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Landfills are the primary option for waste disposal all over the world. Most of the landfill sites across the world are old and are not engineered to prevent contamination of the underlying soil and groundwater by the toxic leachate. The pollutants from landfill leachate have accumulative and detrimental effect on the ecology and food chains leading to carcinogenic effects, acute toxicity and genotoxicity among human beings. Management of this highly toxic leachate presents a challenging problem to the regulatory authorities who have set specific regulations regarding maximum limits of contaminants in treated leachate prior to disposal into the environment to ensure minimal environmental impact. There are different stages of leachate management such as monitoring of its formation and flow into the environment, identification of hazards associated with it and its treatment prior to disposal into the environment. This review focuses on: (i) leachate composition, (ii) Plume migration, (iii) Contaminant fate, (iv) Leachate plume monitoring techniques, (v) Risk assessment techniques, Hazard rating methods, mathematical modeling, and (vi) Recent innovations in leachate treatment technologies. However, due to seasonal fluctuations in leachate composition, flow rate and leachate volume, the management approaches cannot be stereotyped. Every scenario is unique and the strategy will vary accordingly. This paper lays out the choices for making an educated guess leading to the best management option.
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Critical Reviews in Environmental Science and
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Contemporary environmental issues of landfill
leachate: assessment & remedies
Sumona Mukherjee
a
, Soumyadeep Mukhopadhyay
b
, Mohd Ali Hashim
b
& Bhaskar Sen Gupta
c
a
Institute of Biological Sciences, University of Malaya, 50603, Kuala Lumpur, Malaysia
b
Department of Chemical Engineering, University of Malaya, 50603, Kuala Lumpur, Malaysia
c
School of Planning, Architecture and Civil Engineering, Queen's University Belfast, David
Keir Building, Belfast, BT9 5AG, UK
Accepted author version posted online: 12 May 2014.Published online: 12 May 2014.
To cite this article: Sumona Mukherjee, Soumyadeep Mukhopadhyay, Mohd Ali Hashim & Bhaskar Sen Gupta (2014):
Contemporary environmental issues of landfill leachate: assessment & remedies, Critical Reviews in Environmental Science
and Technology, DOI: 10.1080/10643389.2013.876524
To link to this article: http://dx.doi.org/10.1080/10643389.2013.876524
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ACCEPTED MANUSCRIPT
1
Contemporary environmental issues of landfill leachate: assessment & remedies
Sumona Mukherjee
1
, Soumyadeep Mukhopadhyay
2
, Mohd Ali Hashim
2
, Bhaskar Sen Gupta
3
*
Abstract
Landfills are the primary option for waste disposal all over the world. Most of the landfill sites
across the world are old and are not engineered to prevent contamination of the underlying soil
and groundwater by the toxic leachate. The pollutants from landfill leachate have accumulative
and detrimental effect on the ecology and food chains leading to carcinogenic effects, acute
toxicity and genotoxicity among human beings. Management of this highly toxic leachate
presents a challenging problem to the regulatory authorities who have set specific regulations
regarding maximum limits of contaminants in treated leachate prior to disposal into the
environment to ensure minimal environmental impact. There are different stages of leachate
management such as monitoring of its formation and flow into the environment, identification of
hazards associated with it and its treatment prior to disposal into the environment. This review
focuses on: (i) leachate composition, (ii) Plume migration, (iii) Contaminant fate, (iv) Leachate
plume monitoring techniques, (v) Risk assessment techniques, Hazard rating methods,
mathematical modeling, and (vi) Recent innovations in leachate treatment technologies.
However, due to seasonal fluctuations in leachate composition, flow rate and leachate volume,
the management approaches cannot be stereotyped. Every scenario is unique and the strategy
will vary accordingly. This paper lays out the choices for making an educated guess leading to
the best management option.
Keywords: landfill leachate plume, pollution, hazard identification, treatment technologies
1
Institute of Biological Sciences, University of Malaya, 50603, Kuala Lumpur, Malaysia
2
Department of Chemical Engineering, University of Malaya, 50603, Kuala Lumpur, Malaysia
3

BT9 5AG, UK
* Corresponding Author: Dr Bhaskar Sen Gupta; School of Planning, Architecture and Civil Engineering, Qu
University Belfast, Stranmillis Road, David Keir Building, Belfast, BT9 5AG, UK; Phone: +44 78461 12581;
Email: B.Sengupta@qub.ac.uk
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Contents
Contemporary environmental issues of landfill leachate: assessment & remedies ................... 1
Abstract .................................................................................................................................. 1
1 Introduction ........................................................................................................................ 4
2 Landfill leachate: Characteristics and regulatory limits .................................................... 7
3 Leachate plume migration and methods of its monitoring .............................................. 10
3.1 Fate of contaminants in leachate plume .................................................................... 11
3.1.1 Inorganic pollutants ........................................................................................... 12
3.1.2 Organic contaminants ........................................................................................ 16
3.1.3 Biological contaminants..................................................................................... 18
3.2 Monitoring of plume generation and migration: techniques & methodology ........... 19
3.2.1 Hydro-geological techniques for groundwater sampling for geo-chemical analysis
20
3.2.2 Use of stable isotopes to monitor landfill leachate impact on surface waters ... 21
3.2.3 Electromagnetic methods ................................................................................... 24
3.2.4 Electrical methods .............................................................................................. 26
3.2.5 Monitoring the fate of dissolved organic matter (DOM) in landfill leachate .... 30
4 Environmental impact of landfill leachate and its assessment ......................................... 32
4.1 Environmental impact ............................................................................................... 33
4.1.1 Effects on groundwater ...................................................................................... 33
4.1.2 Reduction of soil permeability and modification of soil ................................... 35
4.1.3 Effects on surface water ..................................................................................... 39
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4.2 Hazard assessment of landfill leachate...................................................................... 39
4.2.1 Relative hazard assessment systems .................................................................. 40
4.2.2 Deterministic and stochastic models for monitoring environmental impact of
landfill leachate ................................................................................................................ 47
5 Recent technological developments for landfill leachate treatment and remediation ..... 52
5.1 Application of natural attenuation for leachate remediation ..................................... 54
5.2 Application of biological and biochemical techniques in reactors ........................... 56
5.3 Application of physical and chemical processes for leachate treatment ................... 62
5.3.1 Advance Oxidation Treatments ......................................................................... 62
5.3.2 Adsorption.......................................................................................................... 66
5.3.3 Coagulation-flocculation.................................................................................... 69
5.3.4 Electrochemical treatment ................................................................................. 71
5.3.5 Filtration and membrane bioreactors ................................................................. 73
6 Summary and Discussion ................................................................................................. 75
Acknowledgements .............................................................................................................. 79
References ................................................................................................................................ 80
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1 Introduction
Landfill leachate is defined as any liquid effluent containing undesirable materials percolating
through deposited waste and emitted within a landfill or dump site. Often, its route of exposure
and toxicity remains unknown and a matter of prediction due to extremely complicated
geochemical processes in the landfill and the underlying soil layers (Koshi et al., 2007; Taulis,
2005). The prevalence of landfill waste dumping with or without pre-treatment is on the rise
around the globe due to increasing materialistic lifestyle and planned obsolescence of the
products. According to Laner et al. (2012), in 2008 up to 54% of the 250x10
6
metric tons of
municipal solid waste (MSW) in USA was disposed off in landfills. Also, 77% MSW in Greece,
55% MSW in the United Kingdom, and 51% MSW in Finland was landfilled in 2008 while
about 70% of MSW in Australia has been directed to landfills without pre-treatment in 2002
(Laner et al., 2012). In Korea, Poland and Taiwan around 52%, 90% and 95% of MSW are
dumped in landfill sites, respectively (Renou et al., 2008a). In India, the accumulated waste
generation in four metropolitan cities of Mumbai, Delhi, Chennai and Kolkata is about 20,000
tons d
-1
and most of it is disposed in landfills (Chattopadhyay et al., 2009). Most of the landfill
sites across the world are old and are not engineered to prevent contamination of the underlying
soil and groundwater by the toxic leachate.
Leachate presents high values of biochemical oxygen demand (BOD), chemical oxygen demand
(COD), total organic carbon (TOC), total suspended solid (TSS), total dissolved solid (TDS),
recalcitrant organic pollutants, ammonium compounds, sulfur compounds and dissolved organic
matter (DOM) bound heavy metals which eventually escape into the environment, mainly soil
and groundwater, thereby posing serious environmental problems (Gajski et al., 2012; Lou et al.,
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2009). Around two hundred hazardous compounds have already been identified in the
heterogeneous landfill leachate, such as aromatic compounds, halogenated compounds, phenols,
pesticides, heavy metals and ammonium (Jensen et al., 1999). All of these pollutants have
accumulative, threatening and detrimental effect on the survival of aquatic life forms, ecology
and food chains leading to enormous problems in public health including carcinogenic effects,
acute toxicity and genotoxicity (Gajski et al., 2012; Moraes and Bertazzoli, 2005; Park and
Batchelor, 2002). Broadly speaking, landfill leachate has deep impact on soil permeability,
groundwater, surface water, and nitrogen attenuation all of which will be discussed in Section
4.1.
A leachate is characterized by two principle factors viz., its composition and the volume
generated, both of which are influenced by a variety of parameters, such as type of waste,
climatic conditions and mode of operation. The most important factor influencing landfill
leachate composition is the age of the landfill (Kulikowska and Klimiuk, 2008; Nanny and
Ratasuk, 2002). The regulatory bodies around the world have set specific maximum discharge
limits of treated leachate that has to be maintained prior to the disposal of treated leachate into
any surface water bodies, sewer channels, marine environment or on land to ensure minimal
environmental impact. These are discussed in the Section 2. Monitoring of the contaminated
leachate plume is an arduous but essential task necessary for measuring the extent of spread of
pollution and taking management decisions regarding leachate treatment. A number of
techniques have been followed for the past three decades for leachate plume migration
monitoring, such as hydro-geological techniques for groundwater sampling for geo-chemical
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analysis, use of stable isotopes, electromagnetic methods, electrical methods and bacteriological
experiments, all of which will be discussed in details in Section 3.2.
Assessing the effect of leachate on the environment needs systematic study procedure. The task
is extremely difficult and largely prediction based, due to unpredictability of the soil
environment, groundwater flow and variation of soil permeability in different parts of the world.
However, an educated guess can be taken on the pollution scenario and risk assessment can be
done either by using relative hazard assessment systems or by using stochastic and deterministic
models after gathering background physico-chemical data. Softwares are also used for this
purpose. Section 4.2 describes the procedure of risk assessment of landfill leachate.
Once the landfill leachate plume is monitored and risk assessment has been performed, then the
management decision regarding leachate treatment can be taken. Already some comprehensive
reviews on various leachate treatment technologies have been published (Alvarez-Vazquez et al.,
2004; Deng and Englehardt, 2006; Foo and Hameed, 2009; Kim and Owens, 2010; Kurniawan et
al., 2006b; Laner et al., 2012; Renou et al., 2008a; Wiszniowski et al., 2006). So we have
included a brief but detailed description of only the most recent developments in this field,
mainly in tabular form in Section 5 (Tables 6-12).
This review elucidates the complete leachate management process, beginning with leachate
composition, plume migration, fate of contaminant, plume monitoring techniques, risk
assessment techniques, hazard assessment methods, mathematical modeling up to the recent
innovations in leachate treatment technologies. This paper also steers clear from the topics in
which good reviews are already available and only the most relevant information has been
included.
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2 Landfill leachate: Characteristics and regulatory limits
Landfill leachate can be categorized as a soluble organic and mineral compound generated when
water infiltrates into the refuse layers, extracts a series of contaminants and triggers a complex
interplay between the hydrological and biogeochemical reactions (Renou et al., 2008a). These
interactions act as mass transfer mechanisms for producing moisture content sufficiently high to
initiate a liquid flow (Aziz et al., 2004a), induced by gravitational force, precipitation, surface
runoff, recirculation, liquid waste co-disposal, groundwater intrusion, refuse decomposition and
initial moisture content present within the landfills (Achankeng, 2004; Foo and Hameed, 2009).
The knowledge of leachate characteristics at a specific landfill site is the most essential
requirement for designing management strategy. This knowledge is equally important for
designing containment for new landfill where leachate will be extracted, as well as for managing
the old landfill that lacks proper safeguards installed to contain leachate (Rafizul and Alamgir,
2012). Typical composition of a municipal landfill leachate is given in Table 1.
Two most important factors for characterizing leachate are volumetric flow rate and its
composition. Leachate flow rate depends on rainfall, surface run-off, and intrusion of
groundwater into the landfill (Renou et al., 2008a). According to a number of researchers (Baig
et al., 1999; Christensen et al., 2001; El-Fadel et al., 2002; Harmsen, 1983; Nanny and Ratasuk,
2002; Rapti-Caputo and Vaccaro, 2006; Rodríguez et al., 2004; Stegman and Ehrisg, 1989),
leachate composition is influenced by a number of factors viz., ( i) climatic and hydro-geological
conditions (rainfall, groundwater intrusion, snowmelt); (ii) operational and management issues at
the landfill (compaction, refuse pre-treatment, vegetation cover, re-circulation, liquid waste co-
disposal, etc.); (iii) characteristics of waste dumped in the landfill (particle size, density,
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chemical composition, biodegradability, initial moisture content); (iv) internal processes inside
landfill (decomposition of organic materials, refuse settlement, gas and heat generation and their
transport); (v) age of the landfill. The leachate quality varies, not only from landfill to landfill but
also, between different sampling points at the same landfill site from time to time due to the
variation in the above factors.
Among all the above factors, leachate characterization depending on age may be used for making
initial management decisions since others are too complex to estimate instantly. Although
leachate composition may vary widely within the successive aerobic, acetogenic, methanogenic,
stabilization stages of the waste evolution, four types of leachates can be defined according to
landfill age viz., young, intermediate, stabilized and old as shown in Table 2. However, detailed
management decision may be taken only after considering all the above factors.
The characteristics of the landfill leachate can usually be represented by the basic parameters
COD, BOD
5
, BOD
5
/COD ratio, pH, suspended solids (SS), ammonium nitrogen (NH
4
-N), total
Kjeldahl nitrogen (TKN) and heavy metals. The landfill age was found to have significant effect
on organics and ammonia concentrations (Kulikowska and Klimiuk, 2008). The concentration
and biodegradability of leachate usually decrease with its age. Young leachate fractions have low
molecular weight organic compounds characterized by linear chains, which are substituted
through oxygenated functional groups such as carboxyl and alcoholic groups. Old leachate have
organic compounds with a wide range of molecular weight fractions having complex structures
with N, S and O containing functional groups (Calace et al., 2001). Hence, the management
decision can be generalized and the treatment approach can be chalked out depending on the age
of the landfill.
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Landfill leachates cause enormous harm when they get released into the environment without
proper treatment, as will be discussed in section 4.1. In order to minimize their environmental
impact, regulatory bodies around the world require that the leachate volume is controlled and its
toxicity and contaminant level reduced by using proper treatment technologies (Robinson, 2005).
The regulatory limits of various leachate components in different countries are discussed in
Table 3. India, has specific regulations regarding construction, maintenance and operation of a
landfill and the post closure steps required to be taken for pollution prevention under Schedule
III of the Municipal Solid Wastes (Management and Handling) Rules, 2000. The recent stricter
discharge limits for leachate demands the application of advanced treatment techniques such as
electrochemical treatments, membrane filtrations, advanced oxidations and so on, all of which
involve high installation and operational cost. According to a World Bank (1999) study,
equipment donated by bilateral organizations remains idle due to lack of training or funds for
operation. The regulatory authorities managing landfills inspect the incoming waste but are not
very observant towards the environmental impacts of waste disposal, which results in poor
enforcement of the discharge standards (The World Bank, 1999). The increased private sector
participation in leachate management can lead to better enforcement of standards. Better
incentives such as low taxes, institutional support etc., can draw private sector companies to the
field of leachate management.
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3 Leachate plume migration and methods of its monitoring
It is a well established fact that leachate plumes are formed from landfills with or without liners
and these infiltrate into subsurface aquifers, subsequently forming an even larger plume (Baun et
al., 2004; Bloor et al., 2005; Isidori et al., 2003; Kjeldsen et al., 2002; Slack et al., 2005). The
processes associated with leachate plume formation has also been discussed by other researchers
(Kjeldsen et al., 2002). Leaching tests designed to assess the release of toxic leachate from a
solid waste into the surrounding environment has been earlier reviewed (Scott et al., 2005). A
large number of research has already been done to study the migration of leachate plume through
landfill liners (Baun et al., 2003; Chalermtanant et al., 2009; Edil, 2003; Haijian et al., 2009; Lu
et al., 2011; Varank et al., 2011). Two distinctive routes of landfill leachate transport were
identified by some researchers (Foose et al., 2002; Katsumi et al., 2001). The first route is the
advective and dispersive transport of contaminants through defects in the geomembrane seams
and through clay liner underlying the geomembrane. The second route is the diffusive transport
of organic contaminants through the geomembrane and the clay liner. It was reported that every
10,000 m
2
of geomembrane liner contains 22.5 leaks on an average facilitating the leachate
plume formation (Laine and Darilek, 1993). Chofqi et al. (2004) deduced that there were several
factors that determine the evolution of groundwater contamination, such as (1) depth of the water
table, (2) permeability of soil and unsaturated zone, (3) effective infiltration, (4) humidity and (5)
absence of a system for leachate drainage. Leachate plumes often contain high concentrations of
organic carbon such as volatile fatty acids, humic like compounds and fulvic acids (Christensen
et al., 2001), ammonium (Christensen et al., 2000) and a variety of xenobiotic compounds (e.g.
BTEX compounds, phenoxy acids, phenolic compounds, chlorinated aliphatic compounds and a
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variety of pesticides) (Baun et al., 2004; Kjeldsen et al., 2002). Non-volatile dissolved organic
carbon (DOC), ferrous iron, methane, ammonium, sulfate, chloride, and bicarbonate are also
present in the leachate plume 10500 times higher than natural aquifer conditions (Bjerg et al.,
2003; Christensen et al., 2001).
3.1 Fate of contaminants in leachate plume
The generation of leachate plume depends upon the quantity and quality of leachate, which
varies seasonally depending upon the composition and moisture content of the solid waste,
hydro-geological conditions, climate, local population densities, annual precipitation,
temperature and humidity. All these factors add to the complexity in landfill leachate
characteristics and composition (Christensen et al., 2001; Miyajima et al., 1997). The
contaminant migration greatly depends upon the composition of the leachate or contaminants
entering the ground-water system. Similar contaminants may behave differently in the same
(Abu-Rukah
and Al-Kofahi, 2001). Redox environments were found to vary greatly inside contaminant
plumes due to variation in contaminant load, groundwater chemistry, geochemistry and
microbiology along the flow path (Christensen and Christensen, 2000; van Breukelen et al.,
2003). Existence of redox gradients from highly reduced zones at the source to oxidized zones
towards the front of the plumes was supported by detailed investigation of the terminal electron
acceptor processes (Bekins et al., 2001; Ludvigsen et al., 1999). Some researchers also studied
the steep vertical concentration gradients for contaminants and redox parameters in plume
fringes, where contaminants mix with electron acceptors by dispersion and diffusion processes
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(Lerner et al., 2000; Thornton et al., 2001; van Breukelen and Griffioen, 2004). The fates of
nitrogenous, sulfurous, heavy metals and organic contaminants are discussed under different
paragraphs.
3.1.1 Inorganic pollutants
3.1.1.1 Nitrogenous pollutants
The landfill leachate having NH
4
poses long-term threat of pollution once it escapes into ground
or surface waters (Beaven and Knox, 2000; IoWM, 1999). In the UK, average concentrations of
about 900 mg NH
4
(+NH
3
)N L
-1
have been reported for landfill leachates (Burton and Watson-
Craik, 1998) while legislation probably requires concentrations below 0.5 mg NH
4
N L
-1
for any
discharge in the environment (EA, 2003). The laboratory experiments revealed that most
biological nitrogen removal processes are carried out by the combination of aerobic nitrification,
nitrate reduction, anoxic denitrification and anaerobic ammonium oxidation processes or
(anammox) (Fux et al., 2002; Jokella et al., 2002; Pelkonen et al., 1999). The NH
4
+
in leachate
can undergo sequential bacterial transformation to NO
3
-
under oxidizing environment. Although
NO
3
-
is less toxic than NH
4
+
it still presents a pollution threat and bacterial denitrification to

2
is required under anaerobic conditions, to eliminate it. When oxygen is depleted,
nitrate can be converted to nitrite and finally to nitrogen gas by denitrification. Also, when nitrite
is present under anaerobic conditions, ammonium can be oxidized with nitrite as an electron
acceptor to dinitrogen gas (anammox) (Mora et al., 2004). The attenuation of N pollution
resulting from disposal of organic wastes in landfill sites therefore requires fluctuating redox
conditions favouring the transformations:
NH4+ NO3- N2
. Anaerobic conditions prevent
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the formation of NO
3
-
, so N attenuation by denitrification in landfills is not regarded as a
significant process (Burton and Watson-Craik, 1998). Heaton et al. (2005) acquired data for the
isotope ratios (
13
C/
12
C,
15
N/
14
N and
34
S/
32
S) and dissolved gas (N
2
, Ar, O
2
and CH
4
) composition
of groundwater in and around a landfill site in Cambridgeshire, England. Decomposition of
domestic waste, placed in unlined quarries produced NH
4
+
rich leachate dispersing as a plume
into the surrounding middle chalk aquifer at approximately 20 m below ground level. Few
boreholes around the edge of the landfill extending to the west and north in the direction of
plume flow showed evidence of methanogenesis, SO
4
2-
reduction, and denitrification. The first
two processes are indicative of strongly reducing conditions, and are largely confined to the
leachate in the landfill area. Denitrification does not require such strong reducing conditions and
beyond those strong reducing zones, clear evidence of denitrification comes from data for
elevated
15
N values for NO
3
-
-atmospheric N
2
. This
distribution of redox zones is therefore consistent with an environment in which conditions
become progressively less reducing away from the landfill (Christensen et al., 2001; Heaton et
al., 2005).
3.1.1.2 Reduction of sulfate pollutants
Sulfate reduction is a major process for degradation of organic matters and many anaerobic
subsurface environments have been found to experience this process (Krumholz et al., 1997;
Lovley, 1997; Ulrich et al., 1998). The sulfate reduction is controlled by factors such as
availability of utilizable organic matter as electron donors (McMahon and Chapelle, 1991; Ulrich
et al., 1998), water potential, sediment pore throat diameter, pH and availability of
thermodynamically more favorable electron acceptors (Ludvigsen et al., 1998; Routh et al.,
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2001). In anoxic aquifers, lithologic, climatic, hydrological, and biogeochemical processes
controlling the sulfate supply may determine sulfate reduction (Martino et al., 1998; Ulrich et al.,
1998). Ulrich et al. (2003) undertook field and laboratory techniques to identify the factors
affecting sulfate reduction in a landfill leachate contaminated shallow, unconsolidated alluvial
aquifer. Depth profiles of
35
S-sulfate reduction rates in aquifer sediments revealed a
-like relationship with an apparent K
m
and V
max
of approximately 80 and 0.83

4
-2
day
-1
, respectively. The rate of sulfate reduction was in direct correlation with the
concentration of the sulfate. Near the confining bottom layer of the aquifer, sulfate was supplied
by advection of groundwater beneath the landfill and the reduction rates were significantly
higher than rates at intermediate depths (Ulrich et al., 2003).
3.1.1.3 Heavy Metals (HMs)
Although HMs tend to be leached out of fresh landfill, they later became largely associated with
MSW-derived dissolved organic matter (DOM) which plays an important role in heavy metal
speciation and migration (Baumann et al., 2006; Baun and Christensen, 2004; Li et al., 2009).
Christensen et al. (1996) conducted experiments to determine the metal distribution between the
aquifer material and the polluted groundwater samples (K
d
) and the difference in distribution
coefficients indicated that DOC from landfill leachate polluted groundwater can form complexes
with Cd, Ni and Zn. DOM derived from MSW landfill leachate was observed to have a high
affinity for metals such as Cu, Pb, Cd, Zn and Ni, enhancing their mobility in leachate-polluted
waters (Christensen et al., 1999). However, Ward et al. (2005) deduced that the heavy metal
binding capacities largely fluctuated among various leachates due to variable compositions.
Earlier, it was demonstrated that HMs mobilization was enhanced by reduced pH of the leachate
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with oxygen intrusion in landfill 
and by the presence of large quantity of fatty acids generated at the initial phase of solid waste
degradation (He et al., 2006). In some recent studies, it was revealed that less than 0.02% of
HMs in landfills may leach out over 30 years of land filling (Kjeldsen et al., 2002; Øygard et al.,
2007). Qu et al. (2008) monitored mobility of some heavy metals including Cd, Cr, Cu, Ni, Pb
and Zn released from a full-scale tested bioreactor landfill (TBL) in the Tianziling MSW Landfill
in Hangzhou City, China over the first 20 months of operation. The size of the TBL was
approximately 16,000 m
2
with a combined GCL-HDPE bottom liner, and had four layers of 6
8 m thick MSW layers. At the initial landfill stage, the leachate exhibited high HMs release, high
organic matter content (27,00043,000 g l
1
of TOC) and low pH (56). By the fifth month of
land filling, the methanogenic stage was established, and HMs release was reduced below the
Chinese National Standards. At a landfill age of 0.5 years, 15% of Cr, 25% of Cu, 14% of Ni,
30% of Pb and 36.6% of Zn in solids were associated with amorphous metal oxides and
crystalline Fe oxides. At 1.5 years of filling age, these HMs were largely transformed into
alumino-silicates forms or released with the landfill leachate. Computer modeling revealed that
the humic acid (HA) and fulvic acid (FA) could strongly bind HMs (Qu et al., 2008). Chai et al.
(2012) found strong interactions between HA and Hg. They proposed that the overall stability
constant of Hg(II)HA was determined by the abundant O-ligands in HA. Compared to HA, the
FA having relatively high content of carboxylic groups had a much higher Hg(II)-complexing
capacity. Thus FA played an important role in binding Hg(II) in early landfill stabilization
process.
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3.1.2 Organic contaminants
Organic contaminants in the form of hydrocarbons usually undergoes degradation by bacterial
activity in the vadose zone producing carbonic and organic acids which enhance the mineral
dissolution of the aquifer materials (McMahon et al., 1995). This leads to the production of a
leachate plume with high total dissolved solids (TDS) resulting in the increased groundwater
conductance observed in and around the zones of active biodegradation (Atekwana et al., 2000;
Benson et al., 1997). The acidogenic phase in young landfills is associated with rapid anaerobic
fermentation, leading to the release of free volatile fatty acids (VFA), whose concentration can
be up to 95% of the TOC (Welander et al., 1997). Figure 1 illustrates an anaerobic degradation
scheme for the organic material, measured by COD, inside a sanitary landfill. High moisture
content enhances the acid fermentation in the solid waste (Wang et al., 2003). The methanogenic
phase takes over with the maturity of the landfill. Methanogenic microorganisms converts VFA
into biogas (CH
4
, CO
2
) and in such old landfills, up to 32% of the DOC in leachate consists of
high molecular weight recalcitrant compounds (Harmsen, 1983).
van Breukelen et al. (2003) delineated the leachate plume inside a landfill (Banisveld, The
Netherlands) using geophysical tests by mapping the subsurface conductivity to identify the
biogeochemical processes occurring. Methane was found to form inside the landfill and not in
the plume. Precipitation of carbonate minerals was confirmed by simulation of δ
13
C-DIC
[dissolved inorganic carbon]. Ziyang et al. (2009) investigated the COD compositions in leachate
based on the molecular weight distribution and hydrophobic/hydrophilic partition characteristics
as shown in Figure 2. The COD composition varied over the age of the leachate and the ratio of
TOC/TC decreased over time, indicating decrease in the percentage of organic matters in
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leachate and increase in inorganic substances. Giannis et al. (2008) monitored long-term
biodegradation of MSW in relation to operational characteristics such as air importation,
temperature, and leachate recirculation in an aerobic landfill bioreactor over a period of 510 days
of operation in a lab-scale setup. It was evident from the leachate analysis that above 90% of
COD and 99% of BOD
5
was removed by the aerobic bioreactor. Tuxen et al. (2006) used
microcosm experiments to illustrate the importance of fringe degradation processes of organic
matters within contaminant plumes and identified increased degradation potential for phenoxy
acid herbicide governed by the presence of oxygen and phenoxy acids existing at the narrow
leachate plume fringe of a landfill. Anaerobic processes taking place in a leachate contaminated
alluvial aquifer was studied near Norman Landfill, Oklahama (USA), along the flow path of
aquifer. The center of the leachate plume was characterized by high alkalinity and elevated
concentrations of total dissolved organic carbon, reduced iron, methane, and negligible oxygen,
nitrate, and sulfate concentrations. Occurrence of anaerobic methane oxidation inside the plume
wa
13
C). Methane
13

-order rate
constants ranged from 0.06 to 0.23 per year. Hydro-chemical data suggested a sulfate reducer-
methanogen consortium mediating this methane oxidation. So natural attenuation of organics
through anaerobic methane oxidation was found to be an important process in the plume
(Grossman et al., 2002)
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3.1.3 Biological contaminants
Survival of micro-organisms in groundwater, septic tank and leachate plumes have been
investigated by few researchers (Crane and Moore, 1984; Grisey et al., 2010; Sinton, 1982;
Tuxen et al., 2006). Grisey et al. (2010) monitored total coliforms, Escherichia coli, Enterococci,
Pseudomonas aeruginosa, Salmonella and Staphylococcus aureus for 15 months in groundwater
and leachate beneath the Etueffont landfill (France). They coupled the microbiological tests to
tracer tests to identify the source of contamination. Groundwater was found to have high levels
of faecal bacteria (20,000 CFU 100 mL
1
for total coliforms, 15,199 CFU 100 mL
1
for E. coli
and 3290 CFU 100 mL
1
for Enterococci). Bacterial density was lower in leachates than in
groundwater, except for P. aeruginosa which seemed to adapt favourably in leachate
environment. Tracer tests indicated that bacteria originated from the septic tank of the transfer
station and part of these bacteria transited through waste. Microcosm experiments were used to
measure the fringe degradation of phenoxy acid herbicide across a landfill leachate plume by
microbial activity in lab scale experiments. High spacial resolution sampling at 5 cm interval was
found to be necessary for proper identification of narrow reaction zones at the plume fringes
because samples from long screens or microcosm experiments under averaged redox conditions
would yield erroneous results. The samples were collected by a hollow stem auger drilled down
to the desired level of the cores. The collected cores were sealed with aluminium foil and plastic
stoppers to maintain the redox conditions and stored at 10 °C to be used within 4 days. These
were divided into smaller parts for the microcosm experiments, pore-water extraction, and
sediment analyses, determination of MPN, solid organic matter (TOC), and grain size
distribution. A multi-level sampler installed beside the cores measured the plume position and
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oxygen concentration in the groundwater. Microcosm experiments were performed in 50 mL
sterilized infusion glass bottles, each containing aquifer material from the sediment samples. In
each microcosm, the oxygen concentration was individually controlled to mimic the conditions
at their corresponding depths. The number of phenoxy acid degraders was enumerated by a most
probable number (MPN) method. The results illustrated the importance of fringe degradation
processes in contaminant plumes (Tuxen et al., 2006).
3.2 Monitoring of plume generation and migration: techniques &
methodology
The leachate plume migration have been monitored by using a broad range of techniques and
methods, such as, hydro-geological techniques, electromagnetic techniques, electrical resistivity
and conductivity testing, ground penetrating radars, radioactive tracing systems and microcosm
experiments. Historically, investigations by conventional sampling or electromagnetic methods
were applied only at sites suspected of contamination. However, early detection and monitoring
of leachate plume migration into subsurface is essential for preventing further contamination.
Whatever be the technology, the monitoring wells and their placement is a matter of common
interest, except for electromagnetic techniques. Usually, monitoring wells are constructed at
different depths in and around the landfill site, mostly in the down-gradient of groundwater flow
and the probes and sampling devices are lowered into these wells for measuring various
parameters. This positioning of monitoring wells and a cross section of such a well is shown in
Figure 3. USEPA (2004), in one of its reports, discussed several technologies for detecting the
contaminant leaks in the vadose zone such as advanced tensiometers, cable network sensors,
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capacitance sensors, diffusion hoses, electrochemical wire cables, electrode grids, intrinsic fibre
optics sensors, lysimeters, neutron probes, portable electrical systems, time domain reflectometry
detection cables and wire net designs (USEPA, 2004). Therefore, most of these technologies is
not discussed in this review and the interested readers are advised to access the referred
document. Table 3 gives an overview of the plume monitoring techniques discussed in this
section.
3.2.1 Hydro-geological techniques for groundwater sampling for geo-chemical analysis
The hydro-geological sampling devices had been most frequently used for the past few decades
to collect groundwater samples around leachate plumes to measure and map the plume migration
(Cherry et al., 1983; Chofqi et al., 2004; Christensen et al., 1996; Kjeldsen, 1993; Nicholson et
al., 1983). Cherry et al (1983) used six types of devices for groundwater monitoring to detect
migration of the plume of contamination in the unconfined sandy aquifer at the Borden landfill.
The monitoring devices included (i) standpipe piezometers, (ii) water-table standpipes, (iii) an
auger-head sampler, (iv) suction-type multilevel point-samplers, (v) positive-displacement-type
multilevel point-samplers, and (vi) bundle-piezometers. The last four devices can provide
vertical sample profiles of groundwater from a single borehole. Standpipe piezometers,
multilevel point-samplers and bundle-piezometers were also used by MacFarlane et al. (1983)
for measuring the distribution of chloride, sulfate, electrical conductance, temperature, hydraulic
conductivity, density and viscosity of the leachate & groundwater. The auger-head sampler
yields samples from relatively undisturbed aquifer zones providing a rapid means of acquiring
water-quality profiles for mapping the distribution of a contaminant plume. A suction-type
multilevel sampler consists of twenty or more narrow polyethylene or polypropylene tubes
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contained in a polyvinyl chloride (PVC) casing capped at the bottom. Each tube extends to a
different depth and is attached to a small-screened sampling point that extends through the
casing to draw water from the aquifer of depth of 8 or 9 m when suction is applied. A positive-
displacement multilevel sampler can be used for deeper aquifers since each sampling point is
connected to a positive-displacement pumping device. A bundle-piezometer consists of flexible
polyethylene tubes, fastened as a bundle around a semi-rigid centre-piezometer. In shallow
water-table areas water is withdrawn from each of the tubes and from the PVC piezometer by
suction. In areas with a deep water table, samples are obtained by bailing with a narrow tube
with a check valve on the bottom or by displacement using a double- or triple-tube gas-drive
sampler. Coupling the positive-displacement multilevel sampler or the gas-drive samplers with
the bundle-piezometers is an excellent option for collecting samples that can be filtered and have
preservatives added without the water being exposed to oxygen. The multilevel samplers and
bundle-piezometer can be installed to establish permanent networks for groundwater-quality
monitoring by means of hollow-stem augers in which eight or more polyethylene tubes are
included conveniently in each bundle-piezometer (Cherry et al., 1983).
3.2.2 Use of stable isotopes to monitor landfill leachate impact on surface waters
The uniqueness of isotopic characteristics of municipal landfill leachate and gases (carbon
dioxide and methane) is utilized for monitoring leachate plume migration in groundwater. Few
researchers (Hackley et al., 1996; North et al., 2006; Rank et al., 1995; Walsh et al., 1993)

13
CH
2

18
OH
2
O measurements of
groundwater from landfill monitoring wells to detect leachate infiltration. 
13
C of the CO
2
in

13

13
of the methane fall within a
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range of values representative of microbial methane produced primarily by the acetate-

precipitation values due to the extensive
production of microbial methane within the limited reservoir of a landfill (Hackley et al., 1996).
So monitoring of these isotopic characteristics of leachate provides some insight into its
migration. The biologically mediated methanogenic processes associated with refuse

13
C) in dissolved inorganic carbon

18
O) isotopes of water in landfill leachate (Grossman
et al., 2002)
13
CDIC was also used to investigate the seepage of leachate-contaminated
groundwater into stream water (Atekwana and Krishnamurthy, 2004). Carbon isotopes can also
be used for monitoring biological activity in the aquifers (Grossman, 2002). North et al. (2006)
measured H
2
O using a dual inlet VG SIRA12 mass spectrometer after reduction to H
2
with

13
C of DIC was measured on CO
2
liberated from the sample with 103%
phosphoric acid using a Thermo Finnigan Gas Bench and Delta Plus Advantage mass
spectrometer. The use of compound-specific isotope analysis may also help clarify sources of
contaminants in surface waters, although applications of this technique to landfill leachate are
still being developed (Mohammadzadeh et al., 2005). Vilomet et al. (2001) used strontium
isotopic ratio to detect groundwater pollution by leachate. Natural groundwater and landfill
leachate contamination are characterized by different strontium isotopic ratios (
87
Sr/
86
Sr) of
0.708175 and 0.708457 respectively. Piezometers were used for sampling of groundwater and
The mixing ratios obtained with strontium in groundwater revealed a second source of
groundwater contamination such as fertilizers having
87
Sr/
86
Sr of 0.707859. Pb isotopic ratios
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(
206
Pb/
207
Pb) (Vilomet et al., 2003) and Tritium isotopes (Castañeda et al., 2012) were also used
for the same purpose.
Heaton et al. (2005) determined the changes in N speciation and defined redox conditions in a
leachate plume by using the data for isotope ratios (
15
N/
14
N,
13
C/
12
C and
34
S/
32
S) and dissolved
gas (N
2
, Ar, O
2
and CH
4
) concentrations. Groundwater was sampled in and around a landfill site
in Cambridgeshire, England. They analysed the dissolved gases for determining these isotopic
ratios. The CO
2
gas was collected by using cryogenic trap cooled with dry ice and liquid N
2
and
was analysed for
13
C/
12
C ratios. The other gases such as N
2
, O
2
, Ar and CH
4
, were collected on
activated charcoal cooled in liquid N
2
. Gas yield and their proportions were measured by
capacitance manometer and mass spectrometry respectively.
15
N/
14
N,
13
C/
12
C and
34
S/
32
S ratios
were determined in VG SIRA, VG Optima, and Finnigan Delta isotope ratio mass spectrometers.
In addition to identifying zones of methanogenesis and SO
4
=
reduction, the analysis of the data
indicated processes of NH
4
+
transformation by either assimilation or oxidation, and losses by
formation of N
2
i.e. nitrification & denitrification in a system where there are abrupt temporal
and spatial changes in redox conditions (Heaton et al., 2005). Bacterially mediated
methanogenesis in municipal solid waste landfills cause an enrichment of carbon stable isotope
ratios of dissolved inorganic carbon and hydrogen stable isotope ratios of water in landfill
leachat
.
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3.2.3 Electromagnetic methods
Over the past couple of decades, electromagnetic methods including the resistivity cone
penetration test (RCPT), geophysical exploration such as ground penetrating radar (GPR) and
time domain reflectometry (TDR) have been proposed and developed as potential alternatives to
conventional methods of on-site sampling and laboratory analysis (Atekwana et al., 2000; Börner
et al., 1993; Campanella and Weemees, 1990; Francisca and Glatstein, 2010; Fukue et al., 2001;
Lindsay et al., 2002; Oh et al., 2008; Pettersson and Nobes, 2003; Redman, 2009; Samouëlian et
al., 2005). GPR is one of the most widely used techniques and will be discussed here in brief.
The antenna of GPR transmits and receives high-frequency electromagnetic energy and its
reflections into the subsurface. The transmitted energy reflects at a boundary with sufficient
contrast in dielectric permittivity and the amplitude of such reflection depends on the size of
change in dielectric permittivity across the boundary and proximity of the boundary to the
surface (Figure 4a). The resulting data are presented as a plot, or trace, of amplitude versus two-
way travel-time (TWT), so that a reflection from a boundary is located on the trace at the time
taken for the energy to travel to the boundary and back again (Figure 4b) (Redman, 2009).
Pettersson and Nobes (2003) 
200-MHz antennas for the GPR surveying of contaminated ground at Antarctic research bases.
Readings were taken at 20-cm intervals along straight lines with a time window of 300 ns, and
traces were stacked 16 times to enhance the signal-to-noise ratio. Atekwana et al. (2000)
conducted GPR surveys at the Crystal Refinery located in Carson City, MI constructed in the
1930s releasing hydrocarbons into the subsurface from tanks and pipeline leeks using
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Geophysical Survey Systems, (GSSI) SIR-10A equipment with a 300 MHz bistatic antenna. A
three-scan moving average filter was applied to the data resulting in slight horizontal smoothing.
The GPR study identified three distinct layers; (i) regions of low apparent resistivity, coinciding
with attenuated GPR reflections, (ii) a central region of high apparent resistivity/Low
conductivities with bright GPR reflections below the water table and (iii) an upper GPR reflector
subparallel to the water table, approximately a few meters above the current free product level
and coincident with the top of an oil-stained, light-gray sand layer (Atekwana et al., 2000).
Splajt et al. (2003) investigated the utility of GPR and reflectance spectroscopy for monitoring
landfill sites and found strong correlations between red edge inflection position, chlorophyll and
heavy metal concentrations in grassland plant species affected by leachate contaminated soil.
Reflectance spectroscopy by using spectroradiometer containing contiguous bands at sufficient
spectral resolution over the critical wave range measuring chlorophyll absorption and the red
edge (between 650 and 750 nm) was found to identify vegetation affected by leachate-
contaminated soil. The GPR data identified points of leachate breakout. An integrated approach
using these techniques, combined with field and borehole sampling and contaminant migration
modeling may offer cost-effective monitoring of leachate plume migration. Hermozilha et al.
(2010) combined 3D GPR and 2D resistivity over a heterogeneous media for obtaining
information on landfill structure. They complemented 3D GPR profiling with a constant offset
geometry with 2D resistivity imaging using GPS location techniques to overcome lateral
resistivity variations arising from complexity and heterogeneity of landfill. The 3D GPR was
performed by PulseEcho IV GPR system, using unshielded 100 MHz antennas in 1999 and then
by a Ramac system with a 100 MHz shielded antenna in 2005. ReflexW software was used for
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the GPR data treatment. Boudreault et al. (2010) obtained GPR profiles with a Ramac CU II
system from Mala Geoscience (Mala, Sweden) using 100 MHz center frequency antenna having
a vertical resolution of approximately 33 cm and an actual center frequency of 75 MHz. The
transmitter and receiver antennae were spaced 1 m using a rigid frame in broadside common
offset mode. Data were processed using the REFLEX software from Sandmeier Scientific
Software (Karlsruhe, Germany). No gain was given to the signal in order to compare wave
amplitude between the reflectivity profiles. The two-way travel time was converted to depth
using an average wave velocity of 0.1 m ns
-1
as determined from the wave diffraction patterns
observed in the radar images.
3.2.4 Electrical methods
Geophysical investigation techniques involving electrical conductivity measurements are the
most widely researched of all methods due to easy installation with relatively inexpensive
electrical components. The landfill leachate plumes usually possess elevated ionic load and
enhanced electrical conductivity. So, an aquifer system containing groundwater with a naturally
low electrical conductivity, when contaminated with a leachate plume, will result in a bulk
electrical conductivity anomaly that is readily detectable using both surface, borehole or cross-
borehole electrical resistivity imaging methods (Acworth and Jorstad, 2006).
3.2.4.1 Electrical resistivity and very low frequency electromagnetic induction (VLF-EM)
Benson et al. (1997) conducted electrical resistivity and very low-frequency electromagnetic
induction (VLF-EM) surveys at a site of shallow hydrocarbon contamination in Utah County,
USA. Water chemistry was analyzed through previously installed monitoring wells to enhance
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the interpretation of the geophysical data. The electrical resistivity and VLF data helped map the
contaminant plume by generating the vertical cross-sections and contour maps as an area of high
interpreted resistivity.  also integrated geophysical methods with soil
chemical and hydro-geological methods for investigating groundwater contamination by
leachate. They collected qualitative data from direct current (DC) resistivity geo-electrical
sounding and fast and inexpensive data from VLF-EM survey. The results of VLF-EM method
was expected to have good correlation with those of the DC-resistivity method in which the
signature of a contaminant plume is a low resistivity zone, the depth of investigation being
approximately the same for both methods. The near-surface bodies or discontinuous areas are
more responsive towards galvanic VLF-EM method rather than inductive DC resistivity and thus
simultaneous application of these two methods can very well monitor leachate plume migration.
Al-Tarazi et al. (2008) conducted VLF-EM measurements in a landfill near Ruseifa city at
Jordan with a Geonics EM 16 unit. The transmission from the Russian station (UMS) with a
17.1 kHz and 1 MW power, was used for reliable VLF measurements. They integrated data from
previous DC resistivity study with this VLF-EM data for successfully locating shallow and deep
leachate plume with resistivity less than 20  m
depth. He noticed sign of groundwater contamination resulting in high number of faecal coliform
bacteria and the increase in inorganic parameters such as chloride.
3.2.4.2 Electrical resistivity, cross-borehole tomography and depth-discrete groundwater
electrical conductivity
Acworth and Jorstad (2006) correlated surface resistivity data with cross-borehole tomography
data and depth-discrete groundwater electrical conductivity (Fluid EC) data measured from
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bundled piezometers, to create a continuous, high-resolution image of the distribution of the
leachate plume. Electrical imaging was done using 2 multi-core cables connected to an ABEM
LUND ES464 switching unit slaved to an ABEM SAS4000 Terameter, using the Wenner equi-
spaced electrode configuration. Data were inverted to produce a distribution of true resistivity
using the RES2DINV software. A bundled piezometer with sample tubes at vertical spacing
varying from 0.5 to 1 m was installed to 15 m depth using hollow stem auger technique. Two 15
m strings of 15 gold-plated electrodes in each of them at 1 m intervals were installed one on
either side of the bundled piezometer in a line approximately normal to the groundwater flow
direction and 8 m apart. The strings were then addressed with a current source attached to the top
electrode (1 m depth) in one bore and a current sink in the top electrode in the second bore.
Potential measurements were made between corresponding electrodes at similar depth in the 2
boreholes. The current electrodes were then moved down one position and the process repeated
until the base of the hole was reached. Finally, the results of the cross-borehole tomography
survey demonstrated a strong correlation with the results of the surface resistivity transects and
the groundwater chemistry profiles from the bundled piezometer (Acworth and Jorstad, 2006).
3.2.4.3 Electrode Grids
Applications of electrode grids method in landfill sites essentially rely upon the electrical
conductivity of homogeneous mixtures of soil and landfill leachate, insulating properties of the
geo-membrane liners and ionic concentration of the pore fluid (Frangos, 1997; White and
Barker, 1997). Electrode grid systems cover the entire area beneath a containment unit and can
be used to identify releases and track their migration in the subsurface (USEPA, 2004). The
whole system structurally consists of grid-net electric circuit, electrical conductivity measuring
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sensors adapting two-electrode measurement method, and measuring instruments including
connection system, source meter, and data logger. The electric circuit consists of two arrays of
parallel armored electric wires arranged orthogonally installed in a sub-layer beneath the landfill
liner using simple and durable parts made of high-grade, stainless steel alloy or non-corrosive,
liner compatible conductive HDPE, usually installed during the initial construction of the landfill
facility. One array of electric wires is installed at a specific interval in parallel while the other
array is arranged orthogonally with a same specific interval. Each electrode of two-electrode
sensor is connected to each orthogonal wire at intersections of grid-net electric wires. Finally,
one end of each electric wire forming the grid-net should be connected by branch wires that lead
to a control box of measuring system. The first measurement of electrical conductivity should be
performed to obtain the baseline conditions of the site. Then, electrical conductivity data are
collected with specific time intervals during operation of containment facilities. The location of
contaminant release could be found by searching for deviation points in the distribution of
electrical conductivity (Oh et al., 2008).
3.2.4.4 Electrical resistivity imaging (ERI)
In this process, artificially generated electric currents are supplied to the soil and the resulting
potential difference patterns provide information on the form of subsurface heterogeneities and
their electrical properties as shown in Figure 5 (Kearey et al., 2002). The greater the electrical
contrast between the soil matrix and heterogeneity, the easier is the detection (Samouëlian et al.,
2005). Measurement of electrical resistivity usually requires four electrodes: two electrodes used
to inject the current (current electrodes), and two other electrodes used to record the resulting
potential difference (potential electrodes). Groundwater contamination can also be monitored,
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identified and mapped using an electrical resistivity device 
2001; Samouëlian et al., 2005). Boudreault et al. (2010) performed ERI with a Terrameter SAS
4000 and an ES10-64 switch box with two multiple electrode cables from ABEM (Sundbyberg,
Sweden). Two north-south and four west-east ERI profiles were measured. The electrodes were
pushed into the fill at a regular interval of 1 m to obtain a sufficiently high resolution and a depth
of investigation of about 5 m. A dipoledipole configuration was used to improve the horizontal
sensitivity of the method since the typical urban fill composition has a large short-scale lateral
variability. Robust inversion (with a convergence limit fixed at 1%) of the measured data was
done using the RES2DINV software from Geotomo Software (Boudreault et al., 2010).
3.2.5 Monitoring the fate of dissolved organic matter (DOM) in landfill leachate
Persson et al. (2006) characterized DOM along a groundwater gradient to understand its
interaction with pollutants, such as molecular weight distribution and aromaticity. Groundwater
samples were collected downstream from an old municipal landfill in Vejen, Denmark through
preinstalled Teflon tubes lowered into nitrogen purged iron pipes. The mass spectrometric
analysis of the DOM was carried out on a Micromass Quattro II tandem mass spectrometer
(Manchester, UK), with an electrospray interface, used in the negative ion mode. Estimations of
molecular weight distributions were performed by electrospray ionisation mass spectrometry
(ESI-MS) and size exclusion chromatography (SEC). SEC by Waters Ultrahydrogel 250 column,
a Waters model 2690 LC-pump and a UV-detector at 254 nm was carried out to separate
molecules according to their size rather than their molecular weight. Mass spectrometric results
indicated that in the middle of the gradient, the molecular weight and aromaticity of DOM
decreased to a minimum value while polydispersity increased. However, the aromaticity
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increased to a higher value at the end of the gradient. The molecular weight distribution of DOM
in the groundwater samples as measured with SEC resulted in the same pattern as the mass
spectrometric analysis, showing decreasing molecular weight with increasing distance from the
landfill which can be seen as a process where the DOM gradually becomes more similar to
groundwater fulvic acids (Persson et al., 2006).
Humic substances containing ionizable functional groups such as carboxylic and phenolic groups
exhibit strong affinities toward metal ions (Hernández et al., 2006; Terbouche et al., 2010).
Research on metal binding properties of DOM in the leachate from MSW landfill is lacking. Wu
et al. (2011) utilized fluorescence excitation-emission matrix (EEM) spectroscopy to characterize
the binding phenomenon of DOM with MSW leachate. EEM is a simple, sensitive, non-
destructive technique providing insights into molecular structure of DOM. In combination with a
quenching method, EEM spectroscopy can elucidate the binding properties of metal ions with
DOM (Plaza et al., 2006a, b). However, due to various types of overlapping fluorophores, the
EEM spectra of in situ DOM cannot be easily identified (Henderson et al., 2009). So, a
multivariate chemometric method namely, parallel factor (PARAFAC) analysis, may be used for
decomposing fluorescence EEMs into different independent groups of fluorescent components,
which can then reduce the interference among fluorescent compounds allowing a more accurate
quantification (Engelen et al., 2009). In a recent study, nine leachate samples from various stages
in MSW management were collected and then titrated using four heavy metals (Cu, Pb, Zn and
Cd) as fluorescent quenching agents. Four components with characteristic peaks at Ex/Em of
(240, 330)/412, (250, 300, 360)/458, (230, 280)/340 and 220/432, were identified by the
DOMFluor-PARAFAC model. The results suggested that all the fluorescence EEMs could be
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successfully decomposed by PARAFAC analysis into a four-component model, despite the
dissimilar fluorescence characteristics of the nine leachate samples and the different quenching
effects of different metals at various concentrations. The combination of EEM quenching and
PARAFAC was found to be a useful indicator to assess the potential ability of heavy metal
binding and migration through landfill leachate (Wu et al., 2011).
4 Environmental impact of landfill leachate and its assessment
Leachate is the main toxic compound released from sanitary landfill into the environment,
characterized by high concentrations of numerous toxic and carcinogenic chemicals including
heavy metals and organic matter (Halim et al., 2005). In addition to these chemical mixtures, the
leachates can be contaminated with bacteria, including aerobic, psychrophilic and mesophilic
bacteria, faecal coliforms, and spore-forming-bacteria, including Clostridium perfringens
(Matejczyk et al., 2011). It takes only a small amount of landfill leachate to contaminate large
volume of groundwater, which in turn can contaminate and affect biodiversity and enter the food
chains (Bakare et al., 2007; Garaj-Vrhovac et al., 2009). Multiple chemical exposures may also
pose a higher risk than a single substance. The genotoxic potential of leachates have been
confirmed by several researchers who reported a significant increase in frequencies of
micronuclei, DNA disturbances, sister chromosomal aberrations, chromatid exchanges and also
cut-downs of mitotic indexes in different cell types and model systems (Bakare et al., 2005;
Gajski et al., 2011; Gajski et al., 2012). Different environmental impacts by leachate are being
discussed in the following paragraphs.
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4.1 Environmental impact
4.1.1 Effects on groundwater
Several researchers (Godson and Moore, 1995; Heron et al., 1998; Kerndorff et al., 1992; Lee
and Jones-Lee, 1993; Massing, 1994; Mato, 1999; Mikac et al., 1998; Riediker et al., 2000) have
repeatedly mentioned about the environmental impact of the landfill leachate, particularly on
groundwater quality, regardless of an ideal site selection and a monitoring network design of the
landfill. The danger of leachate infiltration in groundwater is great considering that even the best
liner and leachate collection systems will ultimately fail due to natural deterioration (Needham et
al., 2006; Ouhaldi et al., 2006a, b). In addition, the infiltration of leachate may cause the
variation of groundwater pH and Eh (Rapti-Caputo and Vaccaro, 2006), inducing a metal
dissolution from the subsoil matrix (Prechtai et al., 2008) into the groundwater, even when the
leachate itself is not highly polluted (Kumar and Alappat, 2005; Vadillo et al., 2005). The
presence of organic matter and the modification of pH and redox conditions of the aqueous phase
of the soil may extract awide number of metals, by the dissolution of several mineral species
(Barona et al., 2001; Martinez, 2000; Peters, 1999; Voegelin et al., 2003; Xiaoli et al., 2007).
Risk assessments and environmental regulations for polluted soils are therefore based on batch
extractions of metals, assuming that the results are related to the risk of metal leaching into
ground water or plant uptake (Voegelin et al., 2003). Groundwater quality monitoring systems
being the main indicator to determine the likelihood, and severity of contamination problems, is
of great importance in the overall design of a landfill.
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Van Duijvenbooden and Kooper (1981) investigated the effects of a waste disposal site on the
groundwater flow and groundwater quality in the Netherlands. Measurement of electrical
resistivity and an electromagnetic investigation revealed intrusion of a very large vertical flow
component of landfill plume in the fresh water - salt water boundary at about 40 m depth.
However, local flow patterns indicated an all-sided migration of pollutants into the aquifer (Van
Duijvenbooden and Kooper, 1981). The leachate from the Ano Liosia landfill in Greece was
found to contain high levels of colour, conductivity, TS, COD, NH
3
N, PO
4
3
, SO
4
2
, Cl
, K
+
, Fe
and Pb. The low BOD/COD ratio (0.0960.195), confirmed that the majority of this organic
matter was not easily biodegradable. The sites nearest to the landfill were most polluted,
indicating pollution transfer and the leachate movement through fractures or karstic cavities,
geological and hydrological characteristics of the area under study (Fatta et al., 1999). Mor et al.
(2006) measured concentration of various physico-chemical parameters including heavy metal
and microbiological parameters in groundwater and leachate samples from Gazipur landfill site
near Delhi. The groundwater was found to contain moderately high concentrations of Cl
, NO

,
SO
4

, NH
4
+
, Phenol, Fe, Zn and COD indicating leachate percolation. Interestingly the water
contamination dropped fast with depth up to 30m and further percolation of viscous leachate
became gentler probably due to the hindrance from the solid soil matter (Mor et al., 2006).
Rapti-Caputo and Vaccaro (2006) performed hydrogeological and geochemical monitoring of
two principal aquifer systems, one unconfined, and another confined at 17m depth, below the
landfill of Sant'Agostino in Italy. In the shallower unconfined aquifer, the existence of high
concentration values of K, Na, Cl
-
and SO
4
2-
and heavy metals such as Cr, Ni, Co, Mo and Sr
were found along the flow direction. pH values between 7.16 and 7.9 and redox potential
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    ce of basic water in a reducing environment
favouring the adsorption of ionic substances in soil. The deeper confined aquifer had higher
concentrations of NH
4
+
, Cl
-
, Pb, Cu and Zn than that in the regional aquifer indicating local
diffusion from leachate (Rapti-Caputo and Vaccaro, 2006).
4.1.2 Reduction of soil permeability and modification of soil
Field observations, such as the ponding of leachate at landfills (Nelson, 1995) suggest that some
of the unlined landfills underwent significant reductions in hydraulic conductivity. Other
laboratory and field observations also show that soils can undergo significant reduction in
hydraulic conductivity during leachate permeation (Cartwright et al., 1977; Yanful et al., 1988),
even leading to clogging of leachate collection systems (Brune et al., 1994; Rowe et al., 1997).

continuous biofilms (Rowe et al., 1997; Taylor and Jaffé, 1990) or presence of discontinuous
microbial aggregates in soil pores (Vandevivere and Baveye, 1992), metal precipitation (Rowe et
al., 1997), and gas production by denitrifiers and methanogens (deLozada et al., 1994; Islam and
Singhal, 2004; Taylor and Jaffé, 1990). However, the relative significance of these mechanisms
in controlling the extent of clogging and the dynamics of microbial-metal precipitation
interactions is not yet properly researched.
Continuous flow experiments were conducted by Islam and Singhal (2004) using sand-packed
columns for investigating the relative significance of bacterial growth, metal precipitation, and
anaerobic gas formation on biologically induced clogging of soils. Natural leachate from a local
municipal landfill was amended with acetic acid and then was fed to two sand-packed columns.
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Based on observed transformations the following microbial reactions are assumed to occur in the
columns in presence of acetic acid:
Manganese reduction:
CH3COO- + 4MnO2(s) + 7H+2HCO3- + 4Mn2+ + 4H2O
(1)
Iron reduction:
CH3COO- + 8 Fe(OH)3(s) + 15H+ 2 HCO3-+ 8Fe2+ + 20 H2O
(2)
Sulfate reduction:
CH3COO- + SO42- 2 HCO3- + HS-
(3)
Methanogenesis:
CH3COO- + H2O HCO3- + CH4
(4)
Changes in the observed concentrations of dissolved acetic acid, sulfate, Fe(II), and Mn(II) with
time suggest that methanogenesis and the reduction of manganese, iron, and sulfate occur
simultaneously. Several physical, geochemical, and biological interactions were observed during
leachate transport in soils resulting in a reduction of its permeability. An increase in the substrate
concentration resulted in rapidly increasing pH, inorganic carbon (total dissolved carbonate), and
attached biomass at the column inlet, leading to enhanced precipitation of Fe
2+
, Mn
2+
, and
Ca
2+
at the column inlet thereby decreasing the hydraulic conductivity from an initial value of
8.8×10
3
to 3.6×10
5
cm s
-1
. However, mathematical modeling showed that bioaccumulation and
gas formation played more significant role in reducing hydraulic conductivity, while metal
precipitation had a negligible effect (Islam and Singhal, 2004). In another simulation work by the
same researchers, it was deduced that higher substrate concentrations may increase the extent of
the zone of reduced hydraulic conductivity, but may not lead to further decreasing the
conductivity. Also, finer-grained soils are likely to experience higher conductivity reductions
than larger-grained soils (Singhal and Islam, 2008).
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The percolation of landfill leachate even in absence of a high concentration of a specific
pollutant may induce a strong modification of soil chemical and physical characteristics due to
the alteration of the natural equilibrium between the aqueous phase and the soil matrix. As a
result, a huge amount of cations can be solubilised, thus inducing groundwater pollution. Di
Palma and Mecozzi (2010) performed batch and column experiments for studying metal
mobilization from a soil sampled down gradient of a municipal waste landfill in Northern Italy at
different pH and Eh. At first, the column was washed with distilled water and then a
groundwater, sampled down-gradient in the same site, was used for column leaching. The
concentrations of Fe, Mn, and Ni were evaluated when the pH & Eh were altered. Results
indicated a greater release when acidic conditions were achieved, a positive effect in this case of
the addition of an oxidant and a great Mn mobilization when negative redox potentials were
established. The effect of the addition of oxidant or reductant solutions on soil characteristics
modification during a remediation treatment involving the percolation of an aqueous solution
was investigated. In the case of a pH lowering, the addition of an oxidant such as H
2
O
2
proved to
be effective in decreasing metal dissolution, and could also have a positive effect on aerobic
biological degradation reactions. Conversely, the addition of a reductant, such as dithionite,
strongly enhanced Ni and, mainly, Mn mobilization, even under alkaline conditions (Di Palma
and Mecozzi, 2010).
Chen and Chynoweth (1995) calculated hydraulic conductivities of dry municipal solid waste
(MSW) samples by compacting them in plexiglas columns which were set-up as constant head
permeameters to densities of 160, 320 and 480 kg m

. Water flowed continuously through the
columns under hydraulic gradients of 24·0 m m

. Darcy's equation was used to calculate
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hydraulic conductivity which was found to be time-dependent. The temporal variation was
attributed to varying degrees of saturation due to gas formation and relative movement of fine
particles in the columns. The average hydraulic conductivities at 160, 320 and 480 kg m

were
found to be 9·6 × 10

, 7·3 × 10

and 4·7 × 10

cm s

, respectively. Francisca and Glatstein
(2010) deduced that physicochemical interactions such as changes in the double-layer thickness
and chemical precipitation of carbonates had negligible effect on the hydraulic conductivity of
highly compacted siltbentonite mixtures. However, bioclogging due to accumulated biomass
from bacteria and yeast significantly reduced the hydraulic conductivity and blocked up the soil
pores. The experimental data confirmed the biofilm formation .
Wu et al. (2012) measured water retention curves (WRC) of MSW using pressure plate method
representing the shallow, middle, and deep layers of the landfill and the WRC was found to be
well-reproduced by the van GenuchtenMualem model, which was then used to predict the
unsaturated hydraulic properties of MSW, such as water retention characteristics and unsaturated
hydraulic conductivity. With the increase in the landfill depth and age, the overburden pressure,
the highly decomposed organic matter and finer pore space increased, hence the capillary
pressure increased causing increases in air-entry values, field capacity and residual water
content. Steepness of WRC and saturated water content decreased. The unsaturated hydraulic
properties of MSW showed more silt loam-like properties as the age and depth increased (Wu et
al., 2012).
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4.1.3 Effects on surface water
Yusof et al. (2009) studied the impact of landfill leachate from three different types of landfills,
namely active uncontrolled, active controlled and closed controlled, were characterized, and their
relationships on the river water chemistry. The organic contents in the closed or older landfills
were found to be lower than in the active landfill. Moreover, the higher BOD/COD (0.67) in the
active controlled landfill indicated it to be in the acetogenic phase. Conversely, the lower
BOD/COD (0.16) shown by both the active uncontrolled and the closed controlled landfills is a
typical characteristic of the methanogenic phase of an old landfill (Calli et al., 2005; Fan et al.,
2006). The impact of leachate from an active uncontrolled landfill was the highest, as the organic
content, NH
4
N, Cd and Mn levels appeared high in the river. At the same time, influences of
leachate were also observed from both types of controlled landfills in the form of
inorganic nitrogen (NH
4
N, NO
3
N and NO
2
N) and heavy metals (Fe, Cr, Ni and Mn).
Improper treatment practice led to high levels of some contaminants in the stream near the closed
controlled landfill. Meanwhile, the active controlled landfill, which was located near the
coastline, was exposed to the risk of contamination resulting from the pyrite oxidation of the
surrounding area (Yusof et al., 2009).
4.2 Hazard assessment of landfill leachate
Numerous models and approaches ranging from deterministic water balance analyses such as
Hydrologic Evaluation of Landfill Performance (HELP) (Schroeder et al., 1994) and Flow
Investigation of Landfill Leachate (FILL) (Khanbilvardi et al., 1995) and stochastic simulation
models such as LandSim (GolderAssociates, 1996) 
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Migration with Transformation Products (EPACMTP) (USEPA, 2003) to relative hazard
assessment systems for evaluating landfill hazards have been developed. Each one of these
models and approaches has some advantages and disadvantages. While deterministic and
stochastic models need large amounts of data, involve complex analytical procedures and thus
are time consuming, relative hazard assessment systems, often referred to as hazard
rating/ranking systems, suffer from the subjectivity involved in their scoring methodologies.
However, considering their simplicity, such relative hazard assessment systems are considered to
be more suitable when only a comparative assessment as in the case of priority setting, is the
objective.
4.2.1 Relative hazard assessment systems
In order to comply with the legislations regarding the management of municipal solid waste, it is
necessary to undertake a diagnosis and characterisation of the landfill impacted areas in order to
develop an adequate action plan. However, the remedial and preventive measures cannot be
undertaken at all the existing closed and active landfill sites because of financial constraints. So,
a gradual approach is needed based on a system of prioritization of actions to establish which
landfills need immediate attention for the remediation works. In most cases, the diagnostic
methods made it possible to compare landfills on an environmental basis, but not to take
decisions about their control, closure, capping, or recovery. All of the assessments were related
to the release point, without taking into account the characteristics of their environment (Calvo,
2003).
A number of relative hazard assessment systems for waste disposal sites have been developed
over the past three decades and reported in literature (Singh et al., 2009). Usually, three hazard
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modes are used to evaluate the waste sites: 1. migration of pollutants away from the site via
groundwater, surface water, or air routes, or a combination thereof, 2. fire and explosion
potential, and 3. direct contact with hazardous substances. In most of the systems, site ranking is
based either on the combined score for various routes under migration mode or the score for the
dominant route i.e. the route returning highest score. In course of calculating site hazard, more
information is considered by a system, more accurate is the assessment and evaluation. However,
more data signifies increased complexity, cost, time and chances of error. This reduces the
acceptability of a system among users who always want maximum output with minimum inputs.
Some parameters can be termed as simple parameters that can be determined iwthout any
complex analytical methods such as by site walkover, visual survey, local inhabitant survey,
regional maps of groundwater, soil type, geology etc. The parameters which are difficult to
collect e.g. by field drilling and sampling as well as laboratory testing are considered as complex
parameters. More number of complex parameters in a system reduces its user friendliness. Table
4 lists the number of parameters considered by different hazard rating systems. In this sub-
section, we will discuss mainly four significant hazard rating systems.
4.2.1.1 Leachate Pollution Index (LPI) Method
Kumar and Alappat (2005) discussed about LPI, a quantitative tool having an increasing scale
index based on Delphi technique (Dalkey, 1969), for calculating the leachate pollution data of
landfill sites. In this method, 18 leachate pollutants (e.g. pH, TDS, BOD, COD, heavy metals,
phenolic compounds, chlorides, total colifiorm) were selected for inclusion in the index and were
awarded some significance and pollution weight, that added up to 1.00 for the 18 pollutants.
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The LPI can be calculated using the equation:
n
ii
i1
LPI w p
(5)
Where, LPI = the weighted additive leachate pollution index, w
i
= the weight for the ith pollutant
variable, p
i
= the sub index score of the ith leachate pollutant variable, n = number of leachate
pollutant variables used in calculating LPI and
n
i
i1
w1
. However, when the data for all the
leachate pollutant variables included in LPI are not available, the LPI can be calculated using the
concentration of the available leachate pollutants. In that case, the LPI can be calculated by the
equation:
m
ii
i1
m
i
i1
wp
w
LPI
(6)
where m is the number of leachate pollutant parameters for which data is available.
The procedure for calculating LPI for a given landfill site at a given time involves the following
three steps: Firstly, testing of the 18 leachate pollutants, secondly, calculating sub-index values
(p) based on the concentration of the leachate pollutants obtained during the tests and lastly,
aggregation of sub-index values obtained for all the parameters by multiplying it with the
respective weights assigned to each parameter. For the last step, the above two equations are
used depending upon the situation. High value of LPI indicates higher contamination potential
(Kumar and Alappat, 2005).
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4.2.1.2 Global EnvironmentLandfill Interaction Index or Impact Index (ELI)
Calvo et al. (2005) studied a new methodology for environmental diagnosis of landfill sites. This
methodology was based on the formulation of a general index called Global Environment
Landfill Interaction Index or Impact Index (ELI). In order to calculate this index, some aspects
in each landfill have to be analysed viz, environmental interaction between the release point and
certain affected environmental parameters, environmental values of the surface water,
groundwater, atmosphere, soil and health and operational conditions of the landfill from the
point of view of environment. The rate expression is as follows:
ELI = LI
i
= 
i
×EWC
i
) =
(ERI
groundwater
×EWC
groundwater
)+(ERI
surfacewater
×EWC
surfacewater
)+(ERI
atmosphere
×EWC
atmosphere
)+(ERI
s
oil
×EWC
soil
)+(ERI
health
×EWC
health
) (7)
where
ELI = Global EnvironmentLandfill Interaction Index or Impact Index
ELI
i
= the EnvironmentalLandfill Interaction Index for parameter i
i = the parameters: groundwater, surface water, atmosphere, soil, and health
EWC
i
= the Environmental Weighting Coefficient
ERI
i
= the Environmental Risk Index for the Environmental Effect of parameter i
Ranges of scores are obtained for ELI to classify the overall environmental impact of landfills
as low (0-35), average (31-70) and high (71-105). The ERI aims to gauge the potential for
environmental impact for each observed parameter, reflecting whether or not interaction exists
between the processes in the release point and the characteristics of the environment.
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The E
Focusing on the study of each landfill individually, the ERI enables us to determine which
parameters are most affected by the landfill, making it easier to prioritize suitable control actions.
Analysis of index results provides information about the suitability of the release-point locations
on the basis of which, it would be possible to draw up action plans for the remediation or closure
of the landfill site (Calvo et al., 2005).
4.2.1.3 Hazard rating system by Singh et al. (2009)
Singh et al. (2009) assessed existing site hazard rating systems and came up with a new
groundwater contamination hazard rating system for landfills. The proposed system was based
on source-pathway-receptor relationships and evaluated different sites relative to one another by
the Delphi technique (Dalkey, 1969). The proposed system is more sensitive to the type of waste
and exhibited greater sensitivity to varied site conditions. In this system, 15 parameters are
studied as depicted in Figure 6. Each of them is assigned a best and worst value. The overall
groundwater contamination hazard rating of a waste disposal site was obtained by the following
relationship:
HR,GW = (HS X HP X HR) / SF X 1000
(8)
where H
s
, H
p
and H
R
were the source hazard rating, pathway hazard rating and receptor hazard
rating, respectively; and SF is a scaling factor (equal to 1,000,000). The scaling factor is equal to
the product of the source, pathway, and receptor hazard ratings of a waste disposal site having all
its parameters at the worst values. The overall hazard score obtained from the Equation 8 is
limited to a maximum of 1000 for MSW landfills, 5000 for HW landfills, and 200 for C&D
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waste landfills. The application of different systems to six old municipal solid waste landfills
showed that whereas the existing systems produced clustered scores, the proposed system
produced significantly differing scores for all the six landfills improving decision making in site
ranking (Singh et al., 2009).
4.2.1.4 Assessment of Toxicity Index
Baderna et al. (2011) also proposed an integrated strategy to evaluate the toxicity of the leachate
using chemical analyses, risk assessment guidelines and in vitro assays using the hepatoma
HepG2 cells as a model. Human risk assessment was done based on chronic daily intake (CDI
(mg kg
-1
day)) for each compound, which was calculated using the formula:
CDI = [(Cwater x WI x ED x EF) / (BW x AT)]
(9)
where C
water
 L day
-1
; ED=exposure
duration=30 years; EF=exposure frequency=350 days year
-1
; BW=body weight of the
target=70 kg (adult); AT=exposure average time: 30 years for non-carcinogenic compounds, 70
years (lifetime) for carcinogenic compounds.
The hazard index (HI) was calculated for each compound in order to estimate possible toxic
effects on humans due to the ingestion of leachate-contaminated water, using the formula:
HI=CDI/RfD
(10)
where HI is the hazard index, CDI the calculated chronic daily intake, RfD the reference dose for
the selected compounds (mg kg
-1
day). The RfD is a numerical estimate of a daily oral exposure
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to the human population, including sensitive subgroups such as children, that is not likely to
cause harmful effects during a lifetime (USEPA, 2006).
The assessment of carcinogenic effects was calculated using the cancer risk equation:
CR = CDI×SF
(11)
where CR is the cancer risk, SF the slope factors (kg day mg

): an upper-bound estimate of risk
per increment of dose that can be used to estimate risk probabilities for different exposure levels
(USEPA, 2005).
The ecological risk assessment was based on the dilution scenario used for human risk
assessment. For risk analysis we used traditional risk procedures focused on the Hazard Quotient
defined as follows:
HQ=PEC/PNEC
(12)
where PEC is the predicted environmental concentration (resulting from chemical analysis) and
PNEC the predicted no-effect concentration. The evidences from in vitro studies on HepG2
suggested that leachate inhibited cell proliferation at low doses probably inducing a reversible
cell-cycle arrest that becomes irreversible at high doses. This study confirmed the hypothesis that
cells that survive the initial insult from leachate constituents maintains the potential to proliferate
until the effects on cell metabolism lead to death (Baderna et al., 2011).
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4.2.2 Deterministic and stochastic models for monitoring environmental impact of landfill
leachate
Mathematical models are powerful predictive tools to address issues related to landfill leachate
management. However, inadequate and wrong field data and insufficient understanding of the
complex physico-chemical and biochemical reactions going on in the landfill limit the predictive
capabilities of these mathematical models. So, these models are advised to use for an educated
guesswork and to evaluate the relative importance of selected variables for management purpose.
Numerous mathematical models have been developed since 1980s to simulate the generation and
transport of leachate in landfills (El-Fadel et al., 1996, 1997; Suk et al., 2000). A detailed review
on pre-1995 models was done by El-Fadel et al. (1997). However, these models have their own
disadvantages as a whole (Scott et al., 2005).
4.2.2.1 Assessing the reduction in hydraulic conductivity
Islam and Singhal (2004) came up with a simple mathematical model to assess the total
reduction in hydraulic conductivity in a landfill. It was expressed in terms of the fractional
reduction due to biomass accumulation, metal precipitation, and gas formation, as follows:
Total reduction = 1 - k(t)/k0 = 1 - (1 - (f(x) + g(m)))(1 - h(g))
(13)
where, f(x), g(m), and h(g) are functions for fractional reduction in hydraulic conductivity due to
bioaccumulation, metal precipitation, and gas formation, respectively, k
0
is the initial soil
permeability (L
2
), and k(t) is the soil permeability at time tf(x)+g(m))) represents
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h(g)) acts similarly to the
relative permeability function in representing the effect of gas flow on soil permeability.
The impact of biomass accumulation on the permeability was described using a simple
permeability reduction model proposed by Clement et al. (1996), as follows
f(x) = 1 - (1 - ns/n0)19/6
(14)
where n
s
(=X
s
ρ
k
/ρ
s
) is the volume fraction of the soil-attached biomass (L
3
biomass L
-
3
total), n
0
is the initial soil porosity, X
s
is the microbial mass per unit mass of aquifer solids (M
M
-1
), ρ
k
is the bulk density of aquifer solids (M L
-3
), and ρ
s
is the biomass density (M L
-3
). The
biomass density was estimated as 70 mg-volatile solids cm
-1
(Cooke et al., 1999). Assuming that
approximately 50% of the cellular carbon is protein the biomass density is estimated as 35 mg-
protein cm
-3
. The study suggested that stimulation of anaerobic activity at the base of landfills
might lead to creation of impermeable barriers and pore clogging of leachate collection systems
(Islam and Singhal, 2004).
 developed a mathematical model to simulate landfill leachate behavior and
its distribution throughout the landfill, taking into consideration the hydraulic characteristics of
waste and composition of leachate. The model incorporated governing equations describing
processes taking place during the stabilization of wastes, including leachate flow, dissolution,
acidogenesis and methanogenesis. To model the hydraulic property changes occurring during the
development stage of the landfills, a conceptual modeling approach was proposed. This approach
considered the landfill to consist of columns of cells having several layers. Each layer was
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assumed to be a completely mixed reactor containing uniformly distributed solid waste,
moisture, gases and micro-organisms.
4.2.2.2 Assessment of degradation products of landfill leachate components
Butt et al. (2008) reviewed the advantages and shortcomings of various risk assessment
techniques related to landfill leachate contamination. Also, Butt and Oduyemi (2003) briefly
outlined a holistic procedure for the concentration assessment of the contaminants and a
computer model for the risk assessment of landfill leachate (Butt et al., 2008; Butt and Oduyemi,
2003). Reinhart et al. (1991) used a mathematical mass transport model, the Vadose Zone
Interactive Processes model to describe the fate of organic compounds in sanitary landfills. The
model was used to solve a convective-dispersive equation incorporating the transport and
transformation processes of dispersion, advection, chemical and biological transformation, and
sorption in unsaturated porous media. The model was optimized using input data from laboratory
column operations and the physical/chemical phenomena from the field and it predicted low
mobility of hydrophobic compounds and high mobility of more hydrophilic compounds in the
landfill. Gau and Chow (1998) investigated the characteristics of landfills using different kinds
of waste combinations. COD concentrations of leachate from semiaerobic and anaerobic landfills
were processed by using a numerical method to get a simulation model for the estimation of
variations in the organic pollutants in the leachate. The degradation of the leachate quality was
approximately similar for both types of landfills .
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4.2.2.3 Mathematical simulation and long-term monitoring of leachate components
Ozkaya et al. (2006) simulated the refuse age and leachate components spread out using a
mathematical formula in cells with and without leachate recirculation (C1 & C2 respectively).
The leachate from Odayeri Sanitary Landfill, Istanbul, Turkey was monitored for 920 days by for
the sulfate (SO
4

), chloride (Cl
), COD and BOD. The relationship between these parameters
and refuse age was simulated by a non-linear exponential function:
y=a0+a1e-t+a2te-t
(15)
where a
0
, a
1
and a
2
are unknown constants of the function, the a
0
constant is residual
concentration and y is pollutant concentration at time t as g L
-1
and t is refuse age as months.
This model could predict reaching rate to the peak value of pollutant concentration to ensure
optimization of leachate treatment. Constants in the non-linear equation were solved by the least
squares method, minimizing the total square deviations from the model of the experimental data,
using a MATLAB 7.0 computer program. A good fit was obtained between the measured data
and model simulations. The results showed that there appeared to be little improvement in
leachate quality by leachate recirculation in terms of COD and BOD values, however, it was
determined that the pollution loads more rapidly reached minimum values within the C2 test cell
(Ozkaya et al., 2006)
4.2.2.4 Reliability assessment of groundwater monitoring networks at landfill sites
Monitoring well networks at the landfill sites can be used for detecting leakage plumes. Yenigül
et al. assessed the reliability of groundwater monitoring systems at landfill sites through a
hypothetical problem where the detection probability of several monitoring systems was
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compared by a simulation-based model. A MonteCarlo approach was used to simulate a large
number of contaminant plumes resulting from the failure of the landfill. A single MonteCarlo
realization consists of the following five steps, namely, (i) Generation of a realization of a
random hydraulic conductivity field, (ii) Solution of the steady state groundwater flow model to
determine the velocity field, (iii) Generation of a random leak location, (iv) Solution of the
random walk transport model to determine the concentration field of the contaminant plume until
it reaches the compliance boundary, (v) Check whether the concentration value at a given
monitoring well location exceeds a given threshold concentration (detection limit), to determine
whether a plume is detected or not detected by the monitoring system.
The movement of contaminants in the subsurface was represented by the advectiondispersion
equation (Bear, 1972). The contaminant was assumed to be conservative and to have no
interaction with the solid matrix. The two-dimensional advectiondispersion equation for this
case can be written as:
0
xx xy yx yy
xy
C C C C
D D D D
x y x y
C C C
vv
t x y x y

(16)
where C is the concentration of the contaminant at time t at location (x,y), ν
x
and ν
y
and are
average groundwater flow velocity components in the x and y-directions, respectively, and D
xx
,
D
xy
, D
yx
, D
yy
are the components of the hydrodynamic dispersion tensor (Bear, 1972). The
analysis revealed the lateral dispersivity of the medium as one of the most significant factor
affecting the efficiency of the systems, since it is the primary parameter controlling the size of
the plume. It was also concluded that the reliability of the common practice of three down-
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gradient monitoring wells is inadequate for prevention of groundwater contamination due to
landfills (Yenigül et al., 2005).
4.2.2.5 Computer aided modeling for risk assessment
Hazards can be quantified, simulated and accurate risk analysis can be undertaken by using
computational methods and modelling precise systems, leading to a more effective risk
management. Butt et al. (2008) discussed about some techniques used in landfill risk assessment.
Some computer models and software programme have been described in the Table 5 and their
shortcomings have been pointed out.
5 Recent technological developments for landfill leachate treatment and
remediation
The knowledge of the impact of landfill leachate on the environment has forced authorities to
apply more and more stringent standards for pollution control. In addition, the ever increasing
toxic load in MSW has caused the leachate generated in landfills to become more varied and
complex in composition and thus difficult to treat. For many years, simple biological and
physico-chemical treatments such as aerated lagoons, simple aerobic and anaerobic digesters,
advanced oxidation treatments using ozone or Fenton reagents, adsorption using GAC or PAC,
chemical and electrical coagulation etc., were considered sufficient for treatment and
management of highly concentrated effluents such as landfill leachates. However, it was found
that the simple treatments were insufficient to meet the present stricter effluent disposal
standards targeted towards complete reduction of the negative impact of landfill leachate on the
environment. This implies that new treatment alternatives must be developed. Therefore, in the
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last two decades, a host of new technologies based on membrane filtration, electrochemical
oxidation and combination of different reagents or technologies have been developed as viable
treatment alternative. It was found that integration of age old technologies with advanced
treatment processes yielded excellent treatment efficiency in terms of COD, NH
4
-N, heavy
metals, TOC, DOM etc., removal (Kjeldsen et al., 2002).
Treatment techniques vary depending on the age of the leachate and on the leachate disposal
standards set by the local authorities (Castrillón et al., 2010; Ozturk et al., 2003; Renou et al.,
2008a). Reasonable treatment efficiency can be achieved by using biological treatments for the
removal of COD, NH
3
-N and heavy metals in case of young leachates. However, for treating old
stabilized leachate having low biodegradability, physico-chemical treatments have been found to
be suitable as a refining step for biologically treated leachate. Integrated chemicalphysical
biological processes, in any order, negates the drawbacks of individual processes contributing to
a higher efficacy of the overall treatment (Bohdziewicz et al., 2001; Lin and Chang, 2000).
Due to the climatic conditions and a combination of various physical, chemical and biological
processes occurring in the landfill, the leachate composition can fluctuate over both short and
long periods of time. According to Scott et al. (2005) the variation is particularly pronounced in
an active landfill. Therefore the leachate treatment system must be flexible enough to produce
the same quality effluent despite all the variations (Kochany and Lipczynska-Kochany, 2009). In
spite of different views on the leachate treatment, many experts agree that on-site treatment
facilities are more suitable both in terms of cost and in terms of efficiency.
Many good reviews on leachate treatment technologies have been published over the years
(Alvarez-Vazquez et al., 2004; Deng and Englehardt, 2006; Foo and Hameed, 2009; Kim and
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Owens, 2010; Kurniawan et al., 2006b; Laner et al., 2012; Renou et al., 2008a; Wiszniowski et
al., 2006). So, this section concentrates only on the recent developments in this area post 2005.
Different leachate treatment techniques have been classified as illustrated in Figure 7.
5.1 Application of natural attenuation for leachate remediation
According to USEPA (1999), the amalgamation of different physical, chemical and biological
processes occurring in nature, which can efficiently reduce concentration, toxicity, and/or
mobility of contaminants can be defined as natural attenuation. The application of constructed
wetlands (CW) for natural treatment of leachate has been practised for many years in different
countries with varying degrees of success . CWs
are mainly of two types, free surface water system and subsurface flow system depending on the
nature of wastewater flow. The treatment of wastewater in CWs involves a combination of
biological and biochemical processes (Yalcuk and Ugurlu, 2009). The wetlands provide suitable
milieu for rapid natural attenuation of organic contaminants due to the presence of large variety
of microorganisms, nutrients in the discharging groundwater and a wide range of redox
conditions in the surrounding groundwater or surface water interfaces (Lorah et al., 2009; Tobias
et al., 2001). Microbial communities present in CWs can break down the complex organic
compounds in wastewaters and with age as the microbial population increases in a CW the rate
of organic removal increases (Calli et al., 2006). Fluorescence results reveal the predominance of
bacteria in CWs, including heterotrophic and autotrophic, which are responsible for BOD
5
removal (Sawaittayothin and Polprasert, 2007). However, different treatment plants support
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different bacterial populations and even within a given treatment plant significant variations in
community profile has been observed.
Phytoremediation is an attractive technology for landfill remediation and according to Kim and
Owens (2010), it can stabilize soil while simultaneously remediating landfill leachate. Figure 8
illustrates the interaction between the soil and plant systems for leachate remediation in a CW.
Plants influence the redox potential in planted CWs by supplying oxygen to the soil in the root
rhizospheric zone. Enhanced nitrification by nitrifying bacteria takes place in this zone, thereby
reducing the NH
4
-N concentration in the landfill leachate . The amount
of oxygen in the rhizosphere shows diurnal and seasonal fluctuations depending upon various
factors like photosynthesis, light intensity, stomatal aperture, and temperature 
2012a). The plants that are commonly used in CWs are cattail (Typha latifolia L.), willow-
coppice (Salix sp.), poplars, reed (Phragmites australis Trin ex Steudel), rush (Juncus effusus
L.), yellow flag (Iris pseudacorus L.), and mannagrass (Glyceria maxima) (B
2007; Duggan, 2005; Rosenqvist and Ness, 2004; Wojciechowska et al., 2009; Wojciechowska
and Obarska-Pempkowiak, 2008; Yalcuk and Ugurlu, 2009; Zalesny et al., 2008).
The HM content in leachates from old landfill sites are usually low and do not represent much
difficulty in purification procedures (Christensen et al., 2001; Kjeldsen et al., 2002; Long et al.,
2009). Different biotic and abiotic processes such as complexation, precipitation, flocculation,
adsorption, cation and anion exchange, oxidation and reduction, adsorption, microbial activity
and plant uptake are responsible for heavy metal removal in a CW (Kosopolov et al., 2004; Sinan
Bilgili et al., 2007; Ujang et al., 2005). The mobility and eco-toxicity of HMs depends on the
metal speciation and the fraction of DOM to which it is bound.
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CWs show high BOD
5
, TN and fecal coliforms (FC) removal efficiency of 91%, 96% and more
than 99%, respectively (Bulc, 2006; Mehmood et al., 2009; Sawaittayothin and Polprasert, 2007;
Yalcuk and Ugurlu, 2009). Examples of leachate treatment in CWs and the achieved efficiency is
tabulated in Table 6. According to Picard et al. (2005) about 9899% of nitrogen and phosphorus
removal may be achived in a constructed wetland. Irrespective of the microorganism density and
the type of plants used, the prevailing weather conditions have significant influence on the
treatment capacity of a CW (Akratos and Tsihrintzis, 2007). There are certain drawbacks
associated with the land application of leachate as a phytoirrigant, the most important being high
nitrogen and salinity loadings. Salinity loading due to leachate irrigation can be managed, by
judiciously controlling the leachate application rate and by providing intermittent fresh water
irrigation. According to Smesrud et al. (2011) fresh water irrigation can be 30% of the total
irrigation water supplied.
5.2 Application of biological and biochemical techniques in reactors
Traditionally, landfill leachates have been treated along with sewage in sewage treatment plants.
According to Robinson and Barr (1999), combinations of different biological and physico-
chemical treatment methods for landfill leachate treatment, is more efficient than using any
single treatment system such as Sequential Batch Reactors (SBR), Upflow Anaerobic Sludge
Blanket Reactor (UASB), Anaerobic Digesters, and others. Leachate contains high COD and
NH
4
-N content and some other noxious substances such as heavy metals which are difficult to be
remediated by biological treatments alone (Uygur and Kargi, 2004; Xu et al., 2008).
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In the SBR systems, reaction and sludge settling are completed in the same reactor, sequentially
(Aziz et al., 2011b). The time dependent character of the process facilitates the alteration of SBR
operation cycles in response to variation in waste, which occurs frequently in case of landfill
leachate (Laitinen et al., 2006; Trois et al., 2010). According to Klimiuk and Kulikowska (2006),
the treatment strategy in SBRs maybe designed as follows: dump filling of wastewater into the
SBR over a relatively short period of time, elimination or reduction of aeration and mixing
during filling stage and increasing the volumetric exchange ratio. A long sludge age allows the
growth of slow growing microorganisms in mixed culture of the activated sludge, which
eventually participate in the removal of slow biodegradable substrates. However, for SBRs
operated under aerobic conditions short hydraulic retention time is more favourable as long
hydraulic retention time can cause reduction in biomass concentration due to cell decay (Klimiuk
and Kulikowska, 2006). Many researchers found that the addition of activated carbons like PAC,
GAC and biometric fat cells increased the efficiency of SBRs by effectively removing stable
hydrophobic organic chemical species from biologically treated landfill leachate (Aziz et al.,
2011c; Kargi and Pamukoglu, 2004; Liyan et al., 2009). Neczaj et al. (2007) found that a
pretreatment of landfill leachate by sonication increased COD and nitrogen removal efficiency in
a SBR.
Di Iaconi et al. (2006) proposed an aerobic Sequencing Batch Biofilter Granular Reactor having
high organic removal efficiency of about 80% in terms of COD. Systems with granuar biomass
are known to have up to 15g L
-1
biomass concentrations and conversion capacities of 6-7 kg of
COD m
-3
and relatively low sludge production rates (Di Iaconi et al., 2005). This tretment
technique was further modified by addition of a pre-treatment step for nitrogen removal by
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struvite precipitation, and subsequent biological degradation by ozone which increased nitrogen
removal efficiency (Di Iaconi et al., 2011). Gálvez et al. (2012) and Gálvez et al. (2006) used
submerged biofilter under aerobic and anaerobic conditions for leachate treatment.
Anaerobic digestion is a simple and effective biotechnological process that has been used
extensively to treat organic wastes. Anaerobic processes involve the sequential breakdown of
complex organic compounds by several effectively interacting metabolic groups of
microorganisms (Huang et al., 2003). According to Erses et al. (2008) and Mertoglu et al.
(2006), better organics, nitrogen, phosphorous and alkali metal removal is achieved under
aerobic condition as compared to anaerobic conditions . Co-digestion of sewage and leachate is
an effective leachate treatment option if the leachate is young and the sewage treatment facility
(Garg and Mishra, 2010). Mixing of leachate and sewage
increases the total organic carbon and causes the biogas yield to increase. The biogas yield from
the co-fermentation of sewage sludge and intermediate leachate mixture at the ratio of 20:1 is
13% higher than the biogas yield using sludge alone (Montusiewicz and Lebiocka, 2011).
Single-stage mesophilic mixed anaerobic digestion rector is extensively used for reduction of
organic sludge volume from wastewater treatment processes (Song et al., 2004). Kheradmand et
al. (2010) combined anaerobic digester under meshophilic condition with an activated sludge
unit and achieved 94% and 93% COD reduction at a loading rate of 2.25 g COD L
-1
d
-1
and 3.37
g COD L
-1
d
-1
respectively. The system also achieved heavy metal removal, however ammonia
was not removed by the combined system. A schematic diagram of the laboratory scale
combined anaerobic and aerobic leachate treatment system is shown in Figure 9.
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The Upflow Anaerobic Sludge Blanket (UASB) reactor has been combined with many physical
and chemical treatment techniques for obtaining higher removal efficiencies (Bohdziewicz and
Kwarciak, 2008; Marañón et al., 2006). Bohdziewicz and Kwarciak (2008) combined UASB
with RO while Marañón et al.(2006) effectively combined nitrificationdenitrification treatment
with UASB reactors to obtain the desired removal standards. The moving-bed biofilm reactor
(MBBR) is an effective biological treatment process, which was developed by combining
conventional activated sludge process and fluidized-bed reactor (Chen et al., 2008; Loukidou and
Zouboulis, 2001). Chen et al. (2008) was able to achieve 92-95% COD removal due to
methanogenesis along with 97% NH
4
-N removal in an anaerobic MBBR.
Lab-scale anoxic rotating biological contactor is highly effective for the removal of nitrate from
a mature landfill leachate and is an example of biological attached growth filter technology
(Teixeira and Oliveira, 2000; Wiszniowski et al., 2006). Cortez et al. (2011) was able to achieve
almost 100% nitrate nitrogen removal efficiencies without nitrite or nitrous oxide accumulation,
however the reactor could not achieve the desired carbon removal standards. In this reactor
ammonium is partly converted to nitrite by ammonium oxidizing bacteria and subsequently the
heterotrophic denitrifying bacteria uses nitrite as the final electron acceptor and nitrogen gas is
released as shown in Equation 17 (Hellinga et al., 1999). In some instances Anammox bacteria
converts ammonium and nitrite directly to nitrogen gas as given in Equation 18 (Strous et al.,
1998; van Dongen et al., 2001).
2NO2-+ 6H+ + 6e- N2 + 2OH- + 2H2O
(17)
NH4+ + 1.31 NO2- + 0.066HCO3- + 0.13H+ 1.02N2 + 0.26NO3- + 0.0066CH2O0.5N0.15 + 2.03 H2O
(18)
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Kim et al. (2006) noted that nitrification treatment in a leachate treatment plant was severely
affected due to high free ammonia content of leachate. At high pH the free ammonia
concentration increases which inhibited nitrite oxidizing and ammonia oxidizing bacteria
especially under high NH
4
-N condition.
The coupling of partial nitration process with Anammox is a very economical process, however
Anammox is not suitable for wastewater with COD and NH
4
-N ratio greater than one (van
Dongen et al., 2001; Xu et al., 2010). Berge et al. (2006) experimented with a completely aerobic
nitrification denitrification bioreactor for NH
4
-N removal from landfill leachate and found that
nitrification- denitrification could occur simultaneously in an aerobic landfill cell, without having
two separate anoxic and aerobic cells.
Liang and Liu (2008) combined a partial nitration reactor, Anammox reactor and two
underground soil infiltration systems. The combined system was effective for leachate treatment
and worked stably over a long period of time under the experimental conditions. The
underground soil infiltration system has low construction and operation expenditure. Due to
complex interplay between hydraulic flow and purification processes of filtration, sorption,
chemical reactions, biotransformation, predation and plant uptake, significantly higher
purification can be attained by the underground soil infiltration systems (Van Cuyk et al., 2001).
Underground soil infiltration system is a promising option for advanced treatment of landfill
leachate.
Puig et al. (2011) used microbial fuel cells to treat landfill leachate containing 6033 mg L
1
of
nitrogen and a conductivity of 73,588  cm
1
,
for production of electricity. The microbial fuel
cell had an air-cathode and was run over a period of 155 days. The system was able to remove up
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to 8.5 kg m
3
d
1
of biodegradable organic matter and generated 344 mW m
3
of electrical
energy.
MSW degradation inside a landfill can be enhanced by leachate recirculation as observed by a
number of researchers who used recirculation bioreactors for the purpose of leachate treatment
(Iglesias et al., 2000; Jiang et al., 2007; Jun et al., 2007; Li et al., 2010a). Jiang et al (2007) made
recirculation reactors by packing landfill waste in anaerobic columns, the schematic diagram of
which is as shown in Figure 10. In another experiment Li et al., (2010) used eight years old aged
refuse excavated from Shanghai Refuse Landfill for leachate treatment. In both the cases
excellent organic removal was observed as discussed in Table 7. Han et al. (2011) modified the
aged refuse biofilter by making it semi-aerobic. This new semi-aerobic aged refuse biofilter
reactor showed superior efficacy for nitrogen removal as compared to other aged refuse biofilter
systems. Sometimes the landfills are engineered to act as bioreactor landfills so as to provide a
more controlled means of reduction in greenhouse gases and methane migration (Warith, 2002).
In bioreactor landfills the stabilization and settlement process of MSW is accelerated by
optimizing the conditions for microbial degradation of MSW, this also allows for additional
MSW disposal or faster land reuse (Kelly, 2002). In both aerobic and anaerobic bioreactors,
leachate recirculation increases the moisture content, distributes nutrients and enzymes
between bacteria and the waste, causes pH buffering, dilutes inhibitory compounds, and
distributes methanogens (Bilgili et al., 2007; Sponza and Agdag, 2004). However, there are
certain disadvantages associated with leachate recirculation such as, too much leachate
recirculation can cause ponding, saturation, accumulation of ammonia nitrogen, development of
acidic conditions and/or the inhibition of methanogenesis due to the accumulation of volatile
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fatty acids (Ledakowicz and Kaczarek, 2002; Reinhart and Al-Yousfi, 1996; San and Onay,
2001; Sponza and Agdag, 2004). Hence, internal leachate characteristic in the solid waste landfill
site during recirculation needs to be done by the introduction of monitoring wells (Sormunen et
al., 2008). In bioreactor landfills clog formation during leachate recirculation can be effectively
controlled by methanogenesis of leachate prior to recirculation (Lozecznik et al., 2010). Khire
and Mukherjee (2007) identified the key design variables for leachate recirculation system in a
landfill consisting of vertical wells using the finite-element model HYDRUS-2D numerical
model.
5.3 Application of physical and chemical processes for leachate treatment
5.3.1 Advance Oxidation Treatments
Advanced oxidation processes (AOPs) is used to enhance the bio-treatability of recalcitrant
and/or non-biodegradable organic substances, through the generation of highly reactive chemical
species, such as hydroxyl radicals (
OH) (de Morais and Zamora, 2005; Deng and Englehardt,
2008; Doocey and Sharratt, 2004; Kurniawan and Lo, 2009; Parsons and M.Williams, 2004;
Wang et al., 2006; Wiszniowski et al., 2004; Yu et al., 1998). The
OH breaks the organic
molecules by abstracting a hydrogen atom or by introducing double bonds in the molecule
(Sarria et al., 2002). The
OH
decompose even the most recalcitrant molecules into biodegradable
compounds such as, CO
2
, H
2
O and inorganic ions (Bauer et al., 1999; Gogate and Pandit, 2004a,
b). There are different ways of producing hydroxyl radicals, which enhances the versatility of
AOPs. Some of the methods by which hydroxyl radicals can be generated are: TiO
2
/UV,
H
2
O
2
/UV, Fenton (Fe
2+
/H
2
O
2
), photo-Fenton (Fe
2+
/H
2
O
2
/ UV), electro-Fenton, electro-photo-
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Fenton and ozone (O
3
, O
3
/UV, and O
3
/H
2
O
2
) (Altin, 2008; Atmaca, 2009; Cho et al., 2002;
Frontistis et al., 2008; Hermosilla et al., 2009; Jia et al., 2011; Kurniawan et al., 2006c; Poznyak
et al., 2008; Tizaoui et al., 2007). A disadvantage of some of the AOPs is the high demand for
electric power, which increases the operational cost of the process (Lopez et al., 2004). However,
the introduction of renewable solar energy as the UV photon source has lowered the demand of
electric power (Rocha et al., 2011). This technique is also known as solar photocatalysis. A
combination of AOP and other treatment process, has been found to be an economical as well as
efficient (Kurniawan et al., 2006c).
Meeroff et al. (2012) experimented with a new technique, photochemical iron mediated aeration
(PIMA) process and compared its efficiency with TiO
2
photoctalysis for both real and simulated
leachate. Table 8 illustrates the efficiency of the technique for real landfill leachate. In another
novel approach, Galeano et al. (2011) experimented the applicability of catalytic wet peroxide
oxidation (CWPO) for leachate treatment. It was found that CWPO treatment in the presence of
Al/Fe-pillared clay catalyst was able to remove 50% COD and simultaneously enhance the
biodegradability of the leachate from 0.135 to 0.321 in 4 h of reaction at 18 °C and 72 kPa.
Among the individual AOPs discussed herein, ozonation and/or Fenton oxidation are the most
commonly applied techniques for leachate treatment. Selection of suitable AOP depends on the
leachate characteristics, technical applicability and other parameters such as, effluent discharge
standards, cost-efficiency, regulatory requirements and long-term environmental impacts.
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5.3.1.1 Ozonation
Ozone is known to degrade organic compounds and is effective for the removal of nitrogen,
color and odour (Haapea et al., 2002; Poznyak et al., 2008; Wang et al., 2002). Ozone has a high
oxidation potential (E
0
) of 2.07V as shown in Equation 19, and can be used for the treatment of
contaminated wastewater of high strength (Al-Kdasi et al., 2004; Camel and Bermond, 1998):
O3+2H+ + 2e- O2 +H2O, E = 2.07 V
o
(19)
However, ozonation alone can remove only 35% COD and 50% NH
4
-N from leachate
(Kurniawan et al., 2006a). So, it is applied in conjunction with other treatment techniques for
better efficiency (Kerc et al., 2003). Application of GAC to ozone treatment improved the
process efficiency by accelerating the kinetic rate of the ozone decomposition through the
formation of nascent
OH radicals which have higher oxidation potential of 2.80V as seen in
Equation 20. It can easily oxidize the organic matter present in leachate (Wang et al., 2004).
.OH+H+ + e- H20, E°= 2.80 V
(20)
Ozone is incapable of degrading humic substances (Wang et al., 2004). However, it is highly
suited for ammonia removal as shown in Equation 21 (Kurniawan et al., 2006a):
NH3 + 4O3- NO3- + 4O2 +H2O + H+
(21)
Ntampou et al. (2006) found that ozonation followed by coagulation-flocculation was less
efficient in COD removal as compared to coagulation-flocculation followed by ozonation, which
could reduce COD from an initial value of 1010 mg L
-1
to less than 180 mg L
-1
.
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5.3.1.2 Fenton Oxidation
Treatment of landfill leachate using Fenton process has been widely reported in recent years (de
Morais and Zamora, 2005; Deng and Englehardt, 2006; Gotvajn et al., 2009; Kang and Hwang,
2000; Kim et al., 2001; Pala and Erden, 2004; Stuber et al., 2005; Sun et al., 2009; Zhang et al.,
2005). The mechanism of free radical generation in a Fenton oxidation reaction involves the
following key steps as illustrated in Equations 22 through 27:
2+
Fe + H2O2 OH + OH-
(22)
Fe3+ + H2O2 Fe2+ + OOH + H+
(23)
Fe3+ + OOH Fe2+ + H+ +O2
(24)
OH + Fe2+ Fe3+ +OH-
(25)
OH + OH H2O2
●●
(26)
OH + H2O2 OOH + H2O
●●
(27)
The OH radical can attack and initiate a series of oxidation reactions leading to the degradation
of the organic pollutant as seen in Equation 28:
OH + RH H2O + R further oxidation

●●
(28)
The primary processes involved for leachate treatment by Fenton Reagent are pH adjustment,
oxidation, neutralization, coagulation and precipitation (Kang and Hwang, 2000). According to
Wu et al. (2010) Fenton treatment is highly effective in removal of about 95.8% HS in 24h
period. The photo-Fenton process is much more efficient than heterogeneous TiO
2
,
TiO
2
/H
2
O
2
/UV or homogeneous H
2
O
2
/UV photocatalysis. The initial reaction rate of photo
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Fenton is 20 times higher and leads to almost complete mineralization of the wastewater (Moraes
and Bertazzoli, 2005; Vilar et al., 2011). The H
2
O
2
molecule is cleaved with a quantum yield of
two OH radicals per quanta of absorbed radiation, as shown in Equation 29 (Esplugas et al.,
2002):
H2O2 +hυ 2 OH
(29)
The OH radicals significantly improve the biodegradability. The BOD
5
/COD ratio improves
from 0.13 to 0.37 or 0.42, which is seen to result in an almost total COD and color removal 

.
5.3.2 Adsorption
Adsorption is recognized as one of the most efficient and extensively used fundamental approach
in wastewater treatment processes (Daifullah et al., 2004; Kurniawan et al., 2006b). Traditionally
activated carbon has been used for leachate treatment due to its large porous surface area,
controllable pore structure, thermal stability and low acid/base reactivity (Li et al., 2008;
Méndez-Díaz et al., 2012). Activated carbon has a superior ability to remove a wide variety of
organic and inorganic pollutants dissolved in aqueous and gaseous environments (Chingombe et
al., 2005; Singh et al., 2012).
Activated carbon adsorption was effective for ammonium nitrogen removal from landfill
leachate samples (Foo and Hameed, 2009). The addition of powdered activated carbon (PAC)
improved the performance of biological treatment of leachate (Kargi and Pamukoglu, 2003a, b).
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Lim et al. (2010) used EDTA modified rice husk in a SBR and achieved better COD and
nitrogen removal efficiency as compared to commercially available PAC.
Activated carbons can be prepared from a large variety of carbon-containing materials through
pyrolysis. Large number of agricultural by-products such as sugarcane bagasse, rice straw,
soybean hulls, rice hulls, peat moss, nutshells and other lignocellulosic wastes has been used to
prepare inexpensive and renewable additional source of activated carbons (Ahmedna et al., 2000;
Kadirvelu et al., 2003; Sahu et al., 2010). Activated carbon made from tamarind wood and
chemically activated by zinc chloride was used for the removal of lead and chromium from
wastewater with significant success (Dwivedi et al., 2008; Sahu et al., 2009a; Singh et al., 2008).
Other low cost adsorbents that has been successfully used for heavy metal removal are peat and
rubber wood ash (Hasan et al., 2000; Sen Gupta et al., 2009). These adsorbent may also be used
for the treatment of leachate. A basic two stage process consisting of carbonization followed by
activation is followed for the production of activated carbons. In the first step the carbon content
is enriched for the creation of an initial porosity and second activation stage helps in enhancing
the pore structure (Acharya et al., 2009a; Acharya et al., 2009b). Some reviews have been
published on the preparation of activated carbon, which can be subsequently utilized for leachate
treatment (Demirbas, 2009; Dias et al., 2007).
In addition to activated carbon other materials like clinoptilolite, Zeolite (CV-Z) synthesized
from coal fly ash , limestone, peat, blast furnace slag and pine bark have been utilized for
leachate treatment with good results (Aziz et al., 2004b; Heavey, 2003; Karadag et al., 2008;
Luna et al., 2007; Nehrenheim et al., 2008; Orescanin et al., 2011; Sõukand et al., 2010).
Clinoptilolite has a high NH
4
-N removal efficiency (Hankins et al., 2005). Li et al. (2011b) used
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coal flyash, treated with initiator C for landfill leachate treatment. The efficiency of the above
mentioned adsorbents is discussed in Table 9. Oti et al. (2011) used an iron oxide based
adsorbent Kemiron for the removal of As(V) and As(III) from leachate. Fuller earth beads and
cylinders containing chitosan and sodium silicate as binders was used successfully by Hasan et
al. (2007) for the removal of cesium from wastewater. This can also be replicated for leachate
treatment.
Composite adsorbent media made by combining different materials like zeolite and activated
carbon, carbon and low-cost materials such as limestone or rice husk, carbon waste with Portland
cement as a binder and so on (Azhar et al., 2006; Gao et al., 2005). The combinations of
hydrophilic and hydrophobic groups in the adsorbents make an excellent adsorption system
which can remove both metallic ions and organic substances (Okolo et al., 2000). Studies show
that ammoniacal nitrogen was better adsorbed by composite adsorbents towards than zeolite and
activated carbon (Halim et al., 2010a). Halim et al. (2010b) studied the performance of such
composite adsorbent media via a lab-scale column study which is shown schematically in Figure
11.
Studies have shown that the combination of activated carbon and ozone is a suitable and feasible
option for the treatment of landfill leachate (Fettig et al., 1996; Rivas et al., 2003). Addition of
PAC to activated sludge reactors has shown to enhance the biological treatability of leachate
. Sahu et al. (2009b) used activated rice husk in a three phase modified
multi-stage bubble column reactor and achieved 77.15% and 19.05% lead and BOD
5
reduction
respectively, under optimum conditions. This technique can also be used for leachate treatment,
specifically for the removal of HMs. Li et al. (2010b) applied coagulation flocculation followed
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by adsorption using PAC and obtained 86%, 97.6%, 99.7% and 78%, removal of COD, Pb, Fe
and toxicity respectively under optimum operating conditions.
5.3.3 Coagulation-flocculation
Coagulation and flocculation have been used successfully in treating stabilized and old landfill
leachates and is most effective for colour removal (Kang and Hwang, 2000; Manu and
Chaudhari, 2002; Monje-Ramirez and Velásquez, 2004; Silva et al., 2004). The different types
of coagulation processes include classical chemical coagulation using salts of iron and
aluminium, electrocuagulation and biocoagulation. Four major types of chemical coagulants are
aluminium (III) sulfate (alum), ferric (III) chloride, ferrous (II) sulfate and ferric (III) sulfate.
Studies have shown that ferric (III) sulfate has the highest coagulation efficiency followed by
aluminium (III) sulfate and ferric (III) chloride (Comstock et al., 2010). Tatsi et al. (2003)
worked with three conventional coagulants viz., ferric chloride, aluminium sulfate and lime and
four commercial polyelectrolytes among whom one was anionic, two cationic and another was
non-ionic polymer. He found that although ferric chloride removed 80% COD from partially
stabilized leachate, the removal decreased below 35% when coagulants were added to raw
leachate.
Zouboulis et al. (2004) experimented with bioflocculants produced by the
bacterium Rhizomonas sp. The application of bioflocculant was efficient for the removal of
humic acids from synthetic solutions and reducing COD content from real landfill leachates.
More than 85% humic acid removal was observed at 20 mg L
-1
bioflocculant dose and at pH 7-
7.5.
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Electrocoagulation is a simple and efficient electrochemical method used for the purification of
many types of water and wastewaters and is able to remove large variety of pollutants (Adhoum
and Monser, 2004; Alinsafi et al., 2005; Bayramoglu et al., 2006; Can et al., 2006; Daneshvar et
al., 2006; Ilhan et al., 2008; Kobya et al., 2006; Li et al., 2011a). In electrocoagulation, electric
current destabilizes the suspended, emulsified, or dissolved contaminants in the wastewater
(Emamjomeh and Sivakumar, 2009). Mariam and Nghiem (2010) achieved about 67% TOC and
80% turbidity removals by the electrocoagulation while the removal percent by chemical
coagulation was only 10% TOC and 65% turbidity. The treatment of leachate is easier due to
their high conductivity and chloride content (Labanowski et al., 2010). Several materials have
been used as anode such as Pt, TiO
2
, SnO
2
, Al and Fe. Among them, Al and Fe are most
frequently used (Top et al., 2011). The COD removal for Fe and Al electrodes were 35% and
56% respectively, in 30 min contact time as discussed in Table 11. Fe electrodes transfer higher
numbers of Fe ions into solution leading to higher rate of electrode dissolution, formation of
more sludge with less COD removal. Since, the costs of both Al and Fe electrodes are
comparable, Al electrodes will be a better choice due to its higher efficiency (Ilhan et al., 2008).
However, Bouhezila et al. (2011) estimated a higher operational cost for Al electrode, thus
preferring Fe electrode material.
Coagulation is also used as a pre and post treatment technique for membrane filtration to achieve
higher removal efficiency (Mariam and Nghiem, 2010; Theepharaksapan et al., 2011; Top et al.,
2011). Vedrenne et al. (2012) used chemical coagulation-flocculation with ferric (III) chloride in
conjunction with photo Fenton oxidation and was successful in removing about 56% of COD,
95% TC, 64% NH
4
N, 46% As, 9% Hg and 85% Pb from an aged leachate sample.
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Dissolved air flotation (DAF) technique is used in conjugation with various coagulation-
flocculation techniques to separate the flocculated particles from the wastewater, by bringing the
particles to the surface of the liquid. DAF is also helpful in reduction of BOD
5
, COD and
turbidity (Al-Shamrani et al., 2002a, b; Palaniandy et al., 2010). Studies show that separation by
flotation presents some advantages compared to separation by settling (Pouet and Grasmick,
1995). Adlan et al. (2011) combined chemical coagulation by ferric (III) chloride and DAF for
the treatment of semi-aerobic leachate.
5.3.4 Electrochemical treatment
Stabilized or methanogenic leachates are alkaline and have less than 1% of biodegradable
organic matter as evident by BOD/COD value of 0.004, making electrochemical treatment
techniques more feasible (Tauchert et al., 2006). According to a number of researchers,
electrochemical oxidation of leachate is superior to light-enhanced oxidation, Fenton treatment,
combined UV and O
3
/H
2
O
2
, ultrasound and other physico-chemical processes since it can
efficiently reduce concentrations of organic contaminants, ammonia, and color in leachate
(Gonze et al., 2003; Ince, 1998). Pretreatment techniques, anode materials, pH, current density,
chloride concentration, and additional electrolytes significantly influence the performance of
electrochemical oxidation. During electro-oxidation treatment of leachate, COD reduction can
range from 70% up to >90% and the achieved NH
3
N removal efficiency is almost 100%, under
optimum conditions (Chiang et al., 2001; Ihara et al., 2004).
According toFeng et al. (2003) direct oxidation of organic matter at the anode surface is also
possible. Several anode materials have been used for electrocoagulation, such as boron- doped
diamond binary RuTi oxide-coated titanium anode also called the Dimensional Stable Anode
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(DSA) , Ti/SnO
2
and Ti/PbO
2
, Ti/Pt, graphite and PbO
2
and SnPdRu oxide coated titanium
(SPR), graphite and DSA (Anglada et al., 2011; Cabeza et al., 2007b; Chiang et al., 1995; Cossu
et al., 1998; Feki et al., 2009; Feng et al., 2003; Moraes and Bertazzoli, 2005; Pérez et al., 2010;
Tauchert et al., 2006).
During the electrolysis, the pollutants are degraded either by direct or indirect oxidation
processes as shown in Figure 12 (Chen, 2004; Deng and Englehardt, 2007; Szpyrkowicz et al.,
2001). Deng and Englehardt (2007) found that NH
4
-N removal is higher than COD removal,
indicating the dominance of indirect oxidation during electrolysis reaction. The hypochlorite ion
or hypochlorous acid generated during electrochemical oxidation is the main oxidizing agents:
--
2
2Cl Cl + 2e
(30)
- + -
2
2Cl + H O HClO+ H + Cl
(31)
+-
HClO H +ClO
(32)
The chlorine and hypochlorite oxidize NH
4
+
and are reduced to chloride ions in the process as
given in Equation 33 (Cabeza et al., 2007a; Chen, 2004)
+ + -
4 2 2
2NH + HClO N +2H O+6H +2Cl
(33)
Schoeman et al. (2005) experimented with electrodialysis to desalinate/concentrate the leachate
to effectively reduce the volume pollution control. However, there are two basic drawbacks of
electro-oxidation viz., high energy consumption and possible formation of chlorinated organics
(Deng and Englehardt, 2007). For treating old stabilized landfill leachate, Orescanin et al. (2012)
pre-treated extremely low biodegradable leachate with ozone, followed by simultaneous
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ozonation and electro-oxidation and it was finally subjected to microwave treatment. The
removal percentages obtained were 98.43% colour, 99.48% turbidity, 98.96% suspended solids,
98.80% ammonia, 94.17% COD and 98.56% iron. However, this process uses complex treatment
schedule, high energy and much resource.
5.3.5 Filtration and membrane bioreactors
In recent years advance treatment techniques like, membrane filtrations which were originally
used for of drinking water purification are being applied for leachate treatment. Nanofiltration,
ultrafiltration and reverse osmosis are the major membrane filtration techniques that applied for
leachate treatment. Among them, reverse osmosis is considered to be the most promising
treatment technique available in recent years due to its high removal pollutant efficiency (Chan
et al., 2007; Jenkins et al., 2003; Renou et al., 2008a; Renou et al., 2008b; Ushikoshi et al.,
2002). However, lecahte treatment by involves high pre and post treatment cost and frequent
membrane fouling also affects its performance (Trebouet et al., 2001). It was found that
membrane fouling is increased if the humic acid concentration in the leachate increases 
al., 2012). Frequent membrane fouling in reverse osmosis can be overcome by the application of
vibratory shear-enhanced processing reverse osmosis (VSEPRO) system for treating stabilized
leachate. Leachate containing recalcitrant organics can be effectively treated in a VSEPRO
system due to the shearing force (Chan et al., 2007).
Nanofiltration exhibits treatment characteristics between reverse osmosis and ultrafiltration
(Zouboulis and Petala, 2008). Studies have shown that nanofiltration is highly efficient in
removal of metals like K
+
and Na
+
and boron from landfill leachate (Dydo et al., 2005; Ortega et
al., 2007). Zouboulis and Petala (2008), found that the application of vibratory shear enhanced
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unit (VSEP) on nanofiltration membranes enhanced the treatment efficiency of raw stabilized
leachate. The humic acid removal efficiency was about 97%. The VSEP unit also prevented
membrane fouling by creating shear waves (Zouboulis and Petala, 2008). Xu et al. (2006) found
that Humic substances (HS) in mature leachate from inorganic components could be effectively
removed by ultrafultration.
The addition of successive membrane operations to biological treatments offered new advantage
in the field of landfill leachate treatment (Bodzek et al., 2006) and the combination is called
Membrane Bioreactors (MBR) (Tarnacki et al., 2005). A MBR thus combines the goodness of a
biological reactor and membrane filtration system. The presence of the membrane allows for
long sludge retention time with high organic loading rate and low hydraulic retention time.
According to Robinson (2007) landfill leachate treatment can be highly challenging for MBRs as
high chloride content of the leachate may corrode the membrane system. However Ahmed and
Lan (2012) reported that excellent organics (BOD) and ammonia removal capacity up to 90% or
more can be achieved by MBRs even when dealing with mature or stabilized landfill leachate. In
recent years much attention has been given to MBRs for landfill leachate treatments owing to
their efficiency and small foot-print (Ahn et al., 2002; Alvarez-Vazquez et al., 2004;
Chaturapruek et al., 2005; Melin et al., 2006; Robinson, 2005; Setiadi and Fairus, 2003; Vasel et
al., 2004). Various authors have worked with MBRs obtaining high removal efficiency as cited
in Table 12.
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6 Summary and Discussion
Landfill leachate is extremely toxic due to high concentration of recalcitrant organics and
ammonia nitrogen along with variable quantities of other phosphorus, chlorides, calcium,
magnesium, sulfate, dissolved solids, heavy metals, BTEX and other xenobiotic compounds. In
view of the grave impact of landfill leachate on environment, the regulatory authorities have
been forced to fix increasingly stringent discharge water standards. In developed countries,
directives regarding prevention of leachate seepage into groundwater and soil, collection,
treatment and its disposal exist to some extent. A discussion is provided in Table 3 regarding the
maximum limit of contaminants in treated leachate prior to its disposal into the surrounding
environment. However, due to extreme variation of leachate composition and operating
conditions in different landfills, no guideline or standard operating procedures for leachate
treatment and disposal can be effectively chalked out. While most of the old landfills do not
contain adequate pollution containment mechanisms, these safety considerations are being
integrated into the new landfills during the design phase. So management of old and new
landfills and their troubleshooting should follow different approaches which have been shown in
the Figure 9.
1. Leachate plumes have a widely varying characteristic and composition. Both vertical and
horizontal gradient in redox potential and contaminant concentration dictates the transformation
of nitrogenous, sulfurous, carbonaceous and heavy metal species along the leachate plume.
While amoonium compounds undergo aerobic nitrification, nitrate reduction, anoxic
denitrification and anaerobic ammonium oxidation processes to form harmless nitrogen gas
under fluctuating redox conditions, the sulfate reduction depends on available organic electron
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donors and sulfate electron acceptors. Carbonaceous compounds or organics in the leachate
plume is reflected by the COD which keeps on decreasing over age of the landfill due to natural
anaerobic methane oxidation and natural attenuation. The HMs are found to undergo very less
mobilization as they became stabilized by complexing with DOM, HA and FA.
2. The leachate plume migration can be monitored by using a large number of techniques and
methods. The monitoring techniques are site specific and each landfill site should be carefully
studied before the application of any specific monitoring technique. Construction of monitoring
wells or insertion of hollow stem augers are very common and essential for sampling purposes
and for inserting various probes and electrodes for geo-chemical and electrical monitoring
techniques. Hydro-geological equipment such as piezometers and various samplers are
historically the most used instruments. Isotope mapping and electrical monitoring such as
tomography, ERI, VLF-EM, electrode grid, etc are comparatively new, but very convenient field
techniques. The electromagnetic methods such as GPR, RCPT and TDR can be performed
without monitoring wells and permanent facilities. Sometimes, two or more of these techniques
can be used to complement each other and obtain a clearer picture regarding leachate plume
migration. Bacteriological monitoring can also point out the fringe of the leachate plume by
distinct degradation potentials inside and outside of leachate plume. The suitability of these
different monitoring methods will vary from site to site depending upon groundwater flow, soil
porosity, pore water content, electrical conductivity of soil matrix, soil texture, and logistic
issues.
3. Landfill leachates pose significant risk towards the soil and groundwater environment. It is
well established fact that small amount of leachate can pollute a large volume of groundwater
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once it infiltrates an aquifer by changing its pH and Eh and introducing toxic chemicals. Soil
texture, porosity, permeability and HRT changes mostly due to bioclogging from biomass and
biofilm produced by microbes, gas pocket formation and metal precipitation. Additionally, water
bodies present near landfills may experience higher organic load, inorganic nitrogen content, and
heavy metal concentration.
4. In order to assess the extent of impact of landfill leachate on environment, both qualitative and
quantitative methods are available. However, none of them guarantees an exact assessment of the
actual scenario due to extreme complexity of the leachate plume and soil environment. Relative
hazard assessment systems rank a number of landfills by a comparative rating system to
prioritize the treatment efforts. Around 22 hazard-rating systems have been cited in section 4.2
and four systems have been discussed in details, viz. LPI, E-LI, hazard rating by Singh et al.
(2009) and a toxicity index. All of them stress upon different factors. While some concentrates
on the environment as a whole, some other specializes on the toxic effect of leachate on human
beings. Necessity would decide which hazard rating system is to be used. However, the
subjectivity associated with the scoring system of these hazard rating systems is their main
drawback. In most of the systems, site ranking is based either on the combined score for various
routes under migration mode or the score for the dominant route i.e. the route returning highest
score.
5. Numerous mathematical models that have been developed for different issues related to risk
assessment of landfill leachate are completely dependent on the data input. The results can be
misleading if any input is wrong and the complex chemical and biochemical processes
undergoing in the landfill is predicted wrongly. In this paper, we have reviewed few
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mathematical models for assessing permittivity reduction of soil, degradation of leachate
pollutants, long term fate of leachate components, reliability of groundwater monitoring systems
and also softwares used for modeling purpose. The use of softwares is supposed to be a very
good option. However, in spite of presence of a number of softwares in the market, none is
exactly suitable for leachate plume modeling and a lots of adjustment is required to work with
these generic softwares. These stochastic models should be used for guesswork in case the
leachate composition and biogeochemical and bacteriological processes are fully understood.
Otherwise, the management decisions taken based on the wrong predictions may cost dear.
6. Leachate control systems may include installation of geo-synthetic or other liners at the
bottom of the landfill and leachate collection systems. Treatment of leachate prior to discharge to
surface water is also an integral part of that system (Damgaard et al., 2011). According to the
Department of Environment Food and Rural Affairs (UK) landfills both hazardous and non-
hazardous should have a bottom liner in addition to the geological barrier (DEFRA, 2009). The
danger of leachate infiltration in groundwater is great considering that even the best liner and
leachate collection systems will ultimately fail due to natural deterioration. Nooten et al. (2008)
proposed a semi-passive treatment of leachate during post closure remediation of old landfills,
thereby replacing conventional energy consuming wastewater treatment systems. The system can
also be installed along the gradient of leaking landfills for mitigation of contaminated
groundwater plumes. In another novel approach Ziyang et al. (2011) proposed the introduction of
functional layers embedded in landfill so that leachate strength may be reduced source, thereby
reducing the cost of leachate treatment. Leachate treatment techniques differ depending on the
nature and age of leachate. Biological treatments are most suitable for treatment of young
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leachate while physico chemical treatments like membrane filtration, electrochemical and
advanced oxidation treatments are suitable for stabilized acidogenic leachate. Membrane
filtration in combination with biological treatment was found to be extremely effective.
However, installation of membrane treatment facilities is much expensive than other treatment
techniques. The treatment costs of landfill leachate will vary depending on its capacity and the
composition of waste it has to deal with. Other factors that will contribute towards determining
the treatment cost include the technology employed, the local condition of the site, and the
disposal standards it has to comply with. The total treatment cost will take into account the
construction as well as operational and maintenance costs. While the construction cost usually
depends on the capacity of the landfill and target quality of the effluent, the operation and
maintenance cost will cover manpower, energy, chemicals and maintenance over its lifetime and
even after its closure.
Acknowledgements
The authors are grateful to University of Malaya, Malaysia (Project No:
UMC/HIR/MOHE/ENG/13 and UM-QUB6A-2011) for providing the financial support to carry
out the work.
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80
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