquinone and trihydroxybenzene indicates that a degradation
mechanism acted via decarboxylation. They also studied the action on atrazine, its
derivatives, and other pesticides.
Microalgae contribute to sustainability in environmental conservation by the
photosynthetic fixation of CO2 from atmosphere and gaseous industrial effluents, as well as
through the consumption of different C, N, and P compounds in urban and industrial
wastewaters. These actions of microalgae, either naturally or induced by humans, are possible
due to the different nutritional modes presented by most microalgae. The opposite situations
are autotrophic and heterotrophic nutrition. Diluted OMW from the three-phase system can be
used as nutrient medium for the growth of Scenedesmus obliquus.
This study report the characterization of different wastewaters from modern olive oil
industry. Concretely, the use of Fenton‘s reaction and microalgae to treatment the wastewater
from the two and three phase process, respectively.
2. EXPERIMENTAL
2.1. Wastewater
In the Andalusian provinces of Jaén and Córdoba (Spain) wastewater samples were
collected from several oil mills operating with olive-cleaning and vertical-centrifugation
equipment of different trademarks.
After the analysis and characterization of the samples, one with high chemical-oxygen
demand was collected, and a mixture of olives and olive-oil wastewaters in a proportion of
1:1 (v/v) was prepared in the laboratory. The pH, electric conductivity, and COD were
determined.
Chemical oxidation based on the Fenton‘s reagent was used for the treatment of the
effluents generated by olive-oil mills operating with a continuous two-phase system. Optimal
operating conditions at room temperature included the hydrogen peroxide concentration, as
well as the catalyzer concentration and type, and the coagulant concentration was identified
previously [16,17].
The best catalyst ferric chloride (efficiency and low cost) from among different
compounds was chosen. The COD value of water at the end of the chemical oxidation at
different concentrations (5, 7.5, 10, and 30% w/v) was determined for each catalyst,
maintaining a relationship [catalyst]/[H2O2] = 0.05 (w/v).
Wastewaters from Olive Oil Industry: Characterization and Treatment 205
The [catalyst]/[H2O2] (w/w) was varied between 0.1 and 1.5 with the aim of determining
the optimum value for the maximum degradation of organic matter and phenol compounds.
The oxidation process was completed with pH-neutralization and separation phases (solid-
liquid) to produce the irrigation water.
The experiments were made in a stirred tank reactor. During the experiments, the pH,
temperature, and electric conductivity values were determined in relation to time. 500 mL of
wastewater with a COD = 7.2 g O2/L, electric conductivity = 1.52 mS/cm and a pH = 4.6 was
mixed with the appropriate amount of ferric catalyst and H2O2. The catalyst and H2O2
dissolution were added gradually during the course of oxidation. The mixture was stirred for
2.5 h, which is enough time to complete the oxidation. The reaction is exothermic only when
a catalyst is used. The solid phase and the liquid phase were separated by decanting. The final
COD in the liquid phase was determined and from this value the reduction percentage in this
parameter was calculated.
The freshwater microalga used was Scenedesmus obliquus CCAP 276/3A, supplied by
Culture Centre for Algae and Protozoa, Oban (United Kingdom).
The experiments were performed in stirred batch tank reactors on a laboratory scale. The
photobioreactors, 5 total, were situated in a culturing facility described elsewhere [18,19].
Each reactor had 0.75 L capacity (70 mm in diameter and 200 mm in height) with
thermostatically controlled water circulation, magnetic stirring, and aeration.
The culture medium wastewater was prepared with ultrapure water (Millipore, mod.
Milli-Qplus) for concentrations of 2.5%, 5%, 10% and 20% OMW (v/v). The pH was
adjusted to an initial value of 7.0 and maintained over the course of the culture.
The working temperature was 25ºC. All the cultures were mechanically stirred at 350
rpm and supplied air sterilized by filtration (0.2 m pore size), at a specific rate of 1 v/v/min.
The illumination was continuous, at an intensity of 298 E m-2 s-1 (QSL 2100, Biospherical
Instruments, Inc.). The mean value of the initial biomass concentration was 0.0124 g L-1 and
the standard deviation SD = 0.0080 and the initial cells number was 0.315 109 cell mL-1 and
SD = 0.197 109.
3. RESULTS
3.1. Characterization of OMW
practically fulfil the values demanded in the normative one (The waters should not overcome
the values for the following parameters: pH = 6-9, suspended solids = 600 mg L-1, BOD5 =
1.000 mg O2 L-1, COD = 1.000 mg O2 L-1, Spain legislation). Only one of them has an
inferior pH to 6 units. The suspended solids are always inferior to 600 mg L-1. Only three
samples overcome allowed COD and BOD5.
It is deduced that, in general, most of the present total solids is from mineral character to
the being the superior percentage of ashes to the percentage of organic matter. Only in a case
flotation is detected, that is to say a layer floating whose composition denotes that it is
probably and reasonable oil of the sweat of broken or damaged olives.
A treatment of chemical oxidation, with energetic oxidizers, with a later alkaline correction
of pH followed by a sedimentation or filtration would be enough, to adapt the water to the
demanded requirements. After a short period of natural sedimentation, it would lose a part of the
small fraction of suspended solids and it could be clever for its use in watering.
In the Table 2 are reflected the values of the parameters analyzed for wastewater of
vertical centrifuges of olive oil of the same almazaras for those that we are took waters of
olive laundry. Contrary to that exposed previously for the waters of olive washers, in this case
all the samples overcome the reference values, to exception of the suspended solids since is
not detected sedimentation neither separation of phases. The muddiness of the samples should
be caused by the fine emulsion of oil in water. In all the cases the COD is bigger than the
allowed values.
As expected, in many cases the quantity of total solids is smaller than in the wastewater
of olive washers and in this case the percentage of the organic matter is bigger than the
percentage of the mineral matter. As it was already exposed in the introduction the biggest
quantity in organic matter of these waters it is due to the own composition of the oil and they
should be rich in phenolic compounds, natural and recalcitrant antioxidants to their microbial
degradation and therefore their concentration will be reflected difficultly in the figures of
BOD5 and if on the contrary in those of COD.
It seems therefore that this water is the main wastewater to try. A priori it has been
thought of subjecting it in the first place and for their sour character to processes treatment of
chemical oxidation, with energetic oxidizers, H2O2 with iron salts (Fenton treatment). The
first experiments carried out lead to an important decrease of the colour and the COD, as well
as to an increase of the pH whose value is inside the allowed limits (pH < 6).
The three-phase process, still used in many olive-oil extraction mills, generates a residual
effluent (OMW) that has a high organic load. OMW, a dark brown wastewater, contains
vegetable water from the olive fruit itself, from the washing of the fruit, from the washing of
the olive oil, and from other activities in the mill. This wastewater is markedly acidic.
Although the BOD5 and COD values are 21 and 13 fold greater than those found in the
residual waters produced by the two-phase process, respectively. The filtration of the raw
wastewater decreased the C and N content by 17.5 and 21.4%, respectively. OMW has a high
content in total solids, reaching 66 g L-1 compared to 5 g L-1 for OMW from two phase (Table 3)
and 1.2 g L-1 for untreated, highly loaded urban sewage. Its fatty content accounts for 1.54%.
Table 3 also lists the contents P, K, polyphenols, and carbohydrates provided by Paredes et al.
[20]. The ratios H:C, N:C, O:C, and P:C of the OMW were 0.13, 0.026, 1.22 and 0.005,
respectively, while the ratios of microorganism biomass according to the elemental formula
of Harrison [21] C H1.64 N0.16 O0.52 S0.0046 P0.0054 were 0.14, 0.19, 0.69, and 0.014, indicating a
certain deficiency in N and P in the wastewater. The ratio N:P of the OMW and biomass
5.8:1, 13.6:1 reflected a larger N deficit.
In all the experiments the values of pH, electrical conductivity and temperature was
determined with the course of the chemical oxidation reaction (Figure 1). A temperature
increase was detected in all the experiments, due to the strong exothermal reaction. This
increase over the ambient temperature was determined for the different peroxide
concentrations used. Electrical conductivity also increased with the course of the oxidation
reaction, this increase is logical considering that the amount of catalyst added (Fe source)
increased during the experiment. Moreover, it is important to point out the decrease in the pH
to values around 3.0, which is an optimal value for Fenton‘s reaction (Figure 1).
45
pH, Conductivity (mS/cm) and T ( C)
40
0
35
30
8
6
4
2
0
0 20 40 60
Time (min)
Figure 1 The values of pH, electric conductivity, and temperature variation with the course of chemical
oxidation reaction (□ pH, ● electric conductivity and ∆ temperature) maintaining a rate
[catalyst]/[H2O2] = 0.05 w/v, catalyst used Fe(NH4)2(SO4)6H2O, temperature 25 ºC, [H2O2]initial = 5%
w/v.
Wastewaters from Olive Oil Industry: Characterization and Treatment 209
Table 4. Final values of COD, purification efficiency and sediment solids determined for
the water treated by chemical oxidation reaction (initial conditions:
COD = 7.2 g O2/L, electric conductivity = 1.5 mS/cm, T = 25 ºC).
3500
10
3000
COD (mg O2/L)
8
2500
2000 6
1500
4
[Fe] or
1000
2
500
0 0
0.0 0.2 0.4 0.6 0.8 1.0 1.2
FeCl3/H2O2 (w/w)
Figure 2. Relationship between the values for COD (■), [Fe] (▲) and total phenols (○) of treated water
and the [FeCl3]/[H2O2] relationship. Operating conditions: initial values of COD = 4104 mg O2/L, total
phenols = 290 mg/L, and [Fe] = 5 mg/L.
Table 4 shows the degradation results expressed as the final values of COD and the
depuration efficiency (reduction percentage in the final COD) for ferric chloride salt. Also,
Table 4 shows the sediments solid determined after the neutralization of water treated and
separation test in Imhoff cone during 1 h. The best concentrations of hydrogen peroxide to
work was varied in the range from 5 to 10% w/v, where the depuration efficiency has varied
between 76 and 82%. The highest value of sediment solids (0.8 v/v) has been obtained.
For the determination of the best catalyst/H2O2 relation, a series of experiments was
performed using ferric chloride. The relation was varied between 0.1 and 1.5. Figure 2 shows
a downward trend of all the parameters for the values of the Fe/H2O2 relation between 0.25
and 0.5 (w/w). For this reason, the following experiments were performed with Fe/H2O2
values of 0.25, in order to reduce catalyst consumption.
From Figure 1, it can be deduced that the pH fell from 4.0-5.0 of the treated water to
values of around 3.0, at which the reaction occurs optimally. For this reason, a neutralization
process of the oxidized water was performed in order to adjust the pH to neutrality, as the use
regulations demand. At the same time, the iron ion found was removed as hydroxide, which is
difficult to precipitate. For the precipitation, several flocculants were tested and an oil-based
210 L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al.
anionic one from the Nalco Company was chosen. For the determination of the concentration
needed, several experiments were performed to determine the influence of the flocculant
concentration with relation to the time necessary for settlement in an Imhoff cone, the COD,
and the remaining amount of Fe in the treated water. The resulting values were indicated no
significant influence on the COD, while the Fe concentration decreased to a concentration of
1 mg/L and, from this value, the concentration remained stable.
As mentioned above, the experiment was conducted at a pilot-plant scale (3-5 m3 h-1) in
S.A.T. Olea-Andaluza olive-oil mill factory in Baeza, Jaén (Spain). In this olive-oil mill
factory, function and verification tests of the results found at the laboratory scale were
performed. Only the environmental conditions changed, because, although the plant was
roofed in, it was exposed to the elements (environmental temperature between 1-7ºC during
the mornings).
The plant worked intermittently during the 2004/05 harvest and subsequently was
automated in order to work continuously for the 2005/06 season.
The process in the plant consisted of: 1. Natural sedimentation in independent holding
pools for water from olives and olive-oil washing; 2. Chemical oxidation tank, 3.
Neutralization tank and coagulant addition; 4. Separation of solid and liquid unite by
decanting; 5. Filtering unite.
As explained above, the plant worked with mixtures of approximately 1:1 (v/v) of
wastewater from olives and olive-oil washing. Table 5 shows the values of the parameters
analysed in all the streams.
The working conditions except the temperature (which was the ambient temperature)
were deduced in the laboratory: a relation of [FeCl3]/[H2O2] of 0.25 (w/w), neutralizing agent
NaOH, and coagulant concentration 1 mg/L. Some 4-5 m3 per charge were used
intermittently, the oxidation time being 2.5 h. The final filtration was performed first with
sand and afterwards with a biomass filter (olives stone).
Also, Table 5 shows the same parameters for oxide water at the outlet of the reactor. A
COD value decrease of about 61% and the total phenolic compound content of about 68% can
be deduced. At the same time, there was an expected increase in electrical conductivity
caused by the addition of a catalyst; part of that increase was eliminated in the stage of
decantation and filtration. At the exist of pilot plant practically all the phenolic compounds
were removed, and the Fe decreased by about 78% and COD value by about 76%.
Independently of what happened in the neutralization (Fe(OH)3 precipitation), which in
principle should not carry organic matter, these decreases were due to the filtration process
used, in which the biomass fill of one of the filters (olives stone) acted as an adsorbent of
heavy metals.
The water obtained was taken to a waterproof storage pool for use in irrigation. Today,
the plant is computerized for continuous functioning and its treatment capacity is about 3 m3
h-1 that is, enough to treat the wastewater produced by an olive-oil mill factory with a
capacity of approximately 600.000 kg olives day-1, which corresponds to a medium/large
olive-oil mill factory.
3.3. OMW from Three Phase Process Treatment by Microalgae
For all the experiments, growth curves showed no lag phases, the first phase being
exponential growth followed by a phase of linear increase in biomass with time. In some
experiments, the stationary phase was observed, and even the onset of cell death.
Wastewaters from Olive Oil Industry: Characterization and Treatment 211
*This wastewater formed mixing the wastewaters from olives and olive-oil washing with the ratio of
1 v/v.
The specific growth rates, μ = d(lnx)/dt, during the exponential-growth phases, μm, was
calculated according to Eq. (12)
x (12)
ln a μm t
x0
were plotted against the initial OMW concentration So, expressed in % (v/v), Figure 3.
0.06
%OMW (v/v)
WC
0.05
RL
WF
0.04 UW
0.03
h m
0.02
0.01
0.00
0 5 10 15 20 25
% [S0], v/v
Figure 3. Variation of the maximum specific growth rates with the initial concentration of OMW (WC:
OMW without color, RL: Rodríguez-López medium, WF: OMW without fatty matter, UW: urban
wastewater from secondary treatment as the medium culture). Common conditions: aeration 1 v/v/min,
agitation speed 0 350 rpm and illumination intensity = 298 E m-2 s-1.
The variation of μm with So, appears to indicate an inhibitory effect in the wastewater. This
was to be expected, as the OMW may contain fats, organic acids, phenolic compounds, and
the remains of pesticides, which are known to harm microalgal growth [22].
212 L. Nieto Martínez, Gassan Hodaifa, Mª Eugenia Martínez et al.
From the different inhibition models by substrate and toxicity assayed, the one that best
reproduced the experimental variation observed was the Teissier [23] one of inhibition by
substrate, Eq. (13), solid line in Figure 3.
[S0] x-x0 BOD5, initial BOD5, end BOD5,removal %BOD5,removal BOD5, removal/x-x0
(%OMW v/v) (g L-1) (mg O2 L-1) (mg O2 L-1) (mg O2 L-1) (g g-1)
2.5 0.0156 1100 190 83 82.7 5.3
5 0.0419 2200 1650 550 25.0 13.1
10 0.0876 4400 1900 2500 56.8 28.5
20 0.0577 8800 3100 5700 64.8 98.8
WF 0.0506 2100 849 1251 59.6 24.7
WC 0.113 2100 950 1150 54.8 10.2
Initial concentration of the biomass at t = 0 h was 0.0124 g L-1.
S0 S0
-
μm μm, max e Ki
e Ks (13)
Though the values μm, max = 0.032 h-1 ([S0] = 10 v/v), Ki = 87%, and Ks = 2.83% are
consistent with what observed. At low initial OMW concentration the S. obliquus has a high
affinity for the limiting quantity of the substrate, resulting in a low Ks value. Roughly
speaking it is the division between the lower concentration range where μm is strongly
(linearly) dependent on S0, and the higher range, where μm becomes independent of S0. The
high value of Ki (87% v/v) indicated that the inhibition effect can be observed only in a high
concentration range (cultures with OMW > 10%).
All the cultures received the same aeration level (1 v/v/min), agitation velocity and the
illumination intensity kept the same, 298 E m-2 s-1, but the attenuation of the light, by the
coloration of the medium, was greater the higher the %OMW, and thus the variation
expected, in μm , being light the limiting factor. This fact was confirmed in the control
experiment WC (OMW without color) where the μm value was increased to 0.04 h-1 for the
same culture concentration (5% v/v). But the main factor limiting growth was the fat matter,
where the value of μm registering for WF experiment (OMW without fatty matter) was 0.05 h-
1
similar to that determined for the mineral medium RL. This can be explained may be
determine the greater distortion in the composition of the biomass, increasing the fatty matter
percentage with the augment of %OMW in the culture, of an oily nature, can be adsorbed
onto the cell surface, hampering the access of nutrients. However, in the culture formed with
urban wastewater (UW) it has been determined 0.052 h-1 for μm slightly major than for
mineral medium (RL).
On the other hand, at the end of the cultures, it was determined the contaminant load for
the wastewater after separation from the biomass (Table 6). A greater net reduction of BOD5
was achieved in cultures with 20% OMW (5.7 g O2 L-1). But the higher %BOD5, removed was
determined in the culture with 2.5% OMW. The decline in BOD5 shows the use of organic
compounds by the microalga. The maximum biomass generated values and BOD5, removed are
Wastewaters from Olive Oil Industry: Characterization and Treatment 213
detected in the culture with 10 and 20% OMW, respectively. This circumstance at 20% OMW
may determine the greater distortion in the composition of the biomass, as components of the
undiluted wastewater, of an oily nature, can be adsorbed onto the cell surface, hampering the
access of nutrients (especially O2). The control experiments WF and WC are registering an
increments in the net biomass generation and in the %BOD5, removed (Table 6). This confirms
the effect of an oily nature medium in access nutrients to the cells and the importance of light
in the net biomass generation.
4. CONCLUSIONS
OMW from Two Phase Process Treatment by Fenton Reaction’s
The chemical oxidation (Fenton reaction) studied in this work are able to treat olive mill
wastewater precedent from two phases continuous system. The best removal efficient was
achieved for COD (76%). The phenolic compounds were destroyed. Treated water resulting
from this process is used to irrigate. This process offers a solution for reducing the
environmental effect of wastewaters generated by two-phase centrifugation system of olive-
oil industry. The catalyst used (FeCl3) containing ferric iron ions which lead to savings in the
consumption of oxidizer (to avoid rust ions Fe II to III). The sediments obtained in the
decanter are dredging mud creamy rich in iron. Water obtained are a fully transparent without
odors, phenolic compounds or pesticides.
Obviously, this wastewater without pre-treatment is not an appropriate medium for the
cultivation of the microalga S. obliquus. The strong inhibition of growth during the
exponential phase and nitrogen deficiency necessitate a pre-treatment prior.
REFERENCES
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mill wastes and their valorisation methods. Waste Management, 26, 960-969.
[2] Borja, R., Alba, J. & Banks, C. J. (1997). Impact of the main phenolic compounds of
olive mill wastewaters (OMW) on the kinetics of acetoclastic methanogenesis. Process
Biochem., 32, 121-133.
[3] D‘Annibale, A., Crestini, C., Vinciguerra, V. & Giovannozzi Sermanni, G. (1998). The
biodegradation of recalcitrant effluents from an olive mill by a whit-rot fungus.
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[4] Bisignano, G., Tomaino, A., Lo Cascio, R., Crisafi, G., Uccella, N. & Saija, A. (1999).
On the in-vitro antimicrobial activity of oleuropein and hydroxytyrosol. J. Pharm.
Pharmacol., 51, 971-974.
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[5] Rodis, P. S., Karathanos, V. T. & Mantzavinos, A. (2002). Partitioning of olive oil
antioxidants between oil and water phases. J. Agr. Food Chem., 50, 596-601.
[6] Obied, H. K., Allen, M. S., Bedgood, D. R., Prenzler, P. D. & Robards, K. (2005).
Investigation of Australian olive mill waste for recovery of biophenols. J. Agr. Food
Chem., 53, 9911-9920.
[7] Molina, E. & Nefzaoui, A. (1996). Recycling of olive oil by-products: possibilities of
utilization on animal nutrition. Int. Biodeterior. Biodegrad., 38, 227-235.
[8] Cappasso, R., Evidente, A. & Schivo, L. (1995). Antibacterial polyphenols from olive
mill wastewaters. J. Appl. Bacteriol., 79, 393-398.
[9] Tomati, U., Galli, E., Fiorelli, F. & Pasetti, L. (1996). Fertilisers from composting of
olive mill wastewaters. Int. Biodeterior. Biodegrad., 38, 155-162.
[10] Bigda, R. J. (1996). Fenton‘s chemistry: an effective advanced oxidation process. J.
Adv. Sci. Eng., 6(3), 34-37.
[11] Bossmann, S. H., Oliveros, E., Göb, S., Siegwart, S., Dahlen, E. P., Payawan, L.,
Straub, M. Jr., Wörner, M. & Braun A. M. (1998). New evidence against hydroxyl
radicals as reactive intermediates in the thermal and photochemically enhanced Fenton
reactions. J. Phys. Chem. A., 102, 5542-5546.
[12] Pignatello, J. J., Liu, D. & Huston, P. (1999). Evidence for additional oxidant in the
photoassisted Fenton reaction. Environ. Sci. Technol., 33, 1832-1836.
[13] Haber, F. & Weiss, J. (1934). The catalytic decomposition of hydrogen peroxide by
iron salts. Proc. Roy. Soc., 147, 332-351.
[14] Teel, A. L., Warberg, C. R., Atkinson, D. A. & Watts, R. J. (2001). Comparison of
mineral and soluble iron Fenton's catalysts for the treatment of trichloroethylene. Wat.
Res., 35(4), 977-984.
[15] Rivas, F. R., Beltrán, F. J., Frades, J. & Buxeda, P. (2001). Oxidation of p-
hydroxybenzoic acid by Fenton's reagent. Wat. Res., 35(2), 387-396.
[16] Martínez-Nieto, L., Rodríguez, S., Giménez, J. A., Lozano, J. L., Cobo, A., Ortega, J. &
Hodaifa, G. (2003). Efluentes de la industria del aceite de oliva: contribución al estudio
de la composición y tratamiento de las aguas de lavado de aceituna y de lavado de
aceite. In: Estudio de la composición y tratamiento como subproducto de las aguas de
lavado de aceituna y aceite, 13-44, Ed. Infaoliva, Granada, Spain.
[17] Martínez-Nieto, L., Rodríguez, S., Giménez, J. A., Lozano, J. L., Cobo, A. & Hodaifa,
G. (2004). Procesos oxidativos en el tratamiento de las aguas de lavado de aceituna y de
lavado de aceite. In: Aguas de lavado de aceituna y aceite: procesos de tratamiento,
73-102, Ed. Infaoliva, Córdoba, Spain.
[18] Hodaifa, G. (2004). Aprovechamiento de las aguas residuales de la industria oleícola en
la producción de biomasa de microalgas. PhD Thesis, University of Jaén, Faculty of
Experimental Science, Jaén, Spain.
[19] Sánchez, S., Martínez, M. E. & Espínola, F. (2000). Biomass production and variability
of the microalga Isochrysis galbana in relation to culture medium. Biochem Eng J, 6,
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[20] Paredes, C., Cegarra, J., Roig, A., Sánchez-Monedero, M. A. & Bernal, M. P. (1999).
Characterisation of olive mill wastewater (alpechin) and its sludge for agricultural
purposes. Bioresource Technol., 67, 111-115.
Wastewaters from Olive Oil Industry: Characterization and Treatment 215
Chapter 9
ABSTRACT
Over the past few years one main focus on the research efforts at the Institute for
Sustainable Waste Management and Technology (IAE) has been on possible applications
for reactors with boron doped diamond electrodes (BDD) in the field of (waste) water
treatment. This article deals with the technical construction of the electrodes used
(continuous reactor with a different number of plate electrodes), which were produced by
a spin-off of the institute. The electrodes consist of conductible industrial diamond
particles (< 250 µm), which are mechanically implanted on a fluoride plastic substrate.
These electrodes showed a high mechanical and chemical stability in different test runs.
At the institute, treatment methods for micro pollutants (e.g. pharmaceuticals and
complexing agents) were developed with electrochemical oxidation by BDD. In this case
test runs were made on laboratory scale and technical scale treatment units and
elimination rates up to 99 % were achieved. In this project the analytic is partly provided
by the ―Umweltbundesamt GmbH‖ (UBA), one of the project partners. This agency has
been a project partner in different studies about pharmaceuticals in the ecosystem. These
techniques could also be used for the waste water treatment of alpine cabins. Pilot
projects have been set up. On the basis of these results a follow-up project was launched
last October, in which an alternative treatment process for oil-in-water emulsions and
mixtures was developed by the usage of electrochemical oxidation with BDD. A third
possible application is the disinfection of drinking water from contaminated ground and
*
Corresponding author: E-mail: hannes.menapace@unileoben.ac.at, Phone: +43 3842 402 5105, Fax: +43 3842
402 5102.
218 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
spring water. In this process oxidation agents like ozone or OH radicals produced in situ
by the BDD reactor from the treated water are used to eliminate bacterial contaminants
(for example e. coli) in the water.
1. DIAMOND ELECTRODES
Due to their mechanical and chemical stability boron doped diamond electrodes are well
suited for the treatment of fluid waste and media. During the so called electrochemical
advanced oxidation process oxidizing agents are directly produced out of the organic matrix
of the treated fluid. The chemical structure of the organic matrix and pollutants are degraded
by these agents and the chemical oxygen demand (COD) will be decreased. For instance the
double bonds in the chemical structure of the pollutant are split up and functional groups are
cracked. Thus biodegradability is increased.
Based on experimental research at various universities, several spin-offs have been
established since the year 2000. These companies conduct research, distribute diamond
electrodes, diamond coatings for applications in the field of water and waste water treatment.
In Europe the distribution of such electrodes and reactors is dominated by three companies.
The following chapter gives an overview on three main producers of boron doped
diamond electrodes (BDD-electrodes) regarding treatment of fluid waste.
1.1.1. Adamant
Adamant Technologies SA was founded 2005 in Switzerland. It is a spin-off company of
CSEM, Centre Suisse d‘Electronique et de Microtechnique S.A.. The production facility is
located in the Science and Technology Park Neode, in La Chaux-de-Fonds (NE). The fields
of activities lay mainly in Diamond Coating Technology. Regarding water treatment
applications the so-called Adamant®-Electrodes and complete systems (DiaCell®-
Technology) are available. Additionally, the company is active in water process monitoring.
The diamond coatings are produced by chemical vapor deposition technique (CVD). [1]
1.1.2. Condias
The CONDIAS GmbH is a spin-off of the Fraunhofer Institute for Thin Films and
Surface Technology. The company was founded in 2001 and has its headquarters in Itzehoe,
near Hamburg, Germany. The main products are diamond electrodes with the trade name
DIACHEM®. These electrodes, also produced by the chemical vapor deposition technique
(CVD) on different base materials like Nb, Ta, Ti, Graphite, Si or conductive ceramics, are
primarily used for waste water treatment and electrochemical synthesis. CONDIAS produces
diamond coated areas up to 100 x 50 cm² with diamond layer thickness up to 15 µm. [2]
boron doped diamond electrodes with a layer of titanium oxide. In contrast to the other two
producers of BDD electrodes mentioned, diamond particles up to a size of 250 µm are
mechanically implanted on the metal substrate. In 2006 a new electrode type was developed
and patented. The old model of the titanium substrate was replaced by a film of fluorinated
plastic. The company distributes smaller flow rate reactors for waste water treatment and
disinfection of supply and tap water, but also standard BDD electrodes with a maximum area
of 16 cm x 16 cm. Greater electrode areas are supplied by a special welding methode [3]
In the following chapter the construction of BDD electrodes is explained on the basis of
the pro aqua patent concerning the bipolar diamond electrodes with a substrate of fluorinated
plastic. This type of electrodes was used for the different degradation tests which were
conducted on the Institute for Sustainable Waste Management and Technology (IAE) at the
University of Leoben, Austria. Furthermore, the advantages and disadvantages of
characteristic settings are discussed.
Figure 1. Diamond electrode comprising a support layer of electrically non-conductive material [4]
In the field of waste water treatment and fluid waste disposal, the anodic oxidation
process is a rather new technology, which is not widely used. This treatment method falls into
the category of electrochemical oxidation processes and is an ideal additional treatment step
for conventional disposal systems, especially if no biodegradable substances should be treated
or in presence of toxic chemicals which would inhibit biodegradation processes. [5]
The chemicals commonly used for this purpose are oxygen, hydrogen peroxide, ozone,
permanganate or persulfate. The higher the oxidation potential of the reagent used, the more
efficient the chemical oxidation process is. The most powerful oxidant in water is the
hydroxyl radical with a redox potential of 2.8 V relating to normal hydrogen electrode
(VNHE) [6]. Organic contaminants are degraded into inorganic substances such as H2O, CO2,
and the waste water is additionally disinfected by the agents produced. In general the higher
the oxidation potential, the higher the efficiency of the treatment process. Table 2 shows the
oxidation potential of some chemical substances.
Figure 2. Working range of different electrode materials relating to the potential of the standard
hydrogen electrode [6].
The anodic oxidation can be ascribed to the Advanced Oxidation Process (AOP). This
term comprises all oxidative methods, which have hydroxyl radicals as the main oxidation
agent. Hydroxyl radicals act as the main oxidation agent. As can be seen in the table, the OH.-
radical with an oxidation potential of 2.8 V is a fairly strong oxidation agent in water.
During the anodic oxidation process the boron doped diamond electrodes form the basis
of the process. These synthetic diamond electrodes differ from other electrodes due to their
high mechanical and chemical stability and the manner in which water electrolysis is carried
out. Whereas during the electrolysis of water with conventional electrodes the water molecule
is split into oxygen and hydrogen, during treatment with a diamond electrode, highly reactive
hydroxyl radicals are formed. This is due to the high potential of a diamond electrode.
Previous tests conducted with graphite or carbon bearing electrodes have shown that the use
of these materials is not advantageous as they do not exhibit the necessary potential to
produce hydroxyl radicals [5]. Moreover, signs of wear appear all too readily due to the fact
that, in addition to the formation of oxygen, CO2 formation also occurs leading to the
breakdown of the electrode material. Even electrodes comprised of PbO2, SnO2 or Pt do not
show enough efficiency in the production of OH radicals and are not suiTable due to their low
mechanical and chemical stability [5]. Figure 2 shows the range of activity of the different
electrode materials vs. the hydrogen potential (left side), oxygen (right side). The potential
required to create hydroxyl radicals lies at 2.8.
In contrast, boron doped diamond electrodes work with an energy efficiency of more than
90%. This means that in the anodic oxidation process the OH radicals are formed in an
electro-chemical manner directly from the waste water treated and the impurities are
mineralized or at least transformed into biologically degradable materials, without producing
any further residues or waste. The hydroxyl radicals formed react with the impurities in the
waste water by splitting hydrogen. [6]
H2O OH + e- + H+ (1)
Usability of Boron Doped Diamond Electrodes… 223
Figure 3. Combined chlorine content depending on active electrode area and applied current density,
UB untreated water sample, R3K 9 plates, R2K 6 plates and R1K 3 plates.
The most suitable alternative method is chemical oxidation, aiming for the total
mineralization or the production of harmless or biodegradable compounds by use of oxidants.
By using the electrochemical production of hydroxyl radicals, no additional chemical
substances are necessary. The process can be performed at affordable costs, determined by
the power required for driving the electrochemical process and without the common AOP
drawbacks.
Electrochemical water disinfection by producing disinfecting agents (mainly active
chlorine produced from the naturally dissolved chloride ions) during electrolysis of water is
another common water treatment process. The amount of electrochemical produced oxidizing
agents (for example chlorine content shown in Figure 3) manly depends on the two
parameters active area and current density. Figure 3 shows the content on combined chlorine
for different reactor sizes. Therefore a different number of BDD electrode plates (active area
per plate) were installed in each reactor. To achive a better efficiency of the process, a static
mixer was implemented in downstream of the reactor R1K in the test run R1K SM.
2. FIELDS OF APPLICATION
2.1. Treatment of Pharmaceuticals and Complexing Agents in Waste Water
2.1.1. Introduction
Pharmaceuticals are discharged into the sewer system with human or animal
excrements and finally end up in the municipal sewage plant. Because of their complex
chemical structure some drugs (e.g. Carbamazepine) or chelating agents (for example
224 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
EDTA and NTA) cannot be eliminated using conventional waste water treatment
procedures, so they pass into the aquatic system [7, 8].
As an example the release of pharmaceuticals into surface waters may lead to
increased dissemination of antibiotic resistance [9], endocrine substances like hormones
are suspected to promote feminizing effects on organisms in ecosystems [10].
Complexing agents like EDTA may cause a remobilization of sedimented heavy metals in
surface waters.
While there are already statutory thresholds for EDTA and NTA implemented in
Austria (QZV Chemie OG 2006) [11] according to the EU directive 2000/60/EC [12], a
regulation for pharmaceuticals is expected in the near future.
To be able to meet these requirements two innovative treatment procedures have been
designed. One is the anodic oxidation with boron doped diamond electrodes and the other
one the ozonation. For the second method a new sort of ozone generator is used and the
ozone is injected into the water flow in a venturi injector. The development of both
procedures is in a testing phase. During research at the Institute for Sustainable Waste
Management and Technology at the University of Leoben a new process design was
developed in two steps. First a small lab unit was constructed (flow rates ranged from 3-
80 L/h). Treatment sources included synthetic waste water, cleaned waste water from the
local municipal waste water treatment plant and a wide range of sectoral waste water. The
technical results from this first phase were used for the design and construction of the
technical scale unit.
Figure 4. Sketch of the of the laboratory scale unit - a combination of anodic oxidation and ozonisation.
The first laboratory scale unit (Figure 4) was used in the first project step to determine the
process parameters relevant for further process design and to gain insight on essential parts
and sizes for constructing the tech scale unit. In the second step, the small plant is utilized for
several test series treating different sources of sectoral waste water from hospital and
industry.
The plant follows a modular design concept and consists of two independent segments,
one represents a flow reactor for the anodic oxidation, the other a reaction well with the
attached ozone generator. The parts operate either separately or in combination. Hydraulic
conveyance of the waste water to a particular reaction circuit is provided by diaphragm
pumps (Sera R203-2,4E 3 L/h) and flexible-tube pumps (Gardener, Watson-Marlow 323e
max. 86 L/h).
2.1.3.1.2. Ozonisation
The ozonisation process consists of two steps: the production of O3 and the treatment
reactor for contacting the oxidant with the waste water. This has the advantage of there being
two ways to optimize the treatment plant. In laboratory scale different sorts of production and
insertion were investigated. At the beginning diamond electrodes located on titanium oxide
plates were used, but as these caused technical problems (electrode lifetime, operating
stability) the process was finally substituted by an advanced corona-discharge generator for
the ozone.
Reduction tests were carried out in counter current flow at the beginning, in subsequent
project steps a venturi injector was used to mix waste water and ozone. Similar to the anodic
oxidation, interconnection of sensors is possible.
Figure 5. Tech scale unit; left side: controlling station, right side: reactor unit.
Usability of Boron Doped Diamond Electrodes… 227
2.1.7. Analytics
The process parameters temperature, pH-value and redox potential were directly
monitored by sensors and the anodic oxidation current and voltage were recorded. The
analysis of pharmaceuticals and chelating agents (Table 3) was carried out by the
'Umweltbundesamt' (UBA), where similar projects had been carried out before [10, 13].
During the first preliminary tests of the two technologies applied on the LSU an
analysis of the EDTA elimination was made. A complexometric titration according to
DIN method DIN 38406-3 [14] for determination of calcium and magnesium ions in
water by EDTA was used. Titrating with a calcium solution of defined concentration
gives the possibility to calculate the amount of EDTA in the treated solution. For this
EDTA was added to the untreated water sample. After treatment 50 mL samples were
taken and sodium hydroxide (NaOH, 2 mol/L) and an indicator salt were added. The
sample was then titrated with a calcium chloride solution (CaCl 2, 50 mg/L) until a colour
228 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
change from blue to purple occurred. The concentration of EDTA in the titrated sample
was then calculated according to the DIN standard DIN 38406-3.
2.1.7.1. Pharmaceuticals
For analysis of pharmaceutical compounds 500 mL of the samples were acidified, spiked
with an isotopically marked surrogate standard mixture and subsequently enriched by means
of solid phase extraction. Analytes were eluted using dichlormethane, ethylacetate and
methanol. The resulting extract was concentrated under a gentle stream of nitrogen and
solvents were changed to acetonitrile and water. The final extract was spiked with an internal
standard to follow instrument stability and compensate for matrix effects. Samples were
analyzed by means of liquid chromatography-electrospray ionization-tandem mass
spectrometry. Quantification was performed by external standard method.
mass spectrometry in Single Ion Recording (SIR) mode. Quantification was performed with
internal standard method by means of isotope dilution.
2.1.8. Analysis
During the experiments (Figure 6) we observed a strong dependency of the treatment
success on the applied current density, the reactor surface and the time of contact could be
observed. The results also showed a different degradability of the individual substances as
Carbamazepine showed a better degradability than Diazepam. The degradability of
Complexing agents showed a deterioration of the degrading performance at a low
concentration range (µg/L-Area).
2.1.9. Summary
With the experiments completed thus far, the applicability of the treatments used for
a continuing waste water treatment was proven [15]. Furthermore, the experiments on the
TSU were performed under the most realistic conditions possible to collect data material
for the optimization (e.g. reactor dimension to increase the contact time).
Based on a comparison of the success of a particular treatment of the municipal and
industrial waste water, a statement concerning the applicability for central and decentral
waste water treatment was made. In addition to the treatment success the costs of
investment and the costs of treatment should be considered. Based on 0.07 €/kWh and
after a first estimation for the flow rate with 200 L/h the costs for the treated waste water
depending on current densities from 30.2 to 42.3 mA/cm² will range between 0.16 and
0.60 €/m³.
Figure 6. Actual rates of elimination for the Anti-epileptic drug Carbamazepine, TSU – different
current densities and flow rates.
230 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
proaqua-reactor-1:
Diamond electrodes 4 plates
Active area per electrode 32,5 cm2
Total area 130 cm2
Gap between electrodes 3 mm
max. current density 50 mA/cm2
Feeding electrodes Ru/Ir coated titanium sheet
Housing material Polypropylene
Seals Viton
Flow rate max. 50 L/h
In this case 2.000 mL of emulsion were treated in a batch-system with a flow reactor,
applied with BDD diamond electrodes under a flow rate of 21.8 L/h. A current density of 83.3
mA/cm² was reached by an active electrode surface per plate of 4 x 30 cm² and a constant
Usability of Boron Doped Diamond Electrodes… 231
current of 2.5 A. The duration of the process was limited to 5 h 30 min. Figure 7 shows the
resulting degredation effect by a current density of 83.3 mA/cm².
Figure 7. Degradation of COD for oil-water emulsion after ultra-filtration – 83.3 mA/cm², 21.8 L/h.
Figure 8. Concentration of atrazine and its metabolites in spring water treated with BDD electrodes.
Although atrazine displays only a marginal acute toxizity, studies showed critical effects
in the field of body weight gain, inhibition of ovulation and effects on the heart function.
Furthermore, the substance is suspected to have carcinogenic effects. Hence, for consumers
an ADI value (Acceptable Daily Intake) of 0.005 mg/kg bw (body weight)/d (day) was
defined. Moreover, endocrine effects were detected for atrazine. [17, 18, 19]
Depending on the local factors such as rainfall, soil moisture and the adsorptive
properties of the soil, atrazine finds its way into the groundwater. Due to its higher
mobilization rate the risk of desethylatrazine getting into groundwater is higher in comparison
to atrazine. [20]
Within the scope of a smaller project, degradation tests of atrazine and disethylatrazine
contaminated spring water were conducted at the Institute. In this case the unpurified water
samples were taken from a spring in an area with intensive agricultural usage. For the
treatment a flow reactor according the specifications of Table 4 was used.
Figure 8 shows the degredation effect for atrazine and its metabolites at an applied
current density of 45.5 mA/cm². After treatment (batch as well as continuous operation mode)
of the herbicide contaminated spring water, the concentration of desethylatrazin was below
the threshold according to the Austrian Drinking water directive.
In the alpine region of middle Europe thousands of alpine lodges and mountain inns were
built in exposed positions. In view of the sensible ecosystem and in fact that many lodges
were built above the catchment area of drinking water abstraction, untreated waste water
Usability of Boron Doped Diamond Electrodes… 233
Escherichia coli
Enterococci
Coliformic germs
Colony forming units (CFU) at 22 °C and 37 °C
The treatment of the waste water from the alpine lodge shows the dependence between
disinfection efficiency on the one hand and the parameters flow rate and current density on
the other hand (Table 6).
In the test runs with the contaminated spring water, values for the treated water below the
quantification limits could be achieved. According to the results a continuous operation mode
of the reactor (current density of 36.92 mA/cm² and a flow rate of 43.5 L/h) seems sufficient
for the treatment.
234 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
Table 7 shows microbiotic examinations of waste water effluent from the sewage water
treatment plant in Karlsruhe, Germany before and after treatment by a flow reactor with the
pro aqua diamond electrodes. The analysis was carried out by the Heinrich-Sontheimer-
Laboratory for water technology of the DVGW (German Technical and Scientific Association
for Gas and Water) in Karlsruhe, Germany [24].
4. REFERENCES
[1] Adamant Technologies SA (2009). Specification sheets. http://www.adamantec.com/
[2] CONDIAS GmbH (2009). Specification sheets. http://www.condias.de/. 2009
[3] pro aqua GmbH (2009). Specification sheets. http://www.proaqua.cc. 2009
[4] Schelch, M., Staber, W., Wesner, W., et al. (2007). Method for the production of a
diamond electrode and diamond electrode, Patent: WO 2007/116004 A2/PCT/
EP2007/053337, 18. October.
[5] Kraft, A., Stadelmann, M., et.al. (2003). Anodic oxidation with doped diamond
electrodes: a new advanced oxidation process. Journal of Hazardous Materials, 247-
261.
[6] Tröster, I., et.al. (1998). Electrochemical advanced oxidation process for water
treatment using DiaChem electrodes. Diamond and Related Materials, 640-645.
[7] Ternes, T. (1998). Occurrence of drugs in sewage treatment plants and rivers. Water
Research, 32(11), 3245-3260.
[8] Hohenblum, P., Scharf, S., Gans, O., Moche, W. & Lorbeer, G. (2004). Monitoring of
selected estrogenic hormones and industrial chemicals in ground waters and surface
waters in Austria. Science of the Total Environment, 333, 185-193.
[9] Balcioglu, A. & Ötker, M. (2003). Treatment of pharmaceutical wastewater containing
antibiotics by O3 and O3/H2O2 processes. Chemosphere, Vol. 50, Issue 1, 85-95.
[10] Paumann, R. & Vetter, S. (2003). Hormonwirksame Stoffe in Österreichs Gewässern –
Ein Risiko? – ARCEM-Endbericht, Umweltbundesamt GmbH. Vienna. ISBN 3-85457-
695-1. (Endocrine disrupters in Austria‘s waters – a risk? – Austrian research
cooperation on endocrine modulators)
[11] Federal Republic of Austria: QZV Chemie OG, Qualitaetszielverordnung Chemie
Oberflaechengewaesser – (BGBl. II Nr. 96), 2006 (Quality Target Ordinance,
Chemistry surface waters)
[12] European Community: Directive 2000/60/EC of the European Parliament and of the
Council establishing a framework for Community action in the field of water policy,
23.10.2000
[13] Scharf, S., Gans, O. & Sattelberger, R. (2002). Arzneimittelwirkstoffe im Zu- und Ablauf
von Kläranlagen; BE-201, ISBN 3-85457-624-2Umweltbundesamt GmbH: Vienna.
(Pharmaceutical substances in inflow and effluent of STP´s, Data material, Austrian
federal environmental agency).
[14] DIN 38406-3 (2002). German standard methods for the examination of water, waste
water and sludge – Cations (group E) - Part 3: Determination of calcium and
magnesium, complexometric tritation, E3, German Institute for Standardization.
[15] Menapace, H. M., Diaz, N. & Weiß, S. (2008). Electrochemical treatment of
pharmaceutical wastewater by combining anodic oxidation with ozonation. Journal of
Environmental Science and Health, Part A, 43:8, 961-968.
[16] Umweltbundesamt: Parameterinformationsblatt Atrazin und Desethylatracin.
Datenband Porengrundwasser, Vienna, 2006.
[17] Breckenridge, C. B., Werner, C., Stevens, J. T. & Sumner, D. D. (2008). Hazard
Assessment for Selected Symmetrical and Asymmetrical Triazine Herbicides. In The
Triazine Herbicides, 387-398.
236 Hannes Menapace, Stefan Weiß, Markus Fellerer et al.
[18] Eldridge, J. C. & Wetzel, L. T. (2008). Mode of Action of Atrazine for Mammary
Tumor Formation in the Female Sprague-Dawley Rat. In The Triazine Herbicides, 399-
411
[19] Snedeker, S. & Heather, C. (1999). Critical Evaluation of Atrazine´s Breast Cancer
Risk. Program on Breast Cancer and Environmental Risk Factors in New York State.
Cornell University, Ithaca.
[20] Spark, K. M. & Swift, R. S. (2002). Effect of soil composition and dissolved organic
matter on pesticide sorption. Science of the Total Environment, 298, 1-3, Amsterdam.
[21] Austrian Drinking Water Ordinance, BGBl 304/2001, released on 21.08.2001
[22] Pupunat, L. & Rychen, Ph. (2002). Inactivation of Legionella with the DiaCell® Water
Treatment Technology, Swiss Center for Electronics and Microtechnology Inc, CSEM
Scientific Report.
[23] Furuta, T., Rychen, Ph., Tanaka, H., Pupunat, L., Haenni, W. & Nishiki. Y. (2005).
Application of Diamond Electrodes for Water Disinfection. In Diamond
Electrochemistry, A. Fujishima, et al, Ed.; BKC Inc., Tokyo et Elsevier B.V.,
Amsterdam, 525-542.
[24] Maier, D. (2008). Untersuchung zur Wirkung des von der Fa. pro auqa
Diamantelektroden Produktion GmbH entwickelten AOP-Verfahrens bei der Reinigung
von Karlsruher Abwasser. Karlsruhe: DVGW.
In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 235-253 © 2010 Nova Science Publishers, Inc.
Chapter 10
Veronica Arthurson*
Department of Microbiology, Swedish University of Agricultural Sciences,
Box 7025, 750 07, Uppsala, Sweden.
ABSTRACT
Treatment of wastewater, commonly performed at municipal sewage plants,
generates sanitized water and sewage sludge. Anaerobic degradation of sewage sludge
results in the production of different gases, including the economically valuable methane,
and digested residue (biosolids) with potential value as a crop fertilizer. Traditionally,
digested sewage sludge is disposed either into water, onto or into the earth or into the air.
However, alternative exploitation of digested sewage sludge in agriculture has several
advantages over commercial fertilizers, including environmental aspects benefiting
agricultural sustainability and increased crop yield. Additionally, residue utilization is
nearly always a cheaper option than disposal costs.
Biosolids obtained from the treatment of municipal sewage sludge consist of a
mixture of organic and mineral compounds that significantly affect soil microbial
communities and their biogeochemical activities when applied as a crop fertilizer. The
microorganisms influence soil quality through nutrient cycling, decomposition of organic
matter and maintenance of soil structure, in turn, affecting agricultural and environmental
quality, and subsequently, plant and animal health. Moreover, both soil and residue
normally contain considerable quantities of microorganisms, including both beneficial
and potentially human pathogenic species that may be supported by the new conditions in
*
Corresponding author: . Email: Veronica.Arthurson@mikrob.slu.se., Phone: +46 - 18 - 67 32 12. Fax: +46 - 18 -
67 33 92
238 Veronica Arthurson
the soil. Thus, soil amended with biosolids may present a modified microbial community
composition after some time and, hence, a modified ecosystem function.
At the end of the present chapter, we discuss whether the potential risks of recycling
biosolids to agricultural cropland are acceptable for consumers, producers and scientific
expertise, in view of the resulting alterations in soil microbial diversity, activity and
accompanying functions. Furthermore, optimal ways of managing the recycling process
to achieve the most favourable balance of benefits and risks for the community are
highlighted.
produced in the previous step are degraded by facultative anaerobic and strictly anaerobic
bacteria via a number of fermentative processes. This acid-forming stage results in the
production of carbon dioxide, hydrogen gas, alcohols, organic acids (acetate), some organic-
nitrogen compounds and a few organic-sulfur compounds (Gerardi, 2003). Acetate, either
produced via fermentation of soluble organic compounds or acetogenesis (acids and alcohols
are degraded to acetate), subsequently serves as the main substrate for methane-forming
bacteria, finally resulting in methane gas and carbon dioxide, i.e., biogas (Gerardi, 2003).
Additional gases present at low concentrations in biogas include hydrogen, hydrogen
sulphide, nitrogen and trace levels of hydrocarbons (Williams, 2005). Biogas has a calorific
value of between 20 and 25 MJ/m3 (IEA Bioenergy, 1996), and can either be directly
combusted to produce heat for the digestion process or used in power generation. In the latter
case, the gas needs to be cleaned up to remove corrosive trace gases, moisture and vapors
from the gas stream (Williams, 2005). In addition to the benefits of anaerobic degradation in
terms of renewable energy production and reduction in sewage sludge volume, weight and
putrescibility, the resulting residue (biosolids) may have potential utility as an agricultural
fertilizer. The insoluble solid residue derived from sewage sludge, termed biosolids, generally
consists of organic compounds, macronutrients, micronutrients, non-essential trace metals,
organic micro pollutants and microorganisms (Singh and Agrawal, 2008, Kulling et al.,
2001). Macronutrients in biosolids provide a source of plant nutrients, and the organic content
confers beneficial soil conditioning properties (Logan and Harrison, 1995), emphasizing its
potential value as a crop fertilizer.
to humans through entry into the food chain via crops or grazing animals. Sludge stabilization
by anaerobic or even aerobic digestion usually results in Class B biosolids, which have
limited immediate reuse potential in land application owing to the high content of human
pathogens. On the other hand, class A biosolids are generated as a result of more stringent
treatment, and therefore contain extremely low or non-detectable amounts of microbial
pathogens. Consequently, sewage sludge intended for anaerobic digestion and subsequent
land application requires further treatment (prior to or after the degradation procedure) to
obtain an environmentally acceptable end product of Class A quality. The most well
established ways to achieve this goal are lime treatment, composting, and/or heat treatment,
of which the latter often is the most efficient, resulting in the elimination of most microbial
pathogens. However, bacterial endospores present in digested sewage sludge (Bacillus spp.
and Clostridium spp) are not destroyed by standard heat procedures, and require a highly
expensive method involving at least two separate rounds of heating to ≥ 70°C for ≥ 1 h to
ensure elimination. Primary heating activates the spores into vegetative forms, which
eventually start to metabolize and replicate. The secondary heating step should kill these heat-
labile bacteria, provided the incubation period between the two heating steps is sufficiently
short to prevent the formation of new endospores. Alternatively, irradiation technology
should be considered to meet the requirement for Class A biosolids, and further assessed to
determine whether it presents an effective option to the above sanitization techniques.
However, most endospore-forming bacterial species are indigenously present in soil, and thus
the issue of whether application of heat-treated biosolids actually imposes an increased risk is
debatable. In conclusion, if biosolids are to be applied to agricultural land as a Class B waste
product, site and crop restrictions are necessary. On the other hand, if the residue is further
subjected to one of the above sanitization techniques (e.g., lime treatment, composting or heat
treatment/irradiation), it would be a safer product in terms of risk of disease transmission via
microbial agents.
excess supply of available N resulted in poor quality crops. The authors concluded that for
best yield and crop quality, it is important not to exceed a rate of 5 Mg dry matter ha-1 yr-1 of
liquid or dewatered sludge. In contrast, composted sewage sludge did not cause negative
effects on crop quality when applied at higher rates. Moreover, twelve years of continuous
landspreading of biosolids resulted in improved soil fertility (organic matter and nitrogen
increase), but a small buildup of heavy metals (Cu and Zn). The study by Mantovi and
colleagues supports the significant potential of treated sewage sludge as a crop fertilizer,
although the possible negative effects of organic pollutants and heavy metal content in
sewage sludge intended for agricultural land application need to be considered. Moreover,
plants differ in their abilities to absorb sludge-borne metals from soil (Singh and Agrawal,
2008), which is further influenced by soil properties, including pH, redox potential,
sesquioxide content, organic matter, and application rate of sludge (Hue and Ranjith, 1994).
These parameters should be carefully evaluated and monitored, prior to the application of
biosolids to arable land.
In another field trial, Odlare et al. (2008) studied the changes in soil chemical and
microbiological properties over a period of 4 years, following treatment of soil with different
organic waste products. Very few significant differences in chemical and microbiological
features were detected with several fertilizers of different origins. However, the authors
showed that biogas residue, generated from co-digestion of source-separated household waste
and food residues from restaurants and kitchens, augmented the microbial biomass (analyzed
on the basis of substrate-induced respiration), proportion of active microorganisms, nitrogen
mineralization, and potential ammonia oxidation. Soil treated with anaerobically digested
sewage sludge displayed similar trends of increased microbial biomass, larger proportion of
growing microorganisms and higher degree of nitrogen mineralization, compared to
unfertilized control soil, although the data were not statistically significant throughout the
treatments. Additionally, the study by Odlare et al. (2008) disclosed no negative effects of
organic waste on the soil microbial parameters, and equivalent or better fertilization effects of
all waste products, such as cow manure, pig slurry and mineral fertilizer.. Consistent with
these trends, Banerjee et al. (1997) reported that soil amendment by sewage sludge
augmented microbial, respiratory, and enzyme activities, and higher microbial biomass in
sludge-amended soil is frequently observed (Parat et al., 2005, Dar, 1996). Increase in
microbial biomass when soil microbiota are supplied with C, N and nutrients (i.e. biosolids) is
expected, since most soil microorganisms are heterotrophs that depend on organic carbon for
growth and energy. On the other hand, a decrease in biomass is likely if the biosolids applied
contain toxic chemicals, such as organic pollutants. Sullivan and co-workers (2006b)
observed no changes in total microbial biomass in response to biosolid application, which
was explained as a rapidly acclimatizing soil microbial community whose overall biomass
was sparsely affected by biosolids. Comparable results were reported by García-Gil et al.
(2004).
The group of Epstein (1998) studied the effects of sewage sludge soil application on
water retention, hydraulic conductivity and aggregate stability. In their experiments, both raw
and digested sludge increased the total soil water retention capacity, as well as soil hydraulic
conductivity after 27 days of incubation. Epstein (1998) additionally reported an average of
34% stable aggregates after 175 days of incubation in sludge-treated soil, compared to 17% in
untreated control soil, confirming the beneficial characteristics conferred by sewage sludge.
Long-term application of biosolids evidently improves the physical features of soil, including
242 Veronica Arthurson
saturated and unsaturated hydraulic conductivity, water retention capacity, bulk density, soil
resistance to penetration, total porosity, pore size distribution, aggregation, and aggregate
instability (Tsadilas et al., 2005, Aggelides and Landra, 2000).
Sewage sludge commonly contains high amounts of human pathogenic bacteria excreted
in feces and urine (Dudley et al., 1980, Larsen, 1995, Strauch, 1991). The health risks related
to these pathogens in biosolids spread on agricultural land depend on prior sludge treatments
applied, as well as their ability to maintain virulence during storage and field distribution.
Biosolids subjected to anaerobic digestion and an additional heating/pasteurization step ( ≥
70°C for ≥ 1 h; Class A biosolids) should not contain human pathogenic microorganisms.
However, harmful gut microorganisms frequently remain in class B biosolids. A number of
studies have focused on evaluating the survival of indicator organisms and pathogens
introduced into soil via biosolids. In particular, limited survival of enteric organisms, such as
Escherichia coli, in biosolid-amended arable soil has been documented in laboratory and field
trials (Lang and Smith, 2007). For instance, Lang et al. (2007) investigated the potential
survival of different enteric microorganisms in agricultural soil amended with biosolids, and
reported that dewatered mesophilic anaerobically digested biosolids increased the level of E.
coli in soil. However, the introduced E. coli declined rapidly, and their survival was limited to
three months, irrespective of the environment or timing of sludge application. The authors
concluded that pathogens applied to soil via biosolids decline to background values well
within the cropping and harvesting restrictions imposed when sewage sludge is spread on
farmland. According to traditional Swedish farming practice, anaerobically digested sewage
sludge is spread on agricultural land prior to ploughing in late autumn, which would allow the
three months of potential pathogen survival to pass before harvesting of plants, consistent
with the conclusions of Lang and colleagues (2007). However, the use of sewage sludge
(treated or untreated) on agricultural land in Sweden has been very restricted for some years
now due to the reluctance of the food industry and, hence, the farmers to use it (www.lrf.se).
To improve the quality of the sewage sludge, a Swedish project called ReVAQ, was recently
initiated and the results are evaluated by Urban Water AB (Malmqvist et al., 2006). Instead of
using the sludge as a fertilizer on arable land, de-watered digestion residues derived from
sewage sludge are frequently employed as inorganic matter for land reclamation and/or cover
material at closed landfill sites (Eriksen, 2009). Anaerobic digestion residue obtained from
the degradation of source-separated household waste is frequently used as a biofertilizer in
Sweden (www.avfallsverige.se), as the risk for spreading heavy metals is low.
In a 20-year field study at the University of Arizona (Zerzghi, 2008, Pepper et al.,
2008), the influence of land application of biosolids on the soil microbial community was
evaluated by assessing microbial number, activity and diversity. Class B biosolids had no
adverse affects on soil microbial number (Pepper et al., 2008), but increased microbial
diversity (evaluated through cloning and sequencing of bacterial 16S rRNA genes) and
activity (assessed from common microbial transformations, nitrification and sulfur
oxidation). Moreover, no known pathogens were present in soil sampled 9 months after
the last application of biosolids, and after 20 years of biosolid treatment, fecal coliform
and total coliform counts did not exceed 3 MPN g-1 in the amended plots (Pepper et al.,
2008). Lawlor and co-workers (2000) reported that soil amendment by biosolids affected
overall diversity and biomass only to a limited extent, but had varying effects on
individual species, functions and specific microbial parameters. These results were
consistent with a study by Dennis and Fresquez (1989) reporting shifts in the microbial
community structure as a result of biosolid application. Moreover, Sullivan et al. (2006a)
showed that biosolids negatively affected the presence of arbuscular mycorrhizal fungi
(AMF; measured using specific biomarkers), but augmented the relative abundance of
both Gram-positive and Gram-negative bacteria. The decrease in AMF biomarkers and
increase in bacterial markers may be explained by the concept that soil microbial
communities of infertile ecosystems are frequently dominated by fungi, whereas those of
more fertile, productive ecosystems (such as biosolids) predominantly contain bacteria
(Kourtev et al., 2003, Grayston et al., 2004, Wardle et al., 2004). In addition, Sullivan
and colleagues (2006b) showed that biosolid-amended soil microbial communities were
able to utilize Biolog EcoPlate substrate more quickly (via higher mineralization
activity), compared to microbial communities from control soils. Based on the results, the
authors suggest that the metabolism of at least parts of the soil microbial community
remains elevated for extensive periods after land application of biosolids.
Evaluation of soil chemical properties at the end of the 20-year land application study
in Arizona (Pepper et al., 2008, Zerzghi, 2008) revealed an increase in total and available
soil phosphate concentrations, consistent with data obtained from other similar studies
(Mantovi et al., 2005, Brendecke et al., 1993). In addition, total N was increased in soil
amended with biosolids, and nitrate levels in both biosolid and fertilizer-treated soil
exceeded 10 ppm NO3-N at most soil depths down to 150 cm, indicating a potential risk
for nitrate pollution of ground water, irrespective of the use of biosolids or mineral
fertilizers (Pepper et al., 2008).
Hattori (1988) examined the potential relationship between sewage sludge
decomposition and microbial numbers and activities in soil. The mineralization rate of
organic C and N rapidly increased bacterial numbers and proteinase activity in soil to
maximum levels within the first 3 days, followed by a rapid decline thereafter. Bact erial
numbers were significantly correlated with enzymatic activities and mineralization rates
of organic C and N, indicating that proteinase-producing bacteria certainly contribute to
rapid degradation of proteins in sludge immediately after soil application (Hattori, 1988).
These types of studies would greatly benefit from the development of new molecular
approaches (see review by Arthurson, 2008) providing additional information on bacterial
community structures, metabolic activities and responses to C and N mineralization rates,
and ultimately, correlation of these parameters with the effectiveness and quality of
digested sewage sludge as a crop fertilizer. Molecular methods also have a considerable
244 Veronica Arthurson
CONCLUSION
Do the Benefits of Recycling Biosolids Outweigh the Risks?
The value of biosolids as a fertilizer has been recognized for decades. However, the
related risks of transfer and persistence of human pathogenic microorganisms, disruption of
soil biological processes, altered soil quality and/or fertility, as well as accumulation of
persistent organic pollutants and heavy metals have not been investigated until relatively
recently. Numerous studies have focused on the heavy metals and organic contaminants
present in sewage sludge. However, limited information is available on risk analyses and
strategies to monitor microbial activities in biosolid-amended soil. The human pathogenic
microorganisms present in biosolids naturally vary depending on several factors, and reflect
the incidence of specific diseases in the community where the sewage sludge are derived.
Hence, the environmental risks of land application should be considered for each individual
biosolid product, depending on the specific compositions.
In this chapter, the microbial, physical and chemical properties of biosolid-amended soil
is discussed, focusing on the potential alterations in these parameters following the major
input of organic matter, plant nutrients and toxic elements, potentially present in the
biosolids). The majority of studies performed to date reveal no or limited negative effects on
soil microbial/physical/chemical parameters, and in many cases, show that biosolids exert
equivalent or better fertilization effects, compared to mineral alternatives. As long as the
crucial parameters (heavy metals, organic pollutants, human pathogenic microorganisms) are
carefully monitored in each batch of biosolids and values are within the regulated limitations,
at least Class A biosolids can be further exploited as a suitable crop fertilizer. However,
thorough analyses are required to establish the effects of biosolids applied to soil at the
microbial/enzymatic level. The majority of the publications reviewed in this chapter support
an increase in soil enzymatic processes, such as nitrification, sulphur oxidation and
dehydrogenase activity, as a result of biosolid amendment (Pepper et al., 2008, Zerzghi,
2008). Moreover, land application of biosolids have been shown to positively affect soil
microbial communities, leading to increased soil microbial number, activity and diversity
(Pepper et al., 2008, Zerzghi, 2008). Treatment of land with biosolids (and organic matter in
general) improves soil infiltration, water holding capacity, total porosity and soil aggregation,
and enhances the total organic carbon and nitrogen content (Epstein, 1998, Tsadilas et al.,
2005, Aggelides and Landra, 2000, Zerzghi, 2008). In conclusion, land application of sewage
sludge raises certain crucial environmental issues, such as accumulation of organic pollutants
and heavy metals, persistence of human pathogens, leakage of excessive nutrients, and viral
contamination of surfaces and groundwater. Regulation of these negative parameters by strict
monitoring and further assessment, in conjunction with the several positive effects discussed
above, may allow exploitation of biosolids as a beneficial fertilizer from an
agricultural/environmental point of view. Injudicious amendment can have a potentially toxic
impact, while application of biosolids with adjusted parameters should provide essential
nutrients for agricultural crop growth, and may be a good environmental and economic
alternative to mineral fertilizers.
To minimize risks and optimize benefits from recycling biosolids, the entire recycling
process must be carefully managed in an appropriate manner. For instance, accurate records
246 Veronica Arthurson
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 249-269 © 2010 Nova Science Publishers, Inc.
Chapter 11
ABSTRACT
The purpose of the present study was to design an integrated wastewater treatment
system for a nalla (riverlet) flowing through Indian Institute of Technology Delhi (IITD),
India, besides its cost estimation and comparison with the conventional wastewater
treatment system. The design parameters for integrated aeration-cum-adsorption tank
were worked out for 240 m3 / d flow rate of the wastewater. The important parameters
used for the design included initial COD and BOD concentration in the influent,
treatment time, adsorbent dose, pH, adsorbent particle size and the desired COD and
BOD in the effluent after treatment as prescribed by Central Pollution Control Board,
(CPCB) Delhi, India. All the design parameters of this system were similar to those of
conventional system except for the replacement of aeration tank in conventional system
by the aeration-cum-adsorption tank. The concentration of COD and BOD of the treated
effluent by the integrated system were well within the permissible limits of CPCB
standards (for COD it is 100 ppm and for BOD of 30 ppm) to discharge in the canal for
irrigation purpose. It was worth mentioning here that the adsorbents used in the present
study were based on discarded materials which were available free of cost. Of course, the
cost of their transportation and processing should have been taken into account.
The total cost estimated for the conventional system and the adsorption based system
would be Rs. 198,312 and Rs. 141,275 respectively (including civil work, machinery,
labour, adsorbent and miscellaneous). The cost difference for the two systems would be
approximately Rs 57,037.
This design of integrated system has resulted into saving of cost by 28 % over the
conventional system. Thus, it is a good approach for saving of conventional energy in
*
Corresponding author: Department of Energy and Environmental Sciences, Ch. Devi Lal University, Sirsa,
Haryana, India. E-mail address: rani_sahu@yahoo.com, Telephone no.: +91-9416987345.
250 Rani Devi and R. P. Dahiya
addition to saving the cost of treatment and can be applicable for any country for decen-
tralized sector. Moreover, it is an open ended research and we can recommend more
research by changing the adsorbents types and operating parameters to improve the model.
1. INTRODUCTION
Water is an important commodity for the survival of living species on earth. Quantity of
water utilized has been an index for the quality of life of the people. More water utilized
essentially leads to more wastewater generation, be it in the urban or rural settings. For
improving the quality of life in decentralized sectors, the use of electricity, water supply,
sanitation and other amenities should be raised significantly.
It is well known that the wastewater of domestic origin typically contains organic matter,
pathogens, suspended solids, nutrients (nitrogen and phosphorus) and other pollutants. For
curtailing the environmental and health hazards, there is a need to bring these pollutants down
to the permissible limits for its safe disposal (3, 6, 8, 12). Removal of the organic
contaminants and pathogens from wastewater is of paramount importance (1, 5, 13). The
conventional wastewater treatment technologies as adopted in industrialized nations are
expensive to build, operate and maintain (4, 9, 10), especially for the decentralized communities.
Usually there is an inadequate arrangement for handling this wastewater which either flowing
in open channels or accumulates in low-lying areas or flows through natural open drains.
Ranges of technologies are available for treating these wastewater pollutants. Research
efforts are on (10, 14, 15) for the development of better treatment technologies suited to these
decentralized communities. Keeping in view the requirements of suitability and money, it has
become imperative to find less costly and easily adaptable treatment technologies. Adsorption
based innovative technology (14, 15) proposal having low cost carbonaceous materials
showed good potential, for COD and BOD removal from the domestic wastewater.
The carbon content of adsorbents plays a significant role during the adsorption of organic
impurities. The adsorption capacity increases with the increased carbon content of the
adsorbent and such a trend has been observed by various investigators (7, 16).
Studies have shown that the carbonaceous adsorbents prepared from discarded materials
have potential for reducing the COD and BOD level (14, 15). Use of such adsorbents can
either completely replace or supplement the aeration tank. The aim of this paper is to design
an integrated wastewater treatment system for IIT Delhi nalla by incorporating conventional
as well as adsorption based treatment system. Results of the adsorption experiments reported
in our another paper are utilized in evaluating the design parameters of the adsorption reactor.
We have also calculated and compared the costs of the conventional and integrated
wastewater treatment systems for IIT Delhi nalla.
Technology (IIT), New Delhi, India, spills into a natural nalla (small drain). The drain flows
through IIT campus, meets other open drains on the way and eventually joins the river
Yamuna. For the present investigations, we have collected wastewater samples from five
points across the nalla. The samples were stored at 2-3 0C to avoid any change in their
physico-chemical characteristics. The important physico-chemical characteristics analyzed
were pH, temperature, total solids (TS), total suspended solids (TSS), chemical oxygen
demand (COD) and biochemical oxygen demand (BOD). The pH and temperature of the
wastewater samples were measured on collection site. Total solids, total suspended solids,
COD and BOD were analyzed in laboratory according to the methods prescribed by APHA
handbook (2). The COD and BOD of the wastewater samples were measured in laboratory
before and after treatment with the mixed adsorbent carbon prepared by mixing fly ash, saw
dust, sugarcane bagasse, coconut coir and brick kiln ash in 1:1 ratio). Wastewater was treated
under batch mode by using mixed adsorbent carbon for removing COD and BOD load from
wastewater and the treatment conditions included; treatment time, adsorbent dose, ph, initial
COD and BOD concentrations, agitation speed and adsorbents particle size.
The parameters considered for designing a domestic wastewater treatment plant were,
wastewater flow rate, physico-chemical characteristics of the wastewater, mass loading and
desired characteristics of the effluent. From the wastewater discharge, flowrate was calculated
and its average value was taken for designing the wastewater treatment plant for the nalla.
The average design flow was the daily average wastewater flow rate per unit of wastewater
generated from the source and it was 240 m3/day. The conventional wastewater treatment
system for IIT Delhi nalla could broadly be divided in two units. The first is primary
treatment unit including bar screen chamber, grit chamber and primary clarifier and second is
secondary treatment unit including aeration tank, secondary clarifier and sludge handling
bed. There is a transfer pump between the primary and secondary treatment units.
The nomenclature used in the design equations were: volumetric flow rate as Q; volume,
length, breadth, depth and wall thickness of the tanks as V, L, B, D and d respectively;
retention time as t, wall surface area as A and volume of the building material as Vb with
appropriate subscripts. The concept of common walls led to material saving. Two or more
walls of the chambers could be shared with those of the bar screen chamber, grit chamber,
primary clarifier, aeration tank, sludge drying bed.
Bar Screen Chamber: Volume of the bar screen chamber (VBSC) was calculated by
using equation 1, wastewater flow rate was taken 240 m3/day and retention time of ½ hr.
VBSC = Q ts 1
For calculating the volume of material used, we assumed depth (Dsc) and breadth (Bsc) of
the chamber as 1 m and 1.3 m respectively. The chamber length (Lsc) was then ~ 3.85 m
along the flow direction. Number of walls (n) used in sharing were 2 and the volume of
material used (Vmbsc) was calculate by using the equation numbering 2 as:
252 Rani Devi and R. P. Dahiya
where dsc is the wall thickness with a value of 0.15 m and dfs is the foundation thickness with
a value of 0.2 m for the bar screen chamber.
Grit Removal Chamber: It share a common wall with bar screen chamber. Its bottom
was downward sloping and away from the screen chamber so that the grit could settle at the
end. It was cleaned manually. The important point in the design of the grit basin was that the
flow velocity should neither be too slow nor too high. Flow velocity should be enough to
scour out the settled organic matter and reintroduce it into the flow stream. Such a critical
scouring velocity (vH) is given by the modified Shield‘s formula (Garg; 1998). For grit
particles of 0.2 mm diameter and with sedimentation co-efficient of 0.986, the range of
critical scour velocity was calculated from equation numbering 3 as:
where dg is the diameter of the grit particle, g is acceleration due to gravity and Ss is the
sedimentation coefficient. vH thus calculated was 0.11 to 0.25 m/s.
Like the bar screen chamber, the retention time in this section has been taken as ½ hr for
an average flow rate of 240 m3/day and ¼ hr for peak flow rate of 480 m3/day. We have
evaluated the grit chamber tank volume (Vgc) by using equation 1. Volume of building
material (Vmg) was calculated using equation 2 and also the concept of one wall sharing was
also maintained. The tank has a triangular sloping outlet and a slot.
Primary Clarifier: Since the effluent generation varies throughout the day, it is
necessary to hold the effluent in the primary clarifier for about 4 hours retention time to
homogenize the quality and quantity of the effluent. We assumed the tank depth, sidewall
thickness and thickness of foundation as 3 m, 0.15 m and 0.20 m respectively. We have
calculated the volume of primary clarifier (VCS), surface area (Acs), sidewall (S) and volume
of material used in primary clarifier (Vmcs) by using these assumed parameters.
Aeration Tank: Effluent was transferred from primary clarification tank to the aeration
tank at a rate of 10 m3/hr using transfer pump of 1.5 hp rating. Aeration tank is effective upto
85–90 % reduction of suspended solids, biological oxygen demand (BOD) and chemical
oxygen demand (COD) from wastewater. In the activated sludge process we usually consider
pH, BOD, BOD: N: P ratio and F/M ratio (this ratio is also known as the food to
microorganisms ratio) as 7-7.5, 300 mg/l, 100:5:1 and 0.2:0.5 respectively. The volume of
aeration tank (Va) with 240 m3/day flow rate of wastewater was calculated by using the
equation given as number 4.
Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 253
Va = Q (Bbod) / (Xfm) Y 4
By assuming depth of tank (D) as 3 m, area of the aeration tank (Aa), side of aeration tank
(Sa) and BOD load (Bld) of tank were calculated by using equations numbering 5,6 and 7
respectively as:
Aa = VA / D 5
Sa = Aa 6
Bld = Q.B 7
The oxygen required (O2) for the aeration tank is 1.5 kg/kg of BOD load/day (Bld). So O2
required was calculated as:
We assumed oxygen transfer rate (Otr) for fixed aerator was 1.35 kg/hp/hr. Thus horse
power (H.P.) required for the aerator was calculated from equation numbering 9 as:
To calculate volume of material used in the aeration (Vma) tank, we have area of side wall
(As)2 as (6.324)2 , depth of tank (D) was 3 m and area of foundation (Afa)2 as (6.47)2 m2 and
we assumed thickness of foundation (dfa) as 0.2 m. So Vma was calculated using equation
numbering 10 as:
The amount of sludge required for recirculation was calculated with the help of equation
numbering 11 as:
QR = (QY)/(Xs-Y) 11
where QR is recirculated sludge (m3/d), Q is wastewater flow rate (m3/d), Y is mixed liquor
suspended solid (MLSS) (mg/l) and Xs is activated sludge quantity from secondary clarifier.
Secondary Clarifier: The mixed liquor passed through a sedimentation tank where
separation of the activated sludge from aerated water took place. The settled activated sludge
was removed from the clarifier by gravity and divided into two streams. One stream called
return sludge was sent back to the aeration tank near the inlet. Here, it mixed with the
incoming primary treated effluent and acted as seed for the formation of more activated
sludge and simultaneously maintained the MLSS between 3000-3500 mg/l. The other stream
was excess sludge and was sent to the digester for digestion or percolation off as manure or
for using as land fill. The percolated liquor from the sludge was sent back to the process
stream for treatment. The treated effluent might also be used for horticultural purposes. The
254 Rani Devi and R. P. Dahiya
area (ASC), volume (VSC) and retention time (t) of secondary clarifier tank was calculated by
assuming depth (D) of tank 3.2 m by following equations numbering 12,13 and 14 as:
ASC = (Q + QR) / Lr 12
VSC = ASC D 13
t = VSC / Q 14
Sludge Drying Bed: Sludge formed in the clarifiers is dried in the sludge drying bed for
its final disposal. For designing the sludge drying bed, we assumed breadth (Bdb) of 4 m,
length (Ldb) of 4 m, depth (Ddb) of 1 m, volume of the bed (Vdb) as 16 m3 and wall thickness
of the foundation (ddb) as 0.15 m and thickness of foundation (dfdb) 0.2 m. So the volume of
material (Vmdb) used in the drying beds was calculated considering two wall sharing concept
by following equation numbering 17 as:
From the above calculations, the total volume (VT) of material used in the civil work was
calculated by following equation numbering 18 as:
2.2. Design Parameters for Integrated Wastewater Treatment Plant for IIT
Delhi Nalla
We have worked out the designing of the integrated wastewater treatment plant for IIT
Delhi nalla. Design of aeration tank was modified by incorporating adsorption process by
adding discarded material based mixed adsorbent carbon and it was named as aeration-cum-
adsorption tank. Aeration-cum-adsorption tank was designed on the basis of results obtained
from the adsorption experiments for COD and BOD reduction from domestic wastewater. All
the units of the wastewater treatment system described in section 2.1 for IIT Delhi nalla will
be similar except for the aeration tank which was replaced by aeration-cum-adsorption tank.
Now secondary clarifier was not required and aeration-cum-adsorption tank took its place.
Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 255
The design parameters for integrated aeration-cum-adsorption tank were worked out for 240
m3 / d flow rate of the wastewater.
A vertical partition wall was raised to divide the aeration-cum-adsorption tank in two
parts. A wire mesh filter was placed in the outlet slit of the tank to prevent the adsorbent loss
and also some additional adsorbent added intermittently on daily basis to make up the lose.
Horizontal partition was made in the middle of the tank which acts as a shelf. Discarded
material based mixed adsorbent carbon was added on the horizontal shelves and also at the
tank bottom in both sections. The wastewater coming from the primary settling tank was fed
to the aeration-cum- adsorption tank from its top. A steady flow of air was bubbled through
the liquid from the shelf and bottom of the tank. This agitated and homogenized the COD and
BOD laden liquid and also contributed to their removal through the aeration process. The
volume of the aeration-cum-adsorption tank calculated. The volume of the construction
material used for the aeration-cum-adsorption (Vmaca) tank was calculated by assuming wall
thickness (daca) and foundation thickness (dfaca) of the tank as 0.15 m and 0.2 m respectively.
Total volume of the construction material used was calculated by following equation
numbering 19 as:
The total volume of the material used in constructing the aeration-cum-adsorption based
wastewater treatment system included the material for civil work in all the units of the
treatment system as given in section 2.1 and 2.2. There would be no secondary clarifier with
the aeration-cum-adsorption unit but there were two units of aeration-cum-adsorption tank as
described earlier. The total volume of the construction material (VTmaca) was given by
following equation numbering 20 as:
The estimate was made of approximate cost for the construction of conventional as well
as integrated wastewater treatment plant for a small community having 240 m3/day flow rate.
The cost estimation for the adsorption based wastewater treatment system included the
estimated cost of civil works, electricity, mechanical equipments, aeration equipment and the
cost of adsorbent.
plant. Figure 3 depicted the hydraulic flow rate for each unit of the conventional wastewater
treatment plant for IIT Delhi nalla and Figure 4 showed the sectional drawing for each unit of
the conventional wastewater treatment plant for IIT Delhi nalla. The concept of common
walls led to saving in material. Thus, two or more walls sharing concept could be used for bar
screen chamber, grit chamber, primary clarifier, aeration tank, secondary clarifier and sludge
bed for saving the cost.
Parameters Values
3
Actual flow rate (m /day) 200
Over design (%) 20
3
Average design flow rate (m /day) 240
Peak flow rate (m3/day) 480
Wastewater characteristics
pH 7.1
Total solids (mg/l) 1140
Total Suspended solid (mg/l) 650
COD (mg/l) 1080
BOD (mg/l) 505
Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 257
Figure 1. MainFigure
units for
1: the Main
treatment of the
units for wastewater
the treatment discharged
of the wastewaterindischarged
IIT Delhiinnalla.
IIT Delhi nalla.
Figure 2. Detailed diagram of units for the treatment of the wastewater discharged in IIT Delhi
Taking 240 m3/day flow rate (Q) and ½ hr retention time for this average flow in the
chamber, the volume of bar screen chamber came out to be 5 m3 and material for constructing
the tank walls with a foundation of 0.2 m depth yielded 2.155 m3. Like the bar screen
chamber, the retention time in grit removal chamber had been taken as ½ hr for an average
flow rate of 240 m3/day and ¼ hr for peak flow rate of 480 m3/day and volume of tank
calculated was 5 m3 and total volume of the construction material used was 1.59 m3 when
wall sharing concept was used. Volume of the primary clarifier for 4 hour retention time came
out to be 80 m3 and volume of the material used was 10.20 m3. Two pump sets were provided
using one as functioning and other as standby and both having pumping capacity of 10 m3/hr
and 4.5 m head with 1.5 hp rating as shown in Figure 2.
258 Rani Devi and R. P. Dahiya
Figure 3. Schematics of hydraulic flow rate for each unit of the conventional wastewater treatment plant
for IIT Delhi nalla.
Figure 3: Schematics of hydraulic flow rate for each unit of the conventional wastewater
treatment plant for IIT Delhi nalla.
Figure 4. 4:
Figure Flow diagram to show
Flow diagram to sectional drawing
show sectional for each
drawing unit unit
for each of the
of conventional wastewater
the conventional wastewater
treatment
treatmentplant forforIIT
plant IITDelhi
Delhinalla.
nalla.
Volume of aeration tank was calculated as 120 m3 and area of the aeration tank (Aa), side
of tank (Sa) and BOD load (Bld) of tank were 40 m2, 6.324 m and 72 kg/day respectively. For
the present design, value of oxygen required as calculated from equation 8 was 108 kg/day.
And thus for the aeration tank power of 2.9 hp was required as calculated from equation 9.
But to include stringent requirements, motor power was of 5 hp. Volume of material used in
aeration tank came out to be 14.142 m3 as calculated from equation 10. Quantity of
recirculated sludge as calculated by using equation 11 was 88 m3/day and the ratio for sludge
Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 259
recirculated to wastewater flow rate would be 0.366. The volume of secondary clarifier tank
was 52.48 m3 with retention time of 5.2 hours as calculated by using equation 12, 13 and 14
and volume of material used in secondary clarifier was 7.06 m3. Volume of the sludge drying
bed as calculated by equation 17 was 4.4 m3. The overall design dimensions of conventional
wastewater treatment plant for IIT Delhi nalla are shown in Table 3. The overall design of
this plant was presented by Figure 5. The total volume (VT) of material used for conventional
wastewater treatment for IIT Delhi Nalla as calculated for the civil work was 39.55 m3.
The design parameters for integrated aeration-cum-adsorption tank were worked out for
240 m3 / d flow rate of the wastewater. The important parameters used for the design were
presented in Table 4 and included initial COD and BOD concentration in the influent as 1080
mg/l and 505 mg/l respectively, treatment time of 4 hours, adsorbent dose of 1400 kg, pH 7,
adsorbent particle size of 0.053 mm and the desired COD and BOD in the effluent after
treatment as 100 mg/l and 30 mg/l as prescribed by Central Pollution Control Board (4),
Delhi, India. The dimensions calculated for aeration-cum-adsorption tank were 3 m 4 m
3.5 m as shown in Table 5. The design of the aeration-cum-adsorption tank was shown in
Figure 6. Volume of construction material used for aeration-cum-adsorption tank was 4.6 m3.
The final COD and BOD concentration in the effluent after treatment were 44.6 mg/l and
10.26 mg/l respectively and that was well below the permissible limits of Central Pollution
Control Board, Delhi, India.
Figure 5.Figure
Overall5:dimensions
Overall of wastewater
dimensions of treatment
wastewaterplant for IIT Delhi nalla.
treatment
plant for IIT Delhi nalla.
Vertical
partition 3m
Wastewater inlet 1.5 m
Final treated
effluent to
distribute
Compressed
air
Compressed
air
Horizontal
perforated
Adsorbents Shelf
1.5 m
Compressed
air
The cost estimation for the adsorption based wastewater treatment system described
above was a complex task. It included the estimated cost of civil works, electricity and
mechanical equipments besides aeration equipment and the cost of the adsorbent. The volume
of material used for the civil work of the aeration-cum-adsorption based wastewater treatment
system was 27.54 m3 and for the conventional system as given in section 5.3 was 39.55 m3.
The approximate cost of construction material was taken as Rs. 3000 per m3. The cost of
construction and civil works then came out to be Rs. 118,650 for the conventional system and
Rs. 82,620 for the adsorption based system. Adding 25 % as additional cost including the
labour cost and overhead charges etc., then the cost of construction and civil works came out
to be Rs. 148,312 for the conventional system and Rs. 103,275 for the adsorption based
system. The cost of the mechanical system and electrical motors / pumps for the complete
system could be worked out from the specific choice of the pumps and aerator. It was,
however, expected that the capital cost for these equipments, other than the aerator, would be
approximately Rs. 25,000 for the conventional 240 m3 / d wastewater treatment system. The
aerator together with its motor would cost additional amount of nearly Rs. 25,000. For the
adsorption based unit, the aerator could work with a smaller compressor powered by 1 hp
motor. Its cost would be about Rs. 10,000. It was worth mentioning here that the adsorbents
used in the present study were based on discarded materials which were available free of cost.
Of course, the cost of their transportation and processing should have been taken into account.
For 30 days charge of adsorbents and its make up material an amount of Rs. 3000 might be
required.
The total cost estimated for the conventional system and the adsorption based system
would, therefore, be Rs. 198,312 and Rs. 141,275 respectively. The cost difference for the
two systems would be approximately Rs 57,037.
CONCLUSIONS
It is concluded from this paper that the replacement of aeration tank of conventional
system by aeration-cum-adsorption tank can be a good option for wastewater treatment for
decentralized communities. Due to the use of discarded material based carbonaceous
adsorbents, the COD and BOD concentration has been reduced considerably and these limits
262 Rani Devi and R. P. Dahiya
are well within permissible limits of CPCB. It has also resulted into reduction of cost by
approximately 28 %. Hence the proposed design of domestic wastewater treatment plant can
be adopted for decentralized communities of the world.
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[10] Mazumder, D. & Roy, B. (2000). Low cost options for treatment and reuse of municipal
wastewater, Indian J. Env. Prot., Vol. 20(7), 529-532.
[11] Piet, N. L., Piet, M. V., Lode, S. & Willy, H. V. (1994). Direct treatment of domestic
wastewater by percolation over peat, bark and woodchips, Water Res., Vol. 28(1),
17-26.
[12] Rani Devi, Dahiya, R. P. & Gadgil, K. (2002). Investigation of coconut coir carbon and
sawdust based adsorbents for combined removal of COD and BOD from domestic
waste water, Water and Environmental Management Series, Inter. Water Assoc., 1209-
1218.
[13] Rani Devi & Dahiya, R. P. (2006). Chemical Oxygen Demand (COD) Reduction in
Domestic Wastewater by Fly Ash and Brick Kiln Ash, J. Water, Air and Soil Poll.,
Vol. 174(1-4), 33-46.
[14] Rani Devi, Dahiya, R. P., Ashok Kumar & Vijender Singh (2007). Meeting energy
requirement of wastewater treatment in rural sector, International Journal of Energy
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Integrated Approach for Domestic Wastewater Treatment in Decentralized Sectors 263
[15] Rani Devi & Dahiya, R. P. (2008). COD and BOD removal from domestic wastewater
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 265-284 © 2010 Nova Science Publishers, Inc.
Chapter 12
BIODEGRADATION CHARACTERISTICS
OF WASTEWATERS
ABSTRACT
The objective of this chapter is to put forward an overview of biodegradation
characteristics of wastewaters by emphasizing the significance of COD fractionation.
Recalcitrant COD fractions of effluents can be used as a tool to evaluate whether
discharge standards can be met with a prescribed biological treatment. Moreover, the
appropriate type of biological treatment applicable to the wastewater under investigation
can be addressed and the performance of an existing biological treatment system can be
appraised with reference to inert COD fractions. Besides recalcitrant COD fractions of
segregated industrial effluent streams can be regarded as an essential input of a sound
industrial wastewater management strategy adopting minimization at source philosophy.
Last but not least, data on COD fractions can be used as a solid source of information for
modelling studies that define the design and performance of biological treatment systems.
In this context, COD fractionation data on a wide spectrum of activities ranging from
various industrial sectors to hotels is presented. Segregated industrial wastewater streams
together with domestic sewage and end-of-pipe industrial effluents are evaluated in terms
of their biodegradation characteristics.
1. INTRODUCTION
Research activities focused on the understanding of the biodegradation characteristics of
wastewaters are among the significant developments shaping environmental sciences in the
* e-mail: germirliba@itu.edu.tr
≠ e-mail: orhon@itu.edu.tr
266 Fatos Germirli Babuna and Derin Orhon
last several decades. Characterizing the organic content of effluents via COD parameter has
remarkable advantages, as biodegradable COD provides an electron and energy balance
between organic matter, biomass and oxygen utilized. This requires reliable determination of
biodegradable fraction. Within the existing legislative framework it is possible to meet the
effluent discharge standards defined in terms of COD by applying conventional biological
treatment. In this manner domestic sewage and industrial effluents are handled as if they have
similar characteristics. The effluents directed towards a biological type of treatment are
composed of basically biodegradable and non biodegradable fractions yielding a treated
effluent COD level under the discharge standards. Since all the biodegradable COD fractions
are removed within biological treatment, COD remaining in the treatment plant outlet only
contains the recalcitrant organics initially present in the wastewater itself with an additional
supplement of residual microbial products generated during the course of biochemical
reactions that take place in biological treatment. It should be kept in mind that currently,
compliance with discharge standards depends on an evaluation based on total COD values.
Such a traditional approach can be tolerated for domestic effluents. However it may cause
remarkable negative impacts when applied to industrial effluents that are likely to contain an
array of chemicals with various biodegradation characteristics. Identification of COD
fractions with different biodegradability characteristics has been a turning point in the
appraisal of wastewater characteristics. In this context, biodegradable and recalcitrant COD
fractions have been experimentally assessed. The mentioned COD fractions together with
experimentally determined kinetic and stoichiometric coefficients have been incorporated as
inputs into new models structured to define the design and performance of biological
treatment systems i.e. activated sludge.
From a different perspective, biodegradation characteristics of segregated industrial
effluent streams should be considered as backbone information leading to a sound industrial
management strategy. At least a part of the countless numbers of chemicals used during
manufacturing processes is discharged with the waste streams. Therefore, chemicals having a
recalcitrant nature are introduced into the environment as they by-pass biological treatment
plants without being removed. A new management protocol adopting waste minimization at
source philosophy must be developed for handling such industrial wastewater streams. For
this purpose first of all, auxiliary industrial chemicals generating highly recalcitrant
discharges can be substituted by biodegradable ones. If such an application is not possible due
to case wise reasons, then a specific treatment, preferably not targeting a complete removal
but instead increasing the biodegradable fraction can be prescribed for these non
biodegradable segregated industrial streams. These recalcitrant wastewater streams should be
mixed with others only after passing through the mentioned specific treatment.
In this context this chapter intends to provide an overview of biodegradation
characteristics of wastewaters by presenting data on both domestic sewage and industrial
effluents.
Biodegradation Characteristics of Wastewaters 267
aromatic compounds (i.e. toluene, pyridine, etc.) not oxidized in the COD test can be
considered as insignificant and therefore neglected in most effluents. On the other hand, COD
cannot differentiate biodegradable organics from the non biodegradable compounds that are
also oxidized in the test. However this issue no longer poses a major disadvantage since
separate experimental protocols are developed for the identification of biodegradable and non
biodegradable COD fractions under both aerobic and anaerobic conditions. The details of
these experimental protocols based on either observing batch reactors or respirometric
techniques can be found in literature (Germirli et al., 1991 and 1993; Germirli Babuna et al.,
1998a; Ince et al., 1998; Orhon et al., 1999a; Orhon and Okutman, 2003). COD is often
preferred to BOD and TOC as it provides an electron and energy balance between the organic
matter, biomass and oxygen utilized with the stipulation that its biodegradable fraction is
determined.
Alhough brief evaluation given above clearly recognizes COD parameter as the most
trustworthy method for reflecting the organic content of effluents, the adoption of
BOD5/COD ratio is still used as a preliminary index of biodegradability. It is stated in
literature that the effluents with BOD5/COD ratios higher than 0.4 may be considered as
entirely biodegradable (Chamarro et al., 2001). Table 1 outlines BOD5/COD ratios of selected
industrial wastewaters together with segregated industrial discharge streams carrying some
auxiliary chemicals and domestic sewage.
As evident from the table certain industrial effluents (i.e. one textile plant dealing with
cotton knit fabric manufacturing operations) and almost all of the segregated industrial
streams containing the listed auxiliaries have very low BOD5/COD ratios indicating their
recalcitrant and/or inhibitory and/or toxic nature. An increase observed in BOD5/COD ratio of
a wastewater after being subjected to a partial chemical oxidation is commonly attributed to
increased biodegradation. In other words, the potential of a chemical oxidation process to
improve the biodegradability characteristics of an effluent is often monitored and
consequently appraised by observing the BOD5/COD ratio. However the above given
discussion on BOD parameter clearly indicates the misleading nature of such an evaluation.
Organics or recalcitrant organics or inhibitory substances or toxic compounds, totally or
partly, might be removed by partial chemical oxidation. On the other hand partial chemical
oxidation might generate toxic and/or inhibitory by-products as well (Sharma et al., 2007).
The BOD5/COD ratio presenting the net effect of all these possible mechanisms has to be
handled very carefully and used in a limited manner by considering its deficiencies. Therefore
at this stage, it is recommended to enrich such preliminary rough evaluations by conducting
COD fractionationation experiments that yield more credible information.
The COD parameter reflects the amount of total organics in effluents by establishing
appropriate correlations among substrate, biomass and dissolved oxygen in terms of electron
equivalance. More precise information can be obtained when organics are differentiated in
terms of their biodegradation rates. In this context, the total COD, CT1, in wastewaters can be
divided into of two main components: The total non biodegradable or inert COD, CI1 and the
total biodegradable COD, CS1:
CT1=CS1+CI1
The inert COD can further be categorized in two subgroups identifying soluble and
particulate fractions as soluble inert COD, SI1 and particulate inert COD, XI1:
CI1=SI1+XI1
The total biodegradable COD, CS1 similarly comprises three fractions conveniently
differentiated as readily biodegradable COD, SS1, rapidly hydrolysable COD, SH1, and slowly
hydrolysable COD, XS1:
CS1=SS1+SH1+XS1
CT1
Total Influent COD
CS1 CI1
Total Total
Biodegradable COD Inert COD
ST
Total Soluble COD
SS+SH SI1 SP
Soluble Soluble Inert COD Soluble Inert
Biodegradable COD of Influent Origin Microbial Products
XT
Total Particulate
COD
XH XS XI1 XP
Viable Heterotrophic Particulate Particulate Inert COD Particulate Inert
Biomass Biodegradable COD of Influent Origin Microbial Products
Figure 2. Major soluble and particulate COD fractions in the effluent of biological treatment
The effluent of a properly designed and well operated biological treatment system does
not contain any biodegradable COD fractions as all the biodegradable organics will be
removed within the treatment plant. The particulate inert COD, XI1 and particulate inert
microbial products, XP, are entrapped within the biological sludge and leave the biotreatment
Biodegradation Characteristics of Wastewaters 271
system by sludge wasteage. In this respect among other COD fractions, the soluble inert COD
gains importance since it by-passes the biological treatment system without being involved in
the biochemical reactions. Soluble residual (inert) microbial products, SP, together with the
inert COD of influent origin, SI1, jointly control the magnitude of effluent soluble COD.
Therefore, the effluent soluble COD of a properly designed and well operated biological
treatment plant is composed of SP+SI1.
Experimental protocols are developed for the identification of readily biodegradable
COD, SS1; soluble inert COD, SI1; particulate inert COD, XI1; particulate inert microbial
products, XP and soluble inert microbial products, SP under aerobic conditions (Ekama et al.,
1986; Germirli et al., 1991 and 1993; Orhon et al., 1999a; Orhon and Okutman, 2003).
Similar experimental procedures defining soluble inert COD, SI1; particulate inert COD, XI1;
particulate inert microbial products, XP and soluble inert microbial products, SP under
anaerobic conditions are also available in literature (Germirli Babuna et al., 1998a; Ince et al.,
1998). Once readily biodegradable COD, SS1; soluble inert COD, SI1; and particulate inert
COD, XI1 are experimentally assessed rapidly hydrolysable COD (soluble), SH1, and slowly
hydrolysable COD (particulate), XS1 can be determined from mass balance equations.
The effluent inert COD levels obtained for different industrial wastewaters after passing
through aerobic and anaerobic treatment are tabulated in Table 2. Inert COD contents of
segregated industrial discharges carrying commonly used auxiliary chemicals are also given in
Table 2.
Table 2. The effluent inert COD levels for different industrial wastewaters under
aerobic and anaerobic conditions
The data presented in Table 2 indicates that an anaerobic treatment can be prescribed for
the treatment of wastewaters from the alcohol distillery. On the contrary for the laying step of
laying chicken industry, since the application of aerobic treatment yields a lower effluent
residual COD level than a corresponding anaerobic one; an aerobic type of treatment must be
adopted. Due to the generation of inert metabolic products, it is not possible to obtain an
effluent COD level lower than 3600 mg/l when treating the brewery wastewater under
investigation by means of an anaerobic system. An effluent quality lower than 340 mg/l of
COD can not be obtained when an aerobic biological treatment is prescribed for the
segregated industrial stream carrying the investigated lignosulphonate derivative. Quite a high
percentage of the COD generated by naphtalanesulphonate carrying discharge, namely 87 %,
must be regarded as biorecalcitrant in nature (Germirli Babuna et al., 2007b). Such a
remarkably high recalcitrant fraction indicates the necessity of applying a partial chemical
pre-treatment to this segregated industrial wastewater stream prior to letting it mixed with the
other wastewater sources which will consequently pass through a conventional biological
treatment.
The data on natural and synthetic tannin formulations represent an example that
accentuates the importance of biodegradability while substituting an auxiliary with another
one (Germirli Babuna et al., 2007a). The COD contents (CT1) tabulated for natural and
synthetic tannin display that application of synthetic tannin as an auxiliary chemical must be
preferred since in this case less COD is introduced to the segregated effluent. On the contrary
a correct evaluation can only be obtained from the data on inert soluble COD. When the
segregated discharges containing both of the tannin formulations are comparatively appraised
in terms of the lowest achievable COD levels after biological treatment, natural tannin is
observed to have a lower residual COD of 100 mg/l.
Table 3. (continued)
5. CONCLUSION
COD fractionation providing a new dimension for the biodegradation characteristics of
different fractions is the most useful experimental approach today for the elucidation of
wastewater character and the engineering decisions for appropriate treatment. This approach
also serves as an integral part of currently used mechanistic models. Recognition and
expermental assessment of the recalcitrant portions of COD in wastewaters, together with the
discovery of residual metobic products generation have to be considered as an important
milestone for performance evaluation of applicable biological treatment processes. However,
it is essential to recognize that COD fractionation may become more detrimental than
beneficial, if not properly assessed with the necessary understanding of relevant microbial
mechanisms associated with different experimental techniques and full account of various
analytical uncertainties.
ACKNOWLEDGMENT
Reviewed by Prof. Orhan YENIGUN
Affiliation of the reviewer: Director, Bogazici University, Institute of Environmental
Sciences
276 Fatos Germirli Babuna and Derin Orhon
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Eremektar, G., Karahan Gul, O., Germirli Babuna, F., Ovez, S., Uner, H. and Orhon, D.
(2002) Biological treatability of a corn wet mill effluent, Water Sci. Technol. 45(12),
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Grady, C.P.L., Daigger, G.T. and Lim, H.C. (1999) Biological Wastewater Treatment: 2nd
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Orhon, D., Artan, N., Buyukmurat, M. and Gorgun, E. (1992) The effect of residual COD on
the biological treatability of textile wastewater. Water Sci. Technol. 26(3-4) 815-825.
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278 Fatos Germirli Babuna and Derin Orhon
Chapter 13
ABSTRACT
In the present study, two types of colour removal systems were tested on effluent
samples collected from a coffee pulping factory which discharged on average 15 m3 of
wastewater daily with a colour index of about 2500 OH that was too high for direct
discharge into a river in Kenya. The two colour removal systems used were: (i)
electrolysis combined with wood ash or coffee husks leachate and (ii) electrolysis
combined with phosphate rock solutions at a rate of 0.5 g/l to 4g/l. Phosphate rock is
often used as agricultural liming agent. The surface area of the electrodes was set at close
to 75 m2/m3 of effluent with a current density of 1,200 mA/m2. The experiments were laid
out in a stratified random sampling design and the data were analysed using the Statistical
Package for Social Scientists (SPSS) computer programme version 10.0. Electrolysis
combined with phosphate rock (ELPHOS) proved to be the best process in terms of
power consumption (68% reduction) compared with the 57% reduction by electrolysis
combined with wood ash (ELCAS) and the 58% reduction by electrolysis combined with
*
Corresponding author: Email: lazetiegni@amatala.org
280 L. Etiégni, D. O. Oricho, K. Senelwa et al.
coffee husks ash (ELCHAS). Besides the 100% colour removal, ELPHOS also reduced
other effluent physico-chemical parameters such as BOD, COD, TSS and TS by 79%,
80%, 69%, and 88% respectively. The analysis of ELPHOS treated wastewater showed
that the mill could discharge an effluent that meets local discharge standards for colour
requirements. It is recommended that recycling of the treated water by ELPHOS back to
the factory for cleaning and washing purposes be considered since the quality meets the
requirement for uses of fresh water for cleaning purposes. Furthermore, calculation of
power consumption based on a scale-up batch reactor of 15 m3 proved less expensive to
treat the factory effluent than a set of 12 one 100-L reactors similar to the one used in the
field.
INTRODUCTION
Coffee is a member of the large Rubiacea family, where it constitutes the coffea genus (Coste,
1992). The genus coffea contains approximately 70 species, and the most widely grown are
the arabica and the robusta species (Cambrony, 1992). Today, after petroleum, coffee
euphemistically referred to as ―Black gold‖, is the World's most important traded commodity,
standing above coal, meat, wheat and sugar. The global harvest of coffee, however, is subject
to considerable fluctuations from year to year. These fluctuations are caused by a variety of
factors such as climate-induced fluctuations especially in Brazil, the World larger coffee
producer, the amount of coffee produced and the price charged which are determined by the
commercial policy interests of the producing and purchasing countries. Its cultivation,
processing, trading, transportation and marketing provide employment for a large population
base in all producing countries (Muleta, 2007).
Botanical evidence indicates that coffee plant "Coffea arabica", the variety mostly grown
in Kenya originated from the Abyssinian highlands of central Ethiopia where it was cultivated
by Harrrar tribe but can still grow wild today. Somehow Arab traders got the beans from
Ethiopia across the Red Sea to Yemen around the 6th century AD. Before this, African
Indigenous populations in Ethiopia were using the beans as a solid food: the ripe berries were
squashed, combined with animal fats and shaped into round balls that could be carried and
eaten on long journeys. It wasn't until 1615 that the first shipment of coffee arrived in Europe
at Venice (the European trading headquarters at that time) from Turkey, and coffee houses
quickly spread through Italy to Vienna (in today Austria), then on to most of Europe. France
is credited with the first introduction of coffee to the Americas through its colonies of the
Martinique, the West Indies and the French Guiana where the first coffee plantations were
founded in 1720. In 1727, from these French colonies, coffee found its way to Brazil, where it
became the mainstay of the economy, accounting for 63 percent of the country's exports by
1891. In 1893 coffee from Brazil was introduced to the then British colonies of Kenya and
Tanganyika in East-Africa, only a few hundred kilometres south of where it had originated, in
Ethiopia. Today coffee is grown in most parts of Kenya employing around 6 million people
across a country of about 40 millions inhabitants where it constitutes one of the two major
cash crops as well as an important export commodity, accounting for 15% of total export and
contributing substantially to the country‘s economy. In 2005 for example, Kenya exported
coffee worth a total of US$ 131 million. Kenya coffee beans are considered by gourmet
Batch Treatment for Colour Removal from a Coffee Factory Effluent… 281
coffee lovers as some of the finest in the world. They are ranked with beans from Jamaica's
Blue Mountain or Hawaii's Kona region for quality and flavour.
Coffee making can be divided into two types of processes: green coffee and dried coffee
beans processing. Water is mainly used during washing of the green coffee especially if the
process is a wet one. Wet-processing coffee is a relatively new method of removing the four
layers surrounding the coffee bean. This process results in a coffee that is cleaner, brighter, and
fruitier.
Most countries (e.g. Kenya) with coffee valued for its perceived acidity, process their
coffee using the wet-process (Coffee Research Institute, 2001). In SOCIFINAF, a large coffee
estate in Kenya with a coffee processing factory at Mchana, near Nairobi where the
experiment on colour removal was conducted, wet processing is used, which involves the
following steps: the removal of pulp (pulping) and mucilage (mucilage removal), washing
and fermentation for fully washed coffee as shown in Figure 1. During pick production, the
coffee estate at Mchana factory collects roughly 60,000 kg of green beans per day for a final
output of 3000 kg per day of dry coffee beans ready for export. On average, 15 m3 of
wastewater are generated per day from this factory.
Washing of the fermented coffee which is made of pectin materials has a negative impact
on the effluent parameters. The pectin materials include protopectin (33%), reducing sugars
namely glucose and fructose (30%), non-reducing sugars such as sucrose (20%), cellulose and
ash (17%). After thorough straining, the mucilage is allowed to soften and decompose before
its removal through washing which is generally carried out in large tanks, called washers, by
slowly moving coffee through open-topped channels (Coste, 1992). Wastewater from
fermentation and washing stages is kept in a concrete embankment where solids are allowed
to settle over night before the effluent is released into a near-by river. At Mchana factory, the
treated effluent still has a high colour level of 2,500 oH and a high BOD of 266 mg/L that
must be substantially reduced before any discharge into a near-by river is done according to
new environmental regulations in Kenya. Colour can damage aquatic life in a river through
the inhibition of light penetration into water and may adversely affect the aesthetical value of
any body of water. In addition, unlike the removal of BOD, COD, suspended solids, and
toxicity, which can be successfully reduced through primary treatment followed by biological
treatment, colour is extremely difficult and expensive to remove (UNEP, 1981; Ikehata and
Buchanan, 2000). Therefore there was a need to find an economical method of reducing these
effluent parameters to conform to local discharge standards.
There are several methods of removing colour which include: adsorption, soil media,
coagulation, ultraviolet irradiation, and membrane-based technology such as ultrafiltration
282 L. Etiégni, D. O. Oricho, K. Senelwa et al.
(ASTM,1983; Prasad and Joyce, 1991; Ikeheta and Buchanan, 2000). These technologies
can reduce pollution load and effluents colour to acceptable levels, but most are quite
expensive and very few are in common practice. Electrochemical treatment alone or
combined with wood ash leachate (ELCAS) has recently been tested and proved to be an
effective method of colour reduction from a pulp and paper mill (Etiégni et al. 2004; Orori et
al, 2005). However, over electrical polarisation within electrodes, which normally causes
excess voltage with wastage of electrical energy remains the biggest impediment of this
method (Springer et al., 1995; Ahonen, 2001). Electrical power cost is as high as US¢25.0 per
kWh in Kenya, more than six times what it is in South Africa, the continent‘s industrial
power house. In addition, high pH, high sludge production, high electrical conductivity of the
treated effluent and unavailability of wood ash when needed at some factories may hamper
the successful application of electrochemical methods. Good alternatives to wood ash that
will not only reduce power consumption during electrochemical application, but also be
inexpensive must be identified. In this paper, we report on the potential use of phosphate rock
and coffee husks ash as substitute for wood ash during the removal of colour from a green
coffee processing mill effluent by electrochemical method.
DC power supply
Non-conducting
material
Iron electrodes
-ve
+ve
47cm
47cm
30cm
45cm 65cm
Figure 2. Batch reactor for colour removal from a coffee factory effluent.
Batch Treatment for Colour Removal from a Coffee Factory Effluent… 283
Whenever possible, at least three replications were obtained. Chemical analysis of the
raw and treated coffee factory effluent was carried out for quality assessment using AAS or
flame photometer. Data were analyzed using Analysis of variance, Least significant
difference and Duncan test between different treatments using SPSS statistical package.
Samples of wastewater were collected at the effluent discharge point and promptly
transported for preservation at low temperatures (below 4oC) to avoid biological degradation
before analysis (UNESCO/WHO, 1978; Arudel, 2000). Temperature was measured at the
point of collection. Several other parameters such as alkalinity, colour, turbidity, pH,
biochemical oxygen demand, chemical oxygen demand, total solids, total suspended solids
and electrical conductivity were determined according to TAPPI standard procedures (Tappi
Test Methods, 1992). Colour of the effluent was measured by use of a Loviband colour
comparator.
To Power Supply
Non conducting
material
48.0 cm
Iron Electrodes
2.0 mm thick
30.0 cm
The electro-coagulation experiments were carried out with wood ash (ELCAS),
phosphate rock (ELPHOS) and coffee husks ash (ELCHAS) in a set-up depicted in Figure 2.
Experiments were performed in a monopolar batch reactor, with eighteen sacrificial iron
electrodes connected in parallel (monopolar parallel mode) and kept 20.0 mm apart using a
non-conducting material (Figure 3). Only the outer electrodes were connected to the power
source and anodic and cathodic reactions occurred when the current passed through the
electrodes.
The size of the internal plastic cell was 45 x 65 x 47 cm (width × length × depth) with an
effective volume of 100 litres. The electrodes were immersed 7/8 deep into the effluent
sample to achieve surface area coverage of 75 m2/m3 of wastewater. A power supply pack
with an input of 220 V and variable output of 0-40 V with maximum current of 1.5 amperes
was used as a direct current source during electrolysis. On average, the current density was
maintained at around 1200 mA/m2. After each run, the iron electrodes were rinsed in a water
284 L. Etiégni, D. O. Oricho, K. Senelwa et al.
bath following a scrubbing with a piece of cloth to minimize electrode fouling. For supporting
electrolytes, leaching of various materials i.e. wood ash, coffee husks ash and phosphate rock
was first carried out for 3hrs or 12 hrs (overnight soaked). After that, different volumes of the
leachates were used to yield concentrations of 500, 1000, 1500, 2000, 2500, 3000, 3500, and
4000 g/m3 of wastewater. For each run, the power required for complete colour removal (0oH)
was measured and power consumption calculated using Equation 1.
Phosphate rock (MPR) used in this work came from Northern Tanzania and was chosen
because of its reactivity and availability on local market as a low cost fertilizer. MPR
chemical concentration has been reported by van Kauwenbergh (1991). Wood ash and coffee
husks ash were obtained from the coffee factory.
Power (Watts.hr) = Current (I) * Potential Difference (V) * Time (hr) (1).
RESULTS AND DISCUSION
At the start of each run, the coffee factory effluent was slightly brown. However, with the
application of the current, the effluent turned darker and became more opaque as the reaction
progressed. A layer of foam started forming at the surface of the wastewater probably due to
the production of hydrogen gas. This was followed by the formation of a cloud of flocs and
subsequent rapid decantation.
10000
R/water COD for OSA COD for NSA COD for OSC
100
COD (mg/l)
10
1
0 500 1000 1500 2000 2500 3000 3500 4000
Concentration (g/m3)
For this study, the formation of a cloud was selected as an indication of the end of the
electrocoagulation reaction. Effluent temperature was normal at around 20oC. Before colour
removal, the effluent had a BOD of 851 mg/l, COD of 1845 mg/l and a colour of 2500 oH.
These effluent characteristics were found to be well above the 30 mg/l, 50 mg/l and 15 oH for
BOD, COD and colour respectively, the expected limits set by the National Environment
Management Authority (NEMA); the environmental regulatory body in Kenya.
Batch Treatment for Colour Removal from a Coffee Factory Effluent… 285
The pH of 5.2 was also lower than the recommended 6.5. The results of treated effluent
parameters for the coffee factory are listed in Table 1. ELCAS, ELCHAS and, ELPHOS had
the same reduction range for TS and TSS in the treated effluent from 85 to 89%. This
reduction was, however, not significantly affected by any increase in the concentration of
wood ash, coffee husks ash or phosphate rock leachate. The reduction of solids was probably
due to the fact that the two compounds contained comparable amounts of Ca2+ that acted as a
counter-ion during the experiment. Figure 5 shows BOD reduction of selected treatments and
soaking times.
Raw water
ELCAL ELCAS BH6
ELCAS SH3
ELCAL ELPHOS BH6
ELPHOSSH3
300
BOD (mg/l )
200
100
0
Treatment
The COD values for treated effluent reduced by ELCAS, ELCHAS and ELPHOS (for
both overnight soaked (OS) and not overnight soaked (NS)) are shown in Figure 4. Compared
to the initial raw effluent, the factory‘s COD decreased by between 66 and 91% , 51 and 83 %
and 68 and 94 % through ELCHAS, ELCAS and ELPHOS methods respectively. ELPHOS
was more effective for reducing COD than ELCAS, and the difference was statistically
significant at P = 0.05.
The final COD of 114-594 mg/l was still higher than the local effluent discharge standard
requirements. With the increase in the SE leachate concentration, COD and BOD values
decreased for ELCAS SH3, ELPHOS BH6, and ELCAS BH6. The reduction of COD during
this experiment was slightly higher than the removal rate of 74.0% reported by Orori (2003)
and Orori et al. (2005). The effect of soaking time on COD reduction through ELCHAS,
ELCAS or ELPHOS is shown in Figure 4. The results indicate that as wood ash, coffee husks
ash and phosphate rock were allowed to soak overnight, more removal of COD occurred in
the subsequent experiment. The explaination is that with increase in soaking time, more
supporting electrolyte (SE) was made available in the reactor, thereby increasing the electrical
conductivity of the solution in most cases and helping the electrocoagulation process (Wei-
Lung et al, 2009). Unlike with wood and coffee husks ashes, leaching from phosphate rock
reduced the electrical conductivity of the factory‘s effluent (Table 1). The increased
efficiency of phosphate rock leachate cannot therefore be explained simply through increased
electrical conductivity.
286 L. Etiégni, D. O. Oricho, K. Senelwa et al.
30
OSA NSA OSC NSC OSR NSR
25
20
Power (Whr)
15
10
0
0 1000 2000 3000 4000 5000
Electrolyte Dosage (g/m 3 )
Soaking time seemed to help the reduction of BOD although for example, the difference
between soaking time of 3 and 6 hours was not statistically significant (P> 0.05). This
confirmed findings by Etiégni and Campbell (1991) that generally during wood ash leaching
experiments, more than 90% of materials is expected to leach out after 30 minutes. ELCHAS
reduced BOD by between 47 and 88%, ELCAS between 28 and 78%, while ELPHOS
removed BOD between 51 and 90% and the differences were statistically significant (P<
0.05), although the wide variation in the recorded BOD removal values could not be easily
explained. Wood ash was the primary source of hydroxide ions that enhanced the process for
colour and BOD removals and probably led to reduced power consumption. Low power
consumption could also be attributed to the catalytic properties of metal oxides such as MgO
Mn2O3, Cr2O3, PbO2 found in wood ash and phosphate rock. The catalytic properties of these
metal oxides on the surface, or in the space between the anode and cathode during
electrocoagulation has been recognised in previous studies and might have assisted the
reduction of the time for current flow as reported by Ahonen (2001). Reduction of BOD by
ELPHOS in this experiment was in some instances higher than the values reported by both
Springer et al. (1995) of 70% and Ahonen (2001), and yielded a final BOD value of between
20 and 40 mg/l, the higher end of which is not acceptable by local discharge standards.
Power Reduction
Electrolysis combined with phosphate rock (ELPHOS) proved to be the best in terms of
power consumption (68% reduction) compared to ELCAS (57% reduction) and ELCHAS
(58% reduction), and the differences were statistically significant (P< 0.05) although it also
depended on the quantity of supporting electrolyte (SE) added to the effluent. The effect of
various SE concentrations on the treatment efficiency and power reduction of coffee factory
Batch Treatment for Colour Removal from a Coffee Factory Effluent… 287
effluent is shown in Figure 6. Wood ash, coffee husks ash and phosphate rock used as
supporting electrolytes helped reduce power consumption (Figure 6). Soaking seemed to avail
more supporting electrolyte to the colour removal process, although this may not be true with
phosphate rock. Of the three supporting electrolytes, coffee husks ash had the least impact on
colour removal when not soaked, followed by wood ash and phosphate rock.
The latter showed no significant difference between soaked and not soaked supporting
electrolyte. This stems from its low solubility which has been observed in studies where
phosphate rock was used in agriculture for pH control. When the concentration of the
supporting electrolyte (SE) in solution increased, power consumption reduced probably as a
result of increased conductivity (Figure 7). If the required voltage of the electrocoagulation
reaction is expressed as follows:
The removal of colour may have resulted from the combined effect of Fe2+ and Fe3+
generated in the solution during electrolysis of iron electrodes (Othman et al., 2006). It took
on average between 33 to 54 minutes, depending on the SE concentration, for the effluent
colour to be completely removed. These results could help determine the reactor‘s volume
needed based on the detention time for complete decolourization of the coffee factory‘s
effluent, if there is need for a continuous flow reactor:
Volume
Detention time =
Effluent flow rate
3.5
3.0
2.5
2.0
1.5
1.0
0.5
0.0
ELCAS BH6 ELCAS SH3 ELPHOS BH6 ELPHOS DH3 ELCAL
It shows that the necessary voltage v to access a certain current density had reduced
because of the introduction of the SE, so that the consumed electrical energy had also
decreased (Kashefialasl et al., 2006). SE is said to compress the double layer which in turn
reduces the Zeta potential of the substrate ions and helps their agglomeration or coagulation.
The reduction of power consumption in this experiment was slightly lower than the 80%
reduction reported by Etiegni et al. (2005) or Orori et al. (2005), probably because of the
higher electrode surface coverage of 80 m2/m3 used in these experiments. Higher surface
coverage normally lead to high efficient color removal. The presence of metal ions such as
Ca, Fe, Al may have also helped the coagulation process. Chemical analysis of wood ash in
previous studies has showed a significant presence of several chemical elements such as Ca
and Fe in the form of their corresponding oxides which can act as coagulant when dissolved
in water (Etiégni and Campbell, 1991).
Color Removal
Colour was effectively removed from the coffee factory‘s effluent through
electrocoagulation method. There appeared to be a positive effect of supporting electrolyte as
ELCHAS, ELCAS and ELPHOS removed 100% colour confirming the effectiveness of this
SE in electrocoagulation (Prasad and Joyce, 1991; Koparal and Gütveren, 2002).
ELCHAS, ELCAS and ELPHOS had a negative effect on the treated effluent pH as it
increased by between 27 and 75% for ELCHAS, 19 and 47% for ELCAS and between 9 and
22% for the ELPHOS as shown Table 1. The pH from ELPHOS was lower than ELCAS
probably because of the slow reactivity of phosphate rock compared to wood ash. ELCHAS
had a much higher impact on the coffee effluent final pH.
Aeration had the effect of reducing the factory treated effluent COD (Figure 8). Using
ELCAS followed by over-night aeration, COD was reduced to almost 5 mg/l after an initial
spike to 68 mg/l, while with ELPHOS, COD of the treated effluent remained almost constant
at 45 mg/l. This shows that after treatment with ELCAS, normal aeration often carried out in
most wastewater treatment processes could further reduce the coffee factory wastewater
parameters.
80 ELPHOS
ELCAS
COD (mg/l)
60
40
20
0
1 2 3 4 5 6
Time (hr)
Total Solids
There was a considerable reduction in solids in the treated effluent from the coffee
factory. The reason for such behaviour could be due to the fact that the EC treatment may
have induced the settling velocity of the suspended particles in which more suspended
particle agglomerates cloaked together. The exposure of wastewater to EC treatment would
contribute to a greater ionic charge so that more particles would collide and this would
eventually help in enhancing particles‘ attraction and agglomeration (Othman et al., 2006).
Table 1. Wastewater parameter for raw and treated effluent followed by aeration
(with ash socked for six hours).
600
400
200
0
0 1000 2000 3000 4000
Concentration (g/m 3)
Effluent electrical conductivity (EC) and alkalinity results are also described in Table 1.
The results show that ELCHAS increased EC by between 27 and 39%, ELCAS process
increased slightly the effluent EC by between 8.7 and 21%, while ELPHOS reduced EC by
as much as 51% and their net effect were statistically significant.
An analysis of treated effluent in Table 2 shows that ELPHOS did not substantially
increase minerals, even P for which it is used in agriculture. However ELCAS increased Na
and Mn concentrations in the treated effluent making it unfit for certain uses such as
agricultural irrigation.
600
Alkalinity (eq/m 3)
500
400
300
200
100
0
0 1000 2000 3000 4000
Concentration (g/m 3)
Electrical Conductivity
As mentioned earlier, the electrical conductivity of the treated effluent increased with the
volume of supporting electrolyte. There seemed however to be a maximum in this increase
which shows that maximum EC was achieved at 2000 mg/l or g/m3 except for never-soaked
ash which exhibited a maximum EC at 4000 mg/l (Figure 9). The effect of phosphate rock on
EC was negative. The difference between 2000 and 4000 mg/l was statistically significant
(P< 0.05).
Because of the negative impact of phosphate rock on EC, its overall positive effect on the
removal of colour and other coffee effluent parameters cannot be explained through its impact
on EC. Other factors such as the presence of CaO and MgO in phosphate rock may have
played a role in helping treat the factory‘s effluent (van Kauwenbergh, 1991).
Alkalinity
Alkalinity from ELCHAS increased from 15.0 to 56.0%, while ELPHOS first reduced
alkalinity by as much as 59% (Table 1, Figure 10). The difference in the effect of these two
processes was statistically significant (P< 0.05). The effect of ELCAS on coffee factory‘s
effluent alkalinity was almost insignificant. The alkalinity value of the treated coffee factory
effluent indicates that on average, the treated waste water could be more amenable to
biological treatment if there was further need to improve the effluent characteristics prior to
its discharge.
COST OF TREATMENT
Assuming a discharge of effluent per year of 6000 m3, the average power consumption
for selected best concentrations of SE are shown in Table 3. ELCAS and ELPHOS performed
best at a concentration of 4000 mg/L while ELCHAS at 3000 mg/l. ELPHOS cost was twice
that of ELCAS. The successful treatment of coffee factory‘s effluent using a 100 litre tank
shows that this surface to volume ratio of 75 m2/m3 could be used as a scale-up ratio to full-
scale treatment unit.
Feed Outlet
If a material balance is written around Figure 11, the equation around the reactor is:
Feed in = Feed out + rate of reaction + rate of accumulation ………. …………………. 3
Feed in = Feed out = 0; since no feed is going in and no feed is coming out (batch treatment).
So equation 3 becomes:
(rate of accumulation: dNA/dt) = rate of reaction * volume of the reactor: (-rA)V ……….4
NA = NAO(1-XA)
To do that, what one needs is to plot 1/rA versus XA; the area under the curve will be =
V*t/NAO. Since in our experiment we did not study the reaction kinetic, we cannot get the plot
1/rA versus XA, which will be the ideal way to estimate the volume V. All things being equal
we could use the volume Ratio. But since a direct ratio of Vp/Vlab = 150 is deemed too large,
to get around the problem, we can use flow system of two to three reactors (Constant Stirrer
Tank Reactors) in series. In the absence of the flow system, we will scale-up the reactor
using the surface to volume ratio mentioned earlier.
OPTION ONE
Dimensions of tank before scaling-up
i) Length 1 width 1 height 1 Volume 1
(cm) (cm) (cm) (cm3)
In the laboratory tank reactor, there were 6 cells each with three electrodes occupying a total
surface area of 4.54 m2. To maintain a surface to volume ratio of 75 m2/m3 for a total volume
of 15 m3, we need a total electrode surface area of 1125 m2. We calculate the total length of a
cell in the laboratory reactor as 6.6 cm. Number of cells to fit in the length of the scaled-up
reactor, after removing approximately 5 cm from either ends: (311.0 -10)/6.6 = 45.6 cells
If the voltage per cell is equal to 40 volts, the minimum power required to run
the color removal reaction is equal to:
40 x 225 = 9,000 W or 9.0 kW
Assuming a Transformer and Rectifier with combined efficiency of 85% at 0.95
power factor (pf), the power rating is:
9.0 kW
= 11.15 kW or 11.15 x (54/60) = 10.0
0.85 x 0.95
At US$0.25 the cost of a kWh in Kenya, the factory will be spending 10.0 x 0.25 =
US$ 2.50 per day to treat its wastewater.
Initial capital is may be high due cost of 15 m3 tank, hoists for the electrodes and their
insulators. In addition there will be the need to adapt a welding transformer and a rectifier to
provide a DC reactor current for only 54 minutes per day.
OPTION TWO
No scale-up of the laboratory tank, but setting 12 equal tanks side by side with a
total volume of 1.2 m3
294 L. Etiégni, D. O. Oricho, K. Senelwa et al.
Active surface area of an electrode = 0.126 m2. Since there are twelve (12) positives
electrodes in a tank, the total surface is 0.126 x 12 = 1.51 m2. For a current density of 1.2
Amps/m2, the total current going into a tank is 1.2 x 1.51 = 1.81 Amps. There are twelve
tanks in this option. The amount of current required is 1.81 x12 = 21.74 Amps per batch.
Assuming we maintain the same voltage of 40 volts, the power rating here will be:
40 x 21.74 = 869.6 W or 0.869 kW per batch.
Assuming a Transformer and Rectifier with combined efficiency of 85% at 0.95% power
factor (pf), the power rating will be: 0.869/(0.85 x 0.95) = 1.08 kW
Knowing that there will be 12.5 batches per day of 60 minutes each, the total power
consumed will be: 1.08 x 12.5 = 13.5 kW or 13.5 x (60/60) = 13.5 kWh
Total cost of power consumed = 13.5 x 0.25 = US$ 3.36 per day to run the color removal
reactors. In each of the two options presented above, the costs of control equipment
(pumping, draining, logic control) as well as the cost of labor has not been included. It
appears that the 1st option is the less expensive of the two.
ACKNOWLEDGMENT
The authors wish to acknowledge the assistance of Moi University, Kenya for funding
this project and Ms Abigael Nekessa and Mr. Thuita Moses, graduate students at Moi
University, Department of Soil Science, School of Agriculture and Biotechnology for
providing the necessary information on phosphate rock used in this experiment.
REFERENCES
Ahonen, H. (2001). A Review of General Information on Electrochemical Process
Wastewater Treatment. Finland Laboratories 7p.
Arudel, J. (2000). Sewage and Effluent Treatment. Wiley and Sons, Great Britain, 45-90.
ASTM. (1983). Standard Practice for Coagulation-Flocculation Jar Test of Water. Annual
Book of ASTM Standards. 1101.
Coste, R. (1992). Coffee - The plant and the product. MacMillan Press, London.
Etiégni, L. & Campbell, A. G. (1991). Physical and Chemical Characteristics of Wood Ash
Bioresource Technology, Vol. 37, 173-178.
Etiégni, L., Orori, B. O. & Rajab, M. S. (2004). An Electro-coagulation Method of Color
Removal in Wastewater or Water with Low Power Consumption. International Patent
(PCT/ KE2005/000012).
Ikeheta, Z. P. & Buchanan Q. S. (2000). Decolorization of pulp mill effluents with
immobilized lignin and manganese peroxidase from Phanerochaete chrysosporium.
Environ. Technol., 19, 521-528.
Kashefialasl, M., Khosravi, M., Marandi, R. & Seyyedi, K. (2006). Treatment of dye solution
containing colored index acid yellow 36 by electrocoagulation using iron electrodes.
Int. J. Environ. Sci. Tech., Vol. 2(4), 365-371.
Koparal, A. S. & Gütveren, Ö. B. Ü. (2002). Removal of nitrate fromwater by
electroreduction and electrocoagulation, J. Hazard. Mater, B 89, 83-94.
Muleta, D. (2007). Microbial inputs in coffee (Coffea arabica L.) production systems, South
Western Ethiopia. Implications for promotion of Biofertilizers and Biocontrol agents.
Doctoral thesis, Swedish university of Agricultural Sciences.
Orori, O. B., Etiégni, L., Rajab, M. S., Situma, L. M. & Ofosu-Asiedu, K. (2005).
Decolorization of a pulp and paper mill effluent in Webuye Kenya by a combination of
electrochemical and coagulation methods. Pulp and Paper, Canada, Vol. 106, No. 3, T50-
T55, 21-26.
Orori, O. B. (2003). Colour removal from wastewater of a Pulp and Paper Mill in Kenya, by a
combination of electrochemical and coagulation method. Mphil. Thesis, Department of
Wood Science & Technology, Moi University, Eldoret, Kenya, 146.
Othman, F., Sohaili, J., Ni‘am, F. M. & Fauzia, Z. (2006). Enhancing suspended solids
removal from wastewater using Fe electrodes. Malaysian Journal of Civil Engineering,
18(2), 139-148.
Prasad, D. Y. & Joyce, T. W. (1991). Colour removal from Kraft bleach plant effluents by
Trichoderma sp. TAPPI J., Vol. 65, No. 1, 165-169.
296 L. Etiégni, D. O. Oricho, K. Senelwa et al.
Chapter 14
M. A. Babu*
Department of Environment, Islamic University in Uganda; P.O .Box 2555,
Mbale- Uganda.
ABSTRACT
The major aim of this paper is to review the major problems of water resources in the
developing countries. It is based on problems related to population growth and pollution
and how these are more likely to lead to future conflicts. We know that fresh water is
only 3 % of the total global water and 78% of this is in glaciers. This makes it a scarce
and precious resource which must be sustainably managed. The paper also analyses some
of the already existing and potential conflicts based on water resources. It reviews the
potential threats to Ugandan water resources and problems which are most likely to occur
as a result of these threats. Factors hindering treatment of wastewater as a remedy to
pollution in developing countries have also been discussed. The methodology used in this
paper is based on literature review of the most current issues that affect water resources
world-wide. The review is limited to scientific facts and no political factors affecting
water resources have been included.
It has been found that although Uganda is endowed with 66km2/year of renewable
water resources, population increase, deforestation, degradation of wetlands and pollution
are major threats to its water resources. Problems associated with water quality and
quantities are more likely to result into internal conflicts which are bound to spread
beyond Ugandan borders.
*
Corresponding author: Phone: +256 45 33502, Fax: +256 45 34452, Email: Babumohd@yahoo.com
298 M. A. Babu
INTRODUCTION
Oil is an invaluable resource to mankind; it makes us fly, drive, generate power, till the
land and even drive military hardware. It is a potential threat to the western world and source
of conflict in the world politics. It is believed to be the major destabilizing factor in the
Middle East (Darwish, 1994). It is a luxury that mankind cannot afford to live without.
On the other hand, water is part of us - the human body is composed of 70- 95% water,
the food we eat is water in a different form. It is an essential component that drives the
ecosystem and the food chain. It is a resource with inherent values that cannot be comparable
and measurable. Imagine if water were to be mined like oil, would we live and survive?
Imagine New York City without water for 24 hours! Water is a basic need; it is not a luxury
that mankind can live without.
Water is becoming an increasingly vital resource and is thought to over take oil as the
potential cause of conflict in the Middle East. President Anwar Sadat once said Egypt will
never go to war except when its water resources are threatened. According to UN- reports, the
population in the Middle East is increasing rapidly (Table 1.0) and it is predicted that this will
exert more pressure on the already existing problems associated with water resources.
In the UN report, 18 countries will be on the list of water scarce countries by 2025. These
include: Algeria, Israel/Palestine, Qatar, Saudi Arabia, Somalia, Tunisia, United Arab
Emirates, Yemen, Egypt, Ethiopia, Iran, Morocco, Oman and Syria. Water scarcity can be
described as when country has less than 1,700 m3 per capita, it is said to be experiencing
water stress, while less than 1000 m3 is regarded as water shortage. This list includes most
countries that are already in the volatile Middle East region hence the possibility of conflict
cannot be ruled out. Past experience of the 1960s is testimony to the conflicts. Cross border
raids on water schemes' between Israel, Syria and Jordan culminated into the six day war in
1967 (Darnish, 1994).
Although natural water resources can be replenished through the hydrological cycle, then
one might urge that the natural cycle will take care of the situation. The average renewal rate
for rivers are about 18 days while large lakes and deep aquifer can take up to thousand of
years. The Nubian aquifer in North Africa- known as the world's oldest aquifer were thought
to be filled in past geological years. When depleted, it is not known how long it would take to
recharge (Darnish, 1994). Climate change, pollution and the demand to feed large populations
makes the problem more complex.
It is predicted that the main conflicts in Africa in the next 25 years could be over water. It
is thought that countries will fight each other to have access to this precious and scarce
resource (Russel, 2007). According to the UNDP report as cited by Russel, (2007), 12 African
countries will join the list of water scarce countries. This means that a person will have less
than 1000 m3 of water per year. Water requirement for domestic use depends on lifestyle and
availability. For instance, about 400, 200 and 10 - 20 litres per person per day are consumed
in North America, Europe and Sub-Saharan Africa respectively (World Water Council, 2000).
As can seen, the current consumption in Sub-Saharan Africa is the lowest and it is thought to
worsen in the next 50 years.
The Nile, Niger, Volta and Zambezi basins are believed to be major areas of conflicts.
Other potential water war areas in Southern Africa will involve Botswana, Namibia and
Angola. There are already tensions in these regions.
Water as a Scarce Resource: Potential for Future Conflicts 299
Table 1.0. Prediction of population (in Millions) of some selected Middle East countries.
According to Lester Brown (as cited by Russel, 2007), the head of environmental
research institute Worldwatch, the combined population of Ethiopia, Sudan and Egypt – will
rise from the current 150 million to 340 million in 2050. This will result intense competition
for increasingly limited water resources. "There is already little water left when the Nile
reaches the sea," he says.
Egypt is ready to use force to protect its vital water resources. It is much more concerned
with dams that might be constructed in the Ethiopian highlands which are thought to affect
the flow of the Nile. Egypt is not in compromising position as regards this. In 1989, the
Ethiopian ambassador in Egypt was summoned to explain the presence of Israel hydrologists
around the areas of the Blue Nile. During the same period, the Egyptian parliamentarians
declared their willingness to back their government in taking military action against Ethiopia
if it becomes necessary (Darwish, 1994).
In the Niger basin, Senegal and Mauritania have already fought two short wars in 1987 &
1989 over the Senegal River. The cause of this was that Mauritanian tribesmen searching for
vegetation crossed to the other bank, violating Senegalese sovereignty (Darwish, 1994).
UGANDAN SCENARIO
Uganda is naturally endowed with water resources; renewable water resources are
estimated to be 66km2/year which corresponds to 2800m3/person/year. Open water surface
covers 15% of the total land cover (Figure 2). Uganda also receives annual average rainfall of
900-2000mm of which 7-20% of this recharges the underground water.
300 M. A. Babu
The recharge to ground water is high as compared to the current abstraction volumes
hence there is no over exploitation at current situation. However, it should be noted that
Uganda has 106 towns of which 56 towns have piped water supply. Of the 56 towns, 15 large
towns are mainly supplied with lake or river water. The remaining 41 small towns depend on
high yield boreholes (WWAP, 2006). It should be borne in mind that these towns are rapidly
expanding and the demand for water may result into tapping more of underground water
sources. At same time, there are also 69 other small towns that do not have piped water
supply. If we are to meet the target for the Millennium development goals by supplying water
to the towns, the underground sources may be strained.
Uganda has not yet fitted into the larger picture of water wars. Whether it will feature or
not can be established through, analysing major factors that may drag it into action. In this
context, these are limited to only scientific and not political factors. In our own opinion,
threats to water sources may become future points of conflict as time goes by. The following
anthropogenic activities may be considered as threats to Ugandan water resources:
(a) Deforestation
It is well documented that areas of high altitude with dense rain forest covers receive high
precipitation (WWAP, 2006).The forests provide moisture in the hydrological cycle required
for rainfall formation. At the same time, vegetation cover traps and holds rain water; realising
it slowly hence recharging rivers, streams and underground wells throughout the year.
Water as a Scarce Resource: Potential for Future Conflicts 301
2000000
1800000
1600000
1400000
down stream will be limited to smaller areas that can be sustained by the little water from the
mountains. Competition for the limited wet areas available will cause tension in the rice
growing communities. Of recent, violence linked to possession of wetlands has been reported
in this area. It is also thought that rifts may develop between the communities up and down
streams as they scramble for little water available. Worthy noting, there be will silent
conflicts between man and biodiversity for the limited resources.
(b) Wetlands
Uganda has a total wetland area of 30,000km2. Like forests, wetland degradation is on the
increase. For instance, it is estimated that 45 % of the Nakivubo wetland has been modified or
reclaimed (Emerton et al, 1999). The Nakivubo wetlands as well as those in Eastern and other
parts of Uganda are under pressure. Key activities that degrade wetlands include agriculture,
development, settlement and solid waste disposal.
It has also been found that most of the wetlands of the valley bottoms have been
converted to agricultural use. This has led to change in micro climate and even lowering of
water table as seen in Kabale and Bushenyi (WWAP, 2006).
There are 3 hydrological functions of wetlands which are important in maintaining the
hydrological balance. First, wetlands act as holding water basins slowly releasing water into
rivers and streams thus ensuring continuous flow throughout the year. Secondly, the holding
basins also provide more time for seepage and recharge of underground water and thirdly,
wetlands play a vital role in providing water directly to its beneficiaries. It is estimated that
wetlands directly provide water to 5 million people in Uganda (WSSP, 2001).Hence, off
setting the hydrological balance will lead to drying up of most rivers, streams and wells. The
quantity of underground water sources is expected to decline.
Loss of wetlands may mean loss of many livelihoods. Many wetlands are used for
fishing, grazing, agriculture and for materials used in shelter. Effects of wetland loss are
likely to be felt by the poor rural communities who largely depend on them. Loss of
livelihood may become source to conflicts.
(c) Pollution
This is a major problem mainly facing urban areas. L.Victoria is of major interest since it
is a vital component in the Nile basin, being a resource shared by many riparian countries. It
is the second largest fresh water lake in the world with an area of 69,000km2. The lake is
increasingly receiving pollution from un-treated sewage as well as industrial effluent. Odada
et al., (2004) report that the number of people without sewers in urban areas of L. Victoria
region is high and yet the population growth is over 5- 10%. This raises concern in regard to
the lake quality. According to UNEP-SEO report (2004-2005), the lake is experiencing severe
impact on microbial contamination, eutrophication and suspended solids. It is also seen that
the Congo basin which is within the catchments areas of L.Victoria is under severe
eutrophication (Table 2.0).In fact, it is anticipated that the cost of water treatment for
Kampala city is bound to rise due to increased levels of phytoplankton in the lake. The city
draws its raw water from the lake yet the demand for water supply is fast tracking the rise in
population.
Water as a Scarce Resource: Potential for Future Conflicts 303
Table 2.0. Water pollution in selected water bodies (UNEP, SEO Report, 2004/2005).
Most industries located in Kampala and probably elsewhere in Uganda do not have
effluent treatment systems but drain their wastes into the lakes, rivers and environment. It has
been found that industries release 1,045kg/d BOD5, 96kg/d of nitrogen and 105kg/d of
phosphorous into the lake (WWAP, 2006). The BOD exerts oxygen demand on the water
affecting fish and other aquatic organisms. Nitrogen and phosphorous cause algal blooms,
which may cause skin irritation, production of toxins as well as increasing oxygen depletion
in the lake.
Cholera cases have become common in Kampala and cases of dysentery have increased
from 2300 in 1999 to 8300 in 2002 (WWAP, 2006).
Apart from the industrial effluent into L.Victoria, there are a number of flower farms
being established on the shore line of the lake. Flower farms are known to extensively use
fertilizers and pesticides. If this is left unchecked, water from the lake may become unfit for
consumption or else we pay the heavy price of treating pesticides.
As for the rural areas, small towns are rapidly cropping up and more pit latrines are being
sunk. Underground water will be contaminated and incidences of water borne diseases will be
high. Much as Uganda is endowed with water resources, lack of mitigation to pollution will
result in loss of this precious resource.
304 M. A. Babu
Uganda like many developing countries is experiencing rapid population growth. With
the invention and perfection of the Haber process in the 1960‘s, population and food
production has increased synonymously. The Haber process has greatly improved the
manufacture of nitrogen based fertilizers.
Nitrogen in the fertilizers is transformed to proteins in the food chain and upon
consumption followed by excretion, large amounts of nitrogenous products are released into
the environment (Mulder, 2003). This is known to pollute the receiving water bodies.
Nitrogenous wastes, organic matter and pathogens found in wastewater have affects on public
health and harmful ecological impacts on the environment (Gijzen and Mulder, 2001).
The growing concern of nitrogen, organic and pathogen pollution therefore calls for the
need of wastewater treatment before discharge into natural watercourses. At present, there is
hardly any infrastructure for effective treatment of sewage in developing countries. Municipal
sewerage system coverage and the extent of domestic and industrial wastewater treatment are
inadequate. Treatment level is insufficient in most urban situations (Gijzen et al., 2004). For
example in Latin America, only 14% of collected sewage receives treatment (WHO/UNICEF,
2000). Even when the facilities exist, poor maintenance and operation results in failure of
treatment processes causing pollution of the receiving surface waters (Gijzen et al., 2004).
In Sub-Saharan Africa, 42% of the population (as per 2002) lacked water supply and 64%
lacked basic sanitation (WHO-UNICEF, 2006). Reduction of these proportions will definitely
have environmental effects. Even if these percentages are spread to both urban and rural
communities, the ultimate end of the waste is the environment. This poses a challenge to most
governments.
The major hindrance to sewerage coverage and treatment systems in most developing
countries is the weak economies. Priority in these states is given to security, health and
education. For instance in Uganda, the 1998/1999 budget for sanitation dropped from 46%-
18% while that of education increased from 14-47% (WSP, 2004).
Also the cost of conventional wastewater treatment infrastructure is prohibitive for the
majority of these developing countries, Uganda being inclusive (Gijzen et al., 2004). The
implementation of conventional wastewater collection and treatment in developing countries
to attain EU standards is therefore unrealistic, except maybe in densely populated urban
centres where the average income is much higher. However, this should not be generalised as
most urban areas are facing problems of their own. Population growth in these areas is seen to
out grow the existing wastewater treatment infrastructure (Gijzen and Khonker, 1997; Yu et
al, 1997). Weak economies, corruption, increased demand for urban land and unplanned
development seems to make the expansion of the existing infrastructure nearly impossible
(WWAP, 2006).
Although the millennium development goals emphasize strong water and sanitation
component- which is good, there is a possibility that this will compound the problem of water
pollution further if not well handled. Goal seven for instance proposes reduction of half of the
proportion of people without access to safe drinking water and basic sanitation by 2015 (UN,
2007). It is estimated that 2.6 billion people lack access to sanitation (WHO-UNICEF, 2006).
If this goal is to be achieved; there will be increased generation of wastewater. Mara et al.,
(1992) estimate that 80-90% of water consumed is converted to wastewater. Increased
production of wastewater will put more burdens on the already strained waste management.
Water as a Scarce Resource: Potential for Future Conflicts 305
From this point of view, it is realistic that governments should think of strategies of
solving problems of water pollution if they are to achieve the millennium development goals
and avoid future conflicts. They should make use of low cost treatment technologies through
best approach to improve effluent quality. Much as there is evidence of success of
conventional approach, the concept still needs to be reconsidered from sustainability point of
view (Gijzen et al., 2004).
As such, efforts on improving nutrient removal using natural wastewater treatment
technologies like wastewater stabilization ponds and wetlands that are widely applied in
developing countries become paramount. This will entail protection of the environment from
pollution resulting from wastewater discharge. Governments should reduce the pressure on
the already existing treatment infrastructure by avoiding the centralized approach of sewerage
collection. It is recommended that they delocalize collection systems to many smaller but
effective treatment systems.
It is expected that with the growing population coupled with destruction of forests,
wetlands and increased pollution, the quality and quantity of available water will reduce. This
is most likely to spark conflicts in regions which are already facing water scarcity.
What Do We Learn?
ACKNOWLEDGMENTS
Am grateful to Dr. N. Sarah for sparing her time to read and review this work. Her
contributions have added invaluable knowledge to this paper.
306 M. A. Babu
REFERENCES
Darwish, A. (2004). A Lecture on Environment and Quality of Life, June 1994. Geneva
conference.
Emerton, L., Iyango, L., Luwum, P. & Malinga, A. (1999). The Present Value of Nakivubo
Urban Wetland, Uganda. National Wetlands Conservation & Management Program/
IUCN.
Environmental profile (2005). Forest cover of Uganda. www.rainforests.mongabay.com
Gijzen, H. J, Bos, J. J., Hilderink, H. B. M., Moussa, M., Niessen, L. W. & de Ruyter van
Steveninck, E. D. (2004). Quick scan health benefits and costs of water supply and
sanitation. Netherlands Environmental Assessment Agency. National Institute for Public
Health and the Environment – (MNP-RIVM), The Netherlands
Gijzen, H. J & Mulder, A. (2001). The global nitrogen cycle out of balance. Water, 21, Aug
2001, 38-40.
Gijzen, H. J. & Khondker, M. (1997). An overview of ecology, physiology, cultivation and
application of duckweed, Literature review. Report of Duckweed Research project.
Dhaka, Bangladesh.
Mara, D. D., Alabster, G. P., Pearson, H. W & Mills, S. W. (1992). Waste stabilization ponds,
a design manual for Eastern Africa, Lagoon Technology International Leeds, England.
Mulder, A. (2003). The quest for sustainable nitrogen removal technologies. Wat Sci. Tech.,
48(1), 67-75.
Odada, E. O., Olago, D. O., Kulindwa, K., Ntiba, M. & Wandiga, S. (2004). Mitigation of
environmental problems in L. Victoria, East Africa: Casual chain and policy option
analyses. Ambio, Vol. 33(1-2), Feb. 2004.
Russell, S. (2007). Africa’s Potential Wars, BBC online. www.bbc.com
UN (2007). The Millennium Developments Goals Report 2007, New York.
WHO-UNICEF (2006). Meeting the MDG drinking water and sanitation target: the urban and
rural challenge of the decade, WHO press, Geneva.
World Health Organisation / UNICEF. (2000): Global Water Supply and Sanitation.
Assessment 2000 Report. Geneva, World Health Organisation.
World Water Council (2000). World Water Vision, Making water everybody’s business. The
Use of Water Today, The Hague, Netherlands
WSP, (2004). Strengthening Budget Mechanisms for Sanitation in Uganda.
WSSCC (2004). Resource packs on the water and sanitation Millennium development Goals.
Water supply and sanitation collaborative council, Geneva.
WSSP, (2001). Wetland Sector Strategic Plan 2001-2010. Wetlands Inspection Division,
Ministry of Water, Lands & Environment – Uganda.
WWAP (2006). Uganda National Water Development Report, UN report.
Yu, H., Tay, J. & Wilson, F. (1997). A sustainable municipal wastewater treatment process
for tropical and subtropical regions in developing countries. Wat. Sci. Tech., 35(9),
191-198.
In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 309-323 © 2010 Nova Science Publishers, Inc.
Chapter 15
ABSTRACT
Water is a vital aspect of hemodialysis. During the procedure, large volumes of water
are used to prepare dialysate and to clean and reprocess machines. This paper evaluates
the technical and economical feasibility of recycling hemodialysis wastewater for
irrigation uses, such as watering gardens and landscape plantings. Water characteristics,
possible recycling methods, and the production costs of treated water are discussed in
terms of the quality of the generated wastewater. A cost-benefit analysis is also
performed through comparison of intended cost with that of seawater desalination, which
is widely used in irrigation.
INTRODUCTION
Water is essential to all known forms of life, but this resource is under threat [1].
Growing national, regional, and seasonal water scarcities in much of the world pose severe
challenges for national governments, international development, and environmental policies
[2-5]. In this context, alternative water sources such as wastewater recycling offer a partial
solution by creating fresh water for industry and agriculture [6].
308 Faissal Tarrass, Meryem Benjelloun, and Omar Benjelloun
Hemodialysis, a method for replacing renal function in patients suffering from renal
failure by the removal of excess water and wastes, requires a large volume of water. Water is
used in long-term dialysis facilities to prepare dialysate and to rinse and reprocess dialysis
membranes and machines [7,8]. Assessing the recycling potential of hemodialysis wastewater
must give consideration to both the environmental and economic aspects. This study was
undertaken to analyze hemodialysis wastewater at a Moroccan dialysis facility in order to
examine the feasibility of advanced wastewater treatment for agricultural uses such as
watering gardens and landscape plantings.
IMPACTS OF RECYCLING
Hemodialysis wastewater can enter municipal and natural water systems via residential or
commercial discharges, including hospital effluent. There is a lack of data about the possible
direct and indirect impact of hemodialysis wastewater discharges on the environment.
However, benefits of recycling may result in reduced discharge of wastewater into natural
water bodies and the potential water savings for hospitals.
In terms of economic impacts, it is known that integrating wastewater treatment in
agriculture can bring benefits such as partial cost recovery [1,2]. For hospitals, recycling
wastewater may provide a purchase price reduction [11]. In this study, we analyze the
potential economic benefit of using recycled hemodialysis wastewater for irrigating the
hospital grounds.
Recycling Wastewater After Hemodialysis: … 309
METHODOLOGY
Wastewater Sampling
Water samples were obtained from a single dialysis facility. Using sterile 500-mL bottles,
wastewater was collected from the outflow pipe that drains all hemodialysis sewage
(including waste dialysate and water rejected during treatment by the carbon filters and
reverse osmosis membranes) directly into the municipal sewage line. All samples were placed
in a closed cooler during transit to the laboratory.
Samples for bacteriological testing were processed within 1 to 2 hours. Samples were
cultured using the membrane filtration technique. In brief, membrane filters were placed
aseptically on trypticase soy agar and incubated at 36ºC for 48 hours [12,13]. Total viable
colony counts were documented and isolates identified using standard microbial techniques.
The suitibility of hemodialysis wastewater use in agriculture was evaluated through the
comparison of its characteristics with the Food and Agriculture Organization of the United
Nations (FAO) and the World Health Organization (WHO) standards for wastewater use for
agricultural applications [14,15]. The optimum procedure for treatment to reach standards is
discussed according to the quality of the generated wastewater.
desired effluent quality), and land requirements. The cost report produced by the two
programs includes equipment, operation and maintenance cost (materials and supplies,
energy, and labor).
A cost benefits analysis was also performed in which the cost of water treatment was
compared with that of seawater desalination to produce water of equivalent quality. Data for
cost estimates of water desalination for agricultural applications were based on the
proceedings of the FAO expert consultation on water desalination for agriculture [17].
RESULTS
Hemodialysis Wastewater Composition and Pollution Risk
These results show that apart from an increased but expected conductivity value,
biochemical oxygen demand, Kjedahl nitrogen, chloride sulfate, and phosphorus
concentrations did not exceed the FAO standards, with the exception of the conductivity
value. Bacterial count of the wastewater showed 450 colony-forming units/mL, but coliform
organisms (specifically Escherichia coli species) were undetectable.
The primary challenge for use of hemodialysis wastewater can be its high conductivity.
By contrast, concentrations of dissolved organic substances were under applicable emission
standards for discharges [18,19], and the bacterial count was under WHO standards for
wastewater use in agriculture [14].
Table 2. Technical characteristics of the nanofiltration and reverse osmosis systems for
wastewater recycling
Cost Benefits
Total costs for the two treatment techniques, including capital equipment, operating, and
maintenance costs, are presented in Table 3. Costs were calculated based on current
membrane and equipment prices. Energy, labor, and maintenance were calculated based on
the average Moroccan prices for energy and labor. Membrane life was set at 3 and 4 years for
nanofiltration and reverse osmosis respectively.
312 Faissal Tarrass, Meryem Benjelloun, and Omar Benjelloun
The cost for treating hemodialysis wastewater to achieve a quality that is suitable for
irrigation using nanofiltration and reverse osmosis is 0.70 US$/m3 and 0.74 US$/m3,
respectively.
Table 3. Estimation of costs for wastewater recycling calculated on 288 working days
per year, and 20 working hours per day
For comparaison, the costs associated with various techniques for desalination of
seawater for agricultural use are listed in table 4. With the exception of the latest reserve
osmosis technologies, costs are in excess of 1.15 US$/m3
Given the average cost of 1 US$/m3 for seawater desalination [17], this could result in
cost savings (or benefit) of 20 to 30% in comparison to desalination of seawater (Table 4).
COMMENTS
Arid and semi-arid regions are facing increasingly more serious water shortage problems.
As the population grows in these areas, water is an increasingly valuable and limited resource.
Every effort must be made to use water more efficiently, and new practices are being
developed and implemented in the field of water use and water conservation [27,28].
Hemodialysis represents an environmental challenge, in part due to high water
consumption [8]. In regions with water scarcity, high consumption of water during by
hemodialysis units is a compelling argument supporting wastewater recycling. This paper
discusses the technical and economic feasibility of recycling this type of wastewater for
potential use in irrigation.
Observed values show that organic matters and bacterial biomass were under the
acceptable limits; however, conductivity values exceeded FAO standards [15]. Due to this
high conductivity, wastewater must be treated to accepted standards prior to use for irrigation
[22,23]. Membrane separation has proven to be the preferred treatment process for such high
conductivity wastewaters [22,23,29]. Moreover, this technology has previously been shown
to be efficient and economical in comparison with other approaches [24,25,30,31].
In this study, computer simulations were executed with two models of membrane
treatment (nanofiltration and reverse osmosis) based on the characteristics of the influent and
the desired effluent wastewater quality and assumptions obtained from prior literature
[30,32]. The simulations suggested that both methods showed greater benefit compared to
desalination of seawater, resulting in a cost savings (or benefit) of 20 - 30%.
Membrane separation is a widely used process for the treatment of various types of
wastewater. However, to our knowledge, application of this technology to hemodialysis
wastewater has not been performed. Also, in reviewing the literature, we were unable to
document any engineering application system related to hemodialysis wastewater treatment.
Consequently the result of our analysis calls for further investigations in this area.
In conclusion, due to the high water consumption in hemodialysis, it is essential to study
its potential for recycling. Through analysis and evaluation of the technical and economic
feasibility of hemodialysis wastewater treatment, this study draws attention to this important
but neglected aspect of hemodialysis therapy.
REFERENCES
[1] Rosegrant MW, Cai X, Cline SA. World Water and Food to 2025: Dealing With
Scarcity. Washington DC, International Food Policy Research Institute, 2002
[2] Berrittella M, Hoekstra AY, Rehdanz K, Roson R, Tol RS. The economic impact of
restricted water supply: a computable general equilibrium analysis. Water Res. 2007;
41: 1799-813.
[3] Moe CL, Rheingans RD. Global challenges in water, sanitation and health. J Water
Health. 2006; 4 Suppl 1: S41-S57.
[4] Tal A. Seeking sustainability: Israel's evolving water management strategy. Science.
2006; 313: 1081-4.
314 Faissal Tarrass, Meryem Benjelloun, and Omar Benjelloun
[23] Gerhart VJ, Kane R, Glenn EP. Recycling industrial saline wastewater for landscape
irrigation in a desert urban area. J Arid Environ 2006; 67: 473-86.
[24] Moch I, Chapman M, Steward D. Estimating membrane water treatment costs Membr
Technol 2003; 8: 5-7.
[25] Cote P, Masini M, Mourato D. Comparison of membrane options for water reuse and
reclamation. Desalination 2004; 167: 1-11.
[26] Office National de l‘Electricité: Customer space. Available at: www.one.org.ma.
Accessed 28th February 2008.
[27] Bakir HA. Sustainable wastewater management for small communities in the Middle
East and North Africa. J Envir Manag 2001; 61: 319-28.
[28] Abu-Zeid KM. Recent trends and developments: reuse of wastewater in agriculture.
Envir Manag Health 1998; 2: 79-89.
[29] Marcucci M, Ciabatti I, Matteucci A, Vernaglione G. Membrane technologies applied
to textile wastewater treatment. Ann NY Acad Sci 2003; 984: 53-64.
[30] Noronha M, Mavrov V, Chmiel H. Simulation model for optimisation of two-stage
membrane filtration plants; minimising the specific costs of power consumption. J
Membr Sci 2002; 202: 217-32.
[31] Hafez A, Khedr M, Gadallah H. Wastewater treatment and water reuse of food
processing industries. Part II: Techno-economic study of a membrane separation
technique. Desalination 2007; 214: 261-72.
[32] Moch I, Chapman M, Steward D. Development of a CD-ROM cost program for water
treatment projects. Membr Technol 2003; 6: 5-8.
In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 319-337 © 2010 Nova Science Publishers, Inc.
Chapter 16
ABSTRACT
Heavy metal pollution is a serious problem in many developed and developing
countries. Lead had been recognized as a particularly toxic metal and comes into water
bodies mainly from metallurgical, battery, metal plating, mining and alloy industries. In
order to minimize the impacts of this metal on human health, animals and the
environment, lead-contaminated water and wastewater need to be treated before
discharge to water bodies.
This chapter concerns an investigation of potential usage of corn-processing
wastewater as a new alternative low-cost substrate to produce biosorbent and evaluate
this biosorbent to remove Pb(II) ions from aqueous solutions. For this aim, Rhizopus
oligosporus cultivated on corn-processing wastewater and dried biomass of these fungi
was used as an adsorbent. The adsorption experiments were conducted in a batch process
and the effects of contact time (1-48 hours), initial pH (2-7), initial metal ion
concentration (20-100 mg L-1) and adsorbent dosage (0.5-5 g L-1) on the adsorption were
investigated. Pb (II) ion concentrations before and after adsorption were measured using
Inductively Coupled Plasma-Mass Spectrometry. Maximum adsorption capacity was
achieved at pH 5.0. The isothermal data of dried fungal biomass could be described well
by the Langmuir equation and monolayer capacity had a mean value of 59.88 mg g-1. The
*
Corresponding author: Phone: +90 324 361 00 01/7102, Fax: +90 324 361 00 99, E-mail: ozsoyhd@gmail.com
(H.D. Ozsoy)
318 H. Duygu Ozsoy and J. Hans van Leeuwen
pseudo-second order reaction model provided the best description of the data with a
correlation coefficient 0.99 for different initial metal concentrations. This result indicates
that chemical sorption might be the basic mechanism for this adsorption process and
Fourier Transform Infrared Spectroscopy analyses showed that amide I and hydroxyl
groups play an important role in binding Pb (II).
Because of the high activation capacity of adsorbent and low cost of process dried R.
oligosporus biomass presents a good potential as an alternative material for removal of
Pb (II) ions from the aqueous solutions.
1. INTRODUCTION
Removal of heavy metals from aqueous solutions is one of the major problems in
industrial wastewater treatment because most of them are toxic even at very low
concentrations. The amount of these pollutants in water has been increased with industrial
applications including mining, refining, electroplating and production of textiles, paints and
dyes [1].
Lead had been recognized as a particularly toxic hazardous environmental pollutant and
comes into water bodies mainly from metallurgical, battery, metal plating, lead smelting,
mining and alloy industries. In order to minimize the impacts of this metal on human health,
animals and the environment, lead-contaminated water and wastewater need to be treated
before discharge to water bodies [2].
The conventional methods for removing metals from aqueous solutions include chemical
precipitation, chemical oxidation or reduction, electrochemical treatment, reverse osmosis,
solvent extraction, ion exchange and evaporation. However, these techniques have several
disadvantages such as high chemical cost, low removal efficiency, low selectivity, high
energy requirements, and generation of secondary toxic slurries [3-5]. Therefore removal of
toxic heavy metals in a cost-effective and environment-friendly manner assumes great importance.
Adsorption is a highly effective and economical technique for removal of heavy metals
from aqueous solutions. Commercial activated carbon is a well-known and highly effective
adsorbent, but the high cost of activated carbon limits its use as an adsorbent especially in
developing countries [6]. From this standpoint, numerous investigations were conducted by
scientists in this growing and important field of research for the exploration of alternative
methods using less expensive natural materials [7].
Metal-sorption by various types of biomaterials like metabolically inactive dried biomass
of algae, bacteria and fungi can find useful application for removing metals from solution
because of their unique chemical composition [8-11]. Research indicated that biosorption is a
very effective method to remove metals from the water and wastewater. Cultivation of
microorganisms requires a bioreactor and nutrients such as carbon, nitrogen and trace
elements [12,13]. Therefore, cultivation cost is the most important factor to produce these
biosorbents.
This chapter presents experimental results on removal of Pb (II) ions from aqueous
solution by dried fungal biomass produced from corn-processing wastewater. These results
Pb (II) Ions Removal by Dried Rhizopus Oligosporus… 319
show that it is possible to use food-processing wastewater as a substrate for cultivating the
fungal biomass to reduce operational costs of adsorption processes. Using the wastewaters
would be particularly attractive and cost effective because there are many food-processing
plants in USA and many other countries that could provide suitable industrial wastewater for
cultivating the microbial biomass, such as fungi. The wastewater needs to be treated to
address discharge objectives and therefore, the biomass is produces at minimal cost. The goal
of the method was to investigate the efficacy of dried fungal biomass as adsorbent for
removal of Pb (II) ions from aqueous solution and reduce the treatment costs using another
wastewater as a substrate for fungi.
2. EXPERIMENTAL
The experimental program comprised two phases.
Rhizopus oligosporus was obtained from American Type Culture Collection (Rockville,
MD). The culture was rehydrated and revived in yeast-malt (YM) nutrient broth at 24 ºC. The
revived culture was transferred on to numerous potato dextrose agar (PDA) plates and
incubated at room temperature (24 ºC) for 7 days. Then fungal sporangiospores were
harvested from the surface of PDA plates into sterile distilled water containing 0.85 % (w/v)
saline solution (NaCl) and 0.5 % (v/v) of Tween 80. The harvested cultures were diluted
further to achieve a spore count of 106 to 107 spores/mL, determined by haemocytometer
counts. Glycerin (20 %; v/v) was added to the spore suspension as a cryoprotectant for ultra-
low frozen storage at –75 oC in 2 mL cryo-vials for future use as a bioreactor inoculum.
The inocula were used as a seed in laboratory-scale continuous attached growth tank
reactors using corn-processing wastewater as organic substrate. The wastewater was supplied
from the ADM wet corn milling facility in Cedar Rapids, IA, US. The reactors were operated
at a hydraulic retention time (HRT) of 8 h and solids retention time (SRT) of 2 days. These
HRT and SRT values were found to be optimal for the maximum growth of the Rhizopus
oligosporus [14]. The micro-fungi were growing in the form of attached mycelia and
harvested daily from the bioreactor by natural sloughing off the attachment surface and
subsequent gravity settling. The mycelia were washed with deionised water and dried at 65 ºC
for 24 h. The dried fungal pellets were ground and sieved (0.5 mm< diameter).
The effects of contact time, initial pH, initial metal ion concentration and temperature on
adsorption efficiency were examined through a series of shaker flask tests. After determining
320 H. Duygu Ozsoy and J. Hans van Leeuwen
the optimum conditions, a series of adsorption tests were conducted to determine the isotherm
for Pb (II) ions.
2.3. Chemicals
A 1 g L-1 stock solution of lead was prepared with single reagent grade metal solution
(Claritas, Fisher Chemicals) in deionized water. The metal solution was diluted to appropriate
concentrations as needed and stored at 4 oC until further use. HNO3 and NaOH were obtained
from Fisher Chemicals and used for pH value adjustment.
All sorption tests were conducted using single reagent grade metal to minimize the
variability of metal concentrations and to avoid competitive adsorption of mixed metals on
adsorbent. 100 mL of metal solution was added to each of flask containing 0.1 g (dry weight)
of R. oligosporus. The flasks were placed on an orbital shaker table running at 150 rpm at
30±1 C (except the temperature experiments) until equilibrium was reached. The residual
concentration of Pb (II) ions in the aqueous phase (obtained by centrifugation, 1000 g-10
min) was determined using Inductively Coupled Plasma-Mass Spectrometry (ICP-MS). All
tests were conducted in triplicate.
The concentrations of the Pb (II) ions the in aqueous phase were used to determine the
adsorption capacity of R. oligosporus. Equilibrium sorption isotherms were determined by
mass balance.
The amount of adsorbed Pb (II) ions at equilibrium, qeq (mg g-1) was calculated as follows:
[(Co Ceq )V ]
qeq = (1)
x
where Co and Ceq are the initial and equilibrium concentrations of Pb (II) ions (mg L-1), V
volume of solution and x the weight of sorbent (g).
Measurements were performed with a Hewlett Packard 4500 Series ICP-MS using
external calibration. The instrument was calibrated before each measurement. Operating
parameters are summarized in Table 1.
The Langmuir isotherm was used first to describe observed sorption phenomena. The
Langmuir isotherm applies to adsorption on a completely homogenous surface with negligible
interaction between adsorbed molecules [15,16]. The linear form of the equation can be
written as;
Pb (II) Ions Removal by Dried Rhizopus Oligosporus… 321
Rf power 1200 W
Rf matching 2V
Sample depth 7.8 mm
Plasma gas flow 16 L min-1
Auxiliary gas flow 1.4 L min-
1
Ceq 1 Ceq
= + (2)
qeq bqmax qmax
where Ceq is the equilibrium concentration of Pb (II) ions, qeq is the amount of adsorption at
equilibrium, qmax is the mono-layer capacity, and b is an equilibrium constant of Langmuir.
The Freundlich isotherm (empirical model adsorption in aqueous systems) was also
tested with our experimental data. The linear form of the equation can be written as:
1
lnqeq= lnKf + lnCeq (3)
n
where Kf is the measure of sorption capacity, 1/n is the sorption intensity.
Pseudo-first order and pseudo-second order kinetic models were applied to data to
analyse the sorption kinetics of Pb (II) ions. A simple pseudo first-order equation due to
Lagergren was used by different researchers [17,18]:
kad
log (qeq-qt) = log qeq- (4)
2.303t
where qe and qt are the amount of adsorption at equilibrium and at time t respectively, and kad
is the rate constant of the pseudo first-order adsorption process. A plot of log (qeq-qt) vs. t
would provide a straight line for first-order adsorption kinetics, allowing computation of the
adsorption rate constant, kad.
Ho‘s second-order rate equation, which has been called a pseudo-second order kinetic
expression, has also been applied widely [19,20].
The linear form of the kinetic rate equations can be written as follows:
t 1 1 (5)
t
qt kqeq2 qe q
322 H. Duygu Ozsoy and J. Hans van Leeuwen
where k is the rate constant of sorption (dm3 mg-1 min-1), qe is the amount of metal ion sorbed
at equilibrium (mg g-1), and qt is the amount of metal ion sorbed at time t (mgg-1). The
constants can be determined experimentally by plotting t/qt against t.
2.7. FT-IR Analyses
Fourier Transform Infrared Spectroscopy (FT-IR) analysis in the solid phase was
performed using a Fourier Transform Infrared Spectrometer (Varian 2000 FT-IR). Pure
biosorbent powders were used and spectra of the fungal biomass before and after Pb(II)
sorption were recorded.
3. RESULTS
3.1. The Effect of the Contact Time
The major fraction of Pb (II) ions, adsorbed within the first 6h and dissolve Pb remained
nearly constant afterwards. This suggested that the biosorption process reached saturation
within 6h. For this reason a 6h contact time was used for the further experiments. Figure 1
shows the effect of contact time on adsorption of Pb(II) ions onto the dried R. oligosporus
biomass.
Results of the experiments using 100 mg L-1 Pb (II) solutions and 1 g L-1 adsorbent
showed that efficiencies of adsorption were increased with increasing pH from 2.0 to 6.0
(Figure 2). At the low pH ranges, a high concentration of protons in the solution may have
competed with metal ions in forming a bond with the active sites on the surface of the fungi.
These bonded active sites thereafter became saturated and were unavailable to other cations.
60.00
50.00
40.00
q (mg/g)
30.00
20.00
10.00
0.00
0 5 10 15 20 25 30 35 40 45 50 55
t (hour)
Figure 1. The effects of contact time on adsorption of Pb(II) ions (100 mg l-1) to the dried R.oligosporus
biomass (adsorbent dosage: 1g l-1; pH:5.0; temperature: 30 C).
Pb (II) Ions Removal by Dried Rhizopus Oligosporus… 323
60.00
50.00
40.00 pH2
q (mg/g)
pH3
30.00
pH4
20.00 pH5
10.00 pH6
0.00
0 1 2 3 4 5 6 7
t (hour)
Figure 2. The effects of the initial pH on adsorption of Pb(II) ions (100 mg l-1) to the dried
R.oligosporus biomass (adsorbent dosage:1g l-1; contact time: 6h; temperature: 30 C)
60
50
40
q(mg/g)
30
20
10
0
0 20 40 60 80 100 120 140 160
Co(mg/L)
Figure 3. The effects of initial metal concentration on adsorption of Pb(II) ions to the dead
R.oligosporus biomass (adsorbent dosage:1g l-1; contact time: 6h; pH:5.0).
The adsorption of Pb (II) by the dried R. oligosporus biomass was studied at different Pb
(II) ion concentrations in the range from 20 to 150 mg L-1. Equilibrium sorption capacity of
the dried R. oligosporus biomass increased with increasing initial Pb (II) ion concentrations
(Figure 3). The initial concentration provides an important driving force to overcome all mass
transfer resistance of Pb (II) ions between the aqueous and solid phases. Hence a higher initial
concentration of Pb (II) ions may increase the adsorption capacity.
Experimental results indicated that the efficiency of biosorption was decreased with
increasing adsorbent dosage ranging from 0.5-5 g L-1 (Figure 4). This However the
percentage of the Pb(II) ions biosorption was increased with increasing biomass because of
higher surface area. This can be explained by concentration gradient between the solute
324 H. Duygu Ozsoy and J. Hans van Leeuwen
concentration in the solution and the one in the surface of the adsorbent. When the adsorbent
dosage is higher, there is a very fast adsorption onto the adsorbent surface, which results in a
lower Pb(II) ion concentration in the solution. However, the adsorption sites on the adsorbent
surface remain unsaturated when the Pb(II) ion concentration in the solution drops to a lower
value. Thus, the amount of Pb(II) ions adsorbed onto per unit weight of adsorbent gets
reduced with the adsorbent dosage increasing [25].
Two equilibrium models were employed: The Langmuir and Freundlich isotherm
equations. The correlation coefficient of Freundlich isotherm (R2) was 0.8824 (Figure 5). The
Langmuir model was the best-fit isotherm for adsorption of Pb (II) to the dried R. oligosporus
biomass. Langmuir isotherm model parameters, qmax and b, were estimated from the intercept
and slope of Ceq/qeq vs. Ceq , according to Eq. (2) and obtained as 59.88 (mg g-1) and 0.042
(L g-1), respectively. The correlation coefficient of the Langmuir isotherm (R2) was 0.9820
(Figure 6).
90
80
70
60 0.5
q (mg/g)
50 1
40 3
30 5
20
10
0
0 1 2 3 4 5 6 7
t (hour)
Figure 4. The effects of adsorbent dosage on adsorption of Pb (II) ions to dried R. oligosporus biomass
(contact time: 6h; pH:5.0; temperature: 30 C)
4.5
4
3.5
3
2.5 y = 0.4115x + 2.0629
lnqe
2 R2 = 0.8824
1.5
1
0.5
0
2 2.5 3 3.5 4 4.5 5
lnCe
Figure 5. The Freundlich adsorption isotherm for Pb(II) adsorption on to dried R. oligosporus biomass
(adsorbent dosage:1g l-1; pH: 5.0; initial metal concentration of 20-100 mg l-1; temperature: 30 C).
Pb (II) Ions Removal by Dried Rhizopus Oligosporus… 325
2.500
y = 0.0167x + 0.398
2.000 R2 = 0.982
1.500
Ce/qe
1.000
0.500
0.000
0 20 40 60 80 100 120
Ce (mg/L)
Figure 6. The Langmuir adsorption isotherm for Pb(II) adsorption on to dried R. oligosporus biomass
(adsorbent dosage:1g l-1; pH 5.0; initial metal concentration of 20-100 mg l-1; temperature: 30 C).
20 mg/L
25.00 40 mg/L
60 mg/L
20.00 R2 = 0.995
80 mg/L
R2 = 0.9982
100 mg/L
15.00
t/qt
R2 = 0.9929
10.00 R2 = 0.9924
5.00 R2 = 0.9981
0.00
0 100 200 300 400
t (min)
Figure 7. The plots of pseudo-second order kinetics with respect to different initial Pb(II) ion
concentrations (adsorbent dosage:1g l-1; pH 5.0; temperature:30 C).
Kinetic studies were carried out for biosorption of Pb (II) as a function of contact time at
various initial Pb (II) concentrations ranging from 20-100 mg L-1. Experimental results
indicated that the pseudo-second order reaction model provided the best description of the
data with a correlation coefficient 0.99 for different initial metal concentrations (Figure 7).
Figure 9. FTIR spectra of the dried R. oligosporus biomass after Pb(II) ions adsorption.
Reaction rate constants for pseudo-second order equations are shown at Table 2. The
results indicated that the adsorption system studied follows a pseudo-second order kinetic
model at all time intervals and pseudo-second order rate constants were affected by initial Pb
(II) ions concentration.
Pb (II) Ions Removal by Dried Rhizopus Oligosporus… 327
4. CONCLUSION
To development of an efficient and cost-effective removal process, fungal biomass
produced from food industry wastewater is a good alternative biosorbent. Experimental
results showed that based on the Langmuir coefficients, the total capacity (monolayer
saturation at equilibrium) of the dried R. oligosporus biomass for Pb (II) ions was about 60
mg g-1 (biosorbent dose of 1g/L, 6h contact time, initial Pb (II) concentration of 100 mg/L and
optimum pH of 5.0). Experimental results also indicated that the pseudo-second order
reaction model provided the best description of the data with a correlation coefficient 0.99 for
different initial metal concentrations. The fit of the experimental data to this model suggest
that the process controlling the rate may be chemical sorption. The FTIR analyses showed
that amide I and hydroxyl groups plays an important role in binding of Pb(II).
With the advantage of high metal biosorption capacity, R.oligosporus biomass, produced
from corn-processing wastewater, has the potential to be used as an effective and economic
biosorbent material for the removal of Pb (II) ions from wastewater streams.
ACKNOWLEDGMENTS
Authors are thankful to Dr. Basudeb Saha for technical assistance (ICP-MS
measurements) and Iowa State University for their financial support.
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In: Fluid Waste Disposal ISBN: 978-1-60741-915-0
Editor: Kay W. Canton, pp. 333-353 © 2010 Nova Science Publishers, Inc.
Chapter 17
ABSTRACT
The main objective of this research work is to determine the presence of di(2-
ethylhexyl) phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate, in different
water samples (drinking waters, effluents and surface waters). Different analytical
methods were studied in order to know the best methodology for the quantification of
these compounds. Solid-liquid and liquid-liquid extraction were investigated and finally
the liquid-liquid extraction and analysis by gas chromatography followed by mass
spectroscopy was chosen because of offering the highest recovery rate. In the whole of
this research study, the control of background pollution by reagents and material was
extremely important. The problem of background pollution is more serious in the trace
analysis of phthalates and adipates as a consequence of their presence in almost all
equipment and reagents used in the laboratory.
Respect to the control of the selected plasticizers in the different water samples, bis
(2-ethylhexyl) phthalate and bis (2-ethylhexyl) adipate were detected in drinking water,
effluents and surface waters. On the other hand, diisodecyl phthalate was not detected in
any sample.
INTRODUCTION
Phthalates have been in use for almost 40 years and are used in the manufacture of PVC
and other resins, as well as plasticizers and insect repellents [1]. Plasticizers are used in
building materials, home furnishing, transportation, clothing, and, to a limited extent, in food
packaging and medicinal products [2]. There is also concern regarding the potential health
332 Rosa Mosteo, Judith Sarasa, Mª Peña Ormad et al.
effects of several phthalates because these compounds are used to impart softness and
flexibility to normally rigid PVC products in children‘s toys. Phthalates can enter the
environment through losses during manufacture and by leaching from the final product. This
is because they are not chemically bonded to the polymeric matrix [1]. These compounds
have low water solubilities and tend to adsorb to sediments and suspended solids. Therefore,
they have the potential to leach into their surrounding environment. Certain phthalates have
also shown estrogenic behaviour [3, 4] so they can be classified as potential endocrine disruptors.
Many endocrine disrupting substances, or potential endocrine disrupters (EDCs), were
previously classified as organic micropollutants. Therefore, many substances, such as
foodstuffs, flavonoids, lignans, sterols, fungal metabolites and synthetic chemicals of widely
varying structural classes (e.g. phthalates, PCBs), can interact with hormone receptors and
modulate the endocrine system [5]. Phthalates are classified as EDCs in several lists of
compounds by various organizations [6], for example:
The most of Phthalates and adipates used as plasticizers are being included in list of
priority contaminants of different countries. In the United Stated, the Environmental
Protection Agency established concentration limits in drinking waters for both di(2-
ethylkhexyl)phthalate (6 µg/l) and di(2-ethylhexyl)adipate (0.4 mg/l) [10]. In the European
Union there are not contamination limits for phthalates and adipates, although it is known that
these limits will be established in the near future. The di(2-ethylkhexyl)phthalate is included
in the priority list [11] and identified as priority hazardous substance in the field of water
policy. The guideline value in drinking water proposed by the World Health Organization in
the 1993 Guidelines [12, 14] for the Di(2-ethylhexyl)adipate is 80 µg/l and in the case of
Di(2-ethylhexyl)phthalate the guideline value is 8 µg/l.
Di(2-ethylhexyl)adipate is also known as DEHA. This compound is mainly used as a
plasticizer for synthetic resins such as PVC, but significant amounts are also used as
lubricants and for hydraulic fluids [13]. Reports of the presence of DEHA in surface water
and drinking water are scarce, but DEHA has occasionally been identified in drinking-water
at levels of a few micrograms per litre. As a consequence of its use in PVC films, food is the
most important source of human exposure (up to 20 mg/day).
Di(2-ethylhexyl)phthalate is also known as DEHP. This compound is primarily used as a
plasticizer in many flexible polyvinyl chloride products and in vinyl chloride co-polymer
resins. It has also application as replacement for polychlorinated biphenyls in dielectric fluids
for small (low-voltage) electrical capacitors [13]. This compound has been found in surface
water, groundwater and drinking-water in concentrations of a few micrograms per litre. In
polluted surface water and groundwater, concentrations of hundreds of micrograms per litre
have been reported. Numerous manufactures are trying to substitute the DEHP for the
diisodecyl phthalate, which is a plasticizer less toxic than di(2-ethylhexyl) phthalate (DIDP).
As a consequence of the necessary control of these pollutants in different waters,
analytical methodologies should be established in order to improve the quantification of these
compounds. Methodologies commonly used to analyze organic compounds in trace levels are
based on liquid-liquid or solid-liquid extraction following chromatographic analysis. In this
Control of Plasticizers in Drinking Water, Effluents and Surface Waters 333
study, two methodologies have been developed in order to find the best one for the analysis of
the interest compounds. Both of them are modifications of the 506 and 606 EPA methods
[15, 16].
In this research work the presence of bis (2-ethylhexyl) phthalate, bis (2-ethylhexyl)
adipate and diisodecyl phthalate, plasticizers frequently used by manufacturers, has been
evaluated in different water samples (drinking waters, effluents and surface waters). Different
analytical methodologies, using solid-liquid extraction or liquid-liquid extraction prior to
analysis by GC/MS, were studied in order to know the best methodology for the
quantification of these compounds. A study about the background pollution is carried out
since they are present in the majority of equipment and reagents used in the laboratory.
EXPERIMENTAL PROCEDURE
Analytical Methodology for the Control of DEHP, DIDP and DEHA
Standard of DEHP, DIDP and DEHA purchased from Dr. Ehrenstorfer were used. A
stock solution for each compound is prepared in methanol. In the case of the DEHP, the stock
solution presented concentration of 1000 mg/l, the DIDP 1520 mg/l and the DEHA 1000
mg/l. The chemical structure of DEHP, DIDP and DEHA are shown in Figure 1.
The chromatographic conditions are shown in Table 1.
DEH DEH
P A
DIDP
506 and 606 U.S EPA methods 506 U.S EPA method
Sample (4ºC)
Sample (4ºC)
- Cartridge preparation
Liquid-liquid extraction Solid-liquid (CH2Cl2/MeOH/H2O)
3 extactions with CH2Cl2 extraction - Compound adsorption
506 method:+1extaction C6H6 C18 cartridge - Cartridge dried with air
GC/MS analysis
GC/MS analysis
Figure 2. Diagrams of liquid-liquid extraction (506 and 606 EPA methods) and solid-liquid extraction
(506 EPA method).
The extraction methods used in this study are schematized in the Figure 2. The liquid-
liquid extraction has been carried out taking into account both methods 506 EPA method (use
of NaCl during the process) and 606 EPA method (without NaCl). The solid-liquid extraction
is related to 506 EPA method.
RESULTS
Analysis by GC/MS of Standards
In Figures 3, 4 and 5 are reflected the chromatograms and spectrums for DEHP, DEHA
and DIDP standards, respectively.
Control of Plasticizers in Drinking Water, Effluents and Surface Waters 335
26.07
23.01
A summary of the retention time and the characteristic mass used for the identification of
DEHP, DEHA and DIDP is shown in Table 2.
336 Rosa Mosteo, Judith Sarasa, Mª Peña Ormad et al.
149
85
80 39.64 42.33
75 40.73
43.43
70
65 43.74
44.09
60 45.19
Relative Abundance
55 38.94
37.88
50
46.17
45
37.25 47.45 47.95
40 36.92
35.65 48.62 49.81
31.18 55.63
35 33.94 35.06 50.62 51.47
52.81 54.50
30
25
20
15
10
0
30 32 34 36 38 40 42 44 46 48 50 52 54 56
36 Time (min)
48
(a) Considering the most representative isomers. The quantification is based on the
retention times of these isomers (39.60 min and 41.45 min) and their characteristic
masses (m/z=149).
(b) Considering the area of all the isomers peaks as a whole. The quantification is based
on the manual integration of the total area and the characteristic mass (m/z=149).
For this reason, three calibration curves were established for the DIDP:
Deuterated Anthracene (AD10) (retention time= 16.53 min, m/z= 188) was used as an
internal standard for the quantification of target compounds by GC/MS analysis.
The calibration curves, the linear ranges and the quantification limits obtained for DEHP,
DEHA and DIDP are shown in Table 3.
The background pollution was evaluated in the different stpes involved in the analytical
method used to quantify the compounds of interest (DEHA, DEHP and DIDP). This problem
of background contamination has been more serious in the trace analysis of plasticizers than
in the studies of many other pollutants because these compounds are present in almost all
equipments and reagents used in the laboratory. Different commercial reagents were
evaluated (Carlo Erba, Merck…) with the aim of selecting the commercial brand most
appropriate for carrying out these analyses. The obtained results are shown in Table 4.
4
HPLC quality5 Ultrapure Milli Q water
6
Mineral Water A (bottled in glass)
7
Mineral Water B (bottled in glass)
8
Mineral Water C (bottled in carton)
Modified solid-liquid method
Modified liquid-liquid method
Sample (4ºC)
Sample (4ºC)
- Cartridge preparation
Liquid-liquid extraction
Solid-liquid CH2Cl2/MeOH/H2O df
3 extractions with CH2Cl2 fffffff(Mineral water A)
+1 extraction C6H6 extraction
C18 cartridge - Compound adsorption
-Drying cartridge with N2
Concentration with N2
Concentration with N2
(volume 1 ml)
(volume 1 ml)
The following analytical methodologies have been compared in order to know which one
is the most appropriate for the analysis of DEHP, DEHA and DIDP:
The synthetic samples were prepared by diluting stock solution in mineral water A,
obtaining concentrations of 30 µg l-1 for DEHA, 60 µg l-1 for DEHP and 1.2 mg l-1 for DIDP.
The recovery results obtained for each analytical method are shown in Figure 7.
As it can be observed in Figure 7, the recoveries obtained with the solid-liquid extraction
were very low, probably due to polarity of the adsorbent. As a consequence, this method was
rejected. In the case of liquid-liquid methodologies the recoveries were higher. It is also
noticed that the NaCl addition, used to saturate water and to help the movement of the
compounds to organic phase, improved the liquid-liquid method. The DEHA and DEHP
recoveries were twice and four times higher respectively whereas DIDP recovery increased
slightly. Therefore the liquid-liquid extraction with NaCl addition was selected for the
analysis of the target pollutants.
Figure 7. % Recoveries obtained for DEHP, DEHA and DIDP using different methods of extraction.
340 Rosa Mosteo, Judith Sarasa, Mª Peña Ormad et al.
The main features of the studied mineral waters, taking into account that A, C, D,…are
the brands, are the following:
The obtained results are reflected in Figure 8. It is observed that the tap water was the
most polluted (11.8 µg l-1 DEHA and 29.2 µg l-1 DEHP). The mineral water C bottled in
carton and the mineral water A bottled in glass presented the less concentration of DEHA (1.1
µg l-1) and DEHP (6.9 µg l-1). Considering all the studied samples, the range of DEHA
concentration was 1.1-3.5 µg l-1 and with respect to DEHP the range was 6.9-23.3 µg l-1. The
DIDP compound was not detected in the analyzed waters.
Figure 8. Control of DEHP, DEHA and DIDP in tap and mineral waters.
Furthermore, from Figure 8 it can be concluded that the concentration of the target
pollutants depends on the material of which the bottle is made of. In order to compare in a
better way the concentrations of the target compounds, in Figure 9 are shown the results for
mineral water A and C, bottled in glass, PET and carton. The main difference between
mineral waters was due to the DEHP concentration, which was lower in glass bottle.
Control of Plasticizers in Drinking Water, Effluents and Surface Waters 341
Figure 10. Concentration of DEHP and DEHA in mineral waters bottled in PET.
In Figure 10, the mineral waters bottled in PET are compared. As it can be observed, the
DEHA concentration was very similar for all the analyzed mineral waters. However, the
DEHP concentration showed a great variation. In fact, the DEHP concentration in mineral
waters A and D was twice the concentration in mineral water C and E.
Two different samples of these wastewaters were analyzed and the results obtained are
shown in Figure 11 and it is observed that the DEHP concentration was always higher than
DEHA concentration. DIDP compound was not detected in any sample.
The chromatograms obtained for analyses of groundwater and wastewater II (sampling b)
are reflected in Figure 12 and 13, respectively.
Control of Plasticizers in Drinking Water, Effluents and Surface Waters 343
CONCLUSION
- The study of background pollution during the determination of di(2-ethylhexyl)
phthalate, di(2-ethylhexyl) adipate and diisodecyl phthalate in aqueous samples is
essential and the use of solvent of adequate quality is important as well. Due to the
presence of these plasticizers in almost all equipments and reagents used in the
laboratory, their control in each stage of the analytical methodology is necessary.
- The liquid-liquid extraction with NaCl addition followed by GC/MS was selected for
the analysis of the target pollutants since the highest recoveries were obtained
(80.5% for DEHA, 119% for DEHP and 139.9 % for DIPD).
- The control of the target pollutants in groundwater, wastewaters and mineral waters
indicate that the presence of DEHP was important in all the samples and its
concentration was higher that the DEHA concentration. DIDP compound was not
detected in any sample.
- The detected concentration of DEHP and DEHA is significantly higher in the case of
tap water than in the case of the analyzed mineral waters. In all the samples, the
DEHA concentration always presented a concentration lower than the limit
established by the EPA (0.4 mg/l) for drinking waters [17].
- Respect to the control of these compounds in groundwater and wastewaters, it was
observed that the DEHP concentration is always higher than the DEHA
concentration. The wastewater generated during the process of polymer
transformation by the additives addition is the most polluted by these plasticizers.
344 Rosa Mosteo, Judith Sarasa, Mª Peña Ormad et al.
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INDEX
188, 191, 192, 249, 250, 254, 255, 259, 260, 261,
A 262, 281, 319, 321, 322, 323, 324, 325, 326, 327,
329, 330, 331
abatement, 42, 108
aerobic, xi, 51, 52, 58, 65, 73, 115, 135, 143, 144,
absorption, 9, 87, 88, 93, 95, 98, 122, 136, 137, 181,
147, 149, 152, 156, 160, 162, 163, 168, 170, 172,
183, 185, 195
185, 186, 191, 192, 238, 240, 267, 268, 271, 272,
acceleration, 119, 252
273, 276
accelerator, 95
aerobic bacteria, 58, 185
acceptor, 53, 56, 57, 59, 137, 140, 162
aerosols, 136
accounting, 280
Africa, 46, 47, 280, 282, 297, 300, 306, 308, 317
accuracy, 112, 123
Ag, 330
acetate, 54, 55, 56, 57, 63, 68, 69, 70, 140, 141, 149,
agar, 57, 59, 182, 311, 321
152, 170, 182, 237
age, 12, 191
acetic acid, 53, 68, 69, 140, 157, 201, 262
agents, xiii, 10, 46, 60, 61, 82, 83, 114, 200, 215,
acetone, 61, 69, 201
216, 217, 219, 221, 222, 225, 226, 227, 231, 238,
acetonitrile, 226
240, 297, 329
acid, 10, 13, 17, 20, 45, 53, 55, 57, 62, 63, 68, 69,
aggregates, 7, 83, 87, 239
75, 76, 78, 79, 84, 87, 136, 140, 147, 149, 150,
aggregation, 5, 7, 240, 243
151, 157, 177, 180, 184, 185, 186, 191, 192, 193,
aging, 5
195, 201, 202, 212, 219, 226, 237, 262, 277, 297,
agricultural, xiii, xiv, xv, 30, 49, 199, 213, 230, 235,
330, 331
236, 237, 238, 240, 242, 244, 245, 246, 279, 291,
acidic, x, 6, 9, 17, 30, 36, 50, 56, 60, 62, 80, 81, 83,
304, 310, 311, 312, 314, 316
85, 88, 106, 136, 141, 181, 182, 206, 236
agricultural crop, xiv, 236, 244
acidification, 67
agriculture, ix, xiii, 51, 52, 65, 230, 235, 246, 288,
acidity, 281
291, 303, 304, 309, 311, 312, 313, 314, 316, 317
ACM, 139
air, xiii, 14, 31, 84, 86, 87, 89, 107, 114, 127, 136,
activated carbon, 82, 108, 320
143, 147, 160, 178, 203, 235, 245, 255, 262, 339
activation, xii, xvi, 113, 158, 197, 320
air quality, 136
activation energy, 158
alcohol, 62, 68, 69, 180, 181, 190, 193, 202, 273
active site, 324
alcohols, 53, 54, 59, 61, 62, 141, 237
acute, 114, 121, 130, 154, 230, 276
aldehydes, 61, 62
adaptation, 64, 75, 158, 233, 246
algae, 122, 144, 154, 320
additives, 2, 22, 108, 114, 344, 346
Algeria, 300
adhesion, 7
aliphatic compounds, 54, 201
adipate, xvi, 333, 334, 335, 338, 346, 347
ALK, 54
adjustment, 17, 84, 236, 322
alkali, 61, 62, 85, 92, 95, 108, 109, 139
administration, 126
alkaline, 6, 17, 30, 50, 62, 98, 106, 109, 137, 139,
adsorption, xiv, xvi, 2, 5, 7, 31, 50, 62, 83, 88, 89,
146, 148, 180, 204, 245
95, 96, 108, 113, 114, 115, 181, 183, 185, 187,
alkalinity, 16, 156, 161, 162, 283, 291, 292
346 Index
carbonates, 30, 52, 63, 114 chloride, xii, 23, 28, 107, 197, 202, 207, 221, 225,
carboxyl, 329, 330 311, 313, 334
carboxylates, 53 chlorination, 154
carboxylic acids, 201 chlorine, 12, 18, 19, 221
carcinogenic, 154, 230 chlorobenzene, 201
carrier, 64, 156, 192, 193, 194 chlorophyll, 144
case study, xi, 46, 111, 113, 124, 127, 130, 276, 278 Cholera, 305
casein, 58, 60 chromatograms, 337, 345
cash crops, 280 chromatography, xvi, 226, 333, 347
catalysis, 67 chromium, 9, 16, 17, 18, 19, 20, 21, 67, 109, 114,
catalyst, xii, 109, 136, 197, 202, 203, 206, 207, 208, 262, 330
211, 223 Chromium, 16, 19, 21, 32, 43, 44, 45, 47
catalytic properties, 287 circulation, 63, 115, 127, 203, 230
catchments, 303, 304 civil engineering, 94
catechol, 54, 79, 202 clams, 144
cathode, 5, 8, 10, 11, 16, 17, 23, 287, 288 classes, 183, 242, 334
cation, 7, 9, 15, 106, 107 classical, 51
cattle, 199 classification, 119, 120, 123, 190
cell, 13, 15, 23, 57, 67, 69, 72, 113, 137, 142, 147, clay, 64, 82, 85, 106, 109, 244
154, 185, 187, 188, 192, 203, 209, 211, 231, 240, cleaning, xv, 10, 60, 61, 148, 201, 202, 237, 280, 296
242, 284, 288, 294, 295, 330 cloning, 241
cell death, 209 CO2, xii, 17, 39, 49, 53, 55, 56, 69, 70, 85, 96, 137,
cell membranes, 113 140, 144, 149, 153, 168, 173, 202, 219, 220
cell metabolism, 240 coagulation, ix, 1, 2, 5, 7, 12, 13, 23, 46, 47, 48, 50,
cell organelles, 188 83, 84, 281, 284, 289, 296, 297
cell surface, 185, 187, 188, 211 coagulation process, 289
cellulose, 184, 190, 193, 194, 281, 329 coal, 18, 107, 136, 280
cement, 82 coastal areas, x, 111, 112, 120, 124
ceramics, 64, 82, 108, 109, 110, 216, 218 coatings, 216
CH3COOH, 56, 68 cobalt, 65
CH4, xi, 70, 111, 113 coconut, 251, 262
channels, 250, 281 co-existence, 70
charged particle, 2 coffee, ix, xv, 2, 11, 25, 36, 37, 41, 279, 280, 281,
cheese, 60, 77 282, 283, 284, 285, 286, 287, 288, 289, 290, 291,
chelating agents, 221, 225 292, 293, 296, 297, 303
chemical approach, 130 coke, 136
chemical bonds, x, 82, 104, 236 collagen, 190
chemical engineering, 15 colloidal particles, 2, 4, 5, 7, 13, 142
chemical industry, 199 colloids, 2, 3, 31
chemical oxidation, xii, 197, 202, 204, 206, 207, Columbia, 151
211, 219, 221, 269, 320 combined effect, 288
chemical properties, 241, 242, 243, 246, 247 combustion, 30, 32, 107, 108, 110, 136, 180
chemical reactions, 106, 114 combustion processes, 136
chemical reactivity, 100 commodity, 250, 280
chemical stability, xiii, 95, 215, 216, 217, 220 communities, xiii, 57, 58, 74, 79, 121, 235, 241, 242,
chemical vapor deposition, 216 243, 245, 246, 250, 261, 303, 304, 306, 317
chemicals, 11, 13, 121, 124, 156, 219, 222, 224, 233, community, xiv, 43, 59, 80, 115, 122, 126, 236, 239,
239, 266, 268, 271, 272, 273, 314, 334, 347 241, 242, 243, 245, 246, 255, 303
chicken, 272, 273 competition, 70, 140, 141, 301
chickens, 277 competitive advantage, 145
China, 46, 81, 84, 86, 87, 316 complement, 122, 124, 126
chitin, 329 complexity, ix, 1, 14, 16, 122, 123
compliance, 266
Index 349
components, x, 50, 59, 70, 81, 83, 85, 87, 88, 95, corn, xvi, 229, 276, 319, 320, 321, 330, 331
104, 106, 122, 180, 188, 194, 231, 238, 269, 271 corona, 224
composites, 108 correlation, xvi, 242, 320, 326, 328, 330
composition, xiv, 16, 30, 31, 32, 36, 46, 50, 51, 60, correlation coefficient, xvi, 320, 326, 328, 330
61, 64, 67, 91, 94, 106, 112, 113, 120, 122, 177, correlations, 269
199, 203, 204, 211, 234, 236, 245, 246, 312, 320 corrosion, 10, 11, 23, 45, 114, 122, 140, 170
compost, 146, 244 corrosive, 142, 237
composting, 199, 212, 238 corruption, 306, 316
compressive strength, 83, 84, 88, 94, 95, 96, 98, 100, cost benefits, 312
106 cost of power, 296
computation, 323 cost saving, 164, 313, 314, 315
computer simulations, 315 cost-benefit analysis, xv, 309
concrete, 109, 233, 281 cost-effective, ix, 2, 130, 320, 329
condensation, 23, 180 costs, xii, xiii, xv, 25, 30, 50, 51, 136, 153, 156, 158,
conditioning, 12, 170, 237 163, 164, 165, 168, 169, 172, 175, 199, 221, 227,
conductive, 11, 14, 216, 217, 218 233, 235, 250, 296, 308, 309, 313, 314, 317, 321
conductivity, ix, 1, 12, 14, 22, 23, 27, 29, 31, 32, 43, cotton, 269, 273, 330
127, 202, 203, 206, 207, 208, 219, 226, 228, 232, coupling, 137, 143
233, 239, 282, 283, 286, 288, 291, 292, 311, 313, crack, 92
315, 316 cracking, 61, 62, 88, 217
configuration, 11, 36, 38, 156, 157, 163, 166, 169, crops, 238, 239, 280, 303
171, 173, 284 crude oil, 53, 54, 61, 62, 73, 79, 80, 151
conflict, 300, 302, 303, 316 crustaceans, 122
Congress, 194 crystalline, x, 81, 90, 91, 92, 93, 96, 97, 98, 100,
conservation, 83, 128, 202, 315 102, 103, 104, 106
construction, x, xiii, 82, 83, 199, 215, 217, 222, 255, crystalline solids, 103
257, 259, 261 crystallinity, 91, 92
construction materials, 83 crystallization, x, 81, 84, 90, 98, 102, 103
consultants, 130 crystals, 91, 92, 98, 101, 102, 106
consumers, xiv, 230, 236 cultivation, 149, 154, 182, 188, 190, 211, 229, 242,
consumption, x, xii, xv, 2, 9, 12, 13, 17, 20, 23, 24, 280, 308, 320
25, 26, 27, 28, 36, 41, 43, 44, 59, 83, 102, 144, culture, 51, 52, 58, 59, 63, 65, 70, 137, 148, 160,
153, 169, 202, 207, 211, 230, 279, 282, 284, 287, 182, 184, 185, 192, 203, 210, 211, 213, 267, 321
288, 289, 291, 293, 297, 300, 306, 310, 315, 317 CVD, 216, 218
contact time, xvi, 139, 227, 231, 319, 321, 324, 325, cyanobacteria, 144, 186
327, 328, 329 cyanobacterium, 194
contaminant, 121, 126, 210, 211 cycles, 161, 163
contaminants, ix, xi, xiii, 1, 5, 21, 109, 111, 114, cycling, xiii, 235
120, 122, 126, 216, 219, 243, 250, 334 Cyprus, 277, 278
contaminated soils, 109 cysteine, 57, 58
contamination, 82, 173, 230, 244, 304, 334, 339, 341 cysts, 9
control, xvi, 10, 14, 79, 108, 136, 141, 145, 148, 149, cytochrome, 56
150, 151, 152, 156, 158, 163, 164, 194, 210, 211, cytometry, 147, 246
229, 239, 241, 242, 271, 276, 288, 296, 333, 334, cytoplasm, 185, 188, 192
344, 346
conversion, 73, 77, 83, 137, 138, 140, 142, 143, 144,
147, 149, 154, 180, 236, 293 D
conversion rate, 140
dairy, 50, 53, 58, 60, 63, 75, 80, 143, 151, 277
cooling, 11, 60, 93, 225
dairy industry, 50, 60, 80
cooling process, 93
data set, 128
Copenhagen, 142, 144, 152
database, 87, 92, 102
copper, x, 22, 44, 65, 79, 82, 102, 114, 237, 330, 331
death, 154, 209
copper oxide, x, 82, 102
decane, 54
350 Index
decay, 116 Diamond, viii, 11, 45, 46, 215, 216, 218, 219, 228,
decision makers, 126, 130 233, 234
decisions, 125, 130, 275 diamonds, 10, 222
decomposition, xiii, 54, 55, 59, 68, 69, 70, 74, 85, diaphragm, 223
98, 109, 154, 183, 184, 186, 188, 212, 235, 241 dielectric constant, 3
decomposition reactions, 98 diesel, 61
defects, 113 differentiation, 293
deficiency, 13, 206, 211, 267 diffraction, 30, 83, 88, 92
deficit, 206 diffusion, 21, 58, 106, 173
definition, 59 digestion, 78, 87, 122, 141, 168, 171, 172, 173, 193,
deforestation, xv, 299, 303, 307 199, 236, 238, 239, 240, 244, 253
degradation, xiii, xv, 12, 50, 75, 76, 78, 79, 80, 114, dioxin, 201
115, 121, 124, 145, 146, 181, 183, 185, 191, 198, direct costs, 25
201, 202, 203, 204, 207, 217, 225, 228, 230, 231, discharges, 112, 118, 121, 124, 125, 126, 127, 130,
235, 236, 237, 240, 241, 242, 283, 299, 304, 307 131, 132, 266, 271, 272, 273, 310, 313
degradation mechanism, 185, 191, 202 diseases, 243, 303, 305, 307
degradation process, 114, 121, 236 disinfection, xiii, 215, 217, 221, 231, 232, 233
degrading, 75, 141, 227 disorder, 154
degree of crystallinity, 92 dispersion, xi, 4, 5, 85, 111, 112, 114, 122, 123, 124,
dehydrogenase, 243 125, 126, 127, 128, 130
demobilization, 67 dissociation, 30
denitrification, xi, xii, 153, 154, 155, 156, 157, 159, dissolved oxygen, xi, 55, 135, 144, 145, 147, 150,
160, 161, 163, 164, 165, 167, 168, 170, 171, 173, 157, 158, 159, 162, 166, 189, 267, 269
175, 176, 178 distillation, 61, 62, 179, 314
denitrifying, 67, 143, 151, 155, 156, 158, 161, 168, distilled water, 32, 321
173, 174, 175 distribution, xii, 4, 53, 112, 113, 114, 121, 143, 151,
density, ix, xv, 1, 14, 15, 18, 19, 20, 21, 23, 24, 26, 164, 179, 180, 181, 186, 187, 216, 240
27, 36, 42, 65, 83, 87, 88, 89, 94, 95, 96, 115, diversity, xiv, 236, 241, 243
117, 118, 119, 126, 128, 129, 181, 182, 218, 221, division, 210
223, 224, 227, 228, 229, 231, 232, 240, 279, 284, DNA, 57, 245, 246
289, 294, 295 donor, 56, 57, 69, 139, 140, 142, 155, 158, 173, 174
Department of Energy, 249 donors, 53, 55, 57, 60, 144, 149, 152, 161
Department of Interior, 132 doped, ix, xiii, 10, 42, 215, 216, 217, 220, 222, 233
deposition, 216, 218 doping, 11
deposits, x, 10, 47, 50, 52, 53, 63, 64, 65, 66, 67, dosage, xvi, 20, 30, 319, 325, 326, 327
111, 112, 297 dosing, 15, 18
depreciation, 25 double bonds, 12, 216
derivatives, 12, 75, 202 drainage, 76, 140, 151, 303
desalination, xvi, 309, 312, 314, 315, 316 dream, 307
desert, 317 drinking, x, xiii, xvi, 12, 81, 84, 107, 108, 148, 154,
destruction, 83, 107, 154, 200, 201, 307 175, 177, 215, 231, 233, 306, 308, 333, 334, 335,
detection, 140, 242, 244, 347 346, 347
detention, 27, 33, 35, 288 drinking water, xiii, xvi, 12, 107, 108, 148, 154, 175,
detergents, 46 177, 215, 231, 233, 306, 308, 333, 334, 335, 346,
detoxification, 194, 195, 277 347
developed countries, 82 drought, 310
developing countries, xv, xvi, 82, 299, 306, 307, 308, drugs, 221, 233
319, 320 dry matter, 205, 239
deviation, 203 drying, 108, 251, 254, 259, 261, 304
dextrose, 188, 321 durability, 84, 98, 102
dialysis, 310, 311 duration, 52, 71, 87, 229
diamond, ix, xiii, 10, 11, 42, 215, 216, 217, 220, 222, dust, 251
224, 228, 229, 232, 233 dyes, 8, 320, 330
Index 351
employment, 280
E emulsification, 61
emulsions, xiii, 61, 62, 215
E. coli, 240
endocrine, 222, 230, 233, 334
earth, xiii, 61, 65, 98, 109, 235, 250
endocrine system, 334
ecological, 51, 112, 120, 121, 124, 126, 306
endothermic, 89, 90, 91, 96, 97, 98
ecology, 77, 308
energy consumption, x, xii, 2, 13, 17, 25, 27, 28, 44,
economic disadvantage, 143
153, 169, 297
economics, 142
energy efficiency, 11, 220
ecosystem, xiii, xiv, 120, 126, 215, 231, 236, 242,
energy supply, 136
243, 244, 246, 300
England, 308
ecosystems, 112, 114, 125, 130, 222, 241
environmental conditions, 114, 125, 126, 147, 208
ECSC, 245
environmental effects, 124, 136, 306
effluents, xiv, xvi, 12, 43, 46, 49, 50, 66, 67, 122,
environmental factors, 112
156, 157, 158, 162, 170, 173, 178, 199, 202, 212,
environmental impact, 125, 130, 175
231, 265, 266, 267, 268, 269, 273, 277, 278, 282,
environmental protection, 136
296, 297, 333, 335
Environmental Protection Agency, 130, 133, 147,
Egypt, 300, 301, 307
334, 347
elaboration, xii, 197, 198
environmental regulations, 8, 82, 281
electric charge, 2
environmental resources, xi, 111
electric conductivity, 202, 203, 207
enzymatic, 67, 180, 242, 243
electric field, 13
enzymes, 54, 60, 63, 68, 141, 183, 186, 192, 244
electric power, 22, 41, 42
EPA, 130, 131, 133, 147, 336, 346, 347
electrical conductivity, 12, 27, 31, 32, 43, 206, 208,
epithelia, 122
282, 283, 286, 291, 292
epoxy, 61
electrical power, 17, 136
epoxy resins, 61
electrical properties, 2, 11
equilibrium, 50, 104, 116, 315, 323, 324, 326, 329
electricity, 250, 255, 261
equilibrium state, 104
electrochemical measurements, 45
erosion, 65, 303
electrochemical reaction, 8, 154
Escherichia coli, 231, 232, 240, 245, 313
electrochemistry, 47
esters, 61, 226, 347
electrodeposition, 45
estuarine, 151
electrodes, ix, xiii, xv, 1, 5, 8, 9, 10, 11, 13, 15, 16,
ethanol, 54, 55, 56, 57, 69, 70, 173
17, 18, 20, 21, 22, 23, 25, 26, 27, 40, 41, 42, 43,
Ethanol, 68
45, 46, 215, 216, 217, 218, 219, 220, 222, 223,
Ethiopia, 280, 297, 300, 301
224, 228, 229, 230, 231, 232, 233, 279, 282, 284,
ethyl alcohol, 180
288, 294, 295, 297
ethylbenzene, 54, 78
electroflotation, 7, 11, 14, 16
ethylene, 61
electrolysis, xv, 7, 11, 13, 17, 18, 19, 24, 33, 220,
Eulerian, 123
221, 279, 284, 288
Euro, 165, 314
electrolyte, x, 2, 9, 16, 22, 23, 24, 25, 31, 36, 41, 42,
Europe, 216, 231, 245, 280, 300
43, 48, 286, 287, 288, 289, 292
European Parliament, 233, 347
electrolytes, x, 2, 11, 12, 22, 25, 37, 42, 45, 284, 288
European Union, 334
electromagnetic, 61
eutrophication, 304
electron, 11, 13, 53, 54, 55, 56, 57, 59, 60, 69, 72,
evaporation, xii, 112, 114, 197, 199, 226, 320
77, 87, 92, 137, 139, 140, 142, 144, 149, 155,
evolution, 9, 17, 125
158, 161, 162, 173, 174, 188, 266, 267, 268, 269
examinations, 232
electrons, 9, 22, 41, 53, 55, 72, 92, 106, 163, 267
excess supply, 239
electrophoresis, 7
excrements, 221
electroplating, 11, 66, 109, 320
excretion, 306
electroreduction, 297
exothermic peaks, 91, 96
electrostatic force, 3, 4
Expert System, 123
e-mail, 49, 265
expertise, xiv, 236
emission, 49, 66, 108, 136, 177, 313
352 Index
exploitation, xiii, 56, 66, 112, 122, 147, 244, 302 flavonoids, 334
exports, 280 flexibility, 122, 170, 334
exposure, 124, 290, 307, 331, 334 float, 312
Exposure, 113 flocculation, ix, 1, 2, 5, 7, 22, 31, 84, 115
extraction, x, xi, xvi, 50, 66, 87, 109, 111, 112, 113, flooding, 9, 303
124, 125, 127, 130, 198, 205, 206, 226, 320, 333, flotation, 7, 11, 14, 15, 18, 20, 78, 107, 204
334, 335, 336, 341, 342, 346, 347 flow, xiv, 11, 15, 16, 22, 33, 35, 41, 51, 59, 64, 82,
120, 128, 137, 140, 147, 162, 169, 217, 222, 223,
224, 225, 227, 228, 229, 230, 232, 246, 249, 251,
F 252, 253, 255, 256, 257, 258, 259, 287, 288, 293,
294, 301, 304, 310, 311, 323
fabric, 269, 273
fluctuations, 166, 280
failure, 156, 170, 306, 310
flue gas, xi, 135, 136, 137, 138, 139, 142, 147, 149,
family, 72, 280
237
FAO, 293, 311, 312, 313, 315, 316
fluid, ix, 11, 15, 112, 115, 116, 117, 118, 123, 125,
farmers, 240
129, 216, 219, 224, 233
farming, 63, 240
fluidized bed, 5, 143, 148, 151
farmland, 240
fluorescence, 87
farms, 49, 305
fluoride, xiii, 23, 45, 215
fat, 12, 60, 210, 228
fluorinated, 217
fats, 60, 209, 236, 280
flushing, 60, 61, 62, 303
fatty acids, 55, 60, 62, 75, 76, 141, 236
focusing, 243
February, 198, 277, 278, 316, 317
FOG, 12
feces, 236, 240
food, 12, 49, 60, 67, 199, 202, 238, 239, 240, 242,
Federal Register, 133
252, 280, 300, 306, 317, 321, 329, 331, 333, 334
feeding, 146, 190, 191, 192, 231
food industry, 60, 202, 240, 329
feedstock, 142
food production, 306
feldspars, x, 81, 92, 97, 98, 102, 106
foodstuffs, 334
Fenton‘s reagent, 199, 200, 202
Forestry, 1, 279
fermentation, xii, 54, 67, 68, 69, 70, 179, 180, 182,
forests, 302, 303, 304, 307
190, 191, 193, 194, 237, 281
formaldehyde, 201
fern, 331
fossil fuels, 136, 154
ferrous ion, 200, 201
fouling, 10, 219, 284
fertility, 238, 243, 244, 303
Fourier, xvi, 320, 324
fertilization, 239, 243
fractionation, xiv, 80, 265, 273, 275, 278
fertilizer, xiii, 66, 107, 142, 154, 172, 199, 235, 237,
France, 280, 311
238, 239, 240, 241, 242, 243, 245, 284
freeze-dried, 182
fertilizers, xiii, 50, 62, 63, 66, 235, 238, 239, 241,
freight, 12
244, 246, 305, 306
fresh water, xv, 53, 280, 299, 304, 309
fiber, 12, 78, 278
freshwater, 203
field trials, 240
Freundlich isotherm, 323, 326
fillers, 64
fructose, 281
film, 77, 192, 217, 313
fruits, 198, 303
filter feeders, 115
FTIR, 328, 329, 330
filters, 18, 84, 139, 182, 208, 310, 311
FT-IR, 324
filtration, ix, 2, 5, 13, 14, 46, 74, 148, 174, 187, 193,
FT-IR, 329
203, 204, 206, 208, 229, 311, 313, 317
fuel, 61, 69, 82, 136, 180
financial support, 330
fumaric, 55, 57
Finland, 80, 296
funding, 296
fish, 121, 122, 154, 156, 175, 305
fungal, xvi, 183, 185, 189, 191, 319, 320, 321, 324,
fish production, 175
329, 331, 334
fishing, 304
fungal metabolite, 334
fixation, 75, 102, 154, 202
flame, 283
Index 353
fungi, xvi, 64, 181, 183, 184, 185, 186, 190, 192, grazing, 238, 304
193, 194, 195, 241, 319, 320, 321, 324 Great Britain, 296
fungicide, 142 Greece, 277
fungus, 193, 194, 195, 212 green beans, 281
greenhouse, 82
ground water, xii, 153, 173, 174, 233, 241, 302
G groundwater, xii, 12, 53, 177, 197, 230, 244, 334,
344, 345, 346
gas, x, xi, xvi, 7, 8, 13, 53, 61, 73, 93, 96, 98, 108,
groups, xvi, 2, 12, 55, 57, 59, 65, 67, 69, 70, 71, 74,
111, 112, 113, 114, 122, 124, 125, 127, 128, 130,
88, 91, 92, 106, 123, 127, 142, 144, 181, 185,
132, 135, 136, 137, 138, 139, 142, 143, 147, 148,
190, 216, 226, 236, 244, 320, 329, 330
149, 150, 152, 155, 158, 160, 161, 176, 177, 226,
growth, ix, xii, xv, 2, 20, 50, 51, 55, 56, 57, 59, 64,
236, 237, 285, 323, 333, 335
68, 69, 73, 76, 77, 93, 101, 121, 137, 141, 143,
gas chromatograph, xvi, 226, 333
146, 148, 149, 151, 156, 157, 164, 175, 182, 183,
gaseous waste, 66
184, 191, 192, 197, 199, 202, 209, 210, 211, 239,
gases, xi, xiii, 7, 16, 49, 58, 59, 84, 92, 93, 95, 96,
244, 267, 276, 299, 304, 306, 307, 321, 331
135, 136, 137, 139, 147, 148, 149, 235, 237
growth rate, 121, 156, 157, 164, 184, 209, 210, 276
gasification, 82, 108, 136
growth temperature, 56, 57, 141
Gaussian, 117
guidelines, 125, 126, 316
gel, 187, 189, 190
Gulf of Mexico, 120, 121, 125
gene, 245, 246
gut, 240
gene transfer, 245, 246
generation, 7, 19, 22, 108, 158, 170, 173, 211, 231,
237, 250, 252, 273, 275, 306, 320 H
genes, 79, 241, 242, 246
Geneva, 307, 308, 347 H1, 206
geochemical, 152 H2, 6, 17, 57, 137, 139, 140
geothermal, 53 habitat, 195
Germany, 45, 148, 216, 230, 232 haloalkaliphilic, 151
GFP, 246 halogens, 200
glaciers, xv, 299 handling, 82, 83, 107, 237, 250, 251, 266, 273
glass, 64, 82, 83, 85, 106, 108, 110, 340, 341, 342 harbour, 316
glasses, 109 hardness, 12, 100
global warming, 237 harm, 209
glucose, 54, 182, 183, 184, 185, 186, 189, 192, 193, harvest, 208, 280
196, 281 harvesting, 12, 240
glucose oxidase, 183 Hawaii, 245
glycerin, 184 hazardous substance, 334, 347
glycine, 182, 185, 196 hazardous substances, 347
glycol, 61, 114, 126, 127, 178 hazardous wastes, 109
goals, 122, 302, 306, 307 hazards, 50, 250
gold, 65, 150, 280 health, xiii, xvi, 121, 235, 240, 245, 250, 306, 308,
Gore, 133 315, 319, 320, 333
government, 231, 301, 306, 307, 309 health effects, 334
graduate students, 296 heart, 9, 230
grain, 126 heat, 62, 83, 101, 103, 106, 136, 193, 237, 238, 240
grains, 63, 64 heating, 61, 85, 87, 89, 97, 182, 238, 240
Gram-negative, 80, 241 heavy metal, x, 12, 14, 52, 62, 65, 66, 67, 71, 72, 74,
Gram-positive, 241 76, 77, 78, 82, 83, 84, 86, 87, 101, 102, 103, 104,
granules, 64, 174 105, 106, 107, 108, 109, 112, 140, 141, 148, 151,
graphite, 43, 220 152, 208, 222, 239, 240, 243, 320, 330, 331
grasses, 154 heavy metals, x, 12, 52, 62, 65, 66, 67, 71, 74, 76,
grassland, 245, 246 77, 78, 82, 83, 84, 86, 87, 101, 102, 103, 104,
gravity, 42, 87, 119, 252, 253, 275, 321
354 Index
105, 106, 108, 109, 112, 140, 141, 148, 151, 152, hydrogen sulfide, 78, 107, 137, 150, 151, 152
208, 222, 239, 240, 243, 320, 330, 331 hydrological, 300, 302, 304
height, 203, 294 hydrolysis, 5, 10, 60, 67, 69, 114, 172, 173, 236
hemodialysis, xv, 309, 310, 311, 312, 313, 314, 315 Hydrometallurgy, 79, 150, 152
hemoglobin, 154 hydrophobic, 2, 3
herbicide, 229, 230, 231 hydroquinone, 202
herbicides, 59, 201 hydrothermal, 56, 144
heterotrophic, 143, 146, 155, 156, 161, 173, 202, 267 hydroxide, 7, 8, 10, 18, 31, 43, 62, 208, 226, 287
heuristic, 14 hydroxides, ix, 1, 5, 7, 13, 114
hexane, 54, 226 hydroxyl, xvi, 7, 8, 11, 212, 219, 220, 221, 222, 320,
high temperature, 31, 65, 102, 164, 344 329, 330
highlands, 280, 301, 303 hydroxyl groups, xvi, 320, 329, 330
hips, 27, 45 hypothesis, 83
holistic approach, ix, 1, 14
homogenized, 255
homogenous, 322 I
horizontal gene transfer, 246
IARC, 347
hormone, 334
identification, 59, 102, 140, 147, 242, 268, 271, 338
hormones, 222, 233
IEA, 237, 244
horse, 253
illumination, 203, 210
hospital, 223, 245, 310, 316
images, 98
hospitals, 310
immobilization, 102, 108
host, 242
impact assessment, 124
hot water, 198
implementation, 123, 163, 164, 172, 178, 198, 218,
hotels, xiv, 265
244, 306
household, 239, 240
impurities, 11, 220, 250
household waste, 239, 240
in situ, xi, xiii, 7, 10, 13, 26, 111, 121, 124, 126, 216
households, 59
in vitro, 347
HPLC, 340
in vivo, 347
human, xiii, xvi, 49, 121, 221, 235, 237, 240, 242,
inactivation, 233
243, 300, 319, 320, 334
inactive, 154, 320
human exposure, 334
incidence, 243
humans, 202, 238, 347
incineration, x, 81, 82, 107, 108, 199, 237
humic acid, 78, 245
inclusion, x, 2
humic substances, 48
income, 306
hybrid, 193
incubation, 51, 58, 238, 239
hybridization, 140, 147, 149
incubation period, 238
hydrate, 97, 114
independence, 242
hydration, 2
India, xiv, 44, 249, 251, 259, 316
hydraulic fluids, 334
Indian, xiv, 135, 249, 250, 262
hydro, xi, 2, 53, 59, 61, 62, 65, 77, 78, 79, 111, 112,
indication, 4, 15, 126, 146, 285
113, 114, 115, 121, 124, 237
indicators, 57, 276
hydrocarbon, xi, 75, 111, 121
Indigenous, 280
hydrocarbons, xi, 53, 59, 61, 62, 77, 79, 111, 112,
indium, 25, 44, 297
113, 114, 115, 121, 237
indole, 54
hydrochloric acid, 226
Indonesia, 180
hydrogen, 7, 8, 17, 52, 53, 55, 57, 60, 62, 69, 70, 71,
inducible enzyme, 183
72, 73, 74, 78, 107, 114, 137, 140, 144, 149, 150,
induction, 186
151, 152, 183, 186, 194, 199, 200, 201, 202, 207,
industrial application, xi, 135, 142, 296, 320
212, 219, 220, 231, 237, 285
industrial chemicals, 222, 233, 266, 347
hydrogen gas, 7, 237, 285
industrial combustion, 136
hydrogen peroxide, 183, 186, 194, 199, 200, 201,
industrial sectors, xiv, 265
202, 207, 212, 219, 231
industrial wastes, 51, 52, 65, 74, 109, 136, 201, 277
Index 355
industrialization, xi, 135, 199 iron, ix, 1, 5, 7, 9, 10, 15, 17, 20, 25, 27, 41, 43, 44,
industry, ix, 2, 50, 59, 60, 61, 63, 66, 80, 112, 130, 45, 47, 55, 56, 59, 65, 67, 72, 76, 79, 92, 114,
132, 137, 147, 198, 199, 202, 203, 211, 223, 240, 139, 154, 201, 204, 208, 211, 284, 288, 297, 330
272, 273, 277, 278, 309, 329 irradiation, 238, 281
inert, xiv, 265, 269, 270, 271, 273, 276, 277, 278 irrigation, xiii, xiv, xv, 198, 199, 203, 208, 249, 262,
infants, 154 291, 309, 312, 314, 315, 317
infertile, 241 irritation, 305
infrared, 73, 108 Islamic, 299
infrared light, 73 isoelectric point, 4, 5
infrastructure, 306, 307 isolation, 58, 137
ingestion, 122 isomers, 338, 339
inhibition, 23, 76, 79, 144, 147, 157, 158, 176, 177, isothermal, xvi, 319
186, 209, 210, 211, 230, 236, 281 isotherms, 322, 331
inhibitors, 56, 114, 267 isotope, 126, 227
inhibitory, 79, 137, 138, 209, 269 Israel, 300, 301
inhibitory effect, 209 Italy, 111, 132, 170, 198, 280
initiation, 104 IUCN, 308
injection, 238, 335
innovation, 11
inoculum, 167, 321 J
inorganic, x, 50, 51, 59, 62, 64, 65, 72, 74, 81, 82,
JAG, 183, 193
85, 91, 97, 98, 107, 112, 114, 144, 154, 178, 199,
Japan, 87, 167, 168, 181
219, 240
Japanese, 181
inorganic salts, 62
Jordan, 300
insect repellents, 333
judge, 102
insertion, 224
insight, 107, 222, 223
Inspection, 308 K
instability, 4, 240
instruments, 87 Kenya, x, xv, 1, 2, 43, 44, 46, 47, 279, 280, 281, 282,
insulators, 295 285, 290, 295, 296, 297
integrated unit, 14 ketones, 76
integration, 316, 338 killing, 56
interaction, 88, 115, 154, 161, 224, 322 kinetic energy, 115, 119
interactions, 8, 16, 102, kinetic model, 323, 329
interface, 5 kinetic parameters, 158
intermediaries, 201 kinetics, 37, 108, 152, 177, 212, 278, 323, 327, 331
International Agency for Research on Cancer, 347 King, 131, 179
intrinsic, 91 Kobe, 184, 196
invertebrates, 121, 144 Korean, 45
investment, 25, 227
ion adsorption, 4
ionic, 2, 4, 31, 173, 290 L
ionization, 226, 335
ions, ix, x, xvi, 1, 4, 5, 7, 8, 10, 15, 16, 17, 18, 21, LAB, 191
23, 25, 31, 37, 55, 56, 63, 67, 81, 95, 100, 103, labeling, 91, 99, 103
106, 126, 147, 200, 201, 211, 221, 225, 287, 289, labor, 25, 42, 296, 312, 313, 314
319, 320, 321, 322, 323, 324, 325, 326, 327, 329, labour, xiv, 13, 199, 249, 261
330, 331 lactic acid, 68, 186, 191, 195
IR, 324, 329 lactic acid bacteria, 191
Iran, 300 Lactobacillus, 183, 185, 190, 191, 195
lactose, 60
lagoon, 75, 180
Lagrangian, 123, 127, 128
356 Index
MDA, vii, xii, 179, 180, 181, 182, 183, 184, 185, microorganism, 51, 52, 53, 57, 59, 64, 67, 69, 70,
186, 188, 189, 190, 191, 192 182, 183, 206, 267
measurement, 112, 322 micro-organisms, 115, 245, 246
measures, 117, 119, 125, 126, 136 microscope, 87, 92
meat, 60, 277, 280 microstructures, 92,93, 98
media, xi, 4, 54, 57, 59, 83, 135, 182, 216, 231, 281 Middle East, 300, 301, 317
Mediterranean, xi, xii, 111, 113, 120, 121, 124, 127, migration, 21, 23
197, 198, 199, 245 military, 300, 301
Mediterranean countries, 198, 199 milk, 50, 60, 61, 62
melting, 93, 94, 95, 98 Millennium, 302, 308
membranes, 113, 310, 311 milligrams, 37
Merck, 339 mineral oils, 62
mercury, 65, 76, 114, 237 mineral water, 339, 341, 342, 343, 344, 346
metabolic, 52, 75, 140, 145, 242, 273 mineralization, 52, 74, 76, 141, 201, 221, 239, 242
metabolism, 51, 55, 67, 71, 75, 147, 152, 240, 241 mineralized, 220
metabolites, 77, 229, 230, 231, 334 minerals, 30, 79, 90, 98, 102, 137, 180, 291
metal content, 72, 239 Minerals Management Service, 132
metal hydroxides, 5 mines, 62
metal ions, 5, 8, 15, 16, 21, 95, 106, 147, 201, 289, mining, xvi, 49, 82, 319, 320
324, 331 Minnesota, 79
metal oxide, 82, 85, 223, 287 misleading, 267, 269
metal oxides, 82, 287 missions, 136, 168, 173
metallurgy, 66, 67 mixing, ix, xi, 7, 15, 33, 86, 111, 112, 113, 115, 116,
metals, x, 12, 52, 62, 65, 66, 67, 71, 72, 74, 76, 77, 118, 119, 121, 124, 127, 129, 130, 131, 169, 209,
78, 79, 82, 83, 84, 86, 87, 101, 102, 103, 104, 218, 228, 231, 236, 251
105, 106, 107, 108, 109, 110, 112, 114, 121, 140, MMS, 132
141, 148, 151, 152, 208, 222, 237, 239, 240, 243, mobility, 3, 57, 65, 79, 104, 106
244, 320, 322, 330, 331 MOD, 133
methane, xiii, 69, 82, 141, 170, 172, 191, 235, 236, model system, 193, 196
237, 240 modeling, 278
methanogenesis, 63, 70, 71, 79, 147, 212 models, 120, 122, 123, 124, 126, 131, 209, 266, 275,
methanol, 54, 157, 164, 173, 226, 335 315, 323, 326
methemoglobinemia, 154 modulus, 85
methylation, 76 moisture, 199, 230, 237, 245, 302
methylene, 331 moisture content, 199, 245
Mexico, 120, 121, 125 molar ratio, 160
microaerophilic, 138, 139 molasses, xii, 48, 63, 179, 180, 181, 182, 185, 186,
microalgae, xiii, 198, 202 189, 190, 191, 192, 193, 194, 195, 196
microbes, 180, 192 mold, 181
microbial, xi, xii, xiii, xiv, 75, 78, 79, 115, 135, 137, mole, 293
143, 147, 148, 149, 150, 154, 177, 179, 181, 182, molecular weight, 113, 114, 115, 180, 181, 184, 186
185, 186, 190, 191, 192, 197, 204, 235, 236, 238, molecular weight distribution, 181, 186
239, 241, 242, 243, 244, 245, 246, 247, 266, 270, molecules, 10, 140, 322
271, 275, 278, 304, 311, 321 molybdenum, 65
microbial agents, 238 momentum, 15, 115, 118, 119, 120, 128
microbial communities, xiii, 79, 235, 241, 242, 243, money, 250
245, 246 monolayer, xvi, 319, 329
microbial community, xiv, 115, 236, 239, 241, 245, monomeric, 70
microbiota, 239 Moon, 108
microcosms, 57, 58, 77, 80 Morocco, 300, 309, 310, 312, 316
microflora, 55, 199 morphological, 83, 84, 88, 96, 99
micrograms, 334 morphology, 92, 98, 126, 145
micronutrients, 237 mortality, 154, 175
358 Index
motion, 2, 4, 13, 119 nitrification, xi, 50, 153, 155, 156, 157, 158, 159,
motors, 261 160, 162, 164, 166, 167, 168, 169, 170, 175, 176,
mountains, 304 177, 178, 241, 243
movement, 22, 341 nitrifying bacteria, 156, 162
multidisciplinary, xi, 111, 113, 126, 127, 130 Nitrite, 154, 158, 160, 164
multidrug resistance, 245 nitrogen, xi, xii, 49, 50, 61, 75, 153, 154, 155, 156,
multiplication, 57, 58 158, 159, 160, 161, 162, 165, 168, 170, 173, 174,
municipal sewage, xiii, 82, 221, 235, 244, 311 175, 176, 177, 184, 211, 226, 237, 239, 243, 250,
municipal solid waste, 82, 109 305, 306, 308, 311, 312, 313, 320
mycelium, 188 nitrogen compounds, xii, 49, 153, 154, 161, 237
nitrogen fixation, 75, 154
nitrogen gas, 155, 158, 160, 161, 176
N nitrogen oxides, 161
nitrosamines, 154
Na+, 4, 31
nitrous oxide, 177
Na2SO4, 23, 24, 63
non toxic, 130
NaCl, x, 2, 23, 24, 25, 62, 321, 336, 339, 341, 346
non-enzymatic, 180
Namibia, 300
normal, xii, 12, 61, 157, 179, 180, 190, 219, 238,
naphthalene, 58, 121, 277
285, 289
National Bureau of Standards, 94
norms, 51
native plant, 245
North Africa, 46, 300, 317
natural, x, xi, xiii, 5, 9, 12, 22, 44, 46, 47, 49, 50, 51,
North America, 300
53, 57, 58, 59, 65, 66, 71, 72, 73, 74, 75, 83, 93,
Norway, 121
111, 113, 121, 122, 123, 141, 152, 154, 181, 198,
NTU, 22
204, 241, 250, 251, 273, 277, 300, 306, 307, 310,
nucleic acid, 147
320, 321, 330
numerical tool, 123
natural environment, 49, 50, 51, 53, 57, 58, 65, 66,
nutrient, xi, xiii, 115, 144, 146, 153, 202, 235, 236,
71, 72, 73, 74
246, 307, 321
natural gas, xi, 53, 111, 113, 152
nutrient cycling, xiii, 235
natural resources, 83
nutrients, xiii, 126, 142, 198, 211, 237, 238, 239,
Nb, 216, 218
242, 243, 250, 303, 320
NCA, 46
nutrition, 202, 212
Near East, 277, 278
neck, 93
neem, 331 O
Nepal, 316
Netherlands, 87, 121, 147, 149, 158, 163, 164, 167, observations, xi, 92, 98, 111, 112, 113, 122, 125,
168, 308 126, 130
network, 88, 92, 95, 100, 106 octane, 54
neutralization, 7, 47, 50, 66, 203, 207, 208 odors, 140, 211, 238
New Frontier, 77 offshore, x, xi, 111, 112, 113, 114, 115, 120, 121,
New Jersey, 245 122, 124, 126, 127, 130, 132
New York, 45, 46, 75, 80, 131, 132, 133, 151, 152, offshore oil, x, 111, 112, 127, 130, 132
165, 196, 234, 262, 276, 300, 308 oil, xi, xii, xiii, 11, 44, 50, 53, 54, 57, 61, 62, 73, 77,
Ni, 52, 85, 331 78, 79, 80, 107, 108, 111, 112, 113, 114, 122,
nickel, 65, 66, 67, 109, 114, 330 124, 125, 127, 130, 132, 136, 151, 197, 198, 199,
Nielsen, 53, 78 202, 203, 204, 205, 206, 208, 209, 211, 212, 215,
Niger, 300, 301 228, 229, 239, 244, 245, 300
Nile, 300, 301, 304 oil production, 198, 199
niobium, 65 oil refineries, 50, 62
nitrate, 11, 23, 78, 143, 149, 154, 155, 156, 157, 161, oil refining, 136
162, 163, 171, 173, 174, 175, 177, 241, 297 oil shale, 11
nitrates, 57 oils, 62
nitric oxide, 150, 177
Index 359
olive, ix, xii, 197, 198, 199, 202, 203, 204, 205, 206, Paris, 334, 347
208, 209, 211, 212, 213, 330, 331 Parliament, 233, 347
olive oil, xii, 197, 198, 199, 202, 204, 205, 206, 212 particle density, 83, 87, 88, 89, 95, 96
olives, xii, 197, 198, 202, 203, 204, 208, 209 particles, xiii, 2, 4, 5, 7, 11, 13, 31, 33, 46, 64, 66,
Oman, 300 88, 93, 100, 103, 113, 123, 142, 144, 174, 215,
online, 151, 308 217, 252, 290
Operators, 125 particulate matter, 108, 121
optical, 73, 78, 182 partition, 114, 198, 255
optical density, 182 passivation, 10, 13, 217
optimization, 14, 22, 147, 227, 231 passive, 10, 49, 112
oral, 122 pasteurization, 240
ores, 66, 72, 150 pasture, 307
organelles, 188 pathogenic, xiii, 52, 71, 83, 235, 236, 237, 240, 242,
organic C, 241 243, 245, 246
organic compounds, 10, 52, 53, 54, 55, 56, 57, 58, pathogenic agents, 83
59, 60, 61, 63, 67, 68, 69, 70, 72, 73, 74, 113, pathogens, 82, 175, 238, 240, 241, 244, 250, 306
140, 142, 144, 200, 211, 213, 237, 262, 267, 334 patients, 310
organism, 139, 145 Pb, viii, x, xvi, 21, 82, 84, 86, 101, 102, 103, 104,
organochlorinated, 201 106, 290, 319, 320, 321, 322, 323, 324, 325, 326,
orientation, 120 327, 328, 329, 330, 331
oscillation, 52 PCBs, 201, 334
osmosis, 12, 173, 310, 311, 313, 314, 315, 320 PCT, 36, 46, 233, 296
osmotic, 9 peat, 262, 331
osmotic pressure, 9 pectin, 198, 281
ovulation, 230 per capita, 300
oxalic, 201 percolation, 253, 262
oxalic acid, 201 periodic, 14
oxidants, 221, 222 permeability, 113
oxidation rate, 145, 158 permeable membrane, 121
oxidative, x, 19, 82, 84, 104, 106, 201, 220 permit, 33
oxide, x, 10, 13, 30, 43, 53, 82, 83, 85, 95, 97, 98, permittivity, 3
99, 102, 150, 177, 208, 217, 223, 228, 231, 330 peroxide, 183, 186, 194, 199, 200, 201, 202, 206,
oxide electrodes, 10 207, 212, 219, 231
oxides, x, 13, 30, 81, 82, 83, 84, 85, 88, 95, 96, 100, pesticide, 66, 234, 347
102, 161, 218, 287, 289 pesticides, 59, 66, 202, 209, 211, 305, 340
Oxygen, 47, 145, 151, 160, 178, 219, 262 PET, 342, 344
oxygenation, 59, 67, 150 petrochemical, 43, 50, 53, 54, 57, 58, 61, 74, 170
ozonation, 222, 234, 277 petroleum, 77, 78, 80, 280
ozone, xiii, 49, 109, 154, 216, 219, 222, 224, 225 Petroleum, 61
Ozone, 219 petroleum products, 80
Petrology, 49
pH values, 9, 17, 141, 324
P pharmaceutical, 69, 226, 233, 234
pharmaceuticals, xiii, 215, 222, 225
Pacific, 131, 132
PHB, 202
packaging, 333
phenol, 54, 58, 61, 62, 75, 78, 202, 203
PAHs, 113
phenolic, xii, 193, 197, 198, 204, 208, 209, 211, 212
paints, 66, 320
phenolic acid, 193
Palestine, 300, 301
phenolic compounds, xii, 197, 198, 204, 208, 209,
palletized, 109
211, 212
paraffins, 201
Philippines, 180
parameter, 15, 17, 18, 33, 116, 119, 120, 141, 203,
philosophy, xiv, 265, 266, 271
225, 226, 266, 267, 268, 269, 290
parasites, 246
360 Index
phosphate, x, xv, 2, 12, 36, 39, 43, 44, 47, 241, 279, polymer synthesis, 344
280, 282, 284, 285, 286, 287, 289, 292, 296, 297 polymerization, 12
Phosphate, xv, 24, 279, 284 polymers, 196
Phosphogypsum, 80 polyphenolic compounds, 199
phosphorous, 305 polyphenols, 199, 206, 212
phosphorus, 50, 107, 108, 172, 250, 311, 313 polysaccharides, 188, 329
phosphorylation, 55 polyurethane, 190, 193
photobioreactors, 203 polyurethane foam, 190, 193
photoionization, 347 polyvinyl chloride, 334
photolysis, 113 pomace, 198, 199
photomicrographs, 93 pond, xii, 179, 180, 181
photosynthetic, 12, 78, 137, 143, 150, 202 pools, 208
phototrophic, 77, 142, 144 poor, xii, 146, 179, 184, 239, 303, 304, 306, 310
phylum, 158 population, ix, xv, 2, 67, 146, 280, 299, 300, 301,
physical factors, 14 304, 306, 307, 315
physical properties, 238, 244, 246 population growth, ix, xv, 2, 299, 304, 306, 307
physicochemical, 114, 136, 164, 251, 255 pore, 93, 95, 203, 240
physiological, 59, 246 pores, x, 81, 84, 88, 92, 93, 95, 98, 100, 104, 106
physiology, 53, 145, 308 porosity, 83, 87, 88, 89, 95, 96, 98, 240, 243
phytoplankton, 121, 304 porous, 92, 93, 100, 102, 106, 194
pig, 239 ports, 119, 120
pigments, 195 potassium, 30, 37, 114
pipelines, 61 potato, 27, 45, 188, 321
planar, 88 potatoes, 303
planning, xi, 111 poultry, 276
plants, xi, xii, xiii, 30, 49, 51, 59, 60, 61, 62, 66, 73, poverty, 303
74, 75, 86, 136, 144, 153, 154, 157, 167, 194, powder, 30, 87, 92, 102, 182, 331
197, 222, 233, 235, 236, 239, 240, 244, 245, 246, powders, 324
266, 267, 271, 311, 317, 321 power, xi, xv, 9, 12, 17, 20, 22, 23, 24, 26, 28, 30,
plastic, xiii, 64, 66, 215, 217, 284 36, 41, 42, 62, 136, 169, 221, 223, 228, 233, 237,
plasticizer, 334 253, 258, 279, 282, 284, 287, 288, 289, 291, 293,
plastics, 61, 69, 73 295, 296, 300, 317, 323
platforms, xi, 111, 113, 114, 115, 120, 121, 122, 124, power plants, xi, 30, 62, 136
125, 126, 127, 128, 130 precipitation, 5, 53, 59, 66, 67, 71, 114, 208, 219,
play, xvi, 63, 67, 95, 120, 125, 141, 147, 304, 320 224, 302, 303, 320, 324
Pleurotus ostreatus, 183 preconditioning, 12
ploughing, 240 prediction, ix, 1, 14
Poland, 49 preference, 9, 115
polarity, 10, 11, 14, 115, 219, 224, 229, 341 press, 44, 84, 87, 94, 151, 199, 246, 297, 308
politics, 300 pressure, ix, 2, 9, 61, 65, 136, 174, 246, 300, 303,
pollutants, ix, xi, xiii, 1, 2, 7, 8, 12, 19, 20, 22, 49, 304, 307
50, 51, 57, 58, 59, 61, 62, 63, 67, 69, 74, 83, 107, prevention, 125
108, 116, 117, 123, 130, 135, 200, 201, 215, 216, prices, 25, 201, 313, 314
237, 239, 243, 250, 320, 334, 339, 341, 342, 346 printing, 66
pollution, x, xi, xv, xvi, 2, 12, 13, 42, 49, 50, 51, 64, probability, 92, 98
73, 82, 83, 107, 111, 117, 136, 149, 154, 200, process control, 330
231, 237, 241, 262, 267, 282, 299, 300, 304, 305, producers, xii, xiv, 30, 197, 199, 216, 217, 236, 303
306, 307, 319, 333, 335, 339, 340, 346 production costs, xv, 309
polyamide, 278 productivity, 137
polycyclic aromatic hydrocarbon, 78, 124 profit, 55
Polyelectrolyte, 28, 30 program, 226, 311, 316, 317, 321
polyethylene, 178 prokaryotes, 80
polymer, 66, 145, 334, 344, 346 prokaryotic, 140
Index 361
proliferation, 156, 242 raw material, xi, xii, 61, 83, 84, 85, 87, 88, 89, 91,
propylene, 61 92, 93, 95, 102, 135, 179, 180, 191, 192
protection, 50, 74, 136, 307, 347 raw materials, xii, 83, 84, 85, 87, 88, 89, 91, 92, 93,
protein, 12, 56, 60, 154, 329 95, 102, 179
proteinase, 241 reaction order, 33
proteins, 54, 55, 60, 142, 180, 236, 242, 306 reaction rate, 13, 19, 21, 22, 37, 40, 328
Proteins, 61 reaction time, 21, 296
protocol, 114, 130, 266 reactivity, 100, 109, 284, 289
protocols, 268, 271 reagent, 199, 200, 201, 202, 212, 219, 276, 322
protons, 161, 324 reagents, xvi, 25, 333, 335, 339, 341, 346
protozoa, 237 receptor sites, 8
pseudo, xvi, 14, 18, 320, 323, 327, 328, 329 receptors, 334
Pseudomonas, 145, 185, 190, 193, 194, 246 reclamation, 142, 148, 240, 317
public, 82, 123, 126, 306, 316 Reclamation, 311
public administration, 126 reconditioning, 12
public health, 306 recovery, xvi, 13, 65, 66, 72, 77, 83, 137, 146, 150,
pulp, ix, 2, 8, 12, 17, 20, 23, 40, 41, 43, 44, 47, 61, 151, 172, 212, 244, 262, 278, 310, 316, 333, 341
69, 136, 198, 276, 277, 281, 282, 296, 297 rectification, 61
pulp mill, 296 recycling, xiv, xv, 12, 82, 83, 156, 236, 237, 238,
pumping, 169, 175, 232, 252, 257, 296 244, 280, 296, 303, 309, 310, 313, 314, 315
pumps, 223, 261 redox, 17, 55, 57, 58, 59, 145, 150, 219, 225, 239
purification, xiii, 44, 109, 198, 207, 233 Redox, 226
PVC, 333, 334, 344 reducing sugars, 281
pyrolysis, 82, 107, 108 REE, 65
pyruvic, 53, 55, 57 refineries, xi, 50, 62, 66, 136, 200
refining, 53, 61, 62, 78, 80, 109, 136, 320
refractory, 150
Q refrigeration, 344
regeneration, 82, 137
Qatar, 300
regional, 122, 124, 309
quality of life, 250
regression, 328
quartz, 91, 92, 98, 99, 102, 103
regulation, 67, 130, 222
quinones, 201
regulations, 8, 82, 125, 208, 244, 281
rejection, 114, 178
R relationship, 3, 9, 18, 31, 41, 69, 70, 75, 104, 107,
119, 202, 207, 241
race, 237 relationships, 15, 52, 59, 69
radiation, 73, 87 reliability, 123
rail, 229 remediation, 43, 193
rain, 59, 136, 302 renal, 310
rain forest, 302 renal failure, 310
rainfall, 230, 301, 302 renal function, 310
rainwater, 79 renewable energy, 136, 237, 244
random, xv, 4, 279 reproduction, 154
range, ix, x, 2, 8, 10, 12, 14, 17, 18, 25, 49, 50, 51, reserves, 112
56, 70, 74, 81, 88, 90, 95, 96, 122, 123, 125, 136, reservoirs, x, 53, 59, 78, 111, 113, 114, 246
138, 139, 140, 141, 144, 146, 173, 207, 210, 217, residential, 136, 310
220, 222, 227, 236, 245, 252, 273, 285, 326, 338, residuals, 107, 108
339, 342 residues, 82, 107, 220, 226, 239, 240, 246
rangeland, 246 resins, 61, 62, 333, 334
rare earth, 50, 65 resistance, 10, 25, 27, 77, 101, 140, 222, 240, 242,
rare earth elements, 65 245, 246, 326
resolution, 130
362 Index
resources, ix, x, xi, xv, 2, 81, 82, 111, 123, 177, 299, sea urchin, 121, 122
300, 301, 302, 304, 305, 307 search, 51, 59, 82, 307
respiration, 52, 53, 56, 76, 163, 239 searching, 83, 301
respiratory, 140, 239 seawater, xvi, 7, 113, 114, 115, 120, 121, 126, 127,
restaurant, 11 309, 312, 314, 315
restaurants, 239 security, 306
retention, 27, 139, 150, 158, 164, 168, 173, 239, 251, sediment, 115, 122, 125, 126, 207
252, 254, 257, 259, 321, 338, 339 sedimentation, 5, 15, 61, 65, 83, 107, 113, 114, 204,
Reynolds, 15, 347 208, 252, 253
Reynolds number, 15 sediments, 53, 115, 122, 127, 151, 207, 211, 334
rice, 304 seed, 46, 253, 267, 321
rings, 142 seeds, 44, 45
risk, 130, 170, 230, 233, 236, 237, 238, 240, 241, seeps, 223
242, 243 selecting, 88, 123, 339
risks, xiv, 83, 112, 124, 125, 236, 240, 243, 244, 246 selectivity, 320
river systems, 303 SEM, 87, 92, 98
rivers, 199, 233, 300, 302, 303, 304, 305, 307 semiarid, 245, 246
rods, 5, 151 semi-arid, 245
Rome, 111, 316 semi-arid, 246
room temperature, 87, 108, 146, 178, 202, 321 semi-arid, 246
room-temperature, 87 semi-arid, 315
Royal Society, 47 semiconductor, ix, 2, 11
rubber, 58, 61 Senegal, 301
runoff, 64, 127, 236, 238 sensitivity, 67, 139
rural, 250, 262, 304, 305, 306, 308 sensors, 223, 224, 225
rural areas, 305 separation, 16, 18, 25, 46, 61, 64, 114, 115, 203, 204,
rural communities, 304, 306 207, 211, 253, 315, 317
rust, 211 sequencing, 241
series, 7, 26, 47, 50, 121, 154, 207, 222, 223, 224,
225, 236, 294, 321
S settlers, 168
sewage, xiii, xv, 7, 12, 49, 50, 51, 53, 57, 58, 59, 60,
safe drinking water, 306
61, 62, 63, 64, 66, 67, 69, 70, 71, 72, 73, 74, 75,
safety, 84
77, 82, 107, 108, 109, 149, 176, 206, 221, 232,
saline, 11, 18, 72, 317, 321
233, 235, 236, 237, 239, 240, 241, 242, 243, 244,
salinity, ix, 1, 126, 128, 177, 178
245, 246, 265, 266, 268, 273, 278, 304, 306, 311
salt, 23, 148, 184, 200, 207, 225
Shanghai, 86
salts, 23, 61, 62, 83, 126, 177, 184, 201, 204, 212
shape, 11, 57, 88, 91
saltwater, 120
shaping, 265
sample, xvi, 4, 5, 87, 221, 224, 225, 228, 232, 256,
sharing, 251, 252, 254, 256, 257
267, 284, 333, 345, 346
shear, 2, 3, 120, 128, 129
sampling, xv, 20, 40, 41, 112, 121, 125, 126, 130,
shelter, 304
225, 255, 279, 345
shock, 252
sand, 11, 14, 18, 64, 82, 109, 208, 226, 330
short period, 204
sanitation, 59, 250, 306, 308, 315
shortage, 300, 303, 315, 316
saturation, 15, 160, 324, 329
short-term, 67, 77, 84
Saudi Arabia, 1, 279, 300
sign, 2, 119, 220
savings, 164, 211, 310, 314, 315
silica, x, 12, 81, 92, 106, 109, 330
sawdust, 262, 330, 331
silicate, x, 31, 81, 85, 87, 88, 90, 91, 97, 100, 102,
SBR, 176, 178
106, 109
scaling, 294
silicates, 30, 98, 100, 102, 103, 104
Scanning electron, 87
silver, 65
scarcity, 299, 300, 307, 310, 315, 316
similarity, 15
sea level, 231
Index 363
simulations, 123, 130, 313, 315 speed, 51, 64, 83, 87, 94, 126, 128, 192, 200, 201,
sintering, x, 82, 83, 84, 85, 88, 89, 91, 92, 95, 98, 210, 251, 259
100, 101, 102, 103, 104, 106, 109, 136 sperm, 121
SiO2, x, 39, 81, 83, 85, 86, 88, 89, 90, 91, 92, 93, 94, spheres, 5
97, 102, 106, 109, 110 spills, 251
SIR, 227 spin, xiii, 215, 216
sites, 8, 50, 79, 237, 240, 255, 324, 326 spindle, 140
skeleton, 88 spore, 321
skills, 123 springs, 53
skin, 93, 154, 305 SPSS, xv, 279, 283
slag, 64, 82 Sri Lanka, 47
smelters, 62, 66 SRT, 321
smelting, 320 stability, xiii, 2, 4, 5, 10, 11, 16, 91, 92, 95, 101, 103,
SO2, xi, 96, 135, 136, 137, 138, 139, 147, 148, 151 106, 108, 128, 129, 216, 217, 220, 224, 226, 239
sodium, 31, 57, 62, 85, 87, 92, 114, 139, 225, 226 stabilization, x, 67, 82, 84, 86, 101, 104, 107, 109,
sodium hydroxide, 62, 225, 226 238, 307, 308
software, 87, 123 stages, 61, 63, 64, 65, 68, 69, 71, 74, 90, 147, 188,
soil, xi, xiii, xiv, 36, 50, 54, 66, 80, 82, 109, 135, 198, 281
142, 154, 185, 199, 201, 203, 230, 234, 235, 236, standard deviation, 203
237, 238, 239, 240, 241, 242, 243, 244, 245, 246, standards, xiv, xv, 136, 249, 265, 266, 271, 280, 281,
247, 281, 303, 316 287, 306, 311, 312, 313, 315, 337
soil erosion, 303 Standards, 94, 290, 296, 336
soils, 53, 54, 57, 80, 109, 199, 241, 242, 245, 246 statutory, 222
solar, 141 steady state, 14
solid phase, 64, 203, 226, 324, 326 steel, 10, 53
solid waste, 50, 51, 52, 63, 66, 71, 74, 82, 109, 198, sterile, 311, 321
199, 277, 304 sterols, 334
solidification, 101, 102, 104, 106, 107 stock, 49, 322, 335, 341
sols, 2 stoichiometry, 69
solubility, 16, 39, 65, 104, 106, 110, 115, 288 storage, 49, 199, 208, 240, 321
solvent, 50, 137, 320, 346 storms, 154
solvents, 50, 201, 226 strain, 54, 75, 77, 79, 145, 151, 181, 182, 183, 184,
Somalia, 300 185, 187, 191, 194, 195
sorption, xvi, 66, 108, 234, 320, 322, 323, 324, 326, strains, xii, 54, 58, 76, 141, 179, 182, 183, 185, 189,
330, 331 190, 191, 242
sorption isotherms, 322 strategies, 14, 159, 160, 170, 173, 243, 307, 347
sorption kinetics, 323 stratification, 112, 119, 120, 127, 128, 130
South Africa, 282 streams, xiv, 12, 66, 76, 142, 147, 149, 152, 208,
sovereignty, 301 253, 265, 266, 268, 269, 271, 302, 304, 330
soy, 311 strength, x, 2, 4, 31, 81, 83, 84, 87, 88, 91, 94, 95,
Spain, 153, 197, 198, 202, 204, 205, 208, 212, 213, 96, 98, 100, 102, 106, 176, 193
309, 333, 341 Streptomyces, 185
spatial, 59, 113, 125, 126 stress, 148, 150, 300
speciation, 16, 83, 122 stroke, 231
species, xiii, 5, 7, 16, 33, 54, 56, 57, 68, 69, 72, 107, students, 296
121, 122, 126, 145, 146, 152, 175, 183, 200, 235, subgroups, 269
238, 241, 245, 250, 280, 313 Sub-Saharan Africa, 300, 306
specific adsorption, 5 substances, x, 13, 48, 49, 51, 57, 81, 82, 91, 96, 97,
specific gravity, 87 115, 116, 117, 126, 127, 132, 144, 180, 195, 198,
specific surface, 7 205, 219, 221, 222, 227, 228, 234, 269, 313, 334,
spectrophotometry, 311 345, 347
spectroscopy, xvi, 333 substitutes, 83
spectrum, ix, xiv, 57, 87, 265, 267, 329, 337, 338 substitution, 100, 106, 271
364 Index
toluene, 54, 58, 75, 76, 80, 121, 126, 146, 149, 268
tomato, 303
U
total energy, 136
Uganda, xv, 299, 301, 303, 304, 305, 306, 307, 308
total organic carbon, 107, 243, 267
ultraviolet, 281
total organic carbon (TOC), 267
ultraviolet irradiation, 281
toxic, xi, xvi, 7, 11, 51, 52, 56, 60, 66, 67, 69, 71, 74,
UNDP, 300, 301
79, 104, 111, 112, 113, 120, 122, 126, 130, 141,
UNEP, 281, 297, 304, 305
142, 147, 154, 173, 175, 201, 219, 239, 240, 243,
UNESCO, 283, 297
244, 269, 319, 320, 334, 347
UNICEF, 306, 308
toxic effect, xi, 111, 112, 120, 122, 126, 130, 154
uniform, 252
toxic metals, 67, 104
United Arab Emirates, 300
toxic products, 201
United Kingdom, 203, 297, 334, 347
toxicity, xii, 12, 67, 84, 87, 114, 120, 121, 126, 146,
United Nations, 311, 312, 316
152, 170, 175, 197, 209, 271, 276, 277, 278, 281
United States, 7, 312
toxins, 305
universities, 216
toys, 334
uranium, 79
trace elements, 114, 320
urban areas, 303, 304, 306
tracers, 121, 126, 127
urban centres, 306
tracking, 304
urbanized, 250
trade, 42, 216
urine, 236, 240
trademarks, 202
USEPA, 87, 121, 133, 237, 334
trade-off, 42
trading, 280
trajectory, 120, 125, 128 V
transfer, 11, 13, 103, 140, 154, 201, 242, 243, 245,
246, 251, 252, 253, 267, 326 valence, 4, 92, 106
transformation, 50, 65, 66, 71, 75, 91, 92, 96, 97, validation, 231
102, 103, 114, 344, 346 values, 4, 9, 15, 17, 60, 89, 94, 104, 119, 128, 129,
transformations, 68, 119, 154, 241 141, 158, 180, 201, 203, 204, 206, 207, 208, 210,
transition, 90, 201 211, 232, 240, 243, 266, 267, 286, 287, 300, 315,
transmission, 238 321, 324
transparent, 211 van der Waals, 5, 31
transport, 5, 55, 56, 61, 66, 112, 114, 116, 117, 122, van der Waals forces, 31
123, 124, 126, 272, 277 vanadium, 65
transport processes, 123 vapor, 31, 216, 218
transportation, xiv, 42, 65, 249, 261, 280, 333 variability, 16, 127, 213, 322
traps, 302 variables, 25, 33, 41
travel time, 12 variance, 283
treatment methods, xiii, 12, 50, 59, 215 variation, 18, 88, 92, 106, 147, 207, 210, 287, 344
trial, 239 vegetables, 303
tribal, 307 vegetation, 198, 199, 301, 302, 303
tribes, 54, 55, 71 velocity, 114, 115, 118, 119, 120, 144, 210, 252, 290
trichloroethylene, 201, 212 versatility, 136
tropical areas, 180 vessels, 58, 60
tumours, 154 Victoria, 303, 304, 305, 308
Tunisia, 300 village, 250
turbulence, 16, 114, 115, 116, 120, 124 vinasse, 48
turbulent, 115, 118 vinyl chloride, 334
Turbulent, 131 violence, 304
turbulent mixing, 118 virulence, 240
Turkey, 265, 280, 319 viruses, 9, 237
Turku, 80 viscosity, 3, 93, 95, 98
vitamins, 60
366 Index