university of southampton research repository eprints … · an investigation of recalcitrant...
TRANSCRIPT
University of Southampton Research Repository
ePrints Soton
Copyright © and Moral Rights for this thesis are retained by the author and/or other copyright owners. A copy can be downloaded for personal non-commercial research or study, without prior permission or charge. This thesis cannot be reproduced or quoted extensively from without first obtaining permission in writing from the copyright holder/s. The content must not be changed in any way or sold commercially in any format or medium without the formal permission of the copyright holders.
When referring to this work, full bibliographic details including the author, title, awarding institution and date of the thesis must be given e.g.
AUTHOR (year of submission) "Full thesis title", University of Southampton, name of the University School or Department, PhD Thesis, pagination
http://eprints.soton.ac.uk
This thesis was submitted for examination in November 2009.
An investigation of recalcitrant organic compounds in leachates
Anika Yunus
PhD Thesis, November 2009
UNIVERSITY OF SOUTHAMPTON
FACULTY OF ENGINEERING, SCIENCES AND MATHEMATICS
SCHOOL OF CIVIL ENGINEERING AND THE ENVIRONMENT
An investigation of recalcitrant organic
compounds in leachates
by
Anika Yunus
A thesis submitted for the degree of
Doctor of Philosophy
November 2009
I
UNIVERSITY OF SOUTHAMPTON ABSTRACT
FACULTY OF ENGINEERING, SCIENCES AND MATHEMATICS SCHOOL OF CIVIL ENGINEERING AND THE ENVIRONMENT
Doctor of Philosophy
AN INVESTIGATION OF RECALCITRANT ORGANIC COMPOUNDS IN LEACHATES
By Anika Yunus
Recalcitrant organic compounds remain a key challenge in landfill leachate management as they are resistant to microbial degradation and have potential to damage the water environment. Conventional leachate characterisation methods are time consuming and limited by their inability to provide compositional analysis. This research therefore investigates the characteristics of recalcitrant organic compounds in leachates and undertakes a feasibility study of the possible use of UV absorption and fluorescence spectroscopy for a rapid and economical compositional analysis of recalcitrant organic compounds. Two laboratory experiments are carried out in this regard.
In one experiment, a laboratory scale aerobic biodegradation is carried out on four untreated and two treated leachate samples collected from two UK MSW landfills (Pitsea and Rainham), and leachates are characterised using conventional methods (COD and DOC) as well as UV absorption and fluorescence spectroscopy. It is found that the leachates which have low organic content are easily biodegradable whereas the leachates which have high organic content are not easily biodegradable. UV and fluorescence spectroscopy allow compositional analysis of recalcitrant organic compounds and show that humic and fulvic compounds are the key components of the recalcitrant organic compounds. A rapid biodegradability assessment of different leachates using these novel techniques is in agreement with conventional method showing that a rapid characterisation of leachates using spectroscopic methods is feasible. It is also found that organic compounds in these leachates are aromatic in nature and the leachates containing large amounts of aromatic compounds, condensed aromatic structures are difficult to degrade. In another experiment, the influence of the solid waste component on the development of recalcitrant organic compounds in leachates is investigated by carrying out an anaerobic biodegradation experiment for fresh waste, composted waste, newspaper waste and synthetic waste. Composted waste contributes significantly to the development of the recalcitrant compounds due to the removal of readily biodegradable organic compounds during composting and hence the increased proportion of biologically resistant compounds during the subsequent anaerobic biodegradation. Newspaper waste also contributes significantly to the presence of recalcitrant organic compounds due to the relatively less resistant cellulose and high lignin present in this waste.
This research validates the application of UV and fluorescence spectroscopy for rapid on-site monitoring of landfill leachates that would help scientist and engineers to assess leachate quality, identify various organic compounds and to optimise leachate treatment processes. The analysis performed on Pitsea and Rainham leachates is a promising step towards developing a database of representative information of characteristics of recalcitrant organic compounds in leachates for different UK landfills. This research also provides an understanding of the composition of leachates from different wastes and it can be conclude that disposals enriched with composted and newspaper waste would favour the development of leachates with high concentrations and condensed aromatic structures of recalcitrant organic compounds in a landfill system.
II
Contents Abstract
I
Contents
II
List of figures
V
List of tables
VIII
List of abbreviations
IX
Declaration XI Acknowledgements
XII
Dedication XIII 1 Introduction 1 1.1 Scope of the thesis 1 1.2 Organisation of the thesis 5 2 Background 7 2.1 Introduction 7 2.2 Leachate generation in landfill 7 2.3 Landfill leachate characterisation and the origin of recalcitrant organic
compounds 11
2.3.1 Biochemical Oxygen Demand (BOD) 11 2.3.2 Chemical Oxygen Demand (COD) 11 2.3.3 Total Organic Carbon (TOC) 12 2.4 Composition of recalcitrant organic compounds and its effect on the
environment 14
2.5 Effect of treatment on removal of recalcitrant organic compounds 15 2.6 Limitation of conventional characterisation methods (BOD, COD,
TOC) 19
2.7 Compositional analysis of recalcitrant organic compounds and/or humic substances
20
2.8 Proposed technique of leachate characterisation 23 2.8.1 UV visible spectroscopy 24 2.82 Fluorescence spectroscopy 25 2.9 Summary 30 3 Characterisation of wastes and leachates 32 3.1 Introduction 32 3.2 Aerobic biological leachate treatment 32 3.2.1 Materials (Landfill sampling sites and leachate composition) 32 3.2.2 Experimental methods 34 3.2.3 Preparation of BOD water 34 3.3 Biochemical Methane Potential (BMP) test 35 3.3.1 Material (Pre-test waste sample preparation procedure) 36
III
3.3.2 Experimental procedure 36 3.3.3 Leachate sampling 40 3.3.4 Solid waste sampling 40 3.4 Analytical measurements 43 3.4.1 Biochemical oxygen demand (BOD) analysis 43 3.4.2 Chemical oxygen demand (COD) analysis 43 3.4.3 Carbon analysis 44 3.4.4 Solids content 45 3.4.5 Anion analysis (Cl-, NO2
-, NO3-, PO4
-2, SO4-2) 45
3.4.6 Cation analysis (Na+, NH4+, K+, Mg+2, Ca+2) 45
3.4.7 Elemental analysis 46 3.4.8 Fibre analysis (cellulose, hemi-cellulose and lignin) 46 3.4.9 Gas analysis 47 3.4.10 Spectroscopic characterization of leachate DOC 48 3.4.10.1 UV-visible spectroscopy
3.4.10.2 Fluorescence spectroscopy 48 48
3.5 Carbon mass balance analysis 49 4 Results and Discussion: Aerobic biodegradation of landfill leachates 51 4.1 Introduction 51 4.2 Initial characterisation of landfill leachate samples 52 4.2.1 pH values 52 4.2.2 Organic contents in leachates 52 4.2.3 Solid contents 54 4.2.4 Anions and Cations 55 4.3 Effect of a laboratory scale aerobic biological treatment in the change
of the nature of recalcitrant organic compounds in landfill leachates 58
4.3.1 Conventional COD and DOC analyses 58 4.3.2 Spectroscopic analysis 63 4.3.2.1 UV spectroscopic results 64 4.3.2.2 Fluorescence spectroscopic results 67 4.3.2.3 Fluorescence analyses of humic and fulvic-like
compounds in emission scanning mode 76
4.4 Feasibility assessment of Spectroscopic method 81 4.5 Relationship of spectroscopic methods with COD and DOC 82 4.5.1 Relationship between fluorescence intensity and DOC 82 4.5.2 Relationship between fluorescence intensity and COD 87 4.5.3 Relationship between UV absorbance at 254 nm (UV254) and
COD 87
4.6 Summary 90
5 Results and Discussion: Anaerobic biodegradation of solid waste 91 5.1 Introduction 91 5.2 Characterisation of waste biodegradation in BMP reactors 92 5.2.1 Initial waste characterisation 92 5.2.2 Evolution of biochemical fractions during anaerobic
biodegradation of wastes 94
5.3 Characterisation of leachates formed in BMP reactors 99 5.3.1 pH of leachate samples 99
IV
5.3.2 Carbon contents of leachate samples 100 5.4 Mass balance estimate for the BMP reactors 105 5.5 Spectroscopic analysis 107 5.5.1 Fluorescence spectroscopic results 107 5.5.1.1 Excitation-Emission-Matrix (EEM) spectra of
control samples 107
5.5.1.2 Excitation-Emission-Matrix (EEM) spectra of leachate generated from all waste reactors
111
5.5.1.2.1 Comparative analyses of Excitation-Emission-Matrix (EEM) spectra of leachates generated from all wastes
114
5.5.1.3 Fluorescence analyses of humic and fulvic like materials in leachates in emission scanning mode
127
5.5.2 UV spectroscopic results 131 5.6 Summary
133
6 Conclusion and future work 135 6.1 Conclusion 135 6.2 Future work
141
A Appendix A: EEM spectra of landfill leachates 143 B Appendix B: EEM spectra of leachates from BMP reactors
145
References 156
V
List of figures
2.1 Changes in selected parameters during the phases of landfill stabilization 9
2.2
Fluorescence EEM for a hypothetical leachate sample to show all possible
fluorescence centres (1) Fulvic-like (at 320-370 excitation and 400-470
emission) (2) Humic-like (at 230-260 and 300-370 excitation and 400-470
emission) (3) Tyrosine-like (at 270-280 excitation and 305-312 emission)
and (4) Tryptophan-like (at 270-280 excitation and 340-380 emission)
27
3.1 Analyses carried out on the leachates taken from aerobic reactors 35
3.2 BMP test assay apparatus 38
3.3 Analyses carried out on the leachates taken from BMP reactors 41
3.4 Analyses carried out on the solid samples from BMP reactors 42
4.1 (a) The change in COD over time during aerobic biodegradation for all of
the untreated and treated leachate samples 60
(b) The change in TOC and DOC over time during aerobic biodegradation
for all of the untreated and treated leachate samples 61
c) Aerobic biodegradation of landfill leachates for 30 days (% COD
removal) 62
(d) Aerobic biodegradation of landfill leachates for 30 days (% DOC
removal) 62
4.2 (a) Absorbance at 254 nm (UV254) for all of the untreated and treated
leachate samples 65
(b) % UV254 removal for all of the untreated and treated leachate samples 65
4.3 Excitation-emission matrices (EEM) for landfill leachates before and after
aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like)
69-
70
4.4 Fluorescence spectra in emission scanning mode for all of the untreated
and treated leachate samples
78-
80
4.5 Comparisons of the fluorescence peak with DOC for all of the untreated
and treated leachate samples 84
4.6 Comparisons of the Trp-L (270-280 nm excitation) fluorescence with
DOC for all of the untreated and treated leachate samples 85
4.7 Comparisons between H-L/F-L and Trp-L (270-280 nm excitation)
fluorescence intensity for all of the untreated and treated leachate samples 86
VI
4.8 Comparisons of the fluorescence peaks with COD for all of the untreated
and treated leachate samples 88
4.9 Comparisons of the UV254 with COD for all of the untreated and treated
leachate samples 89
5.1 Cumulative gas production at STP for all of the wastes and control (blank)
in BMP reactors 94
5.2 NDF, ADF and ADL over time for fresh waste, composted waste,
newspaper waste and synthetic waste samples 95
5.3 Cellulose, hemi-cellulose and TC over time for fresh waste, composted
waste, newspaper waste and synthetic waste samples 97
5.4 (C+H)/L over time for fresh waste, composted waste, newspaper waste
and synthetic waste samples 98
5.5 pH of leachates taken from BMP reactors 99
5.6
Variation of total and dissolved organic and inorganic carbon of leachates
with time for fresh waste, composted waste, newspaper waste and
synthetic waste samples
101
5.7 Variation of COD of leachates with time for fresh waste, composted
waste, newspaper waste and synthetic waste samples 103
5.8 Comparisons of TOC with COD of leachates taken from fresh waste,
composted waste, newspaper waste and synthetic waste samples 104
5.9
3D EEM fluorescence spectra of control (blank) reactors at day (a) 5, (b)
10 and (c) 150 showing Tyr-L and Trp-L fluorescence; (d) methanogenic
mineral media at day 0
109
5.10 Variation of Tyr-L and Trp-L fluorescence intensity over time in the
control (blank) reactors 110
5.11 3D EEM fluorescence spectra of fresh waste (FW) leachates at day (a) 50
and (b) 150 116
5.12 3D EEM fluorescence spectra of composted waste (CW) leachates at day
(a) 50 and (b) 150 117
5.13 3D EEM fluorescence spectra of newspaper waste (NW) leachates at day
(a) 50 and (b) 150 118
5.14 3D EEM fluorescence spectra of synthetic waste (SW) leachates at day (a)
50 and (b) 150 119
5.15 A possible degradation pathway of waste organic matter into the 122
VII
recalcitrant organics in leachates
5.16
Comparisons of the fluorescence intensities with DOC for (a) fresh waste
(b) composted waste (c) newspaper waste and (d) synthetic waste
leachates
123-
124
5.17 H-L/F-L ratio and phases in the fresh waste, composted waste and
newspaper waste leachates 126
5.18
Fluorescence spectra of (a, b) fresh waste (c, d) composted waste (e, f)
newspaper waste and (g) synthetic waste leachate samples in emission
scanning mode at excitation wavelengths 230-260 nm and 320-390 nm
129-
130
5.19 The SUVA for fresh waste, composted waste, newspaper waste and
synthetic waste leachates 131
5.20 E4/E6 ratio for fresh waste, composted waste, newspaper waste and
synthetic waste leachates 132
A1 Excitation-emission matrices (EEM) for landfill leachates at day 10 and
20 during aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like)
143-
144
B1
3D EEM fluorescence spectra of control (blank) reactors at day (a) 50,
(b) 60, (c) 70, (d) 80, e) 90, (f) 100, (g) 120 and (h) 140 showing Trp-L
fluorescence at 220-240 nm and 270-280 nm excitation and 340-370 nm
emission
145-
146
B2 3D EEM fluorescence spectra of fresh waste (FW) leachates at day 60, 70,
80, 90, 100, 120 and 140 for control uncorrected and corrected samples
147-
149
B3 3D EEM fluorescence spectra of composted waste (CW) leachates at day
60, 80 90, 100, 120 and 140 for control uncorrected and corrected samples
150-
151
B4 3D EEM fluorescence spectra of newspaper waste (NW) leachates at day
70, 90, 100, 120 and 140 for control uncorrected and corrected samples
152-
153
B5 3D EEM fluorescence spectra of synthetic waste (SW) leachates at day
70, 90, 100, 120 and 140 for control uncorrected and corrected samples
154-
155
VIII
List of tables
2.1 A relative comparison among three methods (BOD, COD, TOC) 12
2.2 Studies on landfill leachate biological treatment (aerobic and
anaerobic) 16
2.3 Studies on landfill leachate integrated physical-chemical-biological
treatment 17
2.4 Studies on landfill leachate treatment with O3 19
2.5 Common methods for the characterization of humic substances 21
3.1 Characteristics of landfills and leachate samples (Information
provided by landfill operators) 33
3.2 Types of waste samples 37
3.3 Breakdown of waste composition for fresh waste samples 37
3.4 Frequency of leachate analysis 38
3.5 Recipe of the methanogenic mineral medium 39
4.1 Chemical composition of landfill leachate samples 53
4.2 Chemical characteristics of leachates from different landfills 57
4.3 Fluorescence properties of landfill leachate samples during aerobic
treatment 73
5.1 Initial composition of solid waste components 93
5.2 Composition of biogas produced from all of the wastes (methane and
carbon dioxide) 105
5.3 Summary of carbon mass balance estimates for waste samples 106
5.4
Intensities and position of the EEM peaks for leachates taken from
fresh waste, composted waste, newspaper waste and synthetic waste
reactors
112-113
IX
List of abbreviations
ACD Acid Detergent Fibre
ADL Acid Digestible Lignin
BOD Biochemical Oxygen Demand
BMP Biochemical Methane Potential
COD Chemical Oxygen Demand
CW Composted Waste
DCOD Dissolved Chemical Oxygen Demand
DOC Dissolved Organic Carbon
DOM Dissolved Organic Matter
EEM Excitation-Emission-Matrix
FAS Ferrous Ammonium Sulphate
FE Final Effluent
FW Fresh Waste
F-L Fulvic-Like
IFE Inner Filtering Effect
Haz Hazardous
H-L Humic-Like
LTP Leachate Treatment Plant
MSW Municipal solid waste
MW Molecular weight
Nap-L Naphthalene-Like
NDF Neutral Detergent Fibre
NMR Non Magnetic Resonance
NW Newspaper Waste
PAE Phthalates
PAH Polycyclic Aromatic Hydrocarbon
PCB Polychlorinated Biphynyls
PVC Polyvinyl Chloride
P2 Phase 2
X
P4 Phase 4
SUVA Specific UV absorbance
SW Synthetic Waste
TC Total Carbon
TCD Thermal Conductivity Detector
TCOD Total Chemical Oxygen Demand
TDS Total Dissolved Solid
TFS Total Fixed Solid
TIC Total Inorganic Carbon
TN Total Nitrogen
TOC Total Organic Carbon
Trp-L Tryptophan-Like
TS Total Solid
TSS Total Suspended Solid
TVS Total Volatile Solid
Tyr-L Tyrosine-Like
UV Ultra Violet
XOM Xenobiotic Organic Matter
XI
DECLARATION OF AUTHORSHIP
I, Anika Yunus declares that the thesis entitled ‘An investigation of recalcitrant organic
compounds in leachates’ and the work presented in the thesis are both my own, and have been
generated by me as the result of my own original research. I confirm that:
this work was done wholly or mainly while in candidature for a research degree at this
University;
where I have consulted the published work of others, this is always clearly attributed;
I have acknowledged all main sources of help;
parts of this work have been published as:
1. Yunus, A., Smallman, D. J., Stringfellow, A., Beaven, R., and Powrie, W. (2008)
Characterisation of landfill leachate DOM using fluorescence spectroscopy. Proceedings of
the Waste 2008 conference, September 16-17, Stratford-upon-Avon, UK
2. Yunus, A., Smallman, D. J., Stringfellow, A., Beaven, R., and Powrie, W. (2008)
Characterisation of the Recalcitrant Organic Matter Formed During the Anaerobic
Biodegradation of Domestic Waste. Proceedings of the Young Water Professionals
conference, April 22-24, University College London, UK
Signed: …………………………….
Date: ………………….
XII
Acknowledgements
I would like to convey my sincere gratitude to my supervisor Professor William Powrie
for his guidance and professional support throughout my graduate research career. His
critical comments and warm support have made my time very productive. I am greatly
indebted to Dr. David J. Smallman, Dr. Anne Stringfellow and Dr. Richard Beaven for
their useful advice, valuable suggestions and feedback. My special thanks are due to Lucy
Ivanova and Julie Vaclavic for their technical assistance in the experimental phases of the
work.
To my parents, I owe everything. A mere “thank you” to acknowledge their sacrifices can
hardly do justice to the immense gratitude I feel. I am also grateful to my inlaw parents for
their continued moral support. I hope that my accomplishments have made them proud. In
this opportunity I would like to convey my gratitude to my beloved husband for providing
me tremendous support, and at the same time keeping me focused and nudging me in the
right direction. His constant support, encouragement and critical comments have led me
through the completion of this work. I look forward to the life we will share together. Last,
but certainly not least, a big thanks goes to my little princess whose inspiration was the
source of strength of my work.
XIII
I dedicate
To my parents
And to my husband
Introduction
1
Chapter 1 Introduction
1.1 Scope of the thesis
Despite recent changes in waste management strategy promoting re-use and recycling, landfill
continues to play an important role in solid waste disposal in the UK. At present more than
62% of municipal solid wastes (MSW) in England are disposed of to landfills (defra, 2007).
One of the potential problems of landfill is the generation of leachates, primarily due to the
inherent moisture of the waste and infiltrating rainwater that passes through the waste layers
(Tchbanoglous et al., 1993). Leachate can contain dissolved and suspended organic
compounds, ammonia-nitrogen, heavy metals, chlorinated organics and inorganic salts (Wang
et al., 2002). Its composition and characteristics depend on the waste composition, climate,
hydrogeological factors and the age of the landfills (Kouzeli-Katsiri et al., 1999; Chae et al.,
2000; Ozkaya et al., 2006). Its characteristics also vary over time with regard to the
biodegradable compounds present (Chu et al., 1994; Ozkaya et al., 2006). Leachate is
therefore regarded as a continuum of organic compounds of different molecular weight (MW)
and structures including biodegradable low MW organic compounds like carbohydrates,
amino acids, organic acids as well as high MW humic and fulvic type molecules (He et al.,
2006; Pelaez et al., 2009). As waste degrades, the concentration of low MW biodegradable
organic compounds in leachates decreases (Calace and Petronio, 1997) and the proportion of
high MW organic compounds increases. These high MW organic compounds are generally
resistant to biodegradation in the landfill over time and often known as recalcitrant organic
compounds.
An influential component of many recalcitrant organic compounds are known as humic
substances (humic and fulvic type molecules) which are heterogeneous mixtures of different
organic materials such as aromatic, aliphatic and phenolic components (Artiola-Fortuny and
Fullar, 1982; Castagnoli et al., 1990 and Christensen et al., 1998). These substances can
significantly affect the accumulation and migration of some priority substances and hence,
play an important role in the natural environment. Weis et al. (1989) reported that humic
Introduction
2
substances can interfere with some other organic compounds and metals affecting the
transport of anthropogenic compounds to the groundwater and the bioavailability of
pollutants. Kang et al. (2002) suggested that humic substances can significantly affect the
behaviour of some pollutants, such as trace metal speciation and toxicity, solubilization and
adsorption of hydrophobic pollutants, disinfection by-product formation, aqueous
photochemistry, mineral growth and dissolution. On the other hand, humic substances can
reduce Hg (II) (mercury) to Hgo thereby providing a pathway for the mobilisation of Hg in the
environment (Stevenson, 1994). Humic-like compounds can also react with the chlorine to
produce carcinogenic chloroform and other undesirable halogenated compounds which might
create considerable problems as well (Stevenson, 1994).
Leachates are usually treated before being discharged to the receiving water. Various
physical/chemical and biological treatment processes have been used for the removal of
organic compounds from landfill leachates (Kettunen and Rintala, 1995; Imai et al., 1998;
Loukidou and Zouboulis, 2001; Uygur and Kargi, 2004; Kargi and Pamukoglu, 2004; Baig
and Liechti, 2001; Monje-Ramirez and Orta de Velásquez, 2004; Bila et al., 2005; Ntampou
et al., 2005). However, neither physical/chemical nor integrated physical-chemical-biological
treatment processes can completely remove the organic compounds contained in leachates
(Germili et al., 1991; Ince et al., 1998; Ozkaya et al., 2006; Bilgili et al., 2008). Although the
remaining amount of organic compound after treatment does not necessarily represent non-
biodegradable fraction of leachate, it has been generally referred as the ‘hard COD’ or
recalcitrant material in the literatures (Ince et al., 1998; Chae et al., 2000; Ozkaya et al., 2006;
Bilgili et al., 2008). In the UK, biological processes are generally used for economical
leachate treatment (Robinson and Knox, 2003) and hence, considerable amount of recalcitrant
organic compounds remains in leachates. Chae et al. (2000), Ozkaya et al. (2006) and Bilgili
et al. (2008) reported that in a biological leachate treatment system, the biodegradable fraction
of leachate can be removed effectively but the recalcitrant fraction passes through the
treatment system almost in an unchanged form. Lema et al. (1988) also reported that the
presence of humic substances make the leachate less amenable to biological treatment.
Therefore, it is essential to understand the evolution and the chemical and structural
composition of recalcitrant organic compounds and/or humic-like components over time.
Several studies have characterised leachates from different landfills of varying ages to
Introduction
3
understand the chemical composition of recalcitrant organics (Kang et al., 2002; Nanny and
Ratasuk, 2002; Fan et al., 2006). Similar studies have also been reported for different UK
landfill sites (Knox, 1983; Robinson and Grantham, 1988; Williams et al., 2002). However
these studies do not provide a chronological and systematic catalogue of the changes in
chemical characteristics of these compounds in course of biodegradation. In particular, in UK
a systematic comparison of the biodegradation associated evolution of the recalcitrant organic
compounds in different landfill sites has not been done. An assessment of the influence of
individual waste types on the recalcitrant organic compounds in leachates is also rare in the
literature. As far as groundwater and surface water pollution is concerned, a chronological
monitoring and characterisation of recalcitrant organic compounds over time is important.
Organic compounds in leachates are usually characterised by several established methods;
biochemical oxygen demand (BOD), chemical oxygen demand (COD), and total organic
carbon (TOC). These methods estimate the organic compound of leachates by oxidising the
constituent organic matter in different ways. In BOD method, biochemical oxidation is used
to estimate the amount of organic compounds, whereas in COD method strong oxidising
agents are used under acidic conditions (APHA, 1998). TOC method uses enhanced oxidizing
ambient such as heat, ultra-violet light and/or strong chemical oxidants. Although these
methods simply aim to estimate organic compounds in leachates, due to the diverse
characteristics of leachate and the different oxidizing conditions in these techniques BOD,
COD and TOC methods do not necessarily give the same estimation. The efficiency of any
leachate treatment is generally evaluated by the percentage reduction of BOD, COD and TOC
over time. However, these techniques are time consuming and laborious and also require
considerable sample preparation (Ellis, 1989). For an investigation on the suitability of
leachates for discharge, compositional analysis has also been carried out in the past through
the prior separation of sub-compounds using degradation methods and later measurements by
spectroscopic methods such as Fourier Transformed Infra-Red (FTIR) and Nuclear Magnetic
Resonance (NMR) (Hayes et al., 1989; Kang et al., 2002; Nanny and Ratasuk, 2002; Fan et
al., 2006). The degradation methods have disadvantages of lengthy sample preparation,
transformations and organic carbon lost. As a result, these techniques of compositional
analysis also cannot be accepted as a rapid technique for onsite characterisation of recalcitrant
materials in leachates. Therefore, there is a need for simple and fast techniques of leachate
Introduction
4
characterisation to acknowledge the role played by recalcitrant organic compounds of
leachates.
In this respect, UV absorption and fluorescence spectroscopy could provide a rapid method
for monitoring and characterizing recalcitrant organic compounds and also for fingerprinting
the other organic pollutants in leachates to optimise leachate treatment processes. The major
advantages of these techniques lie in their rapidity and versatility, their relatively low running
costs and the absence of chemicals and detailed sample preparation. The fluorescence
technique is potentially applicable for monitoring a wide range of constituents such as humic
substance, polyaromatic compounds, xenobiotic compounds etc. (Senesi et al., 1991). The UV
absorption and fluorescence techniques have been widely used in various ecosystems (e.g.
marine, Coble, 1996; groundwater, Baker and Lamont-Black, 2001; sewage-impacted rivers,
Baker, 2001; wastewater and effluents, Westerhoff et al., 2001 and Her et al., 2003). These
studies suggest that UV absorption and fluorescence techniques have a significant potential in
landfill leachate research by fingerprinting various organic compounds as well as analysing
compositional characteristic of the recalcitrant organic compounds in leachates. However,
very little work on the characterisation of leachates using these techniques has been carried
out (Baker and Curry, 2004; Huo et al., 2008; Ham et al., 2008).
In this work, an analysis of the characteristics of recalcitrant organic compounds in leachates
and a feasibility study on the application of UV absorption and fluorescence spectroscopy for
compositional analysis of leachates are performed. The aims of the research are to investigate
the following.
1. The nature of the recalcitrant organic compounds contained in leachates, particularly
its removal in the course of an aerobic biological treatment of leachates collected from
two UK MSW landfills (Pitsea and Rainham) and its evolution in the leachates during
anaerobic biodegradation of solid waste components to understand the influence of
different types of waste on the recalcitrant organic compounds.
Introduction
5
2. The use of UV absorption and fluorescence spectroscopy as a rapid and economical
technique for compositional analysis of the recalcitrant organic compounds with the
aim of assisting and enhancing routinely used analytical methods.
As such, this research performs a systematic comparison of the biodegradation associated
evolution of the recalcitrant organic compounds of leachates from two famous UK landfills
(Pitsea and Rainham) and from leachates generated during biodegradation of different waste
components. These investigations are carried out for the first time using a combination of
conventional methods and potentially new techniques of UV absorption and fluorescence
spectroscopy. This enriches fundamental knowledge of the influence of solid waste
components on the recalcitrant organic compounds and provides deep insight into the
characteristics of these compounds on the investigated UK landfill sites/phases that could be
used to foresee the necessary treatment in different landfill sites. The combined analysis using
conventional methods and UV/fluorescence spectroscopy performs a feasibility study on the
use of UV absorption and fluorescence spectroscopy for a rapid and onsite characterisation of
recalcitrant organic compounds which might find application if regulatory constraints could
be fulfilled. .
1.2 Organization of the thesis
This thesis is divided into six chapters. In this Introduction (Chapter 1) the main objectives of
this research are outlined (section 1.1).
Chapter 2 makes a detailed review of the literature relating to the origin and removal of
recalcitrant organic compounds by various leachate treatment processes. This chapter reviews
the applications and limitations of conventional methods for characterisation of recalcitrant
organic compounds of leachates. Potentially useful techniques of UV absorption and
fluorescence spectroscopy are also presented in this chapter with a focus on effective and
rapid monitoring and characterisation.
Introduction
6
Chapter 3 provides the analytical methods and experimental procedures used to investigate
the nature and composition of the recalcitrant organic compounds in landfill leachates and
also in leachates generated during anaerobic biodegradation of components of solid waste.
Chapter 4 summarizes the experimental results of the characteristics of the recalcitrant
organic compounds in leachates in the course of an aerobic biological treatment process.
Changes in the composition of the recalcitrant organic compounds of leachates during
laboratory scale aerobic biological treatment system was characterised by the conventional
COD and DOC measurements and also by UV absorption and fluorescence spectroscopy.
Chapter 5 summarizes and discusses the experimental results of the characteristics of
leachates formed during an anaerobic biodegradation of waste components. Conventional
COD and DOC measurements and UV absorption and fluorescence spectroscopic results are
reported to evaluate the development of recalcitrant organic compounds in the associated
leachates as the waste degrades.
Chapter 6 proposes the major conclusion drawn from this research, followed by
recommendations for future work.
Background
7
Chapter 2 Background
2.1 Introduction
This chapter presents background studies on the recalcitrant organic compounds in landfill
leachates. The origin, composition and the removal of recalcitrant organics by various
leachate treatment processes are reviewed. The applications and limitations of available
methods to estimate the total amount of recalcitrant organic compounds and to fingerprint
the constituent compounds of recalcitrant organics are discussed. The possible application
of UV absorption and fluorescence spectroscopy in characterising the recalcitrant organics
is assessed by studying published reports of UV and fluorescence applications to marine
and estuarine waters, freshwater, and wastewaters.
2.2 Leachate generation in landfill
Leachates are generated from municipal solid wastes (MSW), which are a heterogeneous
mixture of residential, commercial and some industrial wastes. When disposed of to
landfills, the organic components of the MSW are degraded by a combination of physical,
chemical and microbial processes (Kouzeli-Katsiri et al., 1999; Tchbanoglous et al.,
1993). This leads to the generation of leachates, primarily due to the inherent moisture of
the waste, the degradation process and infiltrating rainwater that pass through the waste
layers (Tchbanoglous et al., 1993). Leachate is a highly contaminated and heterogeneous
wastewater and contains dissolved and suspended organic compounds, ammoniacal-
nitrogen, heavy metals, chlorinated organics and inorganic salts (Wang et al., 2002). The
composition and characteristics of landfill leachates depend on the nature of the disposed
waste, hydrogeological factors, climate and the age of the landfills (Kouzeli-Katsiri et al.,
1999; Chae et al., 2000; Ozkaya et al., 2006). With time a landfill passes through a short
aerobic phase followed by a long anaerobic decomposition phase. Several studies have
described waste biodegradation in landfill as a five-phase process. These are the aerobic
phase, acidogenic phase, acetogenic phase, methanogenic phase and final aerobic phase
(figure 2.1) (Lu et al., 1985; Pohland and Harper, 1985; Tchbanoglous et al., 1993;
McBean et al., 1995). The breakdown processes in the various phases are potentially
Background
8
complex at the microscopic level, involving hundreds of intermediate biochemical
reactions and compounds. Many of these reactions require additional specific synergistic
chemicals, catalysts or enzymes. The biodegradation processes and products relating to
organic matter at the various phases are described below. The nature of these phases plays
a significant role in the composition of leachates.
Lu et al. (1985) and McBean et al. (1995) reported that in the aerobic degradation phase,
aerobic micro-organisms metabolise oxygen in the air trapped within the waste and a
proportion of the organic fraction of the waste to produce carbon dioxide (CO2), water
(H2O) and partly degraded residual organics. Leachate produced during this phase is likely
a result of the moisture squeezed out of the waste during compaction and cell construction
(Lu et al., 1985). It is characterized by the constituent suspended particulate matter, the
dissolution of highly soluble salts initially present in the landfill, and the presence of small
amounts of organic species from aerobic biodegradation (Lu et al., 1985; McBean et al.,
1995). The released energy is in the form of heat and can raise the temperature to 70-90oC
(Lu et al., 1985; McBean et al., 1995). This stage can last for several days or weeks
depending on the availability of oxygen.
In the second stage of the biodegradation process, anaerobic and facultative micro-
organisms hydrolyse and ferment cellulose and/or other materials (proteins, fats etc.)
making them more readily available for the acidogenic bacteria. Acidogenesis is
characterised by the production of acetic acid from the monomers produced in the
preceding stage and other volatile fatty acids which are derived from the biodegradation of
protein, fat and carbohydrate components of the waste (Lu et al., 1985; McBean et al.,
1995). The temperature within the landfill falls to 30-50oC (WMP 26B, 1995). Carbon
dioxide and hydrogen may rise up to 80% and 20% of total volume of gas respectively
(WMP 26B, 1995). The production of different by-products depends on composition of
waste, the environmental condition and the presence of different bacterial species.
Background
9
Figu
re 2
.1 C
hang
es in
sele
cted
par
amet
ers d
urin
g th
e ph
ases
of l
andf
ill st
abili
zatio
n (P
ohla
nd a
nd
Har
per,
198
5)
• Ph
ase
V is
not
show
n •
CO
D –
Che
mic
al O
xyge
n D
eman
d •
TV
A –
Tot
al V
olat
ile A
cids
Background
10
In the acetogenesis phase, acetogenic bacteria convert the long chain fatty acids
(propionic, butyric, valeric and caproic acids) to acetic acid, carbon dioxide and hydrogen.
Lo (1996) reported that the leachate produced in the acidogenic and acetogenic phases is
characterised by an acidic pH (typically 5.0 or 6.0), high concentration of biodegradable
organic compounds and an unpleasant smell. Also because of the low pH values in the
leachate, a number of inorganic compounds, principally heavy metals could be solubilised
during this phase resulting in a potentially chemically toxic leachate. Hydrogen sulphide
may also be produced in this stage as the sulphate compounds in the waste are reduced to
hydrogen sulphide by sulphate reducing micro-organisms (Christensen et al., 1996).
In the following methanogenic phase, the simple organic compounds released by the
acetogenic processes start to be consumed by the methanogenic bacteria to produce
methane and carbon dioxide. The leachate produced in this phase is characterised by a
range of pH of 6.8 to 8.0 (Zehnder, 1978), low concentrations of biodegradable organic
compounds and high levels of ammonia nitrogen (NH3-N) (Lo, 1996). Acetogenic and
methanogenic bacteria form a symbiotic (mutually beneficial) relationship. Acetogenic
bacteria depend on the methanogenic bacteria to remove hydrogen, the accumulation of
which leads to the suppression of acetogenesis whereas methanogenic bacteria depend on
acetogenic bacteria to provide the acetic acid and hydrogen required for methane
production. As methanogenic bacteria are slow growing bacteria, they are unlikely to be
present in the early stage of waste biodegradation. Their initial absence would lead to the
increased partial pressure of hydrogen, which would in turn result in the inhibition of
acetogenesis and accumulation of long chain fatty acids thus lowering the pH, leading to
the further inhibition of methanogenesis.
An additional aerobic phase of biodegradation has been proposed by Christensen and
Kjeldsen (1995). They stated that the final aerobic stage of waste degradation is initiated
by the aerobic micro-organisms which replace the anaerobic forms, and becomes re-
established when the rate of oxygen diffusion through the capping from the atmosphere
becomes greater than the methane and carbon dioxide produced by anaerobic degradation.
Background
11
2.3 Landfill leachate characterisation and the origin of recalcitrant
organic compounds
Landfill leachates are usually characterised by several methods including biochemical
oxygen demand (BOD), chemical oxygen demand (COD) and total organic carbon (TOC).
These methods simply aim at estimating organic compounds in leachates by oxidising the
constituent organic matter and hence, the efficiency of any leachate treatment is evaluated
by the percentage reduction of the BOD, COD and TOC over time. The amount of organic
compounds which are eventually not removed in the course of a treatment processes is
considered as recalcitrant organic compounds and usually estimated by the remaining
percentage of BOD, COD and TOC. The detail processes of BOD, COD and TOC
techniques are discussed below:
2.3.1 Biochemical Oxygen Demand (BOD)
The BOD is essentially a measure of the amount of oxygen required for the biochemical
oxidation of organic compounds usually over 5 to 20 days (Brookman, 1997). In the BOD
test, micro-organisms are used to oxidise and convert the carbonaceous organic matter to
carbon dioxide and water. The presence of toxic substances in leachates can effect (or
even kill) the micro-organisms, even at low concentrations. In addition, the heavy metals
such as lead, copper, mercury or cadmium can inhibit or sometimes completely prevent
the oxidation of the organics present in the leachates (Bourgeois et al., 2001). However,
the BOD test is slow to yield information and takes a relatively long time to complete.
2.3.2 Chemical Oxygen Demand (COD)
The COD is a measure of the oxygen equivalent of the organic matter that is susceptible to
oxidation by a strong chemical oxidant (Bourgeois et al., 2001). The COD test uses
potassium dichromate (K2Cr2O7), acid and heat to oxidise organic carbon to carbon
dioxide and water. The amount of oxidizable organic matter, measured as oxygen
equivalent, is proportional to the potassium dichromate consumed. The major advantage of
the COD test is that the results can be obtained within a relatively short time
(approximately 2 hours). Additionally the presence of toxic substances does not effect
COD measurement. One of the main limitations of the COD test is the inability to
Background
12
differentiate between the biodegradable and non-biodegradable organic compounds. In
addition, the presence of chloride, bromide etc. in the leachates interfere in the COD test,
however added silver nitrates reacts with chlorides, bromides or iodides to produce
precipitates.
2.3.3 Total Organic Carbon (TOC)
The TOC test uses heat, ultra-violet light or a strong chemical oxidant (or a combination
of all three) to oxidise the organic matter. The advantages of TOC method over BOD and
COD are as follows:
1. Quick and easy
2. Chloride and bromide do not interfere
3. Not harmful
4. No requirement of hazardous chemicals (chromium, silver nitrate)
5. No micro-organisms involved
A relative comparison of these three methods is presented in Table 2.1
Table 2.1 A relative comparison among three methods (BOD, COD, TOC)
(BOD) (COD) (TOC)
Analysis
technique
Determination of O2 consumption due to
biochemical oxidation of
sample
Determination of O2 consumed
during chromic acid digestion of
sample
High Temperature Combustion (HTC)
detection of CO2 analysis
with the removal of inorganic carbon
Reagents required
Bacterial seed Chromic acid, silver sulphate,
sulfuric acid, ferrous ammonium sulphate
Phosphoric acid
Time per
analysis 5 to 20 days 2-3 hours 4 – 6 minutes
Accuracy 5-8% 5-8% 2-3%
Interferences Inorganic carbon Nitrite, Sulfide, Chlorides,
Fe, Acids, others Toxic substances
Background
13
The chemical composition of leachates characterised by BOD, COD and TOC has been
studied extensively by various researchers in the past few decades (Qasim and Burchinal,
1970; Chian and DeWalle, 1976; Pohland and Englebrecht, 1976; Lu et al., 1985; Lema et
al., 1988; Chu et al., 1994; Ragle et al., 1995; Chen et al., 1996; Lo, 1996; Christensen et
al., 2001). It is therefore well known that there is a decreasing trend of BOD, COD, TOC
and the ratio of BOD/COD with increasing age of the landfill. Consequently old landfill
leachates have low biodegradability and are correlated with low BOD and COD values
and low BOD/COD ratios. Calace et al. (2001) also reported that the young landfill
leachates have a high fraction (70%) of low molecular weight (< 500 Dalton) and a low
fraction (18%) of high molecular weight (>10000 Dalton) compounds. Conversely old
landfill leachates have low and high molecular weight fractions of about 28% and 67%
respectively. This means that as landfill degradation progresses, the concentration of low
molecular weight biodegradable organic compounds in leachates decreases due to the
fermentation of hydrolysable organic compounds (Calace and Petronio, 1997) and high
molecular weight organic compounds are increased. These high molecular weight organic
compounds are microbially refractory under the preceding condition in the landfill over
time (Chian and DeWalle, 1976; Göbbels and Püttmann, 1996; Calace and Petronio, 1997)
and are often known as recalcitrant organic compounds (Germili et al., 1991; Ince et al.,
1998; Ozkaya et al., 2006).
The generation of recalcitrant organic compounds depends significantly on the
components of the solid waste disposed of to the landfills such as cellulose, hemi-
cellulose, lignin, protein and lipids. Stevenson (1994) suggested that lignin is the main
source of recalcitrant compounds. Tong et al. (1990) reported that lignin comprises a
complex polymer of phenylpropane units, which are cross-linked to each other with a
variety of different chemical bonds and are especially resistant to anaerobic
biodegradation. However, cellulose and hemi-cellulose, which are usually biodegradable
in aerobic and anaerobic environments (Tong et al., 1990), are also accepted to be
recalcitrant compound precursors (Varadachari and Ghosh, 1984; Inber et al., 1989). This
was explained by Komilis and Ham (2003) who reported that the retardation of aerobic
and anaerobic biodegradation of cellulose and hemi-cellulose is primarily due to the
sheathing of cellulose and hemi-cellulose by lignin. This agrees with Khan (1977) who
performed a laboratory experiment on the degradability of cellulose under anaerobic
conditions. He showed that about 40%, 90% and 97% of cellulose in newspaper, office
Background
14
paper and Whatman filter paper were degraded. This was due to the fact that newspaper
was highly lignified whereas office paper and Whatman filter paper were nearly
delignified. Tuomela et al. (2000) reported that lignin degradation is primarily an aerobic
process and white-rot fungi are responsible for most of the lignin decomposition. In an
anaerobic environment such as in landfills, lignin can persist for years and its presence can
significantly reduce the bioavailability of cellulose and hemi-cellulose (Rees, 1980; Wang
et al., 1994; Komilis and Ham, 2003).
Microbial re-synthesis is also an important source of non-biodegradable or recalcitrant
compounds. Pichler and Kogel-Knabner (2000) found that both protein and lipids are re-
synthesized microbially in the MSW environment, resulting in long chain high molecular
weight compounds which are poorly degradable. Dinel et al. (1996) also reported that
proteins may be preserved by encapsulation into recalcitrant cell wall polymers of micro-
organisms and become resistant to biodegradation.
2.4 Composition of recalcitrant organic compounds and its effect on the
environment
Recalcitrant materials in landfill leachates consist mostly of humic substances (Artiola-
Fortuny and Fullar, 1982; Castagnoli et al., 1990 and Christensen et al., 1998). On the
basis of the solubility, Schnitzer and Khan (1972) subdivided humic substances into humic
acids (HA, soluble in basic medium but not soluble in acid medium), fulvic acids (FA,
soluble both in acid and basic medium) and humin (insoluble in water under any pH
condition). Hayes et al. (1989) and Stevenson (1982, 1994) reported that aromatic
structures are the dominant building blocks of the structural core of humic substances and
the compositions of humic substances are dominated by structural moieties (benzene rings,
aliphatic segments, hexose, pentose, and amino acids), functional groups (carboxyl,
carbonyl, hydroxyl, amine, phenolic) and linkages (ester, amide and ether). While these
studies provided some information on the presence of several structural moieties,
functional groups, linkages as well as identifying other aromatic and aliphatic components
among the degradation products, there was no report in the literature revealing any unique
molecular structure that could be considered as an essential building block of humic
substances. It appears that molecules in humic substances differ from one another in
chemical and physical properties. Consequently, humic substances are complex mixtures
Background
15
of macromolecules of varying molecular size, structure and chemical functionality.
Because of this versatility, these substances can significantly affect the accumulation and
migration of some priority substances and hence, play an important role in the natural
environment (Weis et al., 1989; Kang et al., 2002)
Stevenson (1994) stated that these compounds are not physiologically harmful but are
aesthetically unacceptable because they impart a reddish-black colour to potable water and
lakes. Their presence can have both negative and positive impacts. Weis et al., (1989)
reported that humic substances can interfere with some other organic compounds and
metals affecting the transport of anthropogenic compounds to the groundwater and the
bioavailability of pollutants. Kang et al. (2002) stated that humic substances significantly
affect the behaviour of some pollutants, such as trace metal speciation and toxicity,
solubilization and adsorption of hydrophobic pollutants, disinfection by-product
formation, aqueous photochemistry, mineral growth and dissolution. These compounds
can also affect photochemical processes; reduce Hg (II) (mercury) to Hgo thereby
providing a pathway for the mobilisation of Hg in the environment (Stevenson, 1994).
Humic-like compounds can also react with the chlorine to produce carcinogenic
chloroform and other undesirable halogenated compounds which might create
considerable problems as well (Stevenson, 1994). It is therefore essential to understand the
evolution and the chemical and structural composition of recalcitrant organic compounds
and/or humic-like components over time.
2.5 Effect of treatment on removal of recalcitrant organic compounds
Leachates are usually treated before being discharged to the receiving environments.
Various physical, chemical and biological treatment processes have been used extensively
for the removal of organic compounds from landfill leachates. These methods have been
reviewed in detail by Renou et al., (2008).
Table 2.2 summarizes some of the aerobic and anaerobic biological processes and their
efficiency in removing organic compound from landfill leachates. The efficiency of a
treatment process is evaluated by the percentage removal of the BOD, COD and TOC. It
can be seen from Table 2.2 that with the exception of Kettunen and Rintala (1995), a
treatment efficiency of about 80-90% has been achieved by aerobic biological treatment
Background
16
processes. The reported anaerobic treatment efficiency is about 65-75%. Kargi and
Pamukoglu (2004) also reported that it is advantageous to carry out anaerobic treatment
prior to an aerobic process for leachates containing a high COD. Using this approach
Kettunen et al. (1996) achieved a COD removal of 80-90%. It is worth mentioning that the
treatment efficiencies presented in Table 2.2 cannot be directly compared with each other
due to the initial differences in COD values. However, it is generally found that about 10-
20% recalcitrant organic compounds measured by BOD, COD and TOC remain after
biological treatment.
Table 2.2 Studies on landfill leachate biological treatment (aerobic and anaerobic)
Reference Initial leachate
COD (mg/l) Treatment
Overall results
(COD, BOD,
TOC removal)
Gourdon et al.
(1989) 850-1350 Aerobic (Trickling filter) 87% (BOD)
Maehlum (1995) 1182 Aerobic (lagooning) 89% (COD)
Kettunen and
Rintala (1995) 2000-3000
Aerobic (Moving Bed
Biofilm Reactor) 46-64% (COD)
Loukidou and
Zouboulis (2001) 5000
Aerobic (Moving Bed
Biofilm Reactor) 81% (COD)
Zoubouli et al.
(2001) 15000
Anaerobic (Sequencing batch
reactor) 75% (COD)
Kettunen et al.
(1996) 1800
Sequential anaerobic-aerobic
biological reactor (SBR) 80-90% (COD)
Kettunen and
Rintala (1998) 1500-3200
Anaerobic (Up-flow
anaerobic sludge blanket
reactor)
65-75% (COD)
Imai et al. (1998) reported that a single, conventional biological treatment is not effective
in treating leachates with a high concentration of organic matter resistant to
biodegradation and this type of leachate needs additional treatment to become more
biodegradable. According to Imai et al. (1998), one way of doing this is to employ
physical-chemical processes as a pre-treatment prior to a biological treatment or as a final
Background
17
polishing treatment. Table 2.3 summarizes the efficiencies of some of these treatments. It
can be seen that using granular or powder activated carbon (PAC) as an adsorbent a
treatment efficiency of around 90% has been achieved (Morawe et al., 1995; Welander
and Henrysson, 1998; Rivas et al., 2003; Kargi and Pamukoglu, 2004). This indicates that
adsorption was effective as a pre-treatment method for the COD removal.
Table 2.3 Studies on landfill leachate integrated physical-chemical-biological
treatment
Reference Initial leachate
COD (mg/l) Treatment
Overall results
(COD, BOD,
TOC removal)
Morawe et al. (1995) 879-940 Granular activated
carbon as adsorbent 91% (COD)
Welander and
Henrysson (1998) 800-2000
Powdered activated
carbon as adsorbent 96% (TOC)
Kargi and
Pamukoglu (2004) 7000
Powdered activated
carbon as adsorbent 90% (COD)
Rivas et al. (2003) 7400-8800 Activated carbon as
adsorbent 90% (COD)
Bila et al. (2005) 3096 FeCl3 or Al2(SO4)3 as
coagulant 73% (COD)
Ntampou et al.
(2005) 1000
FeCl3 or Al2Cl3 as
coagulant 89% (COD)
Monje-Ramirez and
Orta de Velásquez
(2004)
5000 Fe2(SO4)3 or Al2O3 as
coagulant 78% (COD)
Amokrane et al. (1997) also suggested coagulation as a pre-treatment prior to a biological
process. Aluminum sulfate, ferrous sulfate, ferric chloride and ferric chloro-sulfate were
commonly used as coagulants (Amokrane et al., 1997; Bila et al., 2005). The removal of
organics by coagulation is associated with the removal of humic substances (Ntampou et
al., 2005). There are two mechanisms of removal of humic substances by coagulation; (1)
binding of cationic metal species to the anionic sites resulting in the neutralisation of
Background
18
humic substances and reduction of their solubility and (2) adsorption of humic substances
onto the produces amorphous metal hydroxide precipitates (Ntampou et al., 2005). A
treatment efficiency of 70-80% has been achieved by coagulation (Monje-Ramirez and
Orta de Velásquez, 2004; Bila et al., 2005; Ntampou et al., 2005).
Several researchers have investigated the efficiency of ozonation (O3) for the treatment of
landfill leachates (Huang et al., 1993; Monje-Ramirez and Orta de Velásquez, 2004).
Monje-Ramirez and Orta de Velásquez (2004) reported that chemical oxidation with O3
makes possible the transformation of recalcitrant material into biodegradable forms or
CO2. Table 2.4 shows the achieved efficiencies in various leachate treatments with O3. In
sanitary landfill leachates, oxidation with O3 is not very effective as a single process and
the efficiency is only 30-40%. However, in combination with other methods (biological,
coagulation), O3 enhances the overall efficiency of COD removal, from 44% to 97%.
The above studies suggest that physical-chemical methods as a pre-treatment or final
polishing are sometimes more effective than single biological methods in removing
recalcitrant organic compounds from leachates. Physical-chemical methods are suitable
for small scale laboratory based experiments but are costly in terms of equipment and
infrastructure investments, energy requirement, and large scale chemical usage for landfill
applications and hence they are not economically viable. As a result, biological processes
are the most commonly used methods of leachate treatment in the UK (Robinson and
Knox, 2003), hence considerable amount of recalcitrant organic compounds can remain in
landfill leachates. Therefore, there is a need to understand the characteristics of
recalcitrant organic compounds in leachates. In particular it is essential to understand the
recalcitrant organic compounds in leachates over time and with the nature of waste
components. There appears to have been little if any formal study of this, although
information on the composition of leachates collected from landfills of different ages is
given by Kang et al. (2002); Nanny and Ratasuk (2002) and Fan et al. (2006). It was
mentioned before that leachate composition varies significantly with climate,
hydrogeological factors and the nature of the waste disposed in a landfill. Thus these
studies of leachates from different landfills of varying ages do not form a chronological
and systematic catalogue of the composition of recalcitrant organic compounds with time.
As far as groundwater and surface water pollution is concerned, the systematic
characterisation and monitoring of landfill leachates is imperative.
Background
19
Table 2.4 Studies on landfill leachate treatment with O3
Reference Leachate type and
COD (mg/l) Treatment
Overall results
(COD
removal)
Huang et al. (1993) Old leachate (1610) O3 44%
Imai et al. (1998) Old leachate (126) O3 33%
Baig and Liechti
(2001)
Young leachate
(1585)
Coagulation (FeCl3) +
Aerobic treatment + O3 94%
Monje-Ramirez
and Orte de
Velasquez (2004)
Coagulation treated
leachate (1058) Coagulation + O3 78%
Schulte et al.
(1995)
Biologically treated
leachate (760) O3 97%
Bila et al. (2005) 3096
FeCl3 or Al2(SO4)3 as
coagulant + O3 +
Biological treatment
73%
Ntampou et al.
(2005) 1000
FeCl3 or Al2Cl3 as
coagulant + O3 89%
2.6 Limitation of conventional characterisation methods (BOD, COD,
TOC)
As discussed before, in the leachate treatment processes the amount of organic compounds
that can not be removed in the course of a treatment is described as the recalcitrant organic
compounds and can be detected by the remaining percentage of BOD, COD and TOC. The
disadvantages of these standard characterisation techniques (BOD, COD and TOC) of
leachates include lack of reliability and the difficulty in achieving reproducible results.
Though COD and TOC test results can be obtained within a relatively short time (2 to 3
hrs), BOD test takes a long time (5 to 20 days). Additionally, all these techniques are
laborious and also require sample preparation (Ellis, 1989). Therefore, these conventional
techniques do not provide a method for rapid and effective monitoring of recalcitrant
Background
20
organic compounds in leachates over a time frame that could influence the leachate
treatment system. Therefore, it would be economical if rapid characterisation techniques
of recalcitrant organic compounds could be developed. In addition, these standard
techniques can only be used to estimate the amount of remaining recalcitrant organic
compounds in leachates in the course of a treatment (Tables 2.2, 2.3 and 2.4) but they
cannot be used to identify the degradation nature of some constituents of recalcitrant
organic compounds.
2.7 Compositional analysis of recalcitrant organic compounds and/or
humic substances
Several techniques which were applied in the past for characterizing different components
of humic substances (Schnitzer and Khan, 1978; Stevenson, 1982, 1994; Hayes et al.,
1989a; Weis et al., 1989; Calace and Petronio, 1997; Christensen et al., 1998; Kang et al.,
2002; Nanny and Ratasuk, 2002; Fan et al., 2006) are summarized in Table 2.5.
Elemental composition of humic substances was usually investigated in the past using
chemical/physical or XPS method. Both of these methods have the disadvantage of
considerable sample preparations. XPS has been well established as a powerful tool for the
characterization of surfaces and surface reactions in heterogeneous catalysis, in corrosion
science (Briggs and Seah, 1990), wood science (Johansson et al., 1999) and soil sciences
(Yuan et al., 1998). However, for characterizing the aquatic humic substances in leachates
by XPS method, a number of experimental problems had to be solved. In particular,
reproducible preparations of thin solid layers of dissolved humic substances on an
appropriate substrate surface and the homogeneity of such layers are key challenge of XPS
for this application.
Size and shape of humic particles is usually characterized by Electron Microscopy
although it has been reported that estimation using this technique needs extra care due to a
strong pH dependency (Chen and Schnitzer, (1989). Gel chromatography has been used
for measuring the molecular weight of humic substances (Schnitzer and Khan, 1978).
However, this method has been reported to overestimate the molecular weight due to gel-
solute interactions (Schnitzer and Khan, 1978).
Background
21
Table 2.5 Common methods for the characterization of humic substances
Method Information Reference
Elemental
analysis
Elemental
composition (C, H,
N, S, O)
Christensen et al. (1998); Kang
et al. (2002); Nanny and
Ratasuk (2002); Fan et al.
(2006)
Microscopy Size and shape Chen and Schnitzer (1989)
Acid/Base
titration Proton capacity
Perdue (1998); Abbt-Braun and
Frimmel (2002)
Thermal
degradation Pyrolysis products
Abbt-Braun et al. (1989);
Schulten et al. (2002)
Chemical/
Physical
Oxidative,
reductive
Degradation
products Christman et al. (1989)
X-ray
photoelectron
spectroscopy
(XPS)
Elemental
composition (C, H,
N, S, O)
Bubert et al. (2000)
Fourier
Transformed
Infra-red
(FTIR)
Qualitative
determination of
functional groups
Hayes et al. (1989)
Spectroscopy
Nuclear
Magnetic
Resonance
(NMR)
Quantitative
determination of
functional groups:
aromatics,
aliphatics,
carbohydrates
Kang et al. (2002); Nanny and
Ratasuk (2002); Fan et al.
(2006)
Chromatography Gel
chromatography
Molecular size,
weight
Perminova et al. (1998);
Schmitt-Kopplin et al. (1998)
Background
22
Acid/Base titration methods have also been employed for the determination of acidic
functional groups in the humic substances (Perdue, 1998; Abbt-Braun and Frimmel,
2002). However, the dissociation of two major types of functional groups; carboxyl groups
(-COOH) and hydroxyl group (-OH) were reported to be difficult by this method.
In addition to the characterizing elemental composition and size/shape of humic particles
several methods are also available in the literature for characterizing constituent organic
compounds and/or functional groups in humic substances (Table 2.5). Since the early to
mid-1980s, degradation methods have been applied to produce identifiable sub-
compounds of humic substance for estimation. Thermal pyrolysis, oxidative-reductive and
hydrolytic methods are often used as degradation reactions for sample preparation.
Thurman and Malcolm (1981) also used a method for extraction and isolation of humic
substances from leachate dissolved organic carbon (DOC) for chemical characterisation
which was a variant of degradation method. This method involves the use of XAD resin
for the extraction of humic substances from leachate DOC and the acidification of humic
substances to pH 1.0 for the isolation of humic and fulvic acids. However, there is always
the question of unknown denaturing reactions and of transformations during the
degradation procedure. This sometimes leads to different degradation products formed
from one single precursor due to the involvement of several treatments with acids and
bases to extract humic substances. Smith et al. (1991) reported that during these
procedures a substantial part of DOC (about 30%) may also be lost.
In recent years, FTIR and NMR studies have become the most important spectroscopic
methods of structural characterization of humic substances for identifying various
functional groups such as quantification of partial structures such as aliphatic, O-alkyl,
aromatic and carboxyl carbons. FTIR characterisation is mainly qualitative and NMR
method although provides quantitative information of some functional groups, the
heterogeneous nature of humic substances often results in many overlapping signals that
complicate the interpretation. As a result, to get an unambiguous estimation, these
methods were usually applied after prior separation of identifiable sub-compounds. For
example, Christensen et al. (1998) fractionated hydrophilic fraction of leachate DOC using
the method developed by Thurman and Malcolm (1981) and characterised the hydrophilic
fraction by size-exclusion chromatography, elemental analysis, acid-base titration and
FTIR. It was concluded that although hydrophilic fraction resembles, in many ways to the
Background
23
humic and fulvic acids a substantial portion was also found to be neither humic nor fulvic
acid indicating the potential difficulty of degradation methods.
The qualitative and quantitative characterisation of non-humic fractions such as
carbohydrates, amino acids in humic substances has also been studied in the past. Frimmel
et al. (2002) and Jahnel et al. (2002) fractionated humic substances by hydrolysis with
acids, bases and enzymes and analysed by 13CNMR. They showed that monosaccharides,
like pentoses, hexoses, deoxycarbohydrates and aminosugars make up to 8 mass percent of
the organic carbon of fulvic and humic acids. 13CNMR analyses showed that up to 30%
carbon in humic substances is involved in carbohydrate structures whereas 15NNMR data
showed that up to 90% of nitrogen is attributed to amide nitrogen (Jahnel et al., 2002).
In addition to above mentioned techniques there are two other techniques of compositional
analysis of humic substances which is discussed in the next section.
2.8 Proposed technique of leachate characterisation
The discussions in section 2.6 imply that the conventional methods of leachate
characterisation (BOD, COD and TOC) require considerable sample preparation and
cannot perform the complete compositional analysis of leachates. Among the different
methods of compositional analysis discussed in section 2.7, degradation methods such as
thermal pyrolysis, oxidative-reductive, hydrolysis and/or spectroscopic methods such as
FTIR and NMR could be used to investigate the evolution of some constituent organic
compounds of recalcitrant materials in course of a biodegradation. However, the
degradation process itself can be treated as lengthy sample preparation and obviously has
the disadvantages such as unknown denaturing reactions, transformations and substantial
DOC lost. Although spectroscopic methods FTIR and NMR are interesting, an
unambiguous estimation of constituent organics often requires prior separation of
identifiable sub-compounds using degradation. As a result, these techniques cannot be
accepted as a rapid technique for onsite compositional analysis of recalcitrant materials in
leachates. In this respect, UV and fluorescence spectroscopy might be attractive for
analyses of leachates which do not require any sample preparation. In this section, some
applications of UV and fluorescence spectroscopy in characterising marine and estuarine
waters, freshwater, wastewater etc is presented. These studies show that these techniques
Background
24
have potential for additional landfill leachate characterisation. The key advantage is that
these techniques have the flexibility of rapid estimation of recalcitrant organic compounds
in leachates and have the ability of doing compositional analysis of recalcitrant organics.
2.8.1 UV visible spectroscopy
Ultraviolet-visible spectroscopy (UV-Vis) involves the spectroscopy of photons in the
UV-visible region (200-800 nm wavelengths). When sample molecules are exposed to
light, some of the light energy is absorbed and the electrons are promoted to a higher
energy orbital. An optical spectrometer measures the intensity of light passing through a
sample and compares it with the intensity of light before it passes through the sample. The
amount of light absorbed by the sample is proportional to the concentration of those
organic compounds in the samples which absorb light in the 200 to 800 nm region. Thus
the spectrophotometer records the wavelengths at which absorption occurs, together with
the degree of absorption at each wavelength. An ultraviolet-visible spectrum is essentially
a graph of light absorbance versus wavelength in a range of ultra-violet or visible regions.
MacCarthy and Rice (1985) reported that the absorbance of UV light by a molecule is
mainly due to aromatic ring structures. The aromatic molecules likely to absorb light in the
200 to 800 nm region are referred to as chromophores. Chromophores responsible for
absorbance consist of conjugated double bonds and unbonded electrons like those
associated with oxygen, sulphur and halogen atoms.
Kang et al. (2002) and Weishaar et al. (2003) reported that the different wavelengths of
absorption have many useful applications in characterizing the aromatic carbon content,
aromaticity and molecular weight of the dissolved organic matter (DOM). For example,
UV absorbance at the 280 nm wavelength has been demonstrated to be a useful indicator
of the aromaticity of a sample’s structure (Nanny and Ratasuk, 2002; Kang et al., 2002).
The ratio of the absorbance at 465 nm and 665 nm (E4/E6 ratio) has also been used to
provide an indication of the molecular size of the DOM. High molecular weight
fractions, which have high carbon contents and relatively low numbers of oxygen and
carboxyl groups, are associated with relatively low E4/E6 ratio values. Weishaar et al.
(2003) applied the specific UV absorbance (SUVA), defined as the ratio of UV
absorbance at 254 nm wavelength to DOC (UV254/DOC) to estimate the dissolved
Background
25
aromatic carbon content. It was shown that SUVA was strongly correlated (R2 = 0.97)
with percent aromaticity determined by 13C NMR for several organic matter isolates
from different aquatic environments. Imai et al. (2002), Musikavong and Wattanachira
(2007) and Saadi et al. (2006) also reported SUVA to be an important parameter in
wastewater treatment plants.
UV absorption spectroscopy has also been used to measure the water quality parameters
(Matsche and Strumworher, 1996; Chevalier et al., 2002; Chevakidagarn, 2005).
Different wavelengths have been shown to correlate with BOD, COD and TOC. For
example, Brookman (1997) investigated the absorbance at 280 nm for slurry and farm
effluents and found that it could be useful for the rapid estimation of BOD. Chevalier et
al. (2002) and Chevakidagarn (2005) reported that UV absorbance at 220 and 260 nm
wavelengths correlated well with BOD. UV absorbance at 254 nm wavelength has been
found by most authors to provide a close correlation with COD and DOC (Bari and
Farooq, 1984; James et al., 1985; Matsche and Strumworher, 1996), and suggested UV
absorbance at 254 nm as an excellent surrogate parameter for estimating the aromaticity
of DOM.
2.8.2 Fluorescence spectroscopy
Fluorescence spectroscopy is a type of electromagnetic spectroscopy which analyses
fluorescence from a sample. Fluorescence occurs when a loosely held electron in an atom
or a molecule is excited to a higher energy level by the absorption of light and loses
energy as the electron returns to its original energy level. Some energy within the
molecules is lost through heat or vibration so that emitted energy is less than the exciting
energy; i.e., the emission wavelength is always longer than the excitation wavelength.
Aromatic organic compounds are the principal fluorescence centres in organic matter
because they have consecutive conjugated double or triple bonds (i.e. double or triple
bonds separated by a single bond). Senesi et al. (1991) reported that the content of
substituent groups such as carboxyls and carbonyls (having double bonds) also increases
excitation and emission wavelengths. Molecules containing oxygen or nitrogen atoms,
having lone electron pairs also enhance fluorescence (Senesi et al., 1991). Humic
substances (humic and fulvic type molecules), the essential building block of recalcitrant
Background
26
organic compounds (Senesi et al., 1991; Yamashita and Tanoue, 2003) (described in
section 2.6) are the principal fluorescent organic matter because these molecules contain
aromatic compounds, and a high content of carboxylic groups, polycondensed aromatic
and conjugated structures. Additionally, proteinaceous material exhibits fluorescence
derived from the presence of aromatic amino acids such as phenylalanine, tryptophan and
tyrosine (Coble, 1996).
Lackowicz (1999) reported that the wavelength at which the excitation and emission
occurs is specific to particular molecules. For example, fluorescence peaks at 270-280 nm
excitation and 305-312 nm and 340-370 nm emission can be ascribed to protein-like
substances whereas peaks at 230-260 nm and 300-370 nm excitation and 400-470 nm
emission can be considered as humic substances (Baker, 2002; Baker and Curry, 2004). In
recent years Excitation-Emission-Matrix (EEM) fluorescence spectroscopy has become an
increasingly used technique. It is based on the principle that excitation and emission
wavelengths and fluorescence intensity can be scanned over a range of wavelengths
synchronously and plotted on a single chart, developing a ‘map’ of optical space. Figure
2.2 presents a fluorescence 3D EEM for a hypothetical leachate sample. Fluorescence
peaks may vary in intensity relative to each other depending on the organic matter source,
whereas overall fluorescence intensity variations will reflect changes in concentration.
The important factor in the use of fluorescence spectroscopy is the effect of inner-filtering
often referred to as re-absorption (Larsson, 2007). Re-absorption happens because another
molecule or part of a macromolecule absorbs at the wavelengths at which fluorophore
emits radiation. Another inner filtering effect (IFE) occurs when using concentrated
solutions. Yang and Zhang (1995) defined inner filtering as a shift to longer emission
wavelengths (red shift) due to higher concentrations of fluorophores in the solution and a
shift to shorter emission wavelengths (blue shift) due to decreasing solution
concentrations. Lackowicz (1999) reported that this IFE is important as it can reduce the
fluorescence intensity by 5%. This mechanism is different from quenching, although
quenching reduces fluorescence intensity as well. Lackowicz (1999) defined quenching as
the process by which fluorescence is reduced because the excited molecules lose their
energy via other processes, such as interaction with other molecules rather than emitting
light. Various suggestions have been made about the appropriate correction for inner
filtering effect (IFE) (Senesi, 1990, McKnight et al., 2001). Senesi (1990) suggested
Background
27
dilution as the easiest method to reduce IFE. Zsolnay et al. (1999) and Cox et al. (2000)
stated that reducing the absorbance values of the sample <0.1 at 340 nm could avoid IFE.
However, Tucker et al. (1992), Lackowicz (1999) and McKnight et al. (2001) applied a
mathematical correction to fluorescence data to eliminate the effect of inner filtering.
These methods are based on the measured absorbance of the sample. Since dilution is not
always possible, and corrections based on absorbance are primarily based on the use of a
separate absorbance instrument, neither of these solutions is optimal. In this case, Larsson
et al. (2007) proposed a mathematical correction procedure based on the intensity of
Raman scatter from water. This procedure was found to reduce the error after correction
by up to 50% in comparison with Lackowicz (1999) absorbance correction procedure.
Furthermore, it does not require the use of a separate absorbance measurement, and it is
applicable to on-line and in situ EEM recordings, where the IFE would otherwise cause
problems.
Emission wavelength (nm)
200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
(1)
(2)(3)(4)
Figure 2.2 Fluorescence EEM for a hypothetical leachate sample to show all possible
fluorescence centres (1) Fulvic-like (at 320-370 excitation and 400-470 emission) (2)
Humic-like (at 230-260 and 300-370 excitation and 400-470 emission) (3) Tyrosine-
like (at 270-280 excitation and 305-312 emission) and (4) Tryptophan-like (at 270-280
excitation and 340-380 emission) (Baker, 2002; Baker and Curry, 2004).
Background
28
Fluorescence intensity depends on a number of environmental factors (Hudson et al.,
2007). For example, fluorescence intensity is highly sensitive to pH. Huatala et al. (2000),
Westerhoff et al. (2001) and Patel-Sorrentino et al. (2002) found an increase of
fluorescence intensity (10-40%) in response to an increase in pH. Changes in observed
fluorescence intensity are a result of conformational changes in the humic and fulvic
molecules (Westerhoff et al., 2001) exposing or hiding fluorescent parts of the molecule.
This phenomenon has been related to the fact that at high pH the macromolecule has a
linear structure, and at low pH these structures contract. Patel-Sorrentino et al. (2002) has
suggested that at low pH fluorophores may be situated within the coiled structures and are
masked by non-fluorescent components and accordingly do not contribute to the
fluorescence intensity. To overcome this, Her et al. (2003) suggested that all samples
should be standardised at pH 7.0. Blaser et al. (1999), Sharpless and McGown (1999),
Esteves da Silva et al. (1998), Elkins and Nelson (2001) and Fu et al. (2007) reported that
fluorescence intensity depends on the presence of metal ions (aluminium, iron, copper,
lead, nickel) through the formation of insoluble metal ion complexes. For example,
Reynolds and Ahmad (1995) showed that aluminium and copper caused fluorescence
intensity to be quenched by up to 40%. Temperature is also an important factor in
fluorescence properties (Baker, 2005). Vodacek and Philpot (1987) stated that
fluorescence is inversely related to temperature due to increased collisional quenching at
higher temperatures. Conventionally, the temperature is held constant at 20oC during
fluorescence analysis to avoid any interference from thermal quenching.
The fluorescence spectroscopic technique has been widely used in marine and estuarine
waters, freshwater and wastewater to characterise DOM and to fingerprint the organic
pollutants. The majority of fluorescence research has been carried out in marine waters to
characterise the fluorescence properties of DOM. Coble (1996) examined the variability in
fluorescence of natural DOM, attempting to distinguish between humic substances from
terrestrial and marine sources. Several studies also applied fluorescence spectroscopy as a
tool to determine the biological activity and protein fluorescence in marine environments
(Determann et al., 1998; Parlanti et al., 2000; Yamashita and Tanoue, 2003; Jaffe et al.,
2004). It was found that protein-like fluorescence (Tyrosine and Tryptophan-like) in
marine and estuarine waters derived directly from biological activity (bacterial growth and
death) (Parlanti et al., 2000; Yamashita and Tanoue, 2003).
Background
29
However, the study of the DOM fluorescence in freshwater is not yet as widespread as that
in marine science. Freshwater fluorescence has been investigated with an emphasis on
spatial variation (Baker and Spencer, 2004; Fu et al., 2007). Baker and Spencer (2004)
found that tryptophan-like fluorescence increased in river downstream with the increased
anthropogenic input and urbanization. Similar results were observed in an urban river by
Fu et al. (2007) and on a catchment scale by Baker et al. (2003). Baker (2002) and Baker
et al. (2003) also observed that tryptophan-like fluorescence varied seasonally and was
intense in summer. The authors attributed this to be due to reduced base flow and low
dilution.
A number of studies have applied fluorescence EEM to characterize wastewater DOM in
treatment processes (Reynolds and Ahmad, 1997; Reynolds, 2002; Westerhoff et al., 2001
and Her et al., 2003; Cammack et al., 2004; Elliot et al., 2006). These studies identified
tryptophan-like fluorescence as being most likely to relate to the biodegradable fraction of
wastewater, with a 90% reduction across a treatment process. Fluorescence EEM has been
used to characterise the DOM in sewage impacted rivers (Baker, 2001), groundwater
(Baker and Lamont-Black, 2001), wetland (Maie et al., 2007), lake water (Mostofa et al.,
2005), sea ice (Stedmon et al., 2007) and soil (Ohno and Bro, 2007). The correlation
between fluorescence intensity (FI) and chemical and biological water quality monitoring
parameters such as BOD, COD and TOC (DOC) has been studied in various fields (urban
river, Baker, 2002; Mostafa et al., 2005; Fu et al., 2007; river, wetland, spring, pond and
sewage, Cumberland and Baker, 2007; landfill leachate, Baker and Curry, 2004).
Additionally a number of studies have been carried out to track and characterise the
presence of anthropogenic compounds by fluorescence signature in the freshwater system
that include Fluorescent Whitening Agents (FWAs) from tissue mills and laundry products
(Baker, 2002); a mixture of fluorescent xenobiotic organic matter (XOM), naphthalene
from landfill leachate (Baker and Curry, 2004); material from agricultural effluent (Baker,
2002).
The above studies (section 2.7.1 and 2.7.2) suggest that UV and fluorescence spectroscopy
are adaptable tools that might have potential for application in landfill leachate research.
The major advantages are that these techniques are rapid (1 min per sample), non-
destructive and do not require chemicals or sample preparation (Baker, 2001). Despite
these advantages, the use of UV and fluorescence spectroscopy in characterising
Background
30
recalcitrant compounds in leachates is rare. A limited number of studies are reported. For
example, following Baker and Curry’s (2004) report of fluorescence properties of
leachates, Zheng et al. (2007) and Ham et al. (2008) used this technique to detect
polycyclic aromatic hydrocarbon (PAHs), phthalates (PAEs) and polychlorinated
biphynyls (PCB) in landfill leachates. Similarly the technique has been applied for
characterising leachates from landfills of different ages (Kang et al., 2002; Huo et al.,
2008), and different treatment process (Huo et al., 2008). However, its use to detect the
change in compositions of recalcitrant organic compounds with time and waste
biodegradation has not been reported. In addition, UV and fluorescence spectroscopy
allow differentiation between biodegradable and recalcitrant compounds as well as
fingerprinting the other organic pollutants in leachates.
2.9 Summary
In this chapter, recalcitrant organic compounds in leachates and its removal by various
leachate treatment processes were reviewed. It was found that a significant amount of
recalcitrant organic compounds remain after the widely applied economically viable
biological treatment processes. Applications and limitations of conventional leachate
characterisation methods like BOD, COD and TOC were also discussed. These techniques
usually require considerable sample preparation and cannot provide detailed compositional
analyses of recalcitrant organic compounds. Available methods for investigating
constituent organics of recalcitrant compounds were reviewed. It was found that
degradation methods for compositional analysis have disadvantages such as unknown
denaturing reactions, transformations and substantial DOC lost. Although spectroscopic
methods FTIR and NMR are interesting, an unambiguous estimation of the structure of
constituent organics often requires prior separation of identifiable sub-compounds using
degradation. As a result, these techniques cannot be accepted as rapid and economic
techniques for onsite compositional analysis of recalcitrant compounds in leachates. In
search of an economically viable technique, the application of UV and fluorescence
spectroscopy to marine and estuarine waters, freshwater, and wastewaters were reviewed.
These studies imply that UV and fluorescence spectroscopy have potential for effective
monitoring of recalcitrant organic compounds in leachates which could be applied to
landfill leachate management system to assist and enhance routinely used characterisation
Background
31
methods. In the subsequent chapters, these techniques are used for characterisation of
leachates to perform a feasibility study on the use of UV and fluorescence spectroscopy.
Characterisation of wastes and leachates
32
Chapter 3 Characterisation of wastes and leachates
3.1 Introduction
This chapter presents the detailed experimental procedures used to research the nature of
the recalcitrant organic compounds in leachates. Two laboratory experiments were carried
out in this regard. In one experiment, a laboratory scale aerobic biological treatment
process was carried out on leachate samples collected from two UK MSW landfills, Pitsea
and Rainham in Essex. This experiment provides information on the composition of
recalcitrant organic compounds in various types of landfill leachates at different stages of
treatment process. In the second experiment, Biochemical Methane Potential (BMP) tests
were carried out on four types of solid waste samples (fresh waste, composed waste,
newspaper waste and synthetic waste) to evaluate the effect of the composition of wastes
on the recalcitrant organic compounds in the leachates generated during anaerobic
biodegradation of the waste components. These two experiments aim to fulfil the first
research objective of understanding the nature of the recalcitrant organic compounds in
leachates; particularly how they are removed in the course of a biological treatment
processes as well as how the concentration changes during anaerobic biodegradation of
waste components. Finally the application of potentially useful techniques; UV absorption
and fluorescence spectroscopy for investigating the nature of recalcitrant materials in
leachates is presented. This aims to fulfil the second research objective on the basic
feasibility study of the application of UV absorption and fluorescence spectroscopy for a
rapid and an economical characterisation of the organic material in landfill leachates.
3.2 Aerobic biological leachate treatment 3.2.1 Materials (Landfill sampling sites and leachate composition)
Four untreated and two treated leachate samples were collected from two MSW landfills,
Pitsea and Rainham, to investigate their properties. The leachates were characterized using
a variety of methods (section 3.4). The characteristics of the landfill sites and the collected
leachate samples, and details of the treatment carried out on the leachate samples, are
summarised in Table 3.1. Both sites have been in operation for over 50 years. The Pitsea
Characterisation of wastes and leachates
33
landfill, located 30 miles east of London, has received mainly MSW with commercial,
industrial, hazardous and non-hazardous liquid wastes. The Rainham landfill, also located
east of London on the north bank of the River Thames, contains old phases that have
accepted MSW and hazardous industrial wastes, and more recent phases that contain
predominantly MSW.
Table 3.1: Characteristics of landfills and leachate samples (Information provided by
landfill operators)
Sample Label Nature of waste from which leachate derived Nature of treatment
1 Pitsea (LTP)
Landfill Pitsea Mixed MSW, commercial, industrial and hazardous
wastes, with hazardous liquid wastes until 2002, non hazardous liquid waste thereafter, wastes range in age from 0 to 80 years
Leachate was passed though a treatment plant
consisting of aerobic rotating biological
contactors
2
Pitsea (P4)
Landfill Pitsea, phase 4 Predominantly MSW waste, with some industrial
waste, with liquid non-hazardous wastes, wastes less than 4 years old
The phases not be hydraulically isolated from each other. Large scale leachate recirculation for many years, significant potential of mixing of leachates from different areas.
Untreated
3
Rainham (LTP)
Landfill Rainham, yellow phase Predominantly MSW, some hazardous industrial
wastes, over-tipped by recent non hazardous MSW, waste ~30-60 years old.
The phases not be hydraulically isolated from each other, less potential of mixing of leachates from different areas.
Untreated
4
Rainham (FE)
Landfill Rainham Sample is effluent from leachate treatment plant,
treating “red” and “yellow” phase leachate
Treatment primarily consists of oil separation,
air stripping and settlement of iron sludge, followed
by activated sludge aerobic treatment
5
Rainham (P2)
Landfill Rainham, phase 2 Predominantly domestic MSW and non hazardous
industrial wastes, waste 0-20 years old. The phases not be hydraulically isolated from each
other, less potential of mixing of leachates from different areas.
Untreated
6
Rainham (LTP Haz)
Landfill Rainham, red hazardous Predominantly MSW, some hazardous industrial
wastes. Elevated concentrations of chlorinated aliphatic and
aromatic organic compounds than Rainham (LTP), waste ~30-60 years old.
The phases not be hydraulically isolated from each other, less potential of mixing of leachates from different areas.
Untreated
Characterisation of wastes and leachates
34
3.2.2 Experimental methods
Aerobic biodegradation experiments on the four untreated and two treated leachate
samples were carried out in 1 litre batch glass reactors over a period of 30 days. Air was
supplied by an AQUATEC aquarium pump through 5 mm diameter poly-vinyl chloride
(PVC) tubing via a 3 cm long airstone. The air streams into the reactors were maintained
at a flow rate of 100 ml/min. 500 ml of diluted leachate samples (20 ml leachate samples:
480 ml BOD water) were in contact with air for 30 days at 20oC (The recipe for BOD
water is given in section 3.2.3). 15 ml samples were taken on the 2nd, 6th, 10th, 15th, 20th,
25th, and 30th days of the treatment, filtered through a 25 mm micro glassfibres filter
(Fisher Brand, MF 200), stored in the freezer and treated according to the scheme shown
in figure 3.1. The results were corrected by multiplying by the dilution factor 25.
3.2.3 Preparation of BOD water
Leachate contains more oxygen demanding materials than the amount of dissolved oxygen
available in the air saturated water. Therefore, it is essential to dilute the sample before
incubation to bring the oxygen demand and supply into the appropriate balance and hence
BOD water is used. 1 ml of each of the following four stock solutions was added to 1 litre
distilled water to prepare the BOD water (APHA, 1998).
Reagents
1. Phosphate buffer solution: 8.5g KH2PO4, 21.75 g K2HPO4, 33.4 g Na2HPO4.7H2O
and 1.7 g NH4Cl were dissolved in about 500 ml distilled water and diluted to 1 litre. The
pH was 7.2.
2. Magnesium sulphate solution: 22.5 g, MgSO4.7H2O was dissolved in distilled water
and diluted to 1 litre.
3. Calcium chloride solution: 27.5 g, CaCl2 was dissolved in distilled water and diluted
to 1 litre.
Characterisation of wastes and leachates
35
4. Ferric chloride solution: 0.25 g, FeCl3 was dissolved in distilled water and diluted to 1
litre.
Figure 3.1 Analyses carried out on the leachates taken from aerobic reactors
3.3 Biochemical Methane Potential (BMP) test
Biochemical Methane Potential (BMP) tests were carried out to assess the variability in
the anaerobic biodegradability of the component solid materials. There are several
techniques for BMP testing of solid wastes, although all generally involve the incubation
of a small representative waste sample under controlled anaerobic conditions (usually
mesophilic at 30oC). However, the precise procedures in terms of pre-treatment of the
samples, the inoculum, gas measurement techniques and incubation vary significantly
among the published methods (e.g. Harries et al., 2001; Heerenklage and Stegmann, 2001;
15 ml of leachates taken from aerobic reactors
5 ml for COD analysis
(section 3.4.2) (Jirka and
Carter, 1975).
5 ml for carbon analysis (section
3.4.3).
Unfiltered 2.5 ml samples were preserved by the addition of 25 µl of 100% hydrochloric acid to
remove IC from the samples thus giving the TOC (TOC =
TC – IC) (APHA, 1998).
2.5 ml sample filtered through a 25 mm micro glassfibres filter (Fisher
Brand, MF 200) and preserved by the addition of 25 µl of 100%
hydrochloric acid to remove IC from the samples thus giving the DOC (DOC = DC – IC) (APHA, 1998).
5 ml samples for spectroscopic analysis (UV and fluorescence
spectroscopy). The spectroscopic analysis
did not require any sample preparation
(section 3.4.10).
Characterisation of wastes and leachates
36
Hansen et al., 2004). In this study, the BMP tests were carried out according to the
procedures described by Zheng et al. (2007) and Ivanova et al. (2008).
3.3.1 Material (Pre-test waste sample preparation procedure)
In this experiment, characterization of four types of wastes was undertaken, FW- fresh
waste, CW- composted waste, NW- newspaper waste and SW- synthetic waste (Table
3.2).
The most common method for characterizing a waste is to separate the waste into a
number of different components and determine the percentage of each component. Fresh
waste samples were supplied by the Otterbourne waste transfer station operated by Veolia
Hampshire Ltd (Otterbourne, Winchester, Hampshire, UK), after separation into different
components such as food, grass, newspapers and removal of the plastics and non-
biodegradable components such as nappies, textiles, electrical appliances and construction
material. The sorting analysis result of the fresh waste samples was supplied by the
Otterbourne waste transfer station and is presented in Table 3.3. A sub-sample (20 kg) of
the fresh waste was composted aerobically in the laboratory for 42 days to produce a
sample of composted waste (CW) which was described in Zheng et al. (2007). Synthetic
waste was prepared in the laboratory by mixing food, grass and newspapers wastes
according to the composition of the supplied fresh waste as described in Table 3.3.
100 g of weighted sample of the fresh waste, composted waste, newspaper waste and
synthetic waste were ground to a fine powder and analysed for lignin, cellulose and hemi-
cellulose (section 3.4.8), TC and total nitrogen (TN) (section 3.4.7). For each sample
triplicate experiments were carried out.
3.3.2 Experimental procedure
BMP tests were carried out on each of the four different types of waste samples (FW, CW,
NW and SW) in twelve 1 litre plastic Nalgene bottles (three bottles for each waste
sample). The BMP apparatus is shown in figure 3.2. 100 g of oven dried waste sample was
placed in each reactor bottle. Individual reactor bottles were opened periodically in an
anaerobic cabinet under a nitrogen atmosphere (it is acknowledged that replicate reactors
Characterisation of wastes and leachates
37
would have provided a better statistical data set) to observe the compositional changes of
leachates taken at different stages of the component waste biodegradation with respect to
time of incubation. Solid compositions were determined at the beginning and at the end of
the experimental period for each of the bottled wastes. Gas volumes were measured and
samples of gas were taken at the end of the experiment to analyse methane (CH4) and
carbon dioxide (CO2). Leachate sampling is outlined in Table 3.4.
Table 3.2 Types of waste samples
Waste
Reference Waste description
Fresh waste
(FW)
Fresh waste sample obtained from Otterbourne waste transfer
station operated by Veolia Hampshire Ltd (Otterbourne,
Winchester, Hampshire, UK).
Composted
waste (CW)
Composted waste obtained from the composting of fresh waste for
42 days in the laboratory (Zheng et al., 2007).
Newspaper waste
(NW) Newspaper waste.
Synthetic waste
(SW)
Synthetic waste was prepared in the laboratory same to the
composition as the fresh waste.
Table 3.3 Breakdown of waste composition for fresh waste samples
Component Dry mass (g) % by dry mass
Newspaper/card board 124.15 34.39
Food 105.05 29.10
Grass 131.80 36.51
Total 361.00 100.00
Characterisation of wastes and leachates
38
Figure 3.2 BMP test assay apparatus
Table 3.4 Frequency of leachate analysis
BMP reference Days into experiment Leachate sampling
frequency (Day)
0-20 Every 2nd day
20-70 30, 50, 70 BMP (FW, CW, NW and
SW sample) 70-150 80, 90, 100, 120, 140, 150
To accelerate the degradation of the waste by anaerobic bacteria, 700 ml of a laboratory
prepared methanogenic medium containing mineral nutrients (nitrogen, phosphorous, and
sulphur), trace elements and anaerobic digested sewage sludge (10% by volume) derived
from an anaerobic digester at Millbrook Sewage Works (Southern Water, UK) was added
to each BMP reactor to act as a seed inoculum. The medium was adapted from Florencio
et al. (1995) and is presented in Table 3.5. The final pH of the medium was adjusted to pH
7.83 by adding 2M NaOH.
Characterisation of wastes and leachates
39
Table 3.5: Recipe of the methanogenic mineral medium (Florencio et al., 1995)
Reagents Concentration
(mg/l) Reagents
Concentration
(mg/l)
NH4Cl 280.00 CuCl2.2H2O 0.04
K2HPO4.3H2O 330.00 (NH4)6MoO24.4H2O 0.05
MgSO4.7H2O 100.00 Na2SeO3.5H2O 0.16
CaCl2.2H2O 10.00 CoCl2.6H2O 2.00
FeCl2.4H2O 2.00 AlCl3.6H2O 0.09
H3BO3 0.05 NiCl2.6H2O 0.14
ZnCl2 0.05 EDTA 1.00
MnCl2.4H2O 0.50
The mineral medium was kept under a nitrogen atmosphere for one day to remove oxygen.
The waste samples were then mixed with methanogenic medium and inoculum and sealed
under anaerobic conditions before being placed in a water bath at 30oC and incubated to
promote mesophilic methanogenic conditions. No mechanical mixing of the waste took
place during the test.
In this study the BMP tests were carried out over a period of 150 days although Harries et
al. (2001) showed that virtually all the gas was produced from a similar waste samples in
90 days.
Biogas production was measured by collecting the gas produced from each bottle in an
inverted glass burette containing water acidified with HCl to a pH of 2.0. The acidwater
was displaced as the gas accumulated within the burette allowing the gas volume to be
measured. Acidwater was used to prevent CO2 dissolution. Three control (blank) reactors
containing 700 ml of the mineral nutrient/trace element sewage sludge mix were used to
determine the volume of gas produced from the inoculum alone. The measured volume of
biogas produced was corrected to standard temperature and pressure (STP) using the
ambient temperature and pressure measurements (Eq.3.1).
Characterisation of wastes and leachates
40
)2.273
2.273()3.101
(0 TxPRR a
+= (Eq.3.1)
where, R is the corrected reading, 0R is the uncorrected reading, 101.3 is the atmospheric
pressure at sea level in kPa, aP is the measured atmospheric pressure in the laboratory in
kPa, 273.2 is the temperature in Kelvin at 0oC, and T is the reference temperature in the
laboratory in oC.
3.3.3 Leachate sampling
The BMP test reactors were opened sequentially under a nitrogen atmosphere at various
times to observe compositional changes in the leachates. 20 ml leachate samples were
taken from each bottle periodically, half of which was then filtered through 1.2 µm
Whatman GF/C filters. Equivalent volumes of the methanogenic mineral medium were
then added to the bottles to keep the volume of the liquid unchanged. Unfiltered and
filtered leachate samples were diluted and analysed as shown in figure 3.3.
3.3.4 Solid waste sampling
The composition of the four types of solid waste samples was determined by elemental
and fibre analysis at the beginning and at the end of the experiment to assess changes due
to biodegradation. The procedure for the preparation and analysis of the solid waste
samples is shown in figure 3.4.
Characterisation of wastes and leachates
41
Figure 3.3 Analyses carried out on the leachates taken from BMP reactors
20 ml of leachates taken from BMP reactors
5 ml for COD analysis
(section 3.4.2) (Jirka and
Carter, 1975).
5 ml for carbon analysis (section
3.4.3).
Unfiltered 2.5 ml samples were preserved by the addition of 25 µl of 100% hydrochloric
acid to remove IC from the samples thus giving the TOC (TOC = TC –
IC) (APHA, 1998).
2.5 ml sample filtered through a 25 mm micro glassfibres filter (Fisher
Brand, MF 200) and preserved by the
addition of 25 µl of 100% hydrochloric acid to remove IC from the samples thus giving the DOC (DOC = DC – IC)
(APHA, 1998).
5 ml samples for spectroscopic
analysis (UV and fluorescence
spectroscopy). The spectroscopic
analysis did not require any sample
preparation (section 3.4.10).
5 ml samples for cation analysis
(filtered through a 25 mm micro glassfibres filter (Fisher Brand, MF 200) and
preserved by the addition of 20 µl
of 100% methanesulfonic
acid (MSA) (section 3.4.6).
Characterisation of wastes and leachates
42
Figure 3.4 Analyses carried out on the solid samples from BMP reactors (Ivanova et
al., 2008)
Solid waste sample analysis
At the beginning of the experiment At the end of the
experiment
100 g of each type of waste sample
Divide sample into two sub-samples
10 g samples for elemental analyses
10 g samples for fibre analysis
Degraded wet waste sample + biomass
Oven drying at 70oC and ground to a fine powder Oven drying at 70oC (dried
degraded sample)
Dried sample washed with distilled water through a 2 mm sieve to remove any attached
biomass. (Washed wet degraded waste sample)
Oven drying at 70oC (Washed dried degraded sample) (FW = 82 g, CW = 85 g, NW = 84.5 g, SW
= 81.5 g)
Samples divided into two sub-samples
5 g samples for elemental analyses
5 g samples for fibre analysis
Characterisation of wastes and leachates
43
3.4 Analytical measurements 3.4.1 Biochemical oxygen demand (BOD) analysis
The BOD30 of the landfill leachate samples were analysed using the WTW Oxi-Top
control system. The system uses respirometric measurement whereby if O2 is consumed in
a closed vessel at constant temperature, a negative pressure develops as the carbon dioxide
(CO2) produced is absorbed by the potassium hydroxide (KOH) pellets at the top of the
vessel. The OxiTop measuring head measures and stores this pressure for the whole
duration of the BOD test. 22.7 ml of landfill leachate samples were used for BOD test.
The BOD is calculated using Equation 3.2
)()..(.
)(2
01
12 OpTT
VVV
TROM
BOD mt
m
∆+−
= α (Eq. 3.2)
Where,
M(O2) = molecular weight (32000 mg/mol)
R = gas constant (83.144 mbar/mol-K)
T0 = reference temperature (273.15 K)
Tm = measuring temperature
Vt = bottle volume (ml)
V1 = sample volume (ml)
α = Bunsen absorption coefficient (0.03103)
)( 2Op∆ = difference of the oxygen partial pressure (mbar)
3.4.2 Chemical oxygen demand (COD) analysis
COD measurements of the leachate samples taken from the aerobic biodegradation reactor
and also from the BMP reactors (figures 3.1 and 3.3) were carried out using the micro-
digestion technique (Jirka and Carter, 1975). The Standing Committee of Analysis (SCA)
standard for COD analysis was not used in this study because the SCA method was
limited by a maximum detection limit up to 400 mg/l of COD while in this study
considerably high COD values were presumed. The COD determination provides a
measure of the oxygen equivalent of that portion of the organic matter in a sample that is
susceptible to oxidation by a strong chemical oxidant.
Characterisation of wastes and leachates
44
5 ml of sample volumes (figures 3.1 and 3.3) filtered through a 25 mm glass microfibre
filter (Fisher Brand, MF 200) were refluxed with known amounts of potassium
dichromate and sulphuric acid (COD reagent), and the excess dichromate was titrated with
ferrous ammonium sulphate (FAS). The amount of oxidizable organic matter, measured as
oxygen equivalent, was proportional to the potassium dichromate consumed.
COD is calculated as follows:
COD = 8000 * (VB – VS) * MFAS * Df/sample volume (Eq. 3.3)
Where,
VB = volume of FAS used in titrating the appropriate blanks (ml)
VS = volume of FAS used in titrating the sample (ml)
MFAS = molarity of FAS
Df = dilution factor
8000 = milli-equivalent weight of oxygen x 1000 ml/L
Because COD measures the oxygen demand of organic compounds in a sample of
leachate, it is important that no outside organic material be accidentally added to the
sample to be measured. To control this, a blank sample was created by adding all reagents
(e.g. acid and oxidizing agent) to a volume of distilled water. COD was measured for both
the leachate sample and blank samples, and the two were compared. The COD in the
original sample was subtracted from the COD for blank sample to ensure a true
measurement of organic matter.
3.4.3 Carbon analysis
A high-temperature total organic carbon analyzer (Dohrman Rosemount DC-190, USA)
was used to measure the leachate total carbon (TC), total inorganic carbon (TIC) and total
organic carbon (TOC). The equipment contains a vertical quartz combustion tube packed
with cobalt catalyst. Oxygen flows through it at a rate of 200 ml/min. The furnace was
operated at 800oC. The manual injection mode was used for the leachate samples collected
from the landfills at Pitsea and Rainham. The samples were analysed three times and the
average value taken. The boat sampling mode was used due to the high concentration of
suspended solids in the unfiltered leachate samples collected from the BMP test reactors.
Characterisation of wastes and leachates
45
The equipment utilized a single point calibration which was carried out each time prior an
analysis. TC and TIC standard were prepared using Glycine (NH2CH2COOH) and sodium
bicarbonate (NaHCO3) respectively at a concentration level depending on the
concentration range of the samples.
3.4.4 Solids content
Total solids (TS), volatile solids (VS), total suspended solids (TSS), volatile suspended
solids (VSS), and total dissolved solids (TDS) of landfill leachate samples were carried
out according to the standard methods (APHA, 1998). TS contents were determined by
oven drying at 105oC and VS by oven drying at 550oC. Again the Standing Committee of
Analysis (SCA) standard for TS, VS, TSS, VSS, and TDS analyses was not used due to
the uncertainty of the recommended filter size and manufacturer.
3.4.5 Anion analysis (Cl-, NO2-, NO3
-, PO4-2, SO4
-2)
5 ml landfill leachate samples were filtered through a 25 mm glass microfibres filter
(Fisher Brand, MF 200) and the filtrate were immediately frozen. Anion analysis was
carried out using a Dionex-500 ion chromatograph with an AS9 anion column in
conjunction with an ASRS-1 anion suppressor (Dionex Ltd.). 25 µl volume injections
were applied to the column using a Dionex AS-40 auto-sampler, incorporating a rheodyne
valve. Detection was by a Dionex ED-40 electrochemical detector operating in
conductivity mode. The eluent consisted of 8 mM sodium carbonate and 1 mM sodium
bicarbonate pumped at a flow rate of 1.0 ml/min.
A calibration procedure was carried out prior to each analysis using anion stock solution
prepared from a mixture of NaCl, NaNO2, KNO3, KH2PO4 and MgSO4.7H2O with
concentration levels of 50, 100, 300, 600, 800, 1000 µM.
3.4.6 Cation analysis (Na+, NH4+, K+, Mg+2, Ca+2)
5 ml volumes of filtered leachate samples were used for the cation analysis. This was
carried out using a Dionex-500 ion chromatograph with a CS12A cation column in
Characterisation of wastes and leachates
46
conjunction with a CSRS-II cation suppressor (Dionex Ltd.). The mobile phase was 20
mM methanesulfonic acid pumped at a rate of 1.0 ml/min.
A calibration procedure was carried out prior to each analysis using cation stock solution
prepared from a mixture of NaCl, NH4NO3, KH2PO4, MgSO4.7H2O and CaCl2.2H2O with
concentration levels of 50, 100, 300, 600, 800, 1000 µM.
In this study CEN standard for anion and cation analyses was not used because of the
potential disadvantage of low detection limit for various elements.
3.4.7 Elemental analysis
The total carbon (TC) and total nitrogen (TN) content of the solid waste samples (FW,
CW, NW and SW) at the beginning and at the end of biodegradation process were
determined using a CE Instruments 1112 Flash Elemental Analyser. Prior to CHNS
analysis, solid waste samples were pre-treated according to the procedure described in
section 3.3.4 (figure 3.4). Dried samples were ground to fine powder using a Knifetec
1095 Sample Mill (Foss) and then weighed (2-5 mg) in tin containers. TC and TN contents
of the samples were determined by dry combustion at 900oC in an oxygen atmosphere
with 140 ml/min helium carrier gas and Thermal Conductivity Detector (TCD) detection
of gases produced. L-cysteine, Methionine and Sulphanilamide were used as a standards
and Vanadium pentoxide was used as sample additive to help in oxidation.
3.4.8 Fibre analysis (cellulose, hemi-cellulose and lignin)
Fibre tests provide values for cellulose, hemi-cellulose and lignin content in the waste
(Van Soest et al., 1991). The major methods used to determine fibre fractions are the
Neutral Detergent Fibre (NDF), Acid Detergent Fibre (ADF) and Acid Digestible Lignin
(ADL) tests. The NDF test dissolves readily degradable material such as pectins, sugars,
starch and fats and leaves behind cell wall components of plant material, cellulose, hemi-
cellulose and lignin. The ADF test in contrast, digests less degradable hemi-cellulose and
some proteins leaving a residue of cellulose, lignin and bound nitrogen. The ADL test is a
measure of lignin (Van Soest et al., 1991).
Characterisation of wastes and leachates
47
The components determined by these tests can be summarised as follows:
NDF = Cellulose + Hemicellulose + Lignin + Mineral Ash (Eq. 3.4)
ADF = Cellulose + Lignin + Mineral Ash (Eq. 3.5)
ADL = Lignin + Mineral Ash (Eq. 3.6)
Fibre analysis on the dried solid waste samples (section 3.3.4) was carried out using the
Foss Technology system FibreCap 2021/2023 (Kitcherside et al., 2000) at the beginning
and at the end of the BMP experiment. Prior to analysis solid waste samples were pre-
treated according to the procedure described in section 3.3.4 (figure 3.4). Dried samples
were ground to a fine powder using a Knifetec 1095 Sample Mill (Foss) and different fibre
fractions were analyzed several times. Less than 0.5 g of each sample was prepared by
multiple measurements using ultra sensitive weighting machine which was then used in
each capsule. The NDF test was carried out using chemical procedures as described by
Van Soest et al. (1991), except that the sample was retained in the FibreCap capsule. This
procedure uses α-amylase at two stages in the extraction process to improve the
solubilization of starch. The ADF test involves digesting the waste samples using a
cationic detergent (Cetyltrimethylammoniumbromide, 20 g/l) in 0.5M sulphuric acid for 1
hour after reaching boiling point. The ADL was assessed by further treating the non-dried
and non-ashed residues from previously performed ADF tests (first-step digestion) with
72% sulphuric acid at room temperature for 4 hours to dissolve the cellulose, leaving
lignin as the residue (Effland, 1977). The specific cell wall components (cellulose and
hemi-cellulose) were determined using Eq. (3.7) and Eq. (3.8).
Cellulose and hemi-cellulose can be determined according to the following equations:
Cellulose = ADF – ADL (Eq. 3.7)
Hemi-cellulose = NDF – ADF (Eq. 3.8)
Lignin = ADL (Eq. 3.9)
3.4.9 Gas analysis
Gas samples were collected using hypodermic syringes which were inserted through the
three way valves installed on the top of each BMP reactor. The samples were immediately
analysed.
Characterisation of wastes and leachates
48
The composition of biogas (CH4 and CO2) at the end of the BMP experiment was
determined by gas chromatograph (GC Varian 3800 gas chromatograph) equipped with a
Thermal Conductivity Detector (TCD) and two columns (a Haysep C 80-100 mesh a
Molecular Sieve 13 x 60-80 mesh, Analytical column, UK). Both columns were 1 meter
long, 6 mm diameter and were operated isothermally at 50oC. TCD and injection
temperature were 200oC and 100oC respectively. The injection volume was 0.25 ml.
Carrier gas argon flows through it at a rate of 6 ml/min.
The calibration procedure was carried out prior to each analysis using a calibration
mixture of CH4 and CO2 with concentration levels of 65% and 35% respectively.
3.4.10 Spectroscopic characterization of leachate DOC
5 ml volumes of leachate samples were used for spectroscopic analysis (figures 3.1 and
3.3). Spectroscopic analysis did not require any sample preparation except for the fact that
leachate samples were adjusted to pH 7.0 using 2M NaOH (section 2.8.2) because
fluorescence intensity is highly sensitive to pH change.
3.4.10.1 UV-visible spectroscopy
UV-visible absorption spectra of leachate samples were recorded on a Cecil
Spectrophotometer using a 10 mm quartz cell. Distilled water was used as blank. Leachate
samples were diluted until the absorbance value fell below 0.1/cm at 340 nm wavelength
(Baker, 2005). Scan spectra of the solutions were obtained over a wavelength range of
200-800 nm. Absorbance at 254, 465 and 665 nm wavelengths was also measured (section
2.7.1). SUVA values (l/mg/cm) as a measure of the relative contents of aromatic structures
in the overalls DOM were calculated as (UV254/DOC) x 100 (Weishaar et al., 2003).
3.4.10.2 Fluorescence spectroscopy
Fluorescence Excitation-Emission-Matrix (EEMs) were measured in a standard 10 mm
quartz cell using a Varian Cary Eclipse fluorescence spectrofluorometer equipped with a
temperature controller to enable the measurement of EEMs at precisely controlled
Characterisation of wastes and leachates
49
temperatures. The temperature throughout the study was held constant at 20oC in order to
avoid any interference from thermal quenching (Baker, 2005) (described in section 2.8.2).
Three dimensional EEMs were generated at excitation and emission slit widths of 5 nm
band pass. All samples were scanned in the excitation wavelengths 190-800 nm in 10 nm
steps and emission wavelengths 200-800 nm in 10 nm steps. The spectra of water blank
was obtained in the same conditions and was subtracted from the original spectra of
leachates to eliminate water Raman scatter peaks. Scan speed was 9600 nm/min. Samples
containing high concentrations of leachate were diluted to required dilution as described in
section 3.4.10.1 due to their high fluorescence intensity and 3D maps were built using
Sigma-plot 10 software. Coordinates of the main noticeable peaks were established in
these maps. Peak positions were not affected by this dilution (as tested for selected
samples of this study), indicating that this also avoided any inner-filtering effects (Hudson
et al., 2007).
EEMs are illustrated as the elliptical shape of the contours. The X and Y-axes represent
the emission and excitation wavelength respectively. Contour lines are shown for each
EEM spectrum to represent the fluorescence intensity. The total fluorescence intensity was
calculated by the cumulative integration of the intensity versus emission wavelength data
for any given excitation wavelength range corresponding to different types of materials.
The final intensity value was corrected by multiplying by the dilution factor.
Emission spectra were recorded in emission scanning mode over the range 300-500 nm at
excitation wavelengths 230-260 nm and 320-370 nm. Fluorescence spectra in emission
scanning modes provided important aromatic structural information with the position of
the fluorescence peaks of the different compounds. The total intensity was calculated by
adding the intensities from the intensity versus emission wavelength data for every
excitation wavelength.
3.5 Carbon mass balance estimate
Carbon mass balance estimates for each of the waste samples (FW, CW, NW and SW) at
the end of the BMP tests were carried out using the formula given by Ivanova et al. (2008)
(Eq. 3.10)
Characterisation of wastes and leachates
50
∑∑∑ += daccumulateoutin mmm (Eq.3.10)
This equation states that the sum of the masses flowing into a system should be equal to
the sum of the masses flowing out of the system and the sum of the mass being
accumulated within the system. The BMP test analysis had two input terms, i.e., the total
carbon of the initial waste and of the mineral media; two output terms, i.e., the total carbon
from the produced biogas and final leachate; and two measured accumulated terms, i.e.,
the carbon content in the degraded waste and in the Ca and Mg carbonate precipitates.
The carbon mass balance error can be estimated with the following equation (Ivanova et
al., 2008).
Mass balance error = Expected value – Actual value, (g/Kg DM) (Eq.3.11)
Where,
Expected value = Total carbon of the initial waste + Total carbon of the initial mineral
media, (g/Kg DM) and
Actual value = Total carbon from the produced biogas + Total carbon of the final leachate
+ Total carbon from degraded waste + Carbon precipitated as Ca and Mg carbonates
(CaCO3 and MgCO3).
The error can also be expressed as a percentage of the expected value:
Mass balance error, % = (Expected value – Actual value)/ Expected value x 100
The carbon output produced in the reactors was calculated from the volume of the biogas
produced in each reactor and the CH4 and CO2 concentrations measured in the collecting
burette (Ivanova et al., 2008). Biogas production was measured at 30oC. All biogas
readings were standardized to gas at STP using Eq. 3.1.
The total mass values of leachate carbon were estimated using the volumes of the leachate
in the reactors and the weight of the waste material from which the total carbon was
sourced. The carbon from the precipitation of CaCO3 and MgCO3 was calculated from the
changes in the concentrations of Ca2+ and Mg2+ at the beginning and at the end of the
experiment, the volumes of the leachate in the reactors and the weight of the waste
material (Ivanova et al., 2008).
Results and Discussion: Aerobic biodegradation of landfill leachates
51
Chapter 4 Aerobic biodegradation of landfill leachates
4.1 Introduction
This chapter investigates the nature of the recalcitrant organic compounds in leachates in
the course of a biodegradation by several methods. Landfill leachates were collected from
two UK MSW landfills, Pitsea and Rainham (section 3.2.1) and an aerobic biological
treatment process was carried out (described in section 3.2). At these sites, a range of
hazardous wastes have been co-disposed with municipal solid wastes and there is no
hydraulic isolation between different phases. The collected samples might represent
mixture of leachates generated from a diverse range of waste composition and age and
hence, the effect of a biodegradation process on the evolution of recalcitrant compounds in
these leachates merits investigation. As discussed in Chapter 2, among different
degradation/treatment methods existing in the literature, biological processes are mostly
used in the UK for an economically viable leachate treatment. Therefore, in this study an
aerobic biological treatment process was chosen to investigate the evolution of recalcitrant
organics which would also provide useful information about biological treatment of
leachates in these two landfill sites. The aerobic biodegradation experiments on the
collected leachate samples were carried out in glass reactors over a period of 30 days with
an air supply arrangement. The detail experimental set up is discussed in Chapter 3
(section 3.2.2).
The investigation of the recalcitrant organics was carried out by established methods of
COD and DOC measurements, and by the potential new techniques, UV and fluorescence
spectroscopy. A combined analysis was carried out to investigate the overall
biodegradability of different leachate samples and to investigate the proportion and
individual biodegradation of several organic constituents of leachates in course of the
biodegradation. As such, these investigations provide an insight into the recalcitrant
organic compounds in these landfill leachates.
These investigations therefore aim to:
Results and Discussion: Aerobic biodegradation of landfill leachates
52
1) develop an understanding of the characteristics of the recalcitrant organic
compounds in real leachates during aerobic biological treatment
2) study the possible use of UV and fluorescence spectroscopy for fingerprinting
various organic compounds as well as for analysing the changes in the
compositional characteristics of the recalcitrant organic compounds in the course
of a treatment by assisting routinely used analytical methods
4.2 Initial characterisation of landfill leachate samples
The chemical compositions of four untreated and two treated leachate samples collected
from Pitsea and Rainham landfills are summarized in Table 4.1.
4.2.1 pH values
Table 4.1 shows that the pH of the collected leachates was in the range 7.20-8.42. The pH
of leachate usually increases with time due to decrease of the concentration of free volatile
fatty acids (Chian and Dewalle, 1976; Pohland and Harper, 1985). The reported pH of
acidogenic leachates ranges from 5.6 to 6.9 whereas the pH of methanogenic leachates is
in the range 6.8-8.0 (Lo, 1996; Lu et al., 1985; McBean et al., 1995). The pH values for
the untreated samples in this study therefore indicating a methanogenic state although the
leachates here may come from a mixture of old and new wastes within the sites.
4.2.2 Organic contents of leachates
Table 4.1 shows that all of the untreated and treated leachate samples had BOD30 values in
the range of 85-452 mg/l and COD values in the range of 850-4500 mg/l. As discussed in
chapter 2, BOD/COD values are correlated with the age of the landfill. Cho et al. (2002),
Tatsi et al. (2003), Lopez et al. (2004), Cecen and Aktas (2004) reported that leachates
from young landfills are characterised by high BOD and COD concentrations with values
ranged from 2300 to 25000 mg/l and 10540 to 70900 mg/l respectively, whereas in old
landfill leachates BOD and COD values ranges from 62 to 800 mg/l and 1409 to 3460
mg/l respectively.
Results and Discussion: Aerobic biodegradation of landfill leachates
53
Table 4.1 Chemical composition of landfill leachate samples
Sample 1
(treated)
Sample 2
(untreated)
Sample 3
(untreated)
Sample 4
(treated)
Sample 5
(untreated)
Sample 6
(untreated) Parameter (mg/l)
(except pH) Pitsea
(LTP)a
Pitsea
(P4)
Rainham
(LTP)a
Rainham
(FE)b
Rainham
(P2)
Rainham
(LTP HAZ)c
pH 8.42 7.72 7.20 7.60 8.12 7.40
BOD30d 282 85 113 226 452 282
COD 3300 4500 1100 850 3700 950
BOD30/COD 0.09 0.02 0.10 0.27 0.12 0.30
TC 1026 2728 644 243 2117 854
TIC 351 1477 373 104 1422 778
TOC 675 1251 271 139 695 76
DOC 663 1191 232 125 680 64
TS 17190 32890 7170 8200 10870 7400
TVS 15480 18080 1120 1620 2140 920
TFS 1710 14810 6050 6580 8730 6480
TSS 475 170 167 230 170 150
VSS 100 80 125 110 40 50
TDS 16713 32710 7000 7830 10490 7220
Chloride (Cl-) 675 680 701 242 729 759
Nitrate (NO3-) 0.63 0 1.40 1.0 4.30 5.0
Sodium (Na+) 382 306 325 307 295 298
Ammonium (NH4+) 284 315 343 325 325 341
Potassium (K+) 565 461 481 449 456 474
Magnesium (Mg2+) 57 59 17 11 10 14
Calcium (Ca2+) 696 611 325 254 268 215 a Leachate Treatment Plant b Final Effluent c Hazardous d 30 days BOD
Chian and DeWalle (1976), Chian (1977) and Harmsen (1983) also demonstrated that the
leachates generated from old landfills consist mainly of high molecular weight recalcitrant
organic compounds, which are correlated with low BOD and COD values. Thus the low
BOD and COD values of the collected leachate samples in this study implying significant
leaching from the old wastes in the sites and hence, considerable amount of recalcitrant
organic compounds should be expected to be present. This can also be verified by the
Results and Discussion: Aerobic biodegradation of landfill leachates
54
corresponding BOD/COD ratio which is commonly known as a measure of
biodegradability (Lo, 1996; Chen et al., 1996). The BOD/COD ratios of the collected
leachate samples ranged from 0.02 to 0.30. Studies reported in the literature indicate that
leachates containing high molecular weight recalcitrant organic compounds had
BOD/COD ratios in the range of 0.01-0.40 (Timur and Ozturk, 1999; Marttinen et al.,
2002; Cho et al., 2002; Tatsi et al., 2003; Lopez et al., 2004; Cecen and Aktas, 2004).
Table 4.1 also shows that although Pitsea (P4) leachate had been collected from the phase
where wastes were less than 4 years old (Table 3.1), it had a high pH and low values of
BOD and a low BOD/COD ratio. This may indicate that methanogenic conditions may
have been established at an early stage in this phase. The early establishment of
methanogenic conditions in this phase may be attributed to a high amount of readily
degradable organic waste and the high moisture content (Table 3.1) allowing fast
dissolution of organic compounds and accelerating microbiological decomposition. This
may also be due to the fact that older leachates were mixed in with leachates from this
phase. The early establishment of methanogenic condition has also been reported in the
literature (Lo, 1996; Kulikowska and Klimiuk, 2008). Kulikowska and Klimiuk (2008)
showed that in landfills in Poland, methanogenic conditions had been established at an
early stage with stable COD values of 610 mg/l after about 4 years.
The total carbon content of the untreated leachate Pitsea (P4) was relatively higher than
the untreated leachates collected from Rainham implying that organic and inorganic
compounds in the Pitsea landfill was considerably higher than the Rainham landfill.
However, the treated leachates collected from each of these landfills generally showed
lower total carbon content in comparison to the untreated leachates thereby indicating the
effect of treatment.
4.2.3 Solid contents
The concentrations of total solids (TS), total volatile solids (TVS), total fixed solids (TFS),
total suspended solids (TSS), volatile suspended solids (VSS) and total dissolved solids
(TDS) are presented in Table 4.1. With the exception of TFS and VSS of Pitsea (LTP), the
leachates from Pitsea generally had a higher solids content than the leachates from
Rainham. In the Rainham leachates the TFS contents were higher than the TVS contents
Results and Discussion: Aerobic biodegradation of landfill leachates
55
for all of the untreated and treated leachate samples, whereas for the Pitsea leachates the
trend was reversed. This indicates that the leachates from Rainham contained higher levels
of inorganic solids whereas those from Pitsea contained higher levels of organic solids.
TSS, VSS and TDS contents in all of the untreated and treated leachate samples ranged
from 150 to 475 mg/l, 40 to 125 mg/l and 7000 to 32710 mg/l respectively, which are
significantly higher than the values reported in the literatures (Al-Yaqout and Hamodoa,
2003; Fan et al., 2006). This reflects a high degree of mineralization during active
anaerobic decomposition of the waste in Pitsea and Rainham landfills (Al-Yaqout and
Hamodoa, 2003).
4.2.4 Anions and Cations
The concentrations of chloride (Cl-), NO3- nitrogen, ammonia-nitrogen (NH3-N), sodium
(Na+), potassium (K+), magnesium (Mg2+) and calcium (Ca2+) are presented in Table 4.1.
The values of ammoniacal-nitrogen (NH3-N) observed in all of the untreated and treated
leachate samples were at the high end of the concentration range reported in the literature
(Chu, 1994; Marttinen et al., 2002; Silva et al., 2004). This might be due to the hydrolysis
and fermentation of the nitrogenous fraction of biodegradable substrates (Carley and
Mavinic, 1991). In comparison to ammoniacal-nitrogen (NH3-N), low NO3- nitrogen
values were found for all of the leachates, indicating that the majority of the nitrogen was
in the form of ammonia. The results also show that the values of Na+, K+, and Cl- observed
in all of these leachate samples were in the high concentration range (Fan et al., 2006; Chu
et al., 1994; Al-Yaqout and Hamodoa, 2003). The high values of these salt contents are
correlated with high conductivities and the high TDS values (section 4.2.3). The Mg+2 and
Ca+2 contents of the Pitsea leachates were higher than those of Rainham leachates. This
might be due to higher industrial wastes disposed in Pitsea landfill than in Rainham
landfill (Table 3.1) (Kulikowska and Klimiuk, 2008).
The chemical compositions of leachates from these two UK landfills are compared with
landfills in Taiwan, Hong Kong, Kuwait, USA, Germany, Turkey, Finland and Greece in
Table 4.2 (Fan et al., 2006; Chu et al., 1994; Al-Yaqout and Hamodoa, 2003; Timur and
Ozturk (1999); Marttinen et al. (2002); Tatsi et al. (2003)). The composition of the
Rainham leachates is similar to that of Hong Kong leachates (JB, GDB) although Rainham
is older than the Hong Kong landfill. The low organic contents of the young Hong Kong
Results and Discussion: Aerobic biodegradation of landfill leachates
56
leachates (JB, GDB) are a result of the high temperature and rainfall which enables a
stable methanogenic state to be reached quickly. The Pitsea leachates have higher COD
and solids contents than those of Taiwan, Finland and Hong Kong (JB, GDB), similar to
those of Kuwait and lower than those of the USA, Germany and Turkey. These differences
in leachate composition may be due to the presence of bottom ash in Taiwan landfills,
high rainfall and early methanogenic stages developed in Hong Kong landfills and the
rising water table in Kuwait landfills. The concentrations of chloride, Na+ and K+ of Pitsea
and Rainham leachates are significantly higher than the leachates generated in Taiwan and
Hong Kong (JB, GDB). The high values of these ions are reflected by the high
conductivity in Pitsea and Rainham leachates.
The above results confirm that the chemical composition of leachates may vary
significantly from site to site. Differences were found in the concentrations of organic
matter, anions and cations and in total and fixed solids contents. Leachates from Pitsea had
higher solids, organic matter, sodium, chloride and magnesium contents than leachates
from Rainham in all of the untreated and treated samples. In addition, Pitsea and Rainham
leachates found to have different characteristics (pH, BOD, COD etc.) in comparison to
leachates from other countries. Owing to such a variable characteristics of landfill
leachates, it is essential to conduct a long term monitoring programme to obtain
representative information on leachates and to understand the detailed evolution of the
constituent organic matter.
Results and Discussion: Aerobic biodegradation of landfill leachates
57
Table 4.2 Chemical characteristics of leachates from different landfills
Parameter (mg/l)1
Taiwan Site A, B, C
(Fan et al., 2006)
HK JB, GDB (Chu et al.,
1994)
Kuwait
(Al-yaqout and Hamodoa, 2003)
USA (Al-yaqout and
Hamodoa, 2003)
Germany (Al-yaqout and
Hamodoa, (2003)
Turkey (Timur and
Ozturk, 1999)
Finland (Marttinen
et al., 2002)
Greece (Tatsi et al.,
2003)
Age 10-17 3.5-11 11 - - - - -
Waste type Predominantly MSW, some bottom ash
Predominantly MSW, some Hazardous
MSW and construction/dem
olition waste Not mentioned Not mentioned Not mentioned Not
mentioned Not
mentioned
pH 7.74-7.91 7.6-7.8 7.55 6 6.9 7.3-7.8 7.1-7.6 7.9
BOD 49.6-173.8 30-600 13400 400-45900 10800-11000 84 1050
COD 690-3038 489-1670 158-9400 1340-18100 1630-63700 16200-20000 340-920 5350
TOC 249-1025
TS 3941-9620 920-5580
TVS 498-1580 TFS 398-4010
TSS 34-193 480
VSS 51-166
TDS 3907-9464
NO3-N 2.96-26.7 0.06-179
NH3-N 1120-2500 330-560 940
Na+ 297-3524 132-1190
Ca2+ 15.9-137.5 5.6-122 254.1-2300 70-290
Mg2+ 15.7-163 9-63 5.2-268 233-410 100-270 1 Except pH all parameters are in mg/l
Results and Discussion: Aerobic biodegradation of landfill leachates
58
4.3 Effect of a laboratory scale aerobic biological treatment in the change
of the nature of recalcitrant organic compounds in landfill leachates
4.3.1 Conventional COD and DOC analyses
The change in the nature of organic compounds in all of the untreated and treated leachate
samples during laboratory scale aerobic biological treatment over a period of 30 days is
presented in figure 4.1 (a-d). This investigation was carried out using the conventional
characterisation methods of COD and DOC.
Figure 4.1 (a, b) shows the total and dissolved COD and total and dissolved organic
carbon (TOC, DOC) as a function of time for all of the untreated and treated leachate
samples. The results show a gradual decrease of COD and TOC (DOC) with time
indicating the reduction of organic compounds by biological processes. However, a slight
increase in COD at the beginning of aerobic biodegradation was observed for every
leachate sample (figure 4.1 (a)). Similar results have been reported by Nilsum (1998) and
Bila et al. (2005) and this might be due to a rapid change in the structure of the organic
compounds as a consequence of reactions in the formation of short-term intermediates that
are easily oxidizable in the COD test. Figure 4.1 (a, b) also shows that the COD and DOC
concentrations were high in the untreated leachates Pitsea (P4) and Rainham (P2),
indicating that these two leachates contained significant amounts of organic compounds.
The treated leachate Pitsea (LTP) also had high values of COD and DOC. This suggests
that the treatment applied prior to aerobic biodegradation in the laboratory and the
subsequent aerobic biodegradation were insufficient to remove all of the organic
compounds from this particular leachate. These figures and those in table 4.1 also show
that the organic compounds in the leachates contributing COD and TOC were mostly in
dissolved form.
Figure 4.1 (c, d) shows the percentage removal of COD and DOC in the laboratory scale
aerobic biodegradation experiments as a function of time. The %DOC trend shows that
after 30 days aeration the degradation of different leachates are in the order of Rainham
(LTP) > Rainham (FE) > Pitsea (LTP)> Rainham (LTP Haz) > Pitsea (P4) > Rainham (P2)
> thereby indicating their biodegradability. However, from the %DOC results the
leachates can be divided into two groups in terms of their biodegradation after 30 days.
Results and Discussion: Aerobic biodegradation of landfill leachates
59
The Rainham (LTP) and Rainham (FE) lechates which have low concentrations of organic
compounds could be classed as easily biodegradable leachates whereas the rest of the
leachates which have high concentrations of organic compounds could be classed as not
easily biodegradable. An exception is observed for Rainham (LTP Haz) leachate which
exhibits low biodegradability although this leachate has a low amount of organic
compounds. This might be due to the presence of chlorinated aliphatic and aromatic
organic compounds in this leachate. The %DOC removal found in different leachates was
not the exact replica of the observed %COD removal trend. While the highest %DOC
removal was observed for Rainham (LTP) leachate the lowest %DOC removal was
observed for a different leachate, Rainham (P2). This could be explained by the difference
in the method of estimating non-degraded compounds in these two techniques. Due to the
presence of the strong oxidizing agent, some inorganic fractions of leachate get
incorporated in COD measurements whereas the DOC measurement mostly estimates the
organic fraction of leachate. However, from figure 4.1 (c, d) it can be concluded that the
biodegradability of Rainham (LTP) leachate is the highest after 30 days of aeration
whereas Pitsea (P4) leachate can be accepted as one of the least biodegradable leachates
among the six leachates under consideration.
Results and Discussion: Aerobic biodegradation of landfill leachates
60
0 5 10 15 20 25 3015002000250030003500400045005000 Pitsea (LTP)
Dissolved
CO
D (m
g/l)
Time (days)
Total
0 5 10 15 20 25 30500
100015002000250030003500400045005000
Rainham (LTP)
Dissolved
CO
D (m
g/l)
Time (days)
Total
0 5 10 15 20 25 30
500100015002000250030003500400045005000
Rainham (FE)
CO
D (m
g/l)
Time (days)
Total Dissolved
0 5 10 15 20 25 30100015002000250030003500400045005000 Rainham (P2)
CO
D (m
g/l)
Time (days)
Total Dissolved
0 5 10 15 20 25 30500
100015002000250030003500400045005000
Rainham (LTP Haz)
CO
D (m
g/l)
Time (days)
Total Dissolved
0 5 10 15 20 25 302000
2500
3000
3500
4000
4500
5000 Pitsea (P4)
Dissolved
CO
D (m
g/l)
Time (days)
Total
T
T
UU
U
U
Figure 4.1 (a) The change in COD over time during aerobic biodegradation for all of
the untreated and treated leachate samples (T = treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
61
0 5 10 15 20 25 30300
450
600
750
900
1050
1200
Rainham (P2)
TOC
, DO
C (m
g/l)
Time (days)
TOC DOC
0 5 10 15 20 25 30300
450
600
750
900
1050
1200
Pitsea (LTP)
TOC DOC
(TO
C, D
OC
(mg/
l))
Time (days)0 5 10 15 20 25 30
300
450
600
750
900
1050
1200
Pitsea (P4)
TOC DOC
TOC
, DO
C (m
g/l)
Time (days)
T
UU
U
0 5 10 15 20 25 30
80
120
160
200
240
280
Rainham (LTP)
TOC DOC
TOC
, DO
C (m
g/l)
Time (days)
U
0 5 10 15 20 25 306080
100120140160180200
Rainham (FE)
TOC
, DO
C (m
g/l)
Time (days)
TOC DOC
T
0 5 10 15 20 25 3040
50
60
70
80
90
100Rainham (LTP Haz)
TOC
, DO
C (m
g/l)
Time (days)
TOC DOC
Figure 4.1 (b) The change in TOC and DOC over time during aerobic biodegradation
for all of the untreated and treated leachate samples (T = treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
62
0 5 10 15 20 25 300
102030405060708090
100(c)
% C
OD
rem
oval
Time (days)
Pitsea (LTP)T
Pitsea (P4)U
Rainham (LTP)U
Rainham (FE)T
Rainham (P2)U
Rainham (LTP Haz)U
0 5 10 15 20 25 300
102030405060708090
100(d)
% D
OC
rem
oval
Time (days)
Pitsea (LTP)T
Pitsea (P4)U
Rainham (LTP)U
Rainham (FE)T
Rainham (P2)U
Rainham (LTP Haz)U
Figure 4.1 Aerobic biodegradation of landfill leachates for 30 days (c) % COD
removal; and (d) % DOC removal (T = treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
63
The above results also show an increasing amount of removal of degradable compounds
with time during aerobic biological treatment of different landfill leachate samples.
However, the overall efficiencies and effectiveness of treatment process was found to
depend not only on whether the leachates were untreated or treated but also on the
variation of leachate composition and landfill site characteristics. For example, the treated
leachate Rainham (FE) showed a high percentage removal of COD and DOC in
comparison to the untreated leachate Pitsea (P4) (figure 4.1 (c, d)) while it was expected
that in general a treated leachate would show a low removal of COD and DOC in
comparison to the untreated leachate.
The conventional COD and DOC methods presented in this section were used to estimate
the amount of remaining organic compounds in leachates, and the treatment efficiency was
evaluated by the percentage removal of the COD and DOC over time. However, these
methods were not enough to identify the constituent structure of recalcitrant organic
compounds which might have an effect on the treatment of leachates collected from
different landfill sites. Therefore, UV and fluorescence spectroscopy were used to
characterize the recalcitrant organic compounds by estimating the eventual removal of
different organic constituents of leachate in the course of the biodegradation. These
investigations study the degradation potential of different organic compounds in leachates
and asses the feasibility of using UV and fluorescence spectroscopy for analysing
leachates.
4.3.2 Spectroscopic analysis
In this section the nature and characteristics of recalcitrant organic compounds in landfill
leachates during aerobic biodegradation were evaluated using UV absorbance and
fluorescence spectroscopic techniques. From COD and TOC analyses (figure 4.1 (a, b)) it
was found that the organic matter in the leachate samples was mostly in dissolved form.
Therefore, unfiltered leachate samples were analysed for their spectroscopic
characteristics to obtain the information required.
Results and Discussion: Aerobic biodegradation of landfill leachates
64
4.3.2.1 UV spectroscopic results
UV absorption spectroscopy has many useful applications in characterizing the aromatic
carbon content. Bari and Farooq (1984), James et al. (1985) and Matsche and
Strumworher (1996) reported that UV absorbance at 254 nm wavelength has a close
correlation with aromaticity. McKnight et al. (2001) and Abbt-Braun et al. (2004) showed
a good correlation between UV254 and unsaturated sp2 hybridized carbon atoms by 13CNMR and suggested UV absorbance at 254 nm wavelength (UV254) as an excellent
surrogate parameter for estimating the aromatic organic compounds.
Figure 4.2 (a) shows the change of UV254 absorbance over time during aerobic
biodegradation for all of the leachate samples. The highest initial value of UV254
absorbance was observed in the leachate Pitsea (P4) whereas the lowest value was
observed in the Rainham (FE) leachate. Intermediate values of UV254 absorbance were
observed in Rainham (P2), Pitsea (LTP), Rainham (LTP) and Rainham (LTP Haz)
leachates with decreasing order respectively. This suggests that the Pitsea (P4) leachate
had the highest amount of aromatic organic compounds, with the other leachates having
progressively decreasing amount of aromatic compounds as shown in figure 4.2 (a). It is
worth noting that although Rainham (LTP Haz) leachate contained greater concentrations
of chlorinated aliphatic and aromatic organic compounds (Table 3.1) than Rainham (LTP)
leachate, experimental results showed lower UV254 absorbance values for Rainham (LTP
Haz) leachate than for Rainham (LTP) leachate. This may be related to the lower DOC
value found for the Rainham (LTP Haz) leachate in comparison to the Rainham (LTP)
leachate (figure 4.1 (b)). However, possible other reason will be discussed later (section
4.4.2).
Results and Discussion: Aerobic biodegradation of landfill leachates
65
0 5 10 15 20 25 300
102030405060708090
100
% U
V 254re
duct
ion
Time (days)
Pitsea (LTP)T
Pitsea (P4)U
Rainham (LTP)U
Rainham (FE)T
Rainham (P2)U
Rainham (LTP Haz)U
(b)
(a)
0 5 10 15 20 25 300.00.30.60.91.21.51.82.12.42.73.0
UV 25
4 abs
orba
nce
(cm
-1)
Time (days)
Pitsea (LTP)T
Pitsea (P4)U
Rainham (LTP)U
Rainham (FE)T
Rainham (P2)U
Rainham (LTP Haz)U
Figure 4.2 (a) Absorbance at 254 nm (UV254) and (b) % UV254 reduction for all of the
untreated and treated leachate samples (Dilution factor 25) (T = treated and U =
untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
66
In all of these leachates, the UV254 absorbance decreased with an increasing number of
days of aeration. This suggests the breakdown of the aromatic structures of the constituent
organic matter by an aerobic treatment process (Gottschalk et al., 2000). However, this
might also be attributed to the reaction of oxygen with the unsaturated bonds and aromatic
rings, leading to the splitting of bonds in the organic compounds (Gottschalk et al., 2000).
Figure 4.2 (b) shows the percentage reduction of UV254 absorbance during aerobic
biodegradation for all of the untreated and treated leachate samples. With the exception of
Rainham (LTP Haz) leachate, a high percentage of UV254 absorbance reduction (figure 4.2
(b)) was observed for leachates with low UV254 absorbance values (figure 4.2 (a)). This
suggests that leachates containing low aromatic organic compounds had a high percentage
reduction of UV254 absorbance and vice versa. As discussed before, the deviation for
Rainham (LTP Haz) leachate could be explained again by the possible presence of
chlorinated aliphatic and aromatic organic compounds. The percentage reduction of UV254
absorbance (figure 4.2 (b)) showed the degradation of different leachates in the order of
Rainham (LTP) > Rainham (FE) > Pitsea (LTP) Rainham (LTP Haz) > Rainham (P2) >
Pitsea (P4) which was in good agreement with the percentage removal of DOC observed
in the different leachates (figure 4.1 (d)). Again from the UV254 absorbance reduction
trend, Rainham (LTP) and Rainham (FE) lechates can be accepted as easily biodegradable
leachates whereas the rest of the leachates can accepted as not easily biodegradable.
However, the percentage reduction of UV254 absorbance in different leachates was not the
exact replica of the observed %COD removal trend. This can be explained again by the
fact that some inorganic fractions of leachate get incorporated in COD due to the presence
of strong oxidizing agent while the UV254 absorbance measurable are aromatic organic
compounds. Thus, the fairly good agreement on the degradation trend observed in
different lecahtes by UV254 absorbance and DOC measurements probably indicate that
organic compounds contributing DOC in these leachates were mostly aromatic in nature.
However, the results of UV spectroscopy and the conventional DOC experiments suggest
that leachates containing high aromatic organic compounds usually result in low
percentage removal of DOC in the course of a treatment process.
Results and Discussion: Aerobic biodegradation of landfill leachates
67
4.3.2.2 Fluorescence spectroscopic results
Figure 4.3 shows the Excitation-Emission-Matrix (EEM) maps of the leachate DOM
fractions generated before and after laboratory scale aerobic biodegradation (EEM maps
for other days during aeration are presented in Appendix A). Four distinct zones were
identified on the EEM maps that were indicative of the presence of different organic
substances and are described below. The EEM spectra also show Reyleigh and Raman
scattering peaks originating from the interaction of light and water molecules.
Zone 1 (H-L): Peaks were also observed to be present between 230-260 nm and 360-390
nm excitation and 400-460 nm and 460-480 nm emission respectively and is widely
recognized as a component of the humic-like fractions (H-L) (Coble, 1996; Baker and
Curry, 2004; Cumberland and Baker, 2007), which has also been found in untreated
wastewater (Baker et al., 2004) and urban river water (Fu et al., 2007).
Zone 2 (F-L): A peak, which was present in all of the leachates between 320-350 nm
excitation and 400-440 nm emission, can be attributed to aromatic and aliphatic groups in
the DOM fractions, and is commonly labelled as fulvic-like (F-L) (Coble, 1996; Baker and
Curry, 2004). F-L fluorescence has also been detected in lake water (Mostofa and
Yoshioka, 2005).
Zone 3 & 4 (Protein-like): Peaks were observed at 230-240 nm and 270-280 nm
excitations and 300-320 nm emission, labelled as tyrosine-like (Tyr-L) and at 230-240 nm
and 270-280 nm excitations and 340-370 nm emission labelled as tryptophan-like (Trp-L).
These peaks are attributed to ‘protein-like’ structures (Coble, 1996; Elliot et al., 2006).
‘Protein-like’ structures have also been widely observed in marine water (Coble, 1996;
Ogawa et al., 2001) and river water (Mostofa and Yoshioka, 2005; Fu et al., 2007).
Fluorescence peaks that are ascribed as H-L, F-L and protein-like compounds originate
from the degradation of carbohydrate, proteins, fats etc. present in the waste through
various physical, chemical and microbiological processes (Calace and Petronio, 1997). It
has been suggested that the carbohydrate and protein compounds undergo condensation
that leads to the formation of humic-like (H-L) substances according to the melanoidin
Results and Discussion: Aerobic biodegradation of landfill leachates
68
model (Ikan et al., 1986). Unaltered lignin in the anaerobic environment also contributes
towards the formation of recalcitrant H-L and F-L compounds (Komilis and Ham, 2003).
As discussed in Chapter 2, microbial synthesis taking place in the landfill can also lead to
the formation of in-situ humic and protein-like compounds (Calace and Petronio, 1997;
Pichler and Kogel-Knabner, 2000). Previous studies have demonstrated that protein-like
compounds (Zone 3 and 4) are associated with bacterial activities as well (Cammack et al.,
2004; Elliott et al., 2006). As the H-L and F-L compounds are known as the essential
building block of recalcitrant organic compounds (Artiola-Fortuny and Fullar, 1982;
Castagnoli et al., 1990 and Christensen et al., 1998), the change of fluorescence peak
intensities and peak positioning of recalcitrant H-L and F-L compounds of different
untreated and treated leachates from different landfills in the course of treatment may
provide important information on the compositional characteristics of the recalcitrant
organic compounds of leachates.
Results and Discussion: Aerobic biodegradation of landfill leachates
69
4020
20
2040
40
20
6020
20
20
80604020
20
60
4020
2020
80
60
60
40
40
20
20
20
20
40
40
20
20
20
20
60
60
40
40
60
40
20
20
260
240
220
200180
160140
120
100 80
80
80
100
40
40
60
280260240
220200180160140
1208040
10080
220200180160140
120
100
40100
60
60
60
60
260240220200180160140
12012
0
80
60
280260240220200180160140
100
4020
80
80
80
280
260240
220
200
180160
140120
100
100
40
120
60
100
340
320300
280
260240
220
200180160140
60
120240220200180160
140140
80
40
60
240220200180160
120
40
40
24022020018016014010
060
80
26024022020018016014
0120
2080
4040 40
200180
160140120100
60
40
40
180
160140120100
80
100
20
80
22020018016
0140120
10080
60
40
120
20
180160140120100
40
60
40
160
14012010080
2040
1401201008060
40
20
1008060
20
2040
20
20
100806040
40
20
1008060
40
20
8060
20
20
20806040
20
20
20
806040
20
806040
20
202 0806040604020
2020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (LTP)After aeration
(2)
(1)(3)(4)
(4)20
20
20
20
20
20
2040
40
20
20
40
20
20
20
60
20
4020
20
80
60
60
60
20
20
4020
100
100
2020
40
40
120
80
40
8060
6040
140
20
4020
60404060
80
20
20
20
40
80
40
40
2040
2040
40
806060
80
60
14012010020
40
60
2020
20
100
40
6040
80
20
20
20
60
40
20
120100806020
40
2040
20
2040
40
40
20
60
20
20
4020
20
20
20
20
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
(2)
(1)(3)(4)
(4)
Pitsea (LTP)Before aeration
2020
20
20
20
20
20
20
20
20
20
20
40 40
402020
40
40
40
120
120
100
100
80
80
60
60
4020
60
60
4060
140
20
140120
100
100
80
80
40
40
60
60
40
40
20
20
340320300280260240220200
180
160
160
80
80
60
40
180340320300280260240220
200
100
100
20
200
40
120
120
80
60
340320300280260240220
140
220
160
100
4060
340320300280260240220
180
40
340320300280260240
60
160
60
200
340320300280260240220
200
180
80
200
80
340320300280260240220
220
6040
220
200
60
340320300280260240
60
80
260
80
60
340320300280
60
20040
6040
60
34032030028026024022012040
40
60
34032030028026024022020018016014020
40
60
12040
100
60
20
6040
300280260240220200180160140
40
20
100
4020
20
60
260240220200180160140120
40
80
20
40
40
240220200180160140120100
20
80
20
20018016014012010060
2040
16014012010080
20
20
20
20
12010080604020
20
80604020
20806040
20
6040
20
20
4020
20
2020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (P4)Before aeration
(2)
(1)(4)
(4) 80604020
20
60402010080
4020
8060402020
20
40
806040
40
20
20
40
20
4060
20
8060
40120100
10080
40
220
220
200
200
180
180
160
160
140
140
120
120
100
100
100
80
80
80
60
60
60
4020
40
40
140120
60
240
240
40
80
140
140
240
240
220
220
200
200
180
180
160
16060
10080
60
120
120
100
100
80
80
60
60
40
40
20
20
20
340320300280
280
260
260
260
120
280
280
8060
80
60
340320
320300
300
30080
100140120
60
60
100
100
60
120
60
120
80
140
140
100
320
80
8080
80
340
160
80100140120
10080
4020
100
60
80
32040
120
20
60
340
60
300
40
80
2020 80
340320
40
280
60
340320300
60
220
60
340320300280260240
40
280
8060
60
60
340320300200340320300280260240220
20
100
40
340320300280260240220200180160140120
40
2040
20
40
340320300280260240220200180160140
20
300280260240220200180160140
20
20
2040
24022020018016014012080
402020018016014012010060
4020
12010080
20
40
1008060
20
40
20
100806020
2020
806040604020
20
20
2040
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (P4)After aeration
(2)
(1)(4)
(4)(3)
20
204020
40
2020
2020
20
20
20
40
20
40
20
60
20
8060
20
20
80
80120
100
100
20
4020
160140
4020
20
120
60
4020
140
40
4020
20
140160
2040
40
40
2040
606040
40
20
20
120
40
8020
20
40
100
40
40
20
60
40
60
140
40
206040
160
40
180
1806060
4020
220200
40
6080
260240
4020
280
100
340320300
80100
4020
120140
80
4020
160200180
4020
220
140240
4020
260
120
160140
60
100
120100
40
80
80
6040
60
20
60
6040
1206040
20
40
604020
20
20
20
6040
2040
40
4020
20
20
40204020
4020
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP)Before aeration
(2)
(1)(4)
(3)(4)80604020
4020
4020
10080604020140120
20
12010080604020140
2020
20
20
20
20
20
120100806040
20
14020 40
40
1601401201008060
40240220200180
40
40
40
20
60
40
40
160140120100
80
80
20
20
20
20018016014012010080
60
260240220200180
4040
60
100
18016014012010080
20
60
20
220200180160140120
40
1801601401201008060
20
406040
20
140120100
80
20
100160140120
40
340320300280260240220200180
140120
40
40
160140120100
80
60
6020
6060
40
80
20
80
8040
60
200180160140120
100
100
40
200180160140120
40
120140
6040
60
220200180160100
60
220200180160140120
40
100
100
120
260240220200180160140120
40
60
60
160
160
140
14040
40
16014012010080
60
180
20
60
200
40
16014012010080
80
260240220
60
280340320300
6060
22020018016014012010080
100
120
60
140
100
16014012010080
160
100
80
180
200180160140120100
200
120100
4040
220
1601401201008060
120
60
60
80
100
220200180160140120100804060
60
60
8060
140120100804040
160
40
1401201008060
40
60606040
180
240220200180160140120100806060
20
2001801601401201008040
60
60
40
20018016014012010080
20
80
20
200180160140120100
20
602802602402202001801601401201008040
2001801601401201008060
20
60
10080
20
20
Emission wavength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engh
(nm
)
200
250
300
350
400
450
500
Rainham (LTP)After aeration
(2)
(1)(3)
(4)
(4)
(a)
(b)
(c)
T T
U U
U U
Figure 4.3 Excitation-Emission Matrices (EEM) for landfill leachates before and
after aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like) (dilution factor 25) (T =
treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
70
20
2020
20
20
20
20
20
60402040
20
2040
40
20
60
60
20
1008060
4020
20
20 20
140120100
8080
80
20
200180160
20
100
20
140120
4040
20
120
20
20
40
40
20
40
20
100
2020
40
6040
60
20
20
60404020
20
120140
20
160180
40
4020
200
40
60
20
22040
4020
240
80
300280260
20
4020
320340
60 4040
4020
100
20408060
6040
20
20
80
60
40
604020
20
60
20
40
20
20
20
4040
20
4020
20
80604020
604020
80604020
10080604080604020
20
40
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (FE)Before aeration
(2)
(1)(3)(4)
(4)
16014012010080604020
2020
1601401201008060402022020018020
20
2018016014012010080604020020
20
20
202001801601401201008060
40
40
40
2040
20
60
18016014012010080
60
40260240220200
408060
160
160
140
140
120
120
120
100
100
100
80
80
80
60
60
60
40
40
20
20
120100140
60
180
180
180
180
160
160
140
140
120
120
100
100
100
20
18016014012010080
80 40
80
80
80
60
60
60
40
40
40
20
20
20
340
320300
300
280
280
260
260
260
240
240
240
240
220
220
220
220
200
200
200
200
20
280260240220200
40
320
20
60
20
80
340
8060
200180160140120100
6040
40
40
8060
120
80
320
100
60
340280
40
340320300
80
4080
120100
60100
40
60
20
260
120
40
340320
300280
280
140160
6060
180
40
300
60
200
60
340320
80
220240
60
280260300
4060
100
320
1008040
240
40
60
340320300280260
80
60
60
4060
340320
20
40
40
60
34032030028026040
40
20
340320300280260240220200180160
20
20
34032030028026024022020018016014020
40
120
4020340320300280260240220200180160140
20
340320300280260240220200180160
4020
28026024022020018016014040
20
2020
2001801601401201008060
40
2018016014012010080604040
208060
20
Emission wavelngth (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
ngth
(nm
)
200
250
300
350
400
450
500
Rainham (FE)After aeration
(2)
(1)(3)(4)
(4)
20
20
2020
20
20
20
20
20
20
20
20
40
40
20
40
40
60
60
4020
60 60
40
60
60
40
40
20
60
20
20
20
22020018016014012010080
80
80
80
60
40
80260240220200180160140120
100
100
100
100
240220200180160140120
120
120
100
80
20
80
20
240220200180160140120
140
140
40
120
160
20
260240220200180160140
60
80240220200180160140120
100
180
100
100
40
200
240220200180160140120
60
120
80
24022020018016014014040
100
20
220
4040
240220200180160
80
80
12040
60
80
2402202001801601408040
40
2202001801601401201008040
60
40
60
20
200180160140120100208040
20180160140120100
40
6040
60
4020
16014012010080206040
20
40
40
40
12010080
40
4040
20
20
20
100806020
20
80604020
206040
20
20
404020
20
2020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (P2)Before aeration
(2)
(1)(4)
2040
20
20
20
2040
4020
20
20
20
20
40
2020
40
20
20
6040
40
80
20
40
120
120
100
100
80
80
60
60
4020
60402080
40
140
14040
140
140
120
120
100
100
80
80
60
60
60
60
60
40
40
20
20
340320300280260240220200180160
160
80
40
160
160
40
80
40
60
340320300280260240220200
180
180
10080
60
340320300280260240
220
200
200
100
220
80
80
40
340320300280260
240
120
120
240
140
140
100
60
340320300
280260
260
60
280
60
160
100
340320300
120
60
80
180
8080
200
4020
80
100
32040
60
100
340
80
20040
80
80
80
340320300280260240220
60
18040
60
60
80
3403203002802602402202001204020
60
60
40
100
340320300280260240220200180160140
20
10040
60
20
2040
340320300280260240220200180160
14012016040
40
20
40
40
340320300280260240220200180
40
120
40
40
20
20
30028026024022020018016014060
20
40
20018016014012010080
20
20
60
20
1601401201008040
20
120100806040
20
120100806040
806020
604020
604020
20
404020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (P2)After aeration
(2)
(1)(4)(4)
2040402020
20
40
20
2020
20
20
604080
20
20
20
40
2020
60
100806040
10080
20
20
40
80
100806040
40
40
40
40
40
100806040
80
60
60
1008060
100
100
40
60
4040
1201008060
80
40
120
60
10080
60
6012010080
40
6080
100
140120100
40
60
6040
120
1008060
40
12010080
140
40
40
100
100806040
160
60
120100806040
60
1201008060
40
20
180
60
10080604040
40
80
20
40
1008060
80
6040
4040
60
20
40
60
100806020
60
20
40
80604020
40
2020
80604020
2040
20
80604020
20
20
6040604020
20
20402020
60402020
20604020
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP Haz)Before aeration
(2)
(1)(3)(4)
(4)
100806040
2020
2010080604020
20
140120140120100806040
20
20
20
20180160140120100806040
20404020
20
20
200180160140120100806040
40
320300280260240220
20
80
80
80
60
60
60
40
40
40
20
20
40
20
40
60
40
80
80
12010080
6040
20
60
60
40
40
40
20
20
20
280260240220200180160140120
100
100
220200180160140
100
80
320300280260240220200180160140120
20
40
100
300280260240220200180160140120100
60
20
260240220200180160140120
4020
120
40
60
60
300280260240220200180160140
40
180
1601401201008060
340320300280260240220200
240220200180
40
100
20
60
40
320300280260240220200
180
160140
120
120
40
180
16014012010080
340320300280260240220200
60
200
4040
340320300280260240220
60
200
80
340320300280260240220
60 40
160
40
60
2040
340320300280260240220200180
80
80
120
40
340320300280260240220200180160140
4020
140340320300280260240220200180160
100
120340320300280260240220200180160140100
120
320300280260240220200
180160140120
40
200
60
220340320300280260240
40
100
60
300280260240220200180160140120
40
80
40
20
280260240220200180160140120100
20
80
240220200180160140120100
20
80
4040
300280260240220200180160140120100
20
22020018016014012060
2030028026024022020018016014012010080
4020
20
18016014012060
20
1801601401201008060
2020018016014012010080
20
6026024022020018016014012010080
20160140120100806040
20
40
80602060402020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP Haz)After aeration
(2)
(4)(4)
(3)
(d)
(e)
(f)
U U
U U
T T
Figure 4.3 (contd.) Excitation-emission matrices (EEM) for landfill leachates before
and after aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like) (dilution factor 25) (T =
treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
71
Table 4.3 presents the summary of fluorescence properties for all of the untreated and
treated leachate samples. Leachates Rainham (LTP), Rainham (FE) and Rainham (LTP
Haz) had relatively lower fluorescence intensities in Zones 1 (H-L) and 2 (F-L) and higher
intensities in Zone 3 (Tyr-L) and 4 (Trp-L) than leachates Pitsea (LTP), Pitsea (P4) and
Rainham (P2) (with the exception of Tyr-L compounds in leachate Rainham (LTP Haz)).
This suggests that Rainham (LTP), Rainham (FE) and Rainham (LTP Haz) had relatively
lower concentrations of H-L and F-L compounds and higher concentrations of protein-like
compounds than leachates Pitsea (LTP), Pitsea (P4) and Rainham (P2). As H-L and F-L
compounds are the essential building block of recalcitrant organic compounds, these
results indicate that Pitsea (LTP), Pitsea (P4) and Rainham (P2) leachates contained higher
concentrations of recalcitrant organic compounds than leachates Rainham (LTP), Rainham
(FE) and Rainham (LTP Haz). Table 4.3 and figure 4.3 also show that aerobic
biodegradation decreased the intensities of Zone 1 (H-L) and 2 (F-L) materials by about 4-
100% in all of the untreated and treated leachate samples over 30 days of period. The
decrease in intensities could be explained by the decomposition of H-L and F-L
fluorophores with the aerobic biodegradation (Saadi et al., 2006; Uyguner and Bekbolet
2005) thereby indicating the degradation potential of humic and fulvic like compounds of
different leachates in Table 4.3.
In Rainham (LTP) and Rainham (FE) leachates a comparatively a high reduction (50-
64%) of fluorescence intensities in Zone 1 (H-L) and Zone 2 (F-L) materials was
observed. This indicates that aerobic biodegradation had significantly broken down H-L
and F-L structures in these two leachates. Rainham (FE) was the final effluent of Rainham
(LTP) and Rainham (LTP Haz) (Table 3.1). However, the fluorescence intensities of this
leachate in Zone 1 (H-L) and 2 (F-L) before aerobic biodegradation was almost similar to
the leachates Rainham (LTP) and Rainham (LTP Haz), indicating that the treatment
applied onsite prior to aerobic biodegradation was not effective in removing the
recalcitrant organic compounds. In the case of Pitsea (P4) and Rainham (P2) leachates,
aeration reduced the intensities of Zone 1 (H-L) and 2 (F-L) materials by only about 4-
13%, indicating that a considerable fraction of high molecular weight organic compounds
remained. Although Rainham (LTP Haz) leachate is to be enriched with chlorinated
aliphatic and aromatic organic compounds (Table 3.1), aeration reduced the intensity of
Zone 1 (H-L) materials by 100% after 10 days. This might be attributed to the self-
Results and Discussion: Aerobic biodegradation of landfill leachates
72
quenching within H-L molecules (Chen et al., 2003) or by internal quenching by other
organic or inorganic molecule intermediates formed during the
biodegradation/remineralisation process (Saadi et al., 2006). Senesi (1990) also attributed
this effect to the greater proximity of aromatic choromophores and the consequent greater
probability of deactivation of excited states by internal quenching in higher molecular
weight molecules. This also may explain the low UV254 absorbance observed in this
leachate (section 4.4.1).
Results and Discussion: Aerobic biodegradation of landfill leachates
73
Table 4.3 Fluorescence properties of landfill leachate samples during aerobic
treatment (Total fluorescence intensity was calculated by the cumulative integration of the intensity
versus emission wavelength data for any given excitation wavelength range corresponding to different types
of materials and was corrected by multiplying by dilution factor 25) (SD = standard deviation, n=3)
Zone 1 Zone 2 Zone 3 Zone 4
Humic like fluorescence at 230-260 nm and 360-390 nm excitation and
400-460 nm and 460-480 nm emission
respectively
Fulvic like fluorescence at
320-350 nm excitation and
400-440 nm emission
Tyrosine at 230-240 nm and 270-280
nm excitation
and 300-320 nm emission
Tryptophan at 270-280
nm excitation
and 340-370 nm emission
Tryptophan at 230-240
nm excitation
and 340-370 nm emission
Leachate sample
No. of days after
aeration
FI. Intensity (Mean/SD)
FI. Intensity (Mean/SD)
FI. Intensity (Mean/SD)
FI. Intensity (Mean/SD)
FI. Intensity (Mean/SD)
0 1472200/6.5 907675/7.6 10475/4.3 96575/5.5 122750/5.6 10 1279875/6.3 633775/6.7 10350/4.2 94600/6.5 217275/6.6 20 1247425/5.5 602375/7.0 10300/5.4 93975/6.8 224300/6.0
Pitsea (LTP)
(Treated) 30 1181350/4.6 563475/5.8 7575/5.0 38825/6.9 217950/5.5
% reduction 20 38 28 60
0 1967750/2.2 1045125/2.3 154875/3.8 148725/6.6 10 1843825/2.4 1000000/5.3 14475/3.4 129025/3.7 147750/6.5 20 1797275/3.5 994600/5.4 13650/4.4 118125/4.4 147400/4.4
Pitsea (P4) (Untreated)
30 1718075/3.3 960400/2.3 11900/4.5 104000/7.5 137900/5.8 %
reduction 13 8 18 33
0 1399925/7.6 494875/6.8 31925/5.7 3658750/5.7 1566900/6.6 10 824375/9.0 241550/10.0 31800/9.4 3085000/6.5 11727508.7 20 748475/9.5 215775/10.0 29175/3.7 2821750/9.0 1164700/8.9
Rainham (LTP)
(Untreated) 30 701500/5.2 195750/6.9 27875/7.6 1129250/6.5 1089275/9.0
% reduction 50 60 13 69
0 1347950/4.8 626925/6.9 18375/5.6 1755500/5.5 373425/6.6 10 795850/9.7 337350/8.4 17300/6.6 1753750/9.8 457000/7.6 20 712500/5.7 288050/7.6 14950/7.4 1502000/10.0 428000/4.9
Rainham (FE)
(Treated) 30 562775/7.8 225900/8.4 12400/7.8 505500/10.6 357050/6.0
% reduction 58 64 33 71
0 1943600/2.3 1067300/3.1 105125/10.7 10 1868150/5.5 1047850/2.2 97075/9.4 15475/6.6 20 1821725/4.3 1028600/2.4 96200/4.8 14550/4.6
Rainham (P2)
(Untreated) 30 1779775/3.9 1025850/3.2 76800/5.9 12300/6.8
% reduction 8 4 27
0 1388100/4.8 645600/7.5 8100/4.6 1554750/7.5 217975/8.7 10 379300/2.5 6850/5.5 1024250/5.9 260625/7.8 20 304525/4.5 6650/4.6 912500/6.9 241600/8.8
Rainham (LTP HAZ) (Untreated)
30 259975/5.5 5750/5.6 500000/6.8 240450/5.9 %
reduction 100 60 29 68
Results and Discussion: Aerobic biodegradation of landfill leachates
74
Fluorescence results also show that the intensities of Zone 3 and 4 (270-280 nm
excitation) fluorophores (tyrosine-like and tryptophan-like compounds respectively)
decreased during 30 days of aerobic biodegradation. This reduction can be attributed to the
breakdown of proteineous materials by aeration (Zhang et al., 2008). However, an increase
in fluorescence intensity was observed for Zone 4 (tryptophan-like compounds at 230-240
nm excitation), particularly over the first 10 days for the leachates Pitsea (LTP), Rainham
(FE) and Rainham (LTP Haz) (Table 4.3). This trend may be attributed to the enrichment
of protein-like compounds during DOC biodegradation. It has been reported that proteins
may be preserved by encapsulation into recalcitrant humic molecules (Dinel et al., 1996).
Therefore, it seems likely that the humic bound proteineous compounds might have been
exposed during aerobic biodegradation thereby increasing the fluorescence intensity in this
zone. Table 4.3 also shows that the percentage removal of Trp-L compounds was higher
than for the H-L and F-L compounds in all leachate samples (except for Rainham (LTP
Haz) leachate). It is interesting to note that before biodegradation (at day 0) the intensities
of Tyr-L and Trp-L compounds for leachate Rainham (FE) was significantly lower than
leachate Rainham (LTP). As Rainham (FE) was the final effluent in the treatment of
Rainham (LTP) and Rainham (LTP Haz), it can be said that the applied treatment onsite
was more effective in removing protein-like compounds than H-L and F-L compounds.
This indicates that protein-like compounds were more biodegradable than H-L and F-L
compounds and therefore aerobic biodegradation was more effective in removing Trp-L
compounds than the H-L and F-L compounds. Similar results have been reported by
Ahmad and Reynolds (1995, 1999), Reynolds and Ahmad (1997), and Reynolds (2002).
Their results suggested that Trp-L compounds are the more biodegradable fractions of
DOM and they showed up to 90% reduction in fluorescence intensity of Trp-L compounds
from influent to effluent across a treatment process.
Despite the intensity reduction of the H-L and F-L materials, aeration also caused peak-
shift for DOM fractions of leachates. In particular, aeration shifted the excitation
wavelengths of Zone 2 (F-L) materials of Rainham (LTP) (figure 4.3 (c)), Rainham (FE)
(figure 4.3 (d)) and Rainham (LTP Haz) (figure 4.3 (f)) leachates to longer excitation
wavelengths. This may be attributed to the enhanced oxidation of the Zone 2 (F-L)
materials of these three leachates producing more carbonyl, carboxyl, hydroxyl and amino
groups in the structures of the materials thereby indicting their biodegradability (Uyguner
and Bekbolet, 2005; Zhang et al., 2008). In the case of Pitsea (P4) (figure 4.3 (b)) and
Results and Discussion: Aerobic biodegradation of landfill leachates
75
Rainham (P2) (figure 4.3 (e)) leachates, aeration slightly expanded the EEM peaks of
Zone 1 (H-L) and 2 (F-L) materials to shorter emission wavelengths. These shifts may be
attributed to the following aeration induced changes for the Zone 1 (H-L) and 2 (F-L)
materials (Coble, 1996; Uyguner and Bekbolet, 2005; Zhang et al., 2008):
(1) the decomposition of condensed aromatic molecules to simpler and smaller
aromatic or aliphatic molecules;
(2) the reduction in the degree of π-electron systems resulting for example, from a
decrease the number of aromatic rings or conjugated bonds in chain structures;
(3) the elimination of certain functional groups such as carbonyl, carboxyl, hydroxyl
and amine.
The above results demonstrate that the fluorescence spectroscopy can be used to detect the
presence of different organic compounds in leachates. The relatively low reduction of
florescence intensities (Table 4.3) in H-L and F-L compounds during aerobic
biodegradation simply imply that these materials were the key components of recalcitrant
organic compounds whereas high intensity reduction in protein-like compounds suggest
that protein-like compounds were more biodegradable. Florescence spectroscopic results
also show that leachates from Pitsea exhibited lower intensity reduction of H-L and F-L
compounds in comparison to the Rainham leachates (except for Rainham (P2)), indicating
that Pitsea leachates may in general be less biodegradable than Rainham leachates. This
agrees with the results of conventional COD, DOC, UV254 absorbance analysis. In
addition to studying the biodegradation of different organic compounds in leachates,
florescence spectroscopy can also be used to investigate the structure of the H-L and F-L
compounds (recalcitrant compounds) that might affect biodegradation significantly. This
is addressed in the following section (Section 4.4.3).
Results and Discussion: Aerobic biodegradation of landfill leachates
76
4.3.2.3 Fluorescence analyses of humic and fulvic-like compounds in emission
scanning mode
In this section the structure of fluorescent compounds in leachates during aerobic
biological treatment is discussed. Fluorescence spectra in emission scanning modes
provide information about the aromatic structures with peak positions. Figure 4.4 presents
the fluorescence emission characteristics at specified excitation wavelengths of all of the
untreated and treated leachate samples. All of the leachates exhibited mainly two peaks,
labelled as Peak I at 360 nm and Peak II at 430 nm wavelengths. For all Pitsea leachates,
Peak II at (320-350) nm excitation was the strongest whereas for leachates Rainham (LTP)
and Rainham (FE), Peak I at (230-260) nm excitation was the strongest. Peak II at (360-
390) nm excitation for leachates Pitsea (LTP), Pitsea (P4) and Rainham (P2) also had high
intensities. As compared with the EEM spectrum (figure 4.3), Peak I at (230-260) nm
excitation can be assigned as tryptophan-like fluorophores. Peak II at (230-260) nm and
(360-390) nm excitations can be assigned as H-L and at (320-350) nm excitations as F-L
fluorophores. For leachates, Rainham (LTP) and Rainham (FE), Peak III with low
intensity is also observed at wavelength of (450-470) nm which can not be found in the
EEM spectra (figure 4.3). This Peak III is probably attributed to ‘red shifted’ high
molecular weight H-L fluorophores (Chen et al., 2003).
Kang et al. (2002) reported that the shorter wavelength peaks (360 nm) can be interpreted
to be due mainly to the presence of simple aromatic ring in the molecule, whereas the
peaks in the longer wavelength (430 nm) are due to the condensed aromatic rings and
conjugation of simple aromatic rings. It can be seen from figure 4.4 that leachates Pitsea
(LTP), Pitsea (P4) and Rainham (P2) showed strong intensities at 430 nm (Peak II),
suggesting that these leachates were associated with the presence of linearly condensed
aromatic rings of H-L and F-L molecules. However, it is also observed that leachates
Rainham (LTP) and Rainham (FE) showed the highest peaks at 360 nm (Peak I) (figure
4.4). This can be ascribed to the presence of simple aromatic rings. Figure 4.4 also shows
that aeration reduced the intensities of Peak I and Peak II in all of the leachates. This
decrease can be attributed to the destruction of aromatic ring structures through aerobic
biodegradation. However, aeration also increases the intensities of Peak I for Pitsea (LTP)
leachate. This can be ascribed to the enrichment of the simpler aromatic tryptophan-like
Results and Discussion: Aerobic biodegradation of landfill leachates
77
structures of Pitsea (LTP) leachate during aerobic biodegradation. This result is also
consistent with the EEM spectra (figure 4.3 and Table 4.4).
It was also found that the ratio of intensities of longer to shorter wavelengths for leachate
samples Pitsea (LTP), Pitsea (P4) and Rainham (P2) were higher than those of leachate
samples Rainham (LTP) and Rainham (FE). This implies that leachate samples Pitsea
(LTP), Pitsea (P4) and Rainham (P2) contained aromatic ring of more condensed form
than those for leachate samples Rainham (LTP) and Rainham (FE). This may explain the
relatively higher intensity reduction for peak I and II of Rainham (LTP) and Rainham (FE)
leachates in comparison with Pitsea (LTP), Pitsea (P4) and Rainham (P2) leachates.
The fluorescence spectroscopic results can be correlated with the general trend observed in
the UV spectroscopic results and also with the percentage reduction of COD and DOC.
Spectral analyses (figure 4.4) indicated that Rainham (LTP) and Rainham (FE) leachates
may have contained simple aromatic rings in the molecules. The amount of aromatic
organic compounds was also indicated to be low in these leachates (figure 4.2 (a)). These
leachates eventually exhibited high percentage reduction of COD, DOC (figure 4.1 (c, d)),
UV254 absorbance (figure 4.2 (b)) and H-L and F-L intensities (Table 4.4) during aerobic
treatment. This allows us to conclude that leachates containing low concentration of
aromatic organic compounds and simple aromatic structures may be easily degradable.
Results and Discussion: Aerobic biodegradation of landfill leachates
78
Pitsea (LTP)
320 360 400 440 480 52040
60
80
100
120
140
Peak IIPeak I
(ex. = 230-260 nm)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
Pitsea (P4)
300 330 360 390 420 450 480 510 5400
40
80
120
160
200Peak II(ex. = 320 - 350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
300 330 360 390 420 450 480 510 5400
20
40
60
80
100Peak II
(ex. = 360-390)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
320 360 400 440 480 520 560
20
40
60
80
100
120Peak II(ex. = 320-350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
320 360 400 440 480 520 56040
60
80
100
120
140 Peak II(ex. = 230-260)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
300 330 360 390 420 450 480 510 5400
30
60
90
120 Peak II(ex. = 360 - 390)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
T
U
Figure 4.4 Fluorescence spectra in emission scanning mode for all of the untreated
and treated leachate samples (dilution factor 25) (T = treated and U = untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
79
Rainham (LTP)
Rainham (FE)
300 330 360 390 420 450 480 510 540
14
21
28
35
42
49
56
Peak IIPeak I
(ex. = 320-350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
300 330 360 390 420 450 480 510 5400
20
40
60
80
Peak II
(ex. = 320-350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
320 360 400 440 480 520 560
50
100
150
200
250
300
Peak III
Peak I
(ex. = 230-260)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
320 360 400 440 480 520 5600
50
100
150
200
250
300
Peak III
Peak I
(ex. = 230-260)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
T
U
Figure 4.4 (contd.) Fluorescence spectra in emission scanning mode all of the
untreated and treated leachate samples (dilution factor 25) (T = treated and U =
untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
80
Rainham (P2)
Rainham (LTP Haz)
320 360 400 440 480 520
40
80
120
160
200
Peak II
(ex. = 230-260 nm)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
300 330 360 390 420 450 480 510 5400
40
80
120
160
200
Peak II
(ex. = 360-390)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
360 390 420 450 480 510
50
100
150
200
250
300
Peak II
(ex.= 320 - 350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
300 330 360 390 420 450 480 510 5400
20
40
60
80
Peak II
(ex. = 320-350)
Inte
nsity
Emission Wavelength (nm)
Day 10 Day 20 Day 30
U
U
Figure 4.4 (contd.) Fluorescence spectra in emission scanning mode all of the
untreated and treated leachate samples (dilution factor 25) (T = treated and U =
untreated)
Results and Discussion: Aerobic biodegradation of landfill leachates
81
4.4 Feasibility assessment of spectroscopic method
A feasibility assessment of the rapid characterisation of leachates using UV and
fluorescence spectroscopy can be made if the biodegradation potential of different
leachates obtained by spectroscopic method is compared with the results of well
established COD and DOC methods. The biodegradability of different leachates found
from %UV254 and %DOC reduction experiments were in good agreement with each other.
Both these methods showed that Rainham (LTP) and Rainham (FE) lechates were easily
biodegradable whereas the rest of the leachates were not so easily biodegradable. The
biodegradation study of constituent organic compounds using fluorescence spectroscopy
in different leachates was also in agreement with %DOC removal. Fluorescence
spectroscopy also indicated that after 30 days of aeration Rainham (LTP) and Rainham
(FE) had high degradability whereas Pitsea (LTP), Pitsea (P4) and Rainham (P2) showed
relatively low degradability.
Although %COD removal results were not an exact replica of %DOC removal and
spectroscopic results, in general this can also be accepted as fair agreement, as %COD
removal experiment also showed a high degradability for Rainham (LTP) leachate and low
degradability for Pitsea (P4) leachate. Mild inconsistency between established COD and
novel methods on a case-by-case basis can be attributed to the difference in the method of
estimating non-degraded compounds. As discussed before, due to the presence of a strong
oxidizing agent, some inorganic fractions of leachate get incorporated in COD
measurements whereas the DOC measurement mostly estimates the organic fraction of
leachate. The spectroscopic methods were also organic in nature whereby UV absorption
spectroscopy gave a quick estimation on the degradation of aromatic compounds and
fluorescence spectroscopy indicated degradation of several different organic compounds.
As a result, the good agreement between spectroscopic method and well established DOC
method point to the fact that a rapid assessment of biodegradation potential of leachates
using spectroscopic method was in general reliable. In addition, the UV spectroscopy
allowed the study of the possible biodegradation of aromatic compounds whereas
fluorescence spectroscopy enabled the study of the proportion of humic and fulvic-like
materials and their degradation in leachates. As such, these novel methods provided an
insight into the recalcitrant organic compounds in these landfill leachates.
Results and Discussion: Aerobic biodegradation of landfill leachates
82
4.5 Relationship of spectroscopic methods with COD and DOC
The relationship between fluorescence intensity, UV absorbance at 254 nm wavelength
and chemical and biological leachate characterisation parameters (COD and DOC) are
established in the following section for individual leachates.
4.5.1 Relationship between fluorescence intensity and DOC
The H- L and F- L fluorescence intensities are plotted against DOC and are presented in
figure 4.5 for all of the untreated and treated leachate samples. Linear correlation
coefficients are also presented. The results show that relationship between fluorescence
intensity and DOC varied among leachate samples. Leachate Rainham (P2) gave the best
relation with R2 = 0.97 and 0.94 for H- L and F- L fluorescence respectively, with the
Rainham (FE) leachate giving the weakest relation (H- L, R2 = 0.72 and F- L, R2 = 0.7).
Intermediate values were found for Pitsea (P4), Pitsea (LTP), Rainham (LTP) and
Rainham (LTP Haz) leachates (figure 4.5).
The strongest relationship observed in the Rainham (P2) leachate (figure 4.5) can be
attributed to the presence of significant proportions of H- L and F- L materials, which
dominate the DOC pool. In contrast, the weakest relationship occurred in the Rainham
(FE) leachate, where, H- L and F- L materials probably made up a relatively small
component of the DOC (Table 4.4). The weak relationship observed in this leachate could
be attributed to the relative insignificance of the H-L and F-L fractions as a proportions of
DOC compared with other fluorophores such as Trp-L fractions. Baker (2002)
investigated a small urban catchment where the relationship between F-L and DOC for the
whole catchment was 0.68, with a stronger relationship between these two parameters in a
tributary containing greater proportions of natural DOM but a weaker association between
F-L and DOC in the subcatchments where the anthropogenic influences were strong.
Similarly Cumberland and Baker (2007) determined the F-L to DOC correlation in river,
wetland, spring, pond and sewage samples with strong relationship (R2 = 0.756) between
F-L and DOC in the wetland water samples containing greater proportions of natural
DOM but a poor relationship (R2 = 0.14) between the two parameters in the final treated
sewage effluent. The results of this study in a sense contradict the results of Baker (2002)
Results and Discussion: Aerobic biodegradation of landfill leachates
83
and Cumberland and Baker (2007), where the relationship between H-L/F-L substances
and DOC were observed to be stronger for the anthropogenic influences. This might be
due to the fact that municipal waste might play a significant role as a source for H-L and
F-L compounds.
It is well known that Trp-L fluorescence intensity can be considered as a relic of
anthropogenic material in natural water, leachates and even in treated effluents (Galapate
et al., 1998; Baker et al., 2003, 2004; Reynolds, 2003). The relationship between Trp-L
fluorescence intensity and DOC is presented in figure 4.6 for all of the untreated and
treated leachate samples. When considering the relationship between the Trp-L
fluorescence intensity and H-L/F-L fluorescence intensity (figure 4.7) significant
relationship were also observed. This suggests that both H-L and F-L compounds and
protein-like (Trp-L) compounds played a significant role in the DOC of landfill leachates.
Furthermore, this indicates that the municipal waste might act as the origin not only of the
Trp-L fluorescence substance but also of the H-L and F-L fluorescence substance. That is
why in this study, the H-L/F-L fluorescence- DOC relationship was found to be stronger
for the anthropogenic leachates.
Results and Discussion: Aerobic biodegradation of landfill leachates
84
450 500 550 600 650400000600000800000
100000012000001400000160000018000002000000
H-L peak (R2 = 0.88, P = 0.06) F-L peak (R2 = 0.79, P = 0.11)
Pitsea (LTP)
Inte
nsity
DOC (mg/l)840 900 960 1020 1080 1140 1200
960000
1200000
1440000
1680000
1920000
2160000
H-L peak (R2 = 0.95, P = 0.02) F-L peak (R2 = 0.88, P = 0.06)
Pitsea (P4)
Inte
nsity
DOC (mg/l)
60 80 100120140160180200220240200000
400000
600000
800000
1000000
1200000
1400000 H-L peak (R2 = 0.785 P = 0.11) F-L peak (R2 = 0.77, P = 0.12)
Rainham (LTP)
Inte
nsity
DOC (mg/l)60 70 80 90 100 110 120 130
200000
400000
600000
800000
1000000
1200000
1400000
1600000 H-L peak (R2 = 0.72, P = 0.15) F-L peak (R2 = 0.7, P = 0.17)
Rainham (FE)
Inte
nsity
DOC (mg/l)
510 540 570 600 630 660 690
1000000
1200000
1400000
1600000
1800000
2000000
H-L peak (R2 = 0.97, P = 0.016) F-L peak (R2 = 0.94, P = 0.03)
Rainham (P2)
Inte
nsity
DOC (mg/l)45 48 51 54 57 60 63
300000
400000
500000
600000
700000
800000
F-L peak (R2 = 0.76, P = 0.013)
Rainham (LTP Haz)
Inte
nsity
DOC (mg/l)
T
T
UU
U
U
Figure 4.5 Comparisons of the fluorescence peak with DOC for: Pitsea (LTP), H-L: y =
583187 + 1284x, F-L: y = 312561 + 558x; Pitsea (P4), H-L: y = 1.2e6 + 663x, F-L: y = 786330 + 213x;
Rainham (LTP), H-L: y = 90453 + 5073x, F-L: y = 142004 + 833x; Rainham (FE), H-L: y = 449708 +
4520x, F-L: y = 244141 + 2454x; Rainham (P2), H-L: y = 1.3e6 + 950x, F-L: y = 888865 + 257x;
Rainham (LTP Haz), F-L: y = 391124 + 7058x. (P = probability that R = 0) (T = treated and
U = untreated) (Note: non-zero axes are intercepts)
Results and Discussion: Aerobic biodegradation of landfill leachates
85
450 500 550 600 650400005000060000700008000090000
100000110000120000
Trp-L peak (R2 = 0.55, P = 0.3)
Pitsea (LTP)
Trp-
L In
tens
ity
DOC (mg/l)
60 80 1001201401601802002202401000000
1500000
2000000
2500000
3000000
3500000
4000000 Trp-L peak (R2 = 0.81, P = 0.1)
Rainham (LTP)
Trp-
L in
tens
ity
DOC (mg/l)60 70 80 90 100 110 120 130
600000800000
1000000120000014000001600000180000020000002200000
Trp-L peak (R2 = 0.91, P = 0.05)
Rainham (FE)
Trp-
L in
tens
ity
DOC (mg/l)
510 540 570 600 630 660 69070000
80000
90000
100000
110000
120000 Trp-L peak (R2 = 0.8, P = 0.11)
Rainham (P2)
Trp-
L in
tens
ity
DOC (mg/l)45 48 51 54 57 60 63
600000
800000
1000000
1200000
1400000
1600000
1800000 Trp-L peak (R2 = 0.93, P = 0.04)
Rainham (LTP Haz)
Trp-
L in
tens
ity
DOC (mg/l)
600 750 900 1050 120090000
105000
120000
135000
150000
165000 Trp-L peak (R2 = 0.96, P = 0.03)
Pitsea (P4)
Trp-
L In
tens
ity
DOC (mg/l)
T
T
U
U
U
U
Figure 4.6 Comparisons of the Trp-L (270-280 nm excitation) fluorescence with DOC
for; Pitsea (LTP), y = 82788 + 148x; Pitsea (P4), y = -11344 + 137x; Rainham (LTP), y = -129431 +
17171x; Rainham (FE), y = -560439 + 20110x; Rainham (P2), y = 5753 + 148x; Rainham (LTP Haz) y
= -1.73e6 + 49658x. (P = probability that R = 0) (T = treated and U = untreated) (Note:
non-zero axes are intercepts)
Results and Discussion: Aerobic biodegradation of landfill leachates
86
30000 60000 90000 120000
300000
600000
900000
1200000
1500000
1800000
2100000 H-L (R2 = 0.41, P = 0.40) F-L (R2 = 0.4, P = 0.48)
Pitsea (LTP)
H-L
/F-L
inte
nsity
Trp-L intensity
1500000 2250000 3000000 3750000200000
400000
600000
800000
1000000
1200000
1400000 H-L (R2 = 0.5, P = 0.3) F-L (R2 = 0.48, P = o.3)
Rainham (LTP)
H-L
/F-L
inte
nsity
Trp-L intensity700000 1050000 1400000 1750000
200000
400000
600000
800000
1000000
1200000
1400000
1600000 H-L (R2 = 0.43, P = 0.35) F-L (R2 = 0.4, P = o.37)
Rainham (FE)
H-L
/F-L
inte
nsity
Trp-L intensity
70000 80000 90000 100000 110000
1000000
1200000
1400000
1600000
1800000
2000000
H-L (R2 = 0.79, P = 0.11) F-L (R2 = 0.62, P =0.21)
Rainham (P2)
H-L
/F-L
inte
nsity
Trp-L intensity
100000 120000 140000 160000900000
1200000
1500000
1800000
2100000
2400000 H-L (R2 = 0.98, P = 0.001) F-L (R2 = 0.97, P = 0.01)
Pitsea (P4)
H-L
/F-L
inte
nsity
Trp-L intensity
600000 900000 1200000 1500000270000
360000
450000
540000
630000 F-L (R2 = 0.91, P = 0.05)
Rainham (LTP Haz)
H-L
/F-L
inte
nsity
Trp-L intensity
T
T
U
U
U
U
Figure 4.7 Comparisons between H-L/F-L and Trp-L (270-280 nm excitation)
fluorescence intensity for; Pitsea (LTP), H-L: y = 1.07e6 + 2.84x, F-L: y = 444053 + 2.84x; Pitsea
(P4), H-L: y = 1.22e6 + 4.58x, F-L: y = 797986 + 1.6x; Rainham (LTP), H-L: y = 356587 + 0.21x, F-L: y
= 48426 + 0.09x; Rainham (FE), H-L: y = 332969 + 0.4x, F-L: y = 109036 + 0.2x; Rainham (P2), H-L: y
= 1.4e6 + 5.2x, F-L: y = 923875 + 1.2x; Rainham (LTP Haz), F-L: y = 19888 + 0.4x. (P = probability
that R = 0) (T = treated and U = untreated) (Note: non-zero axes are intercepts)
Results and Discussion: Aerobic biodegradation of landfill leachates
87
4.5.2 Relationship between fluorescence intensity and COD
The fluorescence intensities of H- L and F- L compounds are plotted against COD for all
of the untreated and treated leachate samples in figure 4.8. Linear correlation coefficients
are also shown. The results show that relationship between fluorescence intensity and
COD varied among leachate samples. Leachate Rainham (P2) gave the best relation with
R2 = 0.99 and 0.98 for H- L and F- L fluorescence respectively, with the Rainham (LTP)
leachate giving the weakest relation (H- L, R2 = 0.81 and F- L, R2 = 0.79). Intermediate
values were found for Pitsea (LTP), Pitsea (P4), Rainham (LTP Haz) and Rainham (FE)
leachates (figure 4.8).
Figure 4.8 shows that the fluorescence values decreased with decrease in COD for all of
the leachates. Similar relations between COD and fluorescence intensities of H-L and
protein-like peaks have been demonstrated by Bari and Farooq (1984), Reynolds and
Ahmad (1997), Reynolds (2002), Baker and Curry (2004) and Fu et al. (2007).
4.5.3 Relationship between UV absorbance at 254 nm (UV254) and COD
UV absorbance at 254 nm (UV254) is plotted against COD for all of the untreated and
treated leachate samples in figure 4.9. Linear correlation coefficients are also shown.
Leachate Rainham (LTP) gave the best relation with R2 = 0.99 with the Rainham (FE)
leachate giving the weakest relation (R2 = 0.73). Intermediate values were found for Pitsea
(P4), Rainham (P2), Pitsea (LTP) and Rainham (LTP Haz) leachates (figure 4.9). These
results indicate that chemical oxidation of organic matter might be associated with the
splitting of unsaturated bonds and the dissociation of the aromatic rings and hence
reduction of both COD and UV254. Bari and Farooq (1984) have also demonstrated that
strong correlations between COD and UV absorbance for wastewater.
Results and Discussion: Aerobic biodegradation of landfill leachates
88
1600 2000 2400 2800 3200 3600400000600000800000
100000012000001400000160000018000002000000
H-L (R2 = 0.976, P = 0.012) F-L (R2 = 0.988, P = 0.006)
Pitsea (LTP)
Inte
nsity
COD (mg/l)3200 3600 4000 4400 4800
800000
1000000
1200000
1400000
1600000
1800000
2000000
2200000
H-L (R2 = 0.938, P = 0.03) F-L (R2 = 0.866, P = 0.07)
Pitsea (P4)
Inte
nsity
COD (mg/l)
450 600 750 900 1050 1200200000
400000
600000
800000
1000000
1200000
1400000 H-L (R2 = 0.81, P = 0.1) F-L (R2 = 0.79, P = 0.11)
Rainham (LTP)
Inte
nsity
COD (mg/l)400 500 600 700 800 900
200000
400000
600000
800000
1000000
1200000
1400000
1600000 H-L (R2 = 0.87, P = 0.07) F-L (R2 = 0.87, P = 0.07)
Rainham (FE)
Inte
nsity
COD (mg/l)
2000 2500 3000 3500 4000
1000000
1200000
1400000
1600000
1800000
2000000
H-L (R2 = 0.99, P = 0.0008) F-L (R2 = 0.98, P = 0.012)
Rainham (P2)
Inte
nsity
COD (mg/l)400 500 600 700 800 900 1000
300000
400000
500000
600000
700000 F-L (R2 = 0.91, P = 0.05)
Rainham (LTP Haz)
Inte
nsity
COD (mg/l)
T
T
UU
U
U
Figure 4.8 Comparisons of the fluorescence peaks with COD for; Pitsea (LTP), H-L: y =
975899 + 135x, F-L: y = 274116 + 170x; Pitsea (P4), H-L: y = 1.36e6 + 126x, F-L: y = 847724 + 40x;
Rainham (LTP), H-L: y = 192965 + 867x, F-L: y = -21913 + 369x; Rainham (FE), H-L: y = 18876 +
1404x, F-L: y = -62497 + 726x; Rainham (P2), H-L: y = 1.64e6 + 80x, F-L: y = 983301 + 22x; Rainham
(LTP Haz), F-L: y = -68567 + 680x. (P = probability that R = 0) (T = treated and U =
untreated) (Note: non-zero axes are intercepts)
Results and Discussion: Aerobic biodegradation of landfill leachates
89
1500 2000 2500 3000 3500 40000.5
0.6
0.7
0.8
0.9
1.0 Data R2 = 0.94, P = 0.03
Pitsea (LTP)
UV 25
4 abs
orba
nce
COD (mg/l)2800 3200 3600 4000 4400 4800
1.0
1.2
1.4
1.6
1.8
2.0 Data R2 = 0.97, P = 0.02
Pitsea (P4)
UV 25
4 abs
orba
nce
COD (mg/l)
450 600 750 900 1050 1200 13500.10
0.15
0.20
0.25
0.30 Data R2 = 0.99, P = 0.0004
Rainham (LTP)
UV 25
4 abs
orba
nce
COD (mg/l)300 400 500 600 700 800 900
0.04
0.05
0.06
0.07
0.08
Data R2 = 0.73, P = 0.15
Rainham (FE)
UV 25
4 abs
orba
nce
COD (mg/l)
1800 2250 2700 3150 3600 40500.60
0.75
0.90
1.05
1.20 Data R2 = 0.94, P = 0.03
Rainham (P2)
UV 25
4 abs
orba
nce
COD (mg/l)400 500 600 700 800 900 1000
0.05
0.06
0.07
0.08
0.09
0.10 Data R2 = 0.88, P = 0.06
Rainham (LTP Haz)
UV 25
4 abs
orba
nce
COD (mg/l)
T
U
T
U
U
U
Figure 4.9 Comparisons of the UV254 with COD for; Pitsea (LTP), y = 0.6 + 9e-5x; Pitsea
(P4), y = 0.9 + 1.4e-4x; Rainham (LTP), y = 0.1 + 1.1e-4x; Rainham (FE), y = 0.04 + 4.1e-5x; Rainham
(P2), y = 0.6 + 9.8e-5x; Rainham (LTP Haz) y = 0.05 + 3e-5x. (P = probability that R = 0) (T =
treated and U = untreated) (Note: non-zero axes are intercepts)
Results and Discussion: Aerobic biodegradation of landfill leachates
90
4.6 Summary
In this chapter the behaviour of organic compounds in leachates undergoing an aerobic
biological treatment processes was investigated. Aerobic biodegradation was carried out
over a period of 30 days using four untreated and two treated leachate samples collected
from two UK MSW landfills, Pitsea and Rainham. Leachates were characterised using
conventional methods (COD and DOC) as well as the potentially useful characterisation
techniques of UV absorption and fluorescence spectroscopy. A good agreement was
observed between the spectroscopic method and the well established COD and DOC
methods of studying the biodegradation potential of different leachates. This established
the fact that a rapid assessment of the biodegradation of leachates using spectroscopic
method was in general reliable. In addition, the use of UV and fluorescence spectroscopy
allowed some analysis of leachates and it was suggested that organic compounds in Pitsea
and Rainham leachates were mostly aromatic in nature. Fluorescence spectroscopic result
showed a higher percentage reduction of protein-like (Trp-L) compounds than humic (H-
L) and fulvic (F-L) compounds during aeration confirming that protein-like compounds
were mostly biodegradable whereas H-L and F-L compounds were the key components of
the recalcitrant organic compounds. UV and the fluorescence spectroscopic analyses also
suggested that leachates containing higher concentration of aromatic compounds and
condensed aromatic structures are difficult to degrade. A combined analysis by novel UV
and fluorescence spectroscopy and conventional methods indicated that Rainham (LTP)
and Rainham (FE) is more easily biodegradable whereas Pitsea (LTP), Raniham (LTP
Haz), Pitsea (P4) and Rainham (P2) are not so easily biodegradable.
Results and Discussion: Anaerobic biodegradation of solid waste
91
Chapter 5 Anaerobic biodegradation of solid waste
5.1 Introduction
This chapter investigates the characteristics of recalcitrant organic compounds in leachates
during anaerobic biodegradation of different wastes. In addition, a compositional analysis of
the leachates generated from different waste samples is carried out to understand the nature of
the recalcitrant precursors; humic and fulvic-like compounds and non-recalcitrant materials
that are generated from different types of wastes. This research aims at fulfilling the research
objective of understanding the influence of different types of waste on the generation of
recalcitrant organic compounds and would be informative for waste management practices.
An anaerobic biodegradation experiment was carried out for four types of wastes, i.e., fresh
waste (FW), composted waste (CW), newspaper waste (NW) and synthetic waste (SW) in the
BMP test reactors. The experimental setup and sample preparation techniques are described in
detail in section 3.3. The biodegradation processes took place over a period of 150 days at
approximately 30oC. Anaerobic biodegradation was characterized through the measurement
of the volume of biogas produced, determination of the loss of cellulose, hemi-cellulose and
lignin in the waste samples by the measurements of Neutral Detergent Fibre (NDF), Acid
Detergent Fibre (ADF) and Acid Digestible Lignin (described in section 3.4.8) at the
beginning and at the end of the experiment. Leachates generated in the BMP test reactors
were taken periodically and investigations were performed to determine the composition of
the leachates generated during different stages of waste biodegradation. As in Chapter 4, these
investigations were carried out using conventional methods (i.e., COD and DOC) and using
potentially new technique of UV absorption and fluorescence spectroscopy.
Results and Discussion: Anaerobic biodegradation of solid waste
92
5.2 Characterisation of waste biodegradation in BMP reactors 5.2.1 Initial waste characterisation
Table 5.1 presents the initial composition of the four types of solid wastes. Newspaper waste
and composted waste contained highest (81%) and lowest (48%) percentages of total carbon
respectively, whereas fresh waste and synthetic waste contained intermediate percentages of
total carbon of around 55%. Newspaper waste was mainly composed of 51% cellulose, and
7.4% hemi-cellulose. These results agree with the reported values by Komillis and Ham
(2003). It was also found that composted waste contained the lowest percentages of cellulose
and hemi-cellulose of all of the wastes. This can be attributed to the fact that much of the
readily biodegradable cellulose and hemi-cellulose of the original waste would have been
degraded during composting. However, the percentage of lignin in composted waste and
newspaper waste was higher than in fresh waste and synthetic waste. The hemi-cellulose and
lignin contents of the fresh waste and composted waste found in this experiment were higher
whereas cellulose content was lower than the values reported by Zheng et al. (2007) which
might be due to the variation of the fibre measurements.
Figure 5.1 shows the cumulative biogas volume of fresh waste, composted waste, newspaper
waste and synthetic waste corrected to STP. This figure also shows the cumulative biogas
production in the control (blank) reactors which started to produce biogas from the beginning
indicating the bacterial activity in these reactors. The control (blank) reactors containing the
methanogenic mineral media and sewage sludge were used to determine the volume of biogas
produced from the bacteria alone. The cumulative biogas volume at STP for the control
reactors was 10.5 l/kg after 150 days of anaerobic biodegradation. The biogas produced by
each test waste was determined by subtracting the biogas produced in the control reactor from
the total biogas produced in each reactor. The cumulative biogas volume produced at STP for
fresh waste, composted waste, newspaper waste and synthetic waste were 67.5 l/kg, 24 l/kg,
48 l/kg and 67 l/kg respectively after 150 days of anaerobic biodegradation. Although after
150 days, the cumulative gas production in fresh waste and synthetic waste was the same, the
rate of mineralisation of organic matter in synthetic waste was slower than in fresh waste.
This could be due to the initial lower nitrogen content of synthetic waste (Table 5.1) which
might slow down the mineralisation of cellulose (Francou et al., 2008). Also the total gas
Results and Discussion: Anaerobic biodegradation of solid waste
93
production in composted waste and newspaper waste was lower than in fresh waste and
synthetic waste, probably due to the higher lignin fraction in composted waste and newspaper
waste samples (Table 5.1) and/or transfer into leachates. The low gas production in
composted waste can also be attributed to the fact that much of the readily biodegradable
carbon would have been degraded during composting. Almost 94% and 87% of the total
biogas was produced within 70 days for composted waste and newspaper waste respectively.
This indicates that, in these wastes, most of the bioavailable organic carbon was utilised
within 70 days.
Table 5.1 Initial composition of solid waste components
Chemical
analysis
Fresh
waste
Composted
waste
Newspaper
waste
Synthetic
waste
Fresh
waste
(Zheng et
al., 2007
Composted
waste
(Zheng et
al., 2007
Cellulose, % 25.0 12.0 51.0 30.0 48.6 32.8
Hemi-
cellulose, % 9.5 6.0 7.5 12.0 7.4 2.8
Lignin, % 15.0 20.0 21.0 12.0 8.4 18.6
Total carbon
(TC), % 56.0 48.0 81.0 54.0
Total
nitrogen
(TN), %
2.2 2.5 1.9 1.8
Results and Discussion: Anaerobic biodegradation of solid waste
94
0
10
20
30
40
50
60
70
80
0 50 100 150
Time (days)
Cum
ulat
ive
gas
volu
me
(l/kg
)
Blank
Fresh wasteComposted waste
Newspaper waste
Synthetic waste
Figure 5.1 Cumulative gas productions at STP for all of the wastes and control (blank)
in BMP reactors
5.2.2 Evolution of biochemical fractions during anaerobic biodegradation of wastes
The results of the NDF, ADF and ADL tests for all four types of waste samples at the
beginning and at the end of the anaerobic biodegradation process are shown in figure 5.2 (a-d)
(detailed description in section 3.3.4 and 3.4.8). The NDF content of the waste samples,
which is indicative of the entire fibre fraction of the waste showed a decreasing trend over the
test period for every waste sample indicating cellulose and hemi-cellulose degradation (figure
5.2 (a-d)). The NDF biodegradation rate was 0.61 g /kg DM/day, 0.04 g /kg DM/day, 1.63 g
/kg DM/day and 0.47 g /kg DM/day for fresh waste, composted waste, newspaper waste and
synthetic waste respectively by the end of 150 days. A similar decreasing trend was observed
for ADF biodegradation for all of the waste samples (figure 5.2 (a-d)). ADF represents the
portion of the waste that contains only cellulose and lignin. The average ADF biodegradation
rates were 0.56 g /kg DM/day, 0.003 g /kg DM/day, 1.55 g /kg DM/day and 0.42 g /kg
DM/day for fresh waste, composted waste, newspaper waste and synthetic waste respectively
over the whole period of test. ADL represents the portion of the waste that contains only
Results and Discussion: Anaerobic biodegradation of solid waste
95
lignin. It is interesting to note that all of the four wastes showed a mild increase in the
percentage of lignin fraction after 150 days of biodegradation (figure 5.2 (a-d)). The
maximum increase was for synthetic waste (12.8% to 18%). The increase can be attributed to
the low degradation of lignin and relatively high degradation of cellulose and hemi-cellulose
over time thereby increasing the percentage of ADL. Lignin is known to be non-
biodegradable/recalcitrant in anaerobic environments (Komilis and Ham, 2003).
(a) (b)
(c)
Fresh waste
0
20
40
60
80
100
0 150Time (days)
ND
F, A
DF,
AD
L (%
)
NDFADFADL
Composted waste
0
20
40
60
80
100
0 150Time (days)
ND
F, A
DF,
AD
L (%
)NDFADFADL
Newspaper waste
0
20
40
60
80
100
0 150
Time (days)
ND
F, A
DF,
AD
L (%
)
NDFADFADL
Synthetic waste
0
20
40
60
80
100
0 150
Time (days)
ND
F, A
DF,
AD
L (%
)
NDFADFADL
(d)
Figure 5.2 NDF, ADF and ADL over time for fresh waste, composted waste, newspaper
waste and synthetic waste samples
Results and Discussion: Anaerobic biodegradation of solid waste
96
Average values of cellulose and hemi-cellulose content for the four waste samples at the
beginning and at the end of anaerobic biodegradation were estimated using Eq. 3.7 and Eq.
3.8 (Chapter 3), and are presented in figure 5.3 (a-d). The percentage of cellulose in the BMP
reactors decreased over time at an average rate of 0.77 g /kg DM/day, 0.26 g /kg DM/day, 1.8
g /kg DM/day and 0.82 g /kg DM/day for fresh waste, composted waste, newspaper waste and
synthetic waste respectively over the period of 150 days. Thus, 46%, 32%, 54%, 41% of the
cellulose degraded for fresh waste, composted waste, newspaper waste and synthetic waste
respectively. The highest percentage of cellulose degradation in newspaper waste can be
attributed to the fact that cellulose found in paper is less resistant to biodegradation since
much of its primary structure has already been destroyed during the pulping process (Micales
and Skog, 1997). The comparatively lower percentages of cellulose degradation in fresh waste
and synthetic waste than in newspaper waste were due to the fact that fresh waste and
synthetic waste contained a low percentage of newspaper (35%) (Table 3.3). The lowest
percentage of cellulose degradation in composted waste can be attributed to the fact that most
of the readily biodegradable cellulose in the original waste would already have been degraded
during composting, and the remaining cellulose might be very resistant to biodegradation.
Hemi-cellulose degradation over the period of 150 days for fresh waste, composted waste,
newspaper waste and synthetic waste was 10.5%, 8.3%, 18.7% and 8% respectively (figure
5.3 (a-d)); was less than the cellulose degradation. Thus cellulose is a better indication of
degradation than hemi-cellulose under anaerobic conditions (Komilis and Ham, 2003).
Figure 5.3 (a-d) presents the total carbon (TC) fractions measured for all of the wastes at the
beginning and end of anaerobic biodegradation process. It was found that TC content showed
a decreasing trend over the test period for all of the waste samples. TC represents both organic
and inorganic carbon fractions and cellulose constitutes a significant fraction of the initial TC
in the waste (45% for fresh waste, 25% for composted waste, 63% for newspaper waste and
55% for synthetic waste). Therefore the decreasing trend of TC can be explained by the
cellulose degradation which is consistent with the observed trend of cellulose.
Results and Discussion: Anaerobic biodegradation of solid waste
97
(a) (b)
(d)(c)
Composted waste
0
20
40
60
80
100
0 150Time (days)
Cel
lulo
se,
Hem
icel
lulo
se, T
C (%
D
M) Cellulose
HemicelluloseTC
Newspaper waste
0
20
40
60
80
100
0 150
Time (days)
Cel
lulo
se,
Hem
icel
lulo
se, T
C (%
DM
) CelluloseHemicelluloseTC
Synthetic waste
0
20
40
60
80
100
0 150Time (days)
Cel
lulo
se, H
emic
ellu
lose
TC (%
DM
)CelluloseHemicelluloseTC
Fresh waste
0
20
40
60
80
100
0 150Time (days)
Cel
lulo
se,
Hem
icel
lulo
se, T
C (%
D
M)
CelluloseHemicelluloseTC
Figure 5.3 Cellulose, hemi-cellulose and TC over time for fresh waste, composted waste,
newspaper waste and synthetic waste samples
The (Cellulose + Hemi-cellulose) to Lignin ratio, (C+H)/L may be used to demonstrate the
extent of biodegradation or biodegradation potential (Wang et al., 1994). Figure 5.4 (a-d)
presents (C+H)/L ratios for all of the wastes; a decreasing trend was observed. After 150 days
of anaerobic biodegradation, the (C+H)/L values of fresh waste, newspaper waste and
synthetic waste fell from 2.3, 2.81 and 3.5 to 1.23, 1.2 and 1.58 respectively, while the
(C+H)/L ratio of the composted waste fell from 0.9 to 0.56. This decreasing trend could be
explained by the fact that although lignin in the waste dry mass increased with time, the
cellulose and hemi-cellulose (C+H) degradation was greater than the increase of lignin,
resulting in a decrease of the (C+H)/L ratio (Wang et al., 1994; Hossain et al., 2003). The
initial (C+H)/L ratios for fresh waste, newspaper waste and synthetic waste were between 2.0
Results and Discussion: Anaerobic biodegradation of solid waste
98
and 4.0 which closely matches the values reported in the literature for fresh MSW (Bookter et
al., 1982; Hossain et al., 2003). The low value of (C+H)/L for composted waste was
consistent with the prior decomposition of waste.
(a) (b)
(c) (d)
Fresh waste
0
1
2
3
4
0 150Time (days)
(C+H
)/L
Composted waste
0
1
2
3
4
0 150Time (days)
(C+H
)/L
Newspaper waste
0
1
2
3
4
0 150
Time (days)
(C+H
)/L
Synthetic waste
0
1
2
3
4
0 150
Time (days)
(C+H
)/L
Figure 5.4 (C+H)/L over time for fresh waste, composted waste, newspaper waste and
synthetic waste samples
Results and Discussion: Anaerobic biodegradation of solid waste
99
5.3 Characterisation of leachates formed in BMP reactors
This section presents the chemical characteristics of leachates produced from the anaerobic
biodegradation of all of the wastes in BMP reactors.
5.3.1 pH of leachate samples
The pH of the leachate samples generated from the waste biodegradation in BMP test reactors
is shown in figure 5.5. Initially the pH of the leachates for fresh waste, composted waste,
newspaper waste and synthetic waste were 7.1, 7.6, 7.2 and 7.3 respectively. These values
decreased to minima pH of 4.9 (fresh waste), 5.0 (composted waste), 4.9 (newspaper waste)
and 5.2 (synthetic waste) on day 10. This decrease in pH was probably due to the
accumulation of volatile fatty acids (Chian and DeWalle, 1976). After day 20, a gradual
increase in the pH value for all the wastes leachates was observed, to approximately level
values of 7.1, 7.3, 7.2 and 7.0 for fresh waste, composted waste, newspaper waste and
synthetic waste respectively over the last 80-100 days of the test.
4.5
5
5.5
6
6.5
7
7.5
8
0 20 40 60 80 100 120 140 160
Time (days)
pH
Fresh wasteComposted wasteNewspaper wasteSynthetic waste
Figure 5.5 pH of leachates taken from BMP reactors
Results and Discussion: Anaerobic biodegradation of solid waste
100
5.3.2 Carbon contents of leachate samples
Figure 5.6 (a-d) presents the total organic carbon (TOC), dissolved organic carbon (DOC),
total inorganic carbon (TIC) and dissolved inorganic carbon (DIC) contents of leachate
samples taken from the reactors of each waste sample at different stages of the biodegradation
process. Figure 5.6 (a-d) shows that TOC content increased from an initial value of 1308 mg/l
to 4625 mg/l after 35 days, 1813 mg/l after 5 days, 3888 mg/l after 25 days and 5468 mg/l
after 22 days and then decreased gradually to levels of 705 mg/l, 176 mg/l, 1408 mg/l and
1160 mg/l for fresh waste, composted waste, newspaper waste and synthetic waste
respectively by the end of the test. At the same time, the DOC content of the leachate samples
(which had previously been filtered through 1.2 µm Whatman GF/C filters) increased from
422 mg/l to 4381 mg/l (day 25), 1452 mg/l (day 5), 3821 mg/l (day 22) and 4918 mg/l (day
22) and then decreased (showing the same pattern as TOC) to 412 mg/l, 180 mg/l, 1110 mg/l
and 489 mg/l for fresh waste, composted waste, newspaper waste and synthetic waste
respectively by the end of the test. The initial increase was most likely to have been due to the
degradation of solid organic matter resulting in the formation of a range of soluble organic
matter (Ivanova, 2007).
Figure 5.6 (a-d) also shows that the TIC content decreased from an initial value of 191 mg/l to
67 mg/l (5 days), 57 mg/l (5 days), 38 mg/l (10 days) and 68 mg/l (3 days) and then increased
to levels of 128 mg/l, 556 mg/l, 182 mg/l and 328 mg/l for fresh waste, composted waste,
newspaper waste and synthetic waste respectively by the end of the test. At the same time, the
DIC content of the leachate samples decreased from 141 mg/l to 43 mg/l (day 5), 43 mg/l (day
3), 34 mg/l (day 5) and 44 mg/l (day 3) and then increased (showing the same pattern as TIC)
to 112 mg/l, 476 mg/l, 169 mg/l and 306 mg/l for fresh waste, composted waste, newspaper
waste and synthetic waste respectively by the end of the test. The initial decrease was most
likely to have been due to the consumption of inorganic carbon by bacteria for cell growth
(Ivanova, 2007).
Results and Discussion: Anaerobic biodegradation of solid waste
101
(a)Fresh waste
0
1000
2000
3000
4000
5000
6000
0 50 100 150
Time (days)
TOC
, DO
C (m
g/l)
0
100
200
300
400
500
600
TIC
, DIC
(mg/
l) TOC
DOC
TIC
DIC
Composted waste
0
1000
2000
3000
4000
5000
6000
0 50 100 150
Time (days)TO
C, D
OC
(mg/
l)
0
100
200
300
400
500
600
TIC
, DIC
(mg/
l) TOC
DOC
TIC
DIC
Newspaper waste
0
1000
2000
3000
4000
5000
6000
0 50 100 150
Time (days)
TC, T
OC
(mg/
l)
0
100
200
300
400
500
600
TIC
, DIC
(mg/
l)
TOC
DOC
TIC
DIC
Synthetic waste
0
1000
2000
3000
4000
5000
6000
0 50 100 150
Time (days)
TC, T
OC
(mg/
l)
0
100
200
300
400
500
600
TIC
, DIC
(mg/
l)
TOC
DOC
TIC
DIC
(c) (d)
(b)
Figure 5.6 Variation of total and dissolved organic and inorganic carbon of leachates
with time for fresh waste, composted waste, newspaper waste and synthetic waste
samples (dilution factor 25, 50, 40 and 10 for fresh waste, composted waste, newspaper
waste and synthetic waste respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
102
Figure 5.7 (a-d) presents the total chemical oxygen demand (TCOD) and dissolved chemical
oxygen demand (DCOD) for leachate samples taken from the reactors of each waste sample at
different stages of biodegradation. Figure 5.7 (a-d) shows that TCOD decreased from an
initial value of 982 mg/l to 537 mg/l, 122 mg/l, 215 mg/l and 521 mg/l after 3 days, then
increased gradually to levels of 2392 mg/l, 796 mg/l, 2056 mg/l and 1830 mg/l after 30 days
and then again decreased to levels of 623 mg/l, 295 mg/l, 623 mg/l, 589 mg/l for fresh waste,
composted waste, newspaper waste and synthetic waste respectively by the end of the test. At
the same time, the DCOD of the leachate samples (which had previously been filtered through
1.2 µm Whatman GF/C filters) decreased from an initial value of 537 mg/l to 263 mg/l, 56
mg/l, 122 mg/l and 200 mg/l after 3 days, then increased gradually to levels of 2212 mg/l, 699
mg/l, 1816 mg/l and 1668 mg/l after 30 days and then again decreased (showing the same
pattern as TCOD) to levels of 521 mg/l, 189 mg/l, 498mg/l, 498 mg/l for fresh waste,
composted waste, newspaper waste and synthetic waste respectively by the end of the test.
The initial decrease could be attributed to the consumption of organic matter by the bacteria.
At the beginning some of the organic matter may be used for the bacterial growth, after that
biodegradation started and the increase was likely to be due to the degradation of solid
organic matter resulting in the formation of a range of soluble organic matter (Ivanova, 2007).
Figures 5.6 (a-d) and 5.7 (a-d) show that the organic and inorganic compounds in the
leachates were mostly in dissolved form. However, these results also show that apart from the
variation due to the rise and fall of organic and inorganic carbon, TOC, DOC TCOD and
DCOD values for fresh waste, newspaper waste and synthetic waste leachates were
significantly higher than for composted waste leachates whereas the opposite trend was true
for inorganic carbon content. This could certainly be attributed to the fact that for composted
waste, a large amount of the biodegradable organic carbon present in the original waste would
have been degraded during composting. It was also found that a considerable amount of
organic compounds remained to be removed through biodegradation in fresh waste,
newspaper waste and synthetic waste leachates. To gain a better understanding, TCOD is
plotted against TOC for all of the wastes’ leachates and is presented in figure 5.8 (a-d). The
good relationship observed in the fresh waste, newspaper waste and synthetic waste leachates
could be attributed to the presence of a significant amount of organic compounds in the
leachates which was progressively mineralized. As can be seen the value of R2 for composted
Results and Discussion: Anaerobic biodegradation of solid waste
103
waste leachates was weak (R2 = 0.15). This probably indicates that composted waste
contained very little readily biodegradable organic carbon.
(b)Fresh waste
0
500
1000
1500
2000
2500
3000
0 50 100 150 200
Time (days)
CO
D (m
g/l)
TCOD
DCOD
Composted waste
0
500
1000
1500
2000
2500
3000
0 50 100 150 200
Time (days)
CO
D (m
g/l)
TCOD
DCOD
Newspaper waste
0
500
1000
1500
2000
2500
3000
0 50 100 150 200
Time (days)
CO
D (m
g/l)
TCOD
DCOD
Synthetic waste
0
500
1000
1500
2000
2500
3000
0 50 100 150 200
Time (days)
CO
D (m
g/l)
TCOD
DCOD
(c)
(a)
(d)
Figure 5.7 Variation of total and dissolved COD of leachates with time for fresh waste,
composted waste, newspaper waste and synthetic waste samples (dilution factor 25, 50,
40 and 10 for fresh waste, composted waste, newspaper waste and synthetic waste
respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
104
Fresh waste
R2 = 0.82P < 0.0001
0500
100015002000250030003500400045005000
0 1000 2000 3000
COD (mg/l)
TOC
(mg/
l)
Composted waste
R2 = 0.15P = 0.04
0100200300400500600700800900
0 200 400 600 800 1000
COD (mg/l)
TOC
(mg/
l)Newspaper waste
R2 = 0.58P = 0.003
0500
100015002000
25003000350040004500
0 500 1000 1500 2000 2500
COD (mg/l)
TOC
(mg/
l)
(c)
(a) (b)
Synthetic waste
R2 = 0.52P = 0.009
10001500200025003000350040004500500055006000
500 700 900 1100 1300 1500 1700 1900
COD (mg/l)
TOC
(mg/
l)
(d)
Figure 5.8 Comparisons of TOC with COD of leachates taken from fresh waste,
composted waste, newspaper waste and synthetic waste samples (P = probability that R
= 0)
Results and Discussion: Anaerobic biodegradation of solid waste
105
5.4 Mass balance estimate for the BMP reactors
A mass balance estimate was undertaken to monitor the carbon transfer from the solid to the
liquid and gas phases for all four wastes. After 150 days, all of the BMP bottles were emptied
and samples from the waste and leachate were taken for total carbon analysis. The
composition of the biogas produced (methane and carbon dioxide) from all of the wastes was
investigated at the end of the biodegradation process and presented in Table 5.2. The loss of
carbon by deposition of carbonate precipitates was estimated using the reductions in Ca+2 and
Mg+2 concentrations between the beginning and the end of the test. The summary of carbon
mass balance for waste samples is presented in Table 5.3. Table 5.3 indicates that 91 g, 81 g,
112 g and 71 g of carbon per kg dry wt. were utilized by fresh waste, composted waste,
newspaper waste and synthetic waste respectively during the period of biodegradation. It was
found that 8.5% (fresh waste), 7.0% (composted waste), 10.5% (newspaper waste) and 15%
(synthetic waste) of utilized TC was transferred into the leachate and 39% (fresh waste), 15%
(composted waste), 22% (newspaper waste) and 48% (synthetic waste) into the biogas by the
end of 150-day monitoring period.
The carbon mass balance estimates made for the BMP reactors indicate that the carbon input
exceeds carbon output by 9.8%, 14.4%, 10.3% and 8.2% for fresh waste, composted waste,
newspaper waste and synthetic waste respectively (Table 5.2). However, there are number of
potential sources of error in the carbon mass balances, the most important of which are the
neglect of any carbon utilized for bacterial biomass and growth; and random errors due to
limitations in the measurements devices used in the experiment and the initial sampling.
Table 5.2 Composition of biogas produced from all of the wastes (methane and carbon
dioxide)
Percentage volume of the
total biogas produced Fresh waste
Composted
waste
Newspaper
waste
Synthetic
waste
Methane (CH4) 60 70 67 60
Carbon dioxide (CO2) 38 27 28 35
Results and Discussion: Anaerobic biodegradation of solid waste
106
Table 5.3 Summary of carbon mass balance estimates for waste samples
Parameter Fresh wasteComposted
waste
Newspaper
waste
Synthetic
waste
TCt=0, waste, g/kg DM 560 480 810 540
TCt=0, leachate, g/kg DM 10.5 10.5 10.5 10.5
TCt=ti, waste, g/kg DM 469 399 698 459
TCt=ti, leachate, g/kg DM 7.5 5.7 11.8 10.4
TCt=ti, CaCO3, g/kg DM 1.1 1.65 1.16 1.22
TCt=ti, MgCO3, g/kg DM 0.4 0.45 0.61 0.6
TCt=ti, CH4, g/kg DM 21.9 9.0 17.5 21.6
TCt=ti, CO2, g/kg DM 13.8 3.5 7.3 12.7
Unaccounted C, g/kg DM 56.8 71.2 84 44.9
Mass balance error, % 9.8 14.4 10.3 8.2
Note: DM = dry matter; TCt=0, waste indicates the total carbon in the wastes before the
beginning of the test; TCt=ti, waste indicates the total carbon in the wastes after 150 days of the
test; TCt=0, leachate and TCt=ti, leachate are the TC content (TOC+TIC) of leachate samples taken at
the beginning and at the end of the biodegradation respectively; TCt=ti, CaCO3 and TCt=ti, MgCO3
are the carbon contents as CaCO3 and MgCO3 in the leachate between the beginning and the
end of the biodegradation respectively; TCt=ti, CH4, and TCt=ti, CO2 are the carbon in the biogas
considering methane and carbon dioxide respectively at the end of the test.
The results described in section 5.2 and 5.3 indicate that waste biodegradation and the
characteristics of the leachates generated varies significantly with the nature and composition
of wastes. For example, synthetic waste was prepared according to the composition of fresh
waste (food, grass and paper) but the mineralisation of organic compounds was different in
these fresh and synthetic wastes. It was also indicated that the total carbon transferred to
leachate and gas was relatively low in composted waste (Table 5.3) as the readily
biodegradable organic carbon had already been removed during composting and the
remaining organic carbon was very resistant to anaerobic biodegradation. The biochemical
Results and Discussion: Anaerobic biodegradation of solid waste
107
composition and the gas production trends also showed that fresh waste and synthetic waste
had the highest biodegradation potential of the other three wastes under consideration. Again,
the high percentage of cellulose and lignin also seem to affect the biodegradation of
newspaper waste. Therefore, it can be concluded that the composition and characteristics of
leachates produced during the biodegradation not only depend on the nature of wastes but also
on the anaerobic biodegradability of the waste components. It indicates that these factors
might play an important role in the characteristics of recalcitrant organic compounds in
leachates and hence it is important to investigate the evolution of these compounds in the
leachates associated with the anaerobic biodegradation of wastes. This is presented in section
5.5.
5.5 Spectroscopic analysis
In this section the recalcitrant organic compounds in the leachates generated from different
types of waste in the BMP reactors was studied using fluorescence and UV absorbance
techniques. From TOC and COD analyses (section 5.2.2 and figures 5.6 and 5.7) it was found
that organic matter in the leachate samples was mostly in dissolved form. Therefore the
unfiltered leachate samples were analysed for their spectroscopic characteristics to obtain the
representative information.
5.5.1 Fluorescence spectroscopic results 5.5.1.1 Excitation-Emission-Matrix (EEM) spectra of control samples
Figure 5.9 (a-c) shows 3D EEM spectra generated from fluorescence analyses of samples
taken from control (blank) reactors containing anaerobic seed (section 3.3.2) and
methanogenic nutrient media (Table 3.5) at days 5, 10 and 150 of the anaerobic
biodegradation process (EEM spectra of the other days’ samples are presented in Appendix
B). The nutrient media before seed addition (before biodegradation started) showed no
fluorescence (figure 5.9 (d)). However, in figure 5.9 (a-c), two apparent peaks were observed
that excited at 220-230/270-280 nm and emitted at 300-370 nm for all days. The tryptophan-
like (Trp-L) fluorophore refers to fluorescence centres at 220-230 nm and 270-280 nm
excitation and 340-370 nm emission wavelengths (Baker et al., 2004; Elliot et al., 2006),
Results and Discussion: Anaerobic biodegradation of solid waste
108
whereas the tyrosine-like (Tyr-L) fluorescence centres excite at approximately 220-230 nm
and 270-280 nm and emit at 300-320 nm wavelengths (Coble, 1996; Elliot et al., 2006).
Therefore, these peaks observed in figure 5.9 (a-c) can be attributed to the tyrosine-like (Tyr-
L) and tryptophan-like (Trp-L) compounds. These figures also show that the Tyr-L
fluorescence at 270-280 nm excitation wavelengths was obscured by the Raman line of water
and hence was ignored. These 3D EEM spectra suggest that both Tyr-L and Trp-L
fluorescence might be attributed to the protein-like fluorescence peaks, which were directly
related to the microbial activity of the anaerobic seed in the control reactors. Similar peaks
have also been observed in the river water samples owing to the activity of planktonic bacteria
(Elliot et al., 2006). However, the presence of the Tyr-L and Trp-L fluorescence peaks
ascertained that anaerobic seeds were the possible cause of these ‘protein-like’ peaks
observed in these control samples.
Results and Discussion: Anaerobic biodegradation of solid waste
109
112
121
2132
1
321
31
21
32
321
1
211
2543
2
1
2
1
1
1
321
1
321
1
321321
432
2
2
2121
1
34
321
56
1
21
2
7
1
8
32
109
21
2
1
321
1
1
21
2
2113213213211
324321321
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 5
21
1
3
1
214332
11 4
321
1
4321
765
1
1
1
1
432
65432
765
1
1
2
54321
6543
654321
11
2
4322
43
1
2
98765
43
1
2
1
2
43
2
1
2
2
12
65432
1
6543
1
3
1
2
7654276543
1
2
3
876543
1
1
2
4
4
1
1
54321
5
5432
2
6
1
78
2
765432
3
109
1
131211
3
5432
4
1514
2
1716
6543
5
201918
3
1
5432
3
12
2
765432
112
2
43211
2
4
1
432
1
12
176543212
65432
112
1
654322
1
65431
87654321
654321
321
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 10
1
11
11
21
21111
121
11
21
1
21121
2
21
1
34
21
5
1
6
2
1
1
1
111
1
11
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 150
(a)
54321
2154354321
54321321541
54321
543224354321
54321
5432115432
154321
54321
15432
1
1
154
3
2543313254
2
543
1
1
5432
11
21
5432154321
1
21
154321
1
54321
21
1
543254321
54321
4321
5432
12543
54321
321
154321
321
11
5432115432
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Mineral media Before biodegradation
(b)
(c) (d)
Figure 5.9 3D EEM fluorescence spectra of control (blank) reactors at day (a) 5, (b) 10
and (c) 150 showing Tyr-L and Trp-L fluorescence; (d) methanogenic mineral media at
day 0 (dilution factor 25)
Results and Discussion: Anaerobic biodegradation of solid waste
110
The changes in intensity of the Tyr-L and Trp-L fluorescence of the control (blank) samples
over time are presented in figure 5.10. It was observed that the fluorescence intensities of
these two fluorophores increased from day 5 to day 10 and then gradually decreased up to the
end of the biodegradation process. This could be explained by the fact that the growth of the
anaerobic seeds might be continuing up to day 10, giving the highest peaks at that time. After
that, the numbers of cells decreased and a lower amount of proteins-like substances was
produced over time (assuming that the protein-like substances produced were relative to the
numbers of cells present (Elliot et al., 2006)). This could explain the decreasing trend of the
intensity of Tyr-L and Trp-L fluorescence over time in control reactors.
0 20 40 60 80 100 120 140 1600
10002000300040005000600070008000
Inte
nsity
Time (days)
Tyr-L (220-230 nm) Trp-L (220-230 nm) Trp-L (270-280 nm)
Figure 5.10 Variation of Tyr-L and Trp-L fluorescence intensity over time in the control
(blank) reactor
Results and Discussion: Anaerobic biodegradation of solid waste
111
5.5.1.2 Excitation-Emission-Matrix (EEM) spectra of leachate generated from all waste
reactors
Table 5.4 presents the synthesis and/or utilization of the integrated fluorescence intensities
and position of the EEM peaks of leachates taken at different days of the biodegradation
process from fresh waste, composted waste, newspaper waste and synthetic waste reactors. In
addition to the Tyr-L and Trp-L peaks, two other distinct zones were also identified on the
EEM spectra (shown later) which were indicative of the presence of two types of organic
substances.
Zone 1: A peak was observed in all of the samples between 230-260 nm and 370-390 nm
excitations and 400-460 nm and 460-480 nm emission respectively and is widely recognized
as a component of the humic-like fractions (H-L) (Coble, 1996; Baker and Curry, 2004;
Cumberland and Baker, 2007).
Zone 2: Another peak was also present in all of the samples between 320-340 nm excitation
and 400-440 nm emission, which can be attributed to aromatic and aliphatic groups in the
DOM fractions and is commonly labelled as fulvic-like fractions (F-L) (Coble, 1996; Baker
and Curry, 2004).
It is worth mentioning that the presence of H-L and F-L fluorescence was not observed until
day 50 for leachates generated from all wastes. This might be attributed to the presence of low
molecular weight biodegradable organic compounds in leachates due to the waste degradation
during this period, which was also evidenced by the high values of TOC and COD (figures
5.6 and 5.7). As discussed in Chapter 2, H-L and F-L compounds are the essential building
block of recalcitrant organic compounds. Therefore, relative comparison of the fluorescence
peak intensities and peak positioning of H-L and F-L compounds of leachates generated from
different types of wastes may reveal information on the dependence of the nature and
composition of wastes on the evolution and characteristics of the recalcitrant organic
compounds.
Results and Discussion: Anaerobic biodegradation of solid waste
112
Table 5.4 Intensities and position of the EEM peaks for leachates taken from fresh waste, composted waste, newspaper
waste and synthetic waste reactors (dilution factor 25, 50, 40 and 10 for fresh waste (FW), composted waste (CW), newspaper waste (NW)
and synthetic waste (SW) respectively) (SD = standard deviation, n = 3) Zone 1 (H-L) Zone 2 (F-L) Tyr-L Trp-L
Sample
Day Position
(ex/em) Intensity Mean/SD
Position (ex/em)
Intensity Mean/SD
H-L intensity (column
4+column 6)
Position (ex/em)
Intensity Mean/SD
Position (ex/em)
Intensity Mean/SD Position (ex/em) Intensity
Mean/SD
50 230-260/400-450 3294/7.6 370-390/460-
480 1711/6.5 5005 310-350/400-440 9492/5.4 225-240/330-370 1451/5.5
60 230-260/400-450 2967/9.5 370-390/460-
480 1322/4.4 4289 310-350/400-440 8134/4.4 225-240/330-370 1511/5.0
70 235-260/400-450 2224/4.5 370-390/460-
480 1065/3.0 3289 310-350/400-440 6281/8.4 220-240/330-370 1718/6.4
80 230-270/400-470 2545/5.0 370-390/460-
480 1087/4.8 3632 310-350/400-440 7033/10.0 220-240/330-370 1799/7.4
90 230-270/400-470 3056/8.6 370-400/470-
490 900/7.0 3956 310-360/400-450 7999/9.8 230-240/330-370 198710.0
100 230-270/400-470 4589/8.5 370-400/470-
490 1504/5.5 6093 310-360/400-450 8322/5.8 230-240/330-370 20919.4
120 220-260/400-450 4959/9.5 370-400/470-
500 1211/3.8 6170 310-350/400-450 8136/10.0 240,270-280/350-370 533/8.0
140 230-260/400-440 5234/5.0 370-395/470-
500 1289/4.5 6523 300-355/400-450 7378/6.5 240,270-280/350-370 511/5.5
FW
150 230-260/400-440 5574/6.5 370-395/470-
500 1483/5.2 7057 300-355/400-450 7197/6.6 240,270-280/350-370 434/10
50 230-260/435-460 7323/10.8 350-360/450-
490 791/8.2 8114 300-355/400-450 9410/7.4 270-280/305-
320 911/5.5 230-240/330-400, 270-280/340-370 8400/8.4
60 230-260/435-460 7634/6.5 350-360/450-
490 6745.8 8308 300-355/400-450 10218/5.8 270-280/305-
320 902/10.4 230-240/330-400, 270-280/340-370 7934/4.8
80 230-260/440-460 7699/10.0 350-360/450-
490 935/6.4 8634 300-355/400-450 10624/3.4 270-280/305-
320 9146.8 235-240/330-400, 270-280/340-370 7646/5.4
90 235-260/435-460 7974/10.0 370-380/470-
500 1000/3.2 8974 300-355/400-450 9930/5.2 270-280/305-
320 899/7.4 230-240/330-400, 270-280/340-370 7344/6.0
100 230-260/435-460 10500/6.8 370-380/470-
500 2002/4.2 12502 300-355/400-450 8852/9.8 270-280/305-
320 905/4.5 235-240/330-400, 270-280/340-370 7155/3.8
120 230-260/435-460 10539/6.6 365-375/470-
490 2041/2.8 12580 300-355/400-450 8624/4.5 270-280/305-
320 832/7.6 230-240/330-400, 270-280/340-370 6780/9.5
140 230-260/435-460 10400/7.4 365-375/470-
490 2342/5.5 12742 300-355/400-450 8246/8.0 270-280/305-
320 801/5.8 230-240/330-400, 270-280/340-370 6156/7.4
CW
150 230-260/435-460 10890/7.6 365-375/470-
490 1920/8.0 12810 300-355/400-450 8064/4.5 270-280/305-
320 709/6.0 235-240/330-400, 270-280/340-370 5987/5.5
Results and Discussion: Anaerobic biodegradation of solid waste
113
Table 5.4 (Contd.) Intensities and position of the EEM peaks for leachates taken from fresh waste, composted waste,
newspaper waste and synthetic waste reactors (dilution factor 25, 50, 40 and 10 for fresh waste (FW), composted waste (CW), newspaper
waste (NW) and synthetic waste (SW) respectively) (SD = standard deviation, n = 3)
Sample Zone 1 (H-L) Zone 2 (F-L) Tyr-L Trp-L
Day Position (ex/em)
Intensity Mean/SD
Position (ex/em)
Intensity Mean/SD
H-L intensity (column
4+column 6)
Position (ex/em)
Intensity Mean/SD
Position (ex/em)
Intensity Mean/SD
Position (ex/em)
Intensity Mean/SD
50 320-370/400-450 7452/4.6 270-280/350-
370 605/7.4
70 230-260/400-440 5130/10.0 370-390/460-
480 1682/6.4 6812 300-340/400-440 8414/6.3
270-280/305-
320 199/7.3 220-240/330-
360 2489/6.4
90 230-260/400-445 5547/5.8 370-395/460-
480 1688/9.5 7235 300-340/400-440 8406/6.3 220-240/330-
360 822/6.5
100 230-255/400-445 6761/4.9 370-395/470-
500 1470/4.7 8231 310-360/400-450 8423/5.5 270-280/355-
370 549/7.8
120 230-260/400-445 7031/6.8 370-395/470-
500 1505/5.8 8535 310-360/400-450 8550/5.9 270-280/355-
370 479/7.2
140 230-265/400-445 7818/5.8 370-405/470-
500 1863/5.8 9681 310-360/400-450 9185/6.2
235-240/350-370, 270-
280/350-370 449/6.8
NW
150 230-265/400-445 8196/10.4 370-400/470-
500 2123/6.8 10319 300-355/400-450 9245/5.4
235-240/350-370, 270-
280/350-370 426/6.3
50 270-310/390-440 2449/5.6
70 300-350/390-440 3807/7.3 270-280/350-
380
90 265-315/390-450 3604/7.4 230-245/350-
370
100 265-315/390-450 3202/6.5 230-245/345-
370 791/6.6
120 265-315/390-450 3600/6.8 230-245/345-
370 921/7.6
140 265-320/370-460 4409/6.6 230-245/340-
370 916/7.7
SW
150 265-320/370-460 5524/5.4 230-245/340-
370 1320/5.5
Results and Discussion: Anaerobic biodegradation of solid waste
114
5.5.1.2.1 Comparative analyses of Excitation-Emission-Matrix (EEM) spectra of
leachates generated from all wastes
Figures 5.11 to 5.14 show the EEM spectra of the leachate samples taken from the fresh
waste, composted waste, newspaper waste and synthetic waste reactors respectively at days
50 and 150 of the biodegradation process (EEM spectra of the other days’ samples are
presented in Appendix B). These figures illustrate the presence of different peaks observed in
leachates. The control corrected EEM spectra are also presented in these figures (assuming
that microbial activity of the anaerobic seeds is similar in the control reactors as well as in the
fresh waste, composted waste, newspaper waste and synthetic waste reactors).
Table 5.4 show that for fresh waste leachates, the intensities of H-L and F-L compounds
decreased from day 50 to day 70. This decrease could be explained by the release of proteins
into the leachates and their subsequent utilisation. In fresh waste leachate a high content of
protein is expected due the presence of significant amount of food and grass in fresh waste
(65%, Table 3.3). Protein usually contains methylene (-CH2) functional groups which
contributes to lipophilicity and it has been reported that the presence of lipophilic extractives
could favour absorbing H-L and F-L compounds (Wu, 2002; Chen, 2003). Therefore, the
intensity decrease of H-L and F-L compounds from day 50 to day 70 can be explained by the
fact that a considerable amount of generated H-L and F-L compounds remains undetected by
the fluorescence spectroscopy due to the sorption by lipophilic extractives present in the fresh
waste leachate. However, this decrease of H-L/F-L fluorescence intensity was not observed in
newspaper waste and synthetic waste leachates (Table 5.4). This could be attributed to the
higher cellulose content of newspaper waste and synthetic waste compared to fresh waste
(Chen, 2003). It is known that cellulose and hemi-cellulose contribute little to the overall
sorption capacity as lipophilic extractives are small in these polymers (Chen, 2003). However,
composted waste contained a lower percentage of cellulose and hemi-cellulose and a higher
percentage of lignin than fresh waste, which could contribute towards the sorption of organics
(Table 5.1). Nevertheless, a decrease of H-L/F-L fluorescence intensity was not observed in
this leachate. This might be due to the fact that the sorption has been compensated by a
significant H-L and F-L compounds generation during anaerobic biodegradation of
composted waste.
Results and Discussion: Anaerobic biodegradation of solid waste
115
It was also observed from figures 5.12 (a-b) and Table 5.4 that Zone 1 (H-L) fluorophores of
composted waste leachates were in the long emission wavelengths. This suggested that H-L
fluorophores of composted waste leachates were more aromatic and high molecular weight
than those of the other leachates (Hudson et al., 2007). In composted waste, readily
biodegradable organic carbon would have been degraded during composting. As a result,
subsequent anaerobic biodegradation of the remaining organic matter would result in more
biologically resistant H-L compounds which might be aromatic in nature. This explains the
shift of H-L fluorophores to the longer emission wavelengths in this leachate. In addition,
figure 5.14 (a-b) showed that in comparison to other leachates, Zone 2 (F-L) fluorophores of
synthetic waste leachates were at shorter excitation and emission wavelengths. This indicated
that Zone 2 (F-L) fluorophores of synthetic waste leachates might contain a lower proportion
of aromatic rings than fresh waste, composted waste and newspaper waste leachates.
Results and Discussion: Anaerobic biodegradation of solid waste
116
21 4
11
213
1
1321
1
1
1
1
5432
1
2
1
1
1
1
432
11
2765
2
3
3
2
2
1
1
2
1
2
4
4
2
4
4
3
3
543
1
2
2
1
1
171615141312111098765
5
76
5
2
2
32
3
171615141312111098765
387654
1514131211109876
3
2
6
3
3
4
16151413121110987
2
6
3
5
1716151413121110987
4
6
21
5
5
2
181716151413121110987
66
5
4
5
4
2019181716151413121110987
3
7
3
5
6
6
151413121110988
4
1514131211109
2
7
4
33
16151413121110985
33
2
171615141312111098764
3
3
141312111098765
2
4
7
12111098765
8
1
4141312111098765
9
5
4
1110
23
131211109876
12
3
2
1110987654
13
12
2
2
2
10987654312
21
3
9876543
2
21
2
287654322
2
3
8765432
21
27654311
165432
1
2
2
1543
1
65432
1
1
165432543211
2
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 50
11
1
1
1
1
1
1
1
2
2
1
1
2
1
3
1
11
4
32
21
3
1
12
2
2
3
1
5
1
2
1
1
1
1
2
322
3
1
2
4
654
3
1
1
2
3
1
21
2
15432
1
1
21
1
1
1
1
1
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 50
(1)
(2)
(1)
213
1
212131
1
1
432
1
21
1
1
32
11
54
2
3
3
2
2
1
1
1
2
2
11
1
1
3
3
2
2
432
1
1
15141312111098765
4
4
65
5
2
32
3
1615141312111098764
3
2
654
1312111098765
2
5
3
3
3
4
3
15141312111098766
3
4
16151413121110987
4
5
21
4
4
16151413121110987
65
5
55
2
6
3
2019181716151413121110987
3
6
3
1413121110987
5
7
4
32
3
1413121110986
232
14131211109875
3
1514131211109876
1
4
222
121110987653
2
10987654
1
4
2
1312111098765
6
4
3
122
111098765
2
7
310987654
11
1
2
29876543
2
1
76543
1
2
1
2
765432
176543
2
2
2
1
1543
1
11
1
54321
1
14321
432
1
115432
3211
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 150
12
1
1221
1
3211
1
2
11
1
1
43
11
2
2
1
1
21
1
3
3
1
2
2
2
1
1
1
13121110987654
43
4
2
2
141312111098765
2
3
43
111098765
4
4
2
2
5
3
3
131211109876
3
5
2
3
14131211109876
54
1
4
3
1514131211109876
5
4
2
5
3
4
4
2
4
1817161514131211109876
3
5
2
1211109876
2
6
23
1211109875
22
111098764
3
3
1312111098765
1
3
2
2
10987654
1
3
2
2
987654
1
3
1
2
1110987654
2
3
21
9876542
1
9876543
1
1
1
1
17654321
154321
1
15432
1
1
54321
1432
1
1
321
1
14321
32
1
4321211
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedFW day 150
(1)
(1)
(2)
(b)
(a)
Figure 5.11 3D EEM fluorescence spectra of fresh waste (FW) leachates at day (a) 50, (b)
150 (dilution factor 25)
Results and Discussion: Anaerobic biodegradation of solid waste
117
21
1
11
1
3214
1
1
1
3214
1
1
1
54321
1
2
1
1
1
1
2
2
654
3
1
987
1
1
2
3
1
4
4
1
8765
51
1
65432
131211109
1
6
5432
2
1
10987
32
5432
1
1
6
5432
2
1
2
2
3
7
9876541
2
6
2
543
2
2
2
22
5
654
3
37654
3
3
2
654
23
3
2
6
62
7654
7
2
2
876543
8
2
1
9
5432
1110
1
34
12
5432
1413
1
5
1615
23
765432
191817
2
1
5
6
20
4
5432
7
21
2
5
3
22
5
87654
8
1 2
4
65432
2
4
3
76543
3
243
1
2
2 1
3
54322
4
1
76543
1
229876543
1
2
276543
2
18765432
109876543198765432
2
1
143
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedCW day 50
11
1
12 1
1
1
2
1
432
1
15
4
3
3
1
876
1
1
4
1
1
1
65
11
1
4
1
1
1
4
1
2
1
3
21
3
1
121
4
1
5
1
2
6
1
78
1
9
1
1
3
10
1
1211
1
2
1
13
1
2
32
1
1
11
1
1
2
2
2
2
1
1
111
3211
232
11
432321
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedCW day 50
(1)
(2)
(a)
11
1
1
1
1
1
1
3212
1
1
1
3
2
1
54
1
2
1
3
36
5
4
4
1
1
1
987
1
5
1
76
2
1
1
5
1
1
2
6
2
11
2
5
1132
5
1
4
1
2
3
5
21
1
3
6
3
1
78
2
1
9
1
10
3
2
21
111312
1
21
4
151416
23
2
21
1
1
17
432
4
2
1
1
3
2
21
1
3
2
1
1
21
1
21432
132
14321
654322431
12
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedCW day 150
211
1
1
1
22
1
32
1
14 3
3
765
1
3
1
1
1
54
11
4
11
3
1
1
1
2
211
45
1
6
2
7
1
8
1
1 910
1
3
11
21
11
12
32
2
11
22
11
3211
21321321
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedCW day 150
(1)
(2)
(b)
Figure 5.12 3D EEM fluorescence spectra of composted waste (CW) leachates at day (a)
50 and (b) 150 (dilution factor 50)
Results and Discussion: Anaerobic biodegradation of solid waste
118
21
11
321
11 2
21
1
213
13
2
1
31
432
1
2
11
1
1
432765
321
3
22
1
1
1
34
2
254
3
2323
432
1
1 1
9876
4
1
323
3
5432
2
1
987654
3
43
5432
254
1
34
32
4
4
5432
1
1
2
65
432
87
1
5
4
2
4
432
6
1
5
1110987
34
543
2
53
2
3
23
43
76
54
2
5643
5
3
543
32
2
4
5431
45
54321
5
65432
321
1
2
5
2
5432
6
6547
1
4
21
3
5432
7
1
2
21
5432
8
3
1
9
2
432
1
1
2
765432
2
1
432
12
1
1
1
432
3
1
2
1
2
2
432
32
2
3
1
109876543
21
12
13
1
1
11
132
1
1
11
321321
1
11
1
65432
1
1
15432
121
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedNW day 50
1
1
1
321
1
21
12
1323
1
321
12
11
11
1
32
2654
32123
1
2
432
2323
21
765
2
1
21
1
1
2
3
2
32
1
1
76543
432
32
154
1
34
32
3
4
32
1
1
5
4
32
2
765
44
2
1
1
109876
2
2
2
3
32
542
1
5
32
7654 3
1
5
3
633
32
32
1
2
4
4
321
3
454
32
2
1
2
2
44
432
321
1
3
2
32
54376
1
12
321
3
1
3
1
32
2
1
1
2
1
1
2
1
22
5432
1
1
21
1
21
12
1
11
21
32
1
3
98765432
1
1
2
21
121
1
1
1
2121
1
1
1
54321
1
1
1432
11
1
21
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedNW day 50
(2)
12
12
122131
1
432
1
1
1
1
1
1
32
11
4
2
3
3
2
2
1
1
21
1
11
3
3
3 2
2
2
1
1
15141312111098765
4
4
4
5
2
32
171615141312111098764
2
2
43
3
12111098765
3
5
3
3
2
33
15141312111098766
2
4
4
3
3
1716151413121110987
5
21
4
4
2
1716151413121110987
65
5
6
3
5
2
4
212019181716151413121110987
3
6
3
5
1413121110987
2
7
3
1413121110986
32
32
4
4
2
1312111098751514131211109876
1
4
3
3
322
12111098765
1
3
22
10987654
2
1
42
2
2
112111098765
2
4
22
2
1
1
10987653
1
1
1
9876542
18765432
5431
1
1654322
1
5431
1432
1
1432
1
1
14321
32
1
11432
11
2111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedNW day 150
(1)
(1)(2)
213
12
2132131
1
1
1
432
1
1
1
1
1
432
11
65
2
1
2
3
3
2
2
1
1
2
2
4
4
111
4
4
3
3
43
2
3
2
2
1
1
171615141312111098765
5
2
2
65
5
32
3
18171615141312111098765
6543
14131211109876
3
6
3
32
4
1716151413121110987
4
2
7
3
5
18171615141312111098
6
21
5
5
2
1918171615141312111098
76
6
3
7
4
5
5
3
4
5
232221201918171615141312111098
4
7
3
3
4
1615141312111098
6
8
5
15141312111097
232
151413121110985
3
4
17161514131211109876
1
4
223
3
1413121110987654
2
111098765
1
421
2
141312111098765
3
52
3
1
21
2
1211109876
2
3
1
11109876542
2 1
4
98765432
1
76543
1
2
1
1
765432
176543
2
2
1
165431
1
1
154321
1
15432
15432
1
1154323211
111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedNW day 150
(a)
(b)
Figure 5.13 3D EEM fluorescence spectra of newspaper waste (NW) leachates at day (a)
50 and (b) 150 (dilution factor 40)
Results and Discussion: Anaerobic biodegradation of solid waste
119
32
1
5
1
1
243
1
321
432
13
3
54
542134
14321
54321
1
4
5
11
54321
1
1
1
4323
4321
1
765
54
1
4
2
2
2
1
3
1
3
3
3
5432
1
3
3
2
2
32
12
1
1
1
121110987654
4
654
2
4
21
2
131211109876
5
1
5
2109876543
3254
2
1211109876
21
5
1
43
2
22
131211109876
3
811109
9876543
546
15141312
151413121110
4
1
3
3
12111098765
21
4
2
10987654
4
121110987
6
5
2
6
54
55
1413121110987
76
8
3
109
3
4
1211109845
6
1514131211
5
6
465
14131211109874
5476
45
5
141312111098765
1
4
41
54
987656
13121110987653
2
5
1
54321
1
4
131211109876543
1
3421
1413121110987654
4
5
3
3
52
14131211109876
1
4
1
3
543
141312111098765
21
3
34
1413121110987654
32
1
4
22
21
141312111098765
1
3
4
1
12
13121110987654
2
111098765
321
3
11110987654
1
432
2
1
1
1
9876543
21
2
43
876543
12
216543
1
1
165432
4321321
1
54321
2
1
2
3
23
3
54212
21
122
3
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedSW day 50
(2)54321
21
87
21
1
43
2
54321
543
24
3
9876
1
521354
165432
543214
87
11
21
1
11
765432
11
8
1
5432
121
3
765432
1098
654
2
5
3
3
3
2
2
1
4
2
4
4
4
1
543
2
2
4
4
3
3
3
654
3
2
2
2
2
1
1
1
1098765
5
10987
5
32
4
10987
6
21
6
1
31098765
4365
1
1
3
10987
213
6
1
54
1
4
3
2
10987
24
31098765
651
5
2
7
4
4
10987
6
321
6
21098765
109
8
78
23
65
66
109
2
587
5
5109
6
3
7
4
65
8
5
7
109
6
6
7
76
6
6
10987
7
56
5
9871066 6
109876
8
5
3
7543217
109876
9
4
5
10
32
109876
5
4
6
38
4
8
10987
43
6
510
10987
5
3
5432
4
4
3
1098765
543
52
4
1
32
4
109876
2
6
2
523
10987
1
5
2
21
4
109876
1
432
4
2
4
11098765
5763
4
3
4
4
1098765
32
4
2
65
2
1098765
21
4
11098765
1
21109876543
43215
11
211
98765432
12
1
3
2
23
654
65422
31
11
22
3
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedSW day 50
65432
321
10
321
1
1
48765
2
7654321
987654
327
6
1098
98543216987
1
11
8765432
1
987654321
2
8
109
21321
1
1
1
1098765432
1
21
132
2
987654
3
121
7
8765432
2
14131211109
111098
2
9
5
5
5
4
4
4
3
3
3
2
2
2
1
1
1
6
2
7
6
6
6
11098654
2
6
6
5
5
5
4
4
4
3
3
3
6543
324
2
2
2
1
1
1
201918171615141312111098
7
7
121110987
4
8
1
34
3
4321
4
3
20191817161514131211
10
9
21
10
4
1
201918171615141312111098765
765
4
98
5
20191817161514131211
4321
10
1
2
3687
5
5
20191817161514131211
76
16
4
20191817
1817161514131211109876
11109812
20193
8
2
8
8
3
6
6
2019181716151413121110
9
4321
9
3
201918171615141312111098
7
89
2019181716151413
12
1110
4
12
5
911108
10
2019181716151413
7141312
11
16
7
20191817
6
8
20191817161587109
4
11109
12
8
8
12
20191817161514138
1011
987131211
910
20191817161514131211109
2
9111087
8
1
11
87
18171615141312
8
20191817161514131211109
9
7
2
3
10987654321
1
4
8
201918171615141312111098
7
6
3
6
2
3
9874321
1
20191817161514131211109
87
8
10
2
5
76 5
10114
20191817161514131211
32
9
1
1098765
2019181716151413121110
4321
6
6
587
8
2019181716151413121110987
654321
8
3
4321
20191817161514131211109
1
65
9
2
2143
20191817161514131211108
43
20191817161514131211109
654321
7
1
2
201918171615141312111098
2
987654
5
325
1
2
17161514131211109876
4321
5
2
8765
1514131211109876
2143
5
11
1211109876
321
2
2 11
109876543
1
87654321
9
11
654322
10987654 32
543
132
4
21
645
65
10987434
4321
21
12
54 34
65
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedSW day 150
(2)765432
321
321
1
1
1
87654
2
7654321
987654
327
6
1098
1
9854321
1
6
1
987
1
1
1
98765432
10987654321
1
8
10
212
321
1
1
1
1098765432
21
132
21
9876543
1 2
2
7
9876543210
1098
2
5
5
5
4
4
4
3
3
3
2
2
1
76
23
7
6
6
6
21098654
2
6
6
6
5
5
5
4
4
41
76543
324
4
3
3
3
2
2
2
1
1
1
1098
8
7
7
7
7
1
1
1098
8
1
4
4
321
4
109
324
1
1098765
765 98
5 564321
2
4387
2
5
576
1
109876
1098
3
8
1
6
6
10
9
9 1
4
43215
1
1
3
109
8 7
89
45
9108
10
3
88
6
98710
4
10109
7
9
9
1098
10
10
9
987
9
108
10
10
8
9
10987654321
109
10
7
4
798
432110
1098
10
3
5
9 8
43251098
78
3
54321
7109
8
987654321
8
4
2
321
109
65
9
2
2
4
1
3214
4
10
1
9
2 1
5
10
654321
7
3
1098
987654
6
54365
34
10987
54321
5
2
9876
2
3
109876
21
4
5
1
1109876
32
2
2 1
1
109876543
1
876543219
11
65432
1
2
10987654 32
543
1432
4
2
5
1
6
765
1109874
1
3
4321
21
2254 3
465
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedSW day 150
(a)
(b)
Figure 5.14 3D EEM fluorescence spectra of synthetic waste (SW) leachates at day (a) 50
and (b) 150 (dilution factor 10)
Results and Discussion: Anaerobic biodegradation of solid waste
120
Table 5.4 shows that the intensity of H-L compounds gradually increased from day 70 to the
end of the biodegradation for all leachates while a variation in the intensity of F-L compounds
was observed. The changes in the intensities of H-L and F-L compounds indicated the change
in concentration of these compounds during anaerobic biodegradation of the wastes. It is
worth noting that the formation of H-L and F-L compounds followed the trend reported by
Artiola-Fortuny and Fullar (1982) that initially leachates contain high levels of F-L
compounds and as time progresses, the fractions of H-L compounds increase and the fractions
of F-L compounds may decrease or remain constant depending on the biodegradation
processes. As discussed before, H-L and F-L compounds are the key components of the
recalcitrant organic compounds. Therefore, it can be said that this increase of fluorescence
intensity of H-L and F-L compounds was an indication of the increase in the proportion of the
recalcitrant organic compounds in leachates during the anaerobic biodegradation. The
increase of recalcitrant organic compounds in leachates during the anaerobic biodegradation
could be attributed to the results of the relatively less degradable lignin and the generation of
recalcitrant products both from cellulose degradation and possibly microbial synthesis. The
other reason could be the release of less degradable material from the wastes as a result of
breakdown of the waste and the degradation of the readily degradable portion. A possible
degradation pathway of the waste components that leads to the generation of recalcitrant
organics is presented in figure 5.15. Table 5.4 also shows that the fluorescence intensities of
H-L and F-L compounds in composted waste and newspaper waste leachates are significantly
higher than in fresh waste and synthetic waste leachates. This indicates that composted waste
and newspaper waste leachates contained relatively higher concentrations of recalcitrant
organic compounds than fresh waste and synthetic waste leachates. This could be attributed to
the fact that composting significantly degraded readily biodegradable organic carbon in
composted waste and hence a subsequent anaerobic biodegradation of the remaining organic
matter produced more biologically resistant compounds. For newspaper waste leachates, high
percentage cellulose and lignin could contribute to the generation of biologically resistant
compounds. The gas production trend (figure 5.1) also showed that about 94% and 87% of the
total gas was produced within 70 days for composted waste and newspaper waste
respectively. This implies that most of the biologically available carbon was utilised within 70
days with the remaining organic carbon being relatively resistant to biodegradation. All of
these factors might play an important role in the formation of higher proportions of
Results and Discussion: Anaerobic biodegradation of solid waste
121
recalcitrant organic compounds in composted waste and newspaper waste leachates and hence
high intensities of H-L and F-L compounds were observed in these leachates. To understand
better the relative proportions of H-L and F-L compounds in the dissolved organic compounds
in leachates, the intensities of H-L and F-L compounds were plotted against the DOC for
every leachate sample (from day 50 onwards) in figure 5.16 (a-d). As can be seen from figure
5.16 (a-d), newspaper waste leachates gave the best relation between the H-L and F-L
fluorescence intensity and DOC with R2 = 0.94 and 0.89 respectively. The strongest
relationship observed in newspaper waste leachates indicated that significant proportions of
H- L and F- L compounds dominated the remaining DOC from day 50 onwards. In contrast,
the weakest relation observed in the fresh waste leachates (H- L, R2 = 0.32 and F- L, R2 =
0.27) indicating that H- L and F- L compounds made up a relatively small component of the
remaining DOC and significant amounts of biodegradable organic compounds were still
present. Composted waste leachate also showed good relationship between these two
parameters (H- L, R2 = 0.90 and F- L, R2 = 0.60). These results were indicative of the
significant amount of recalcitrant organic compounds in newspaper waste and composted
waste leachates as well. It was also observed from figure 5.16 (a-d) that for newspaper waste
leachates, both H-L and F-L compounds were the significant portions of remaining DOC
whereas for fresh waste and composted waste leachates, H-L compounds dominated the DOC.
Results and Discussion: Anaerobic biodegradation of solid waste
122
Figure 5.15 A possible degradation pathway of waste organic matter into the
recalcitrant organics in leachates
Waste components Carbohydrates (cellulose, hemi-cellulose), lignin, proteins
Sugars
Transformation by micro-organisms
Humic substance (H-L and F-L compounds)
Relatively less degradable lignin
Amino compounds
Microbially synthesized proteins (Tyr-L and Trp-L)
Microbial synthesis
Results and Discussion: Anaerobic biodegradation of solid waste
123
(a)
(b)
(Fresh waste)
(Composted waste)
1000 1500 2000 2500
4000
6000
8000
10000 H-L peak (R2 = 0.32)
Inte
nsity
DOC (mg/l)1000 1500 2000 2500
6000
7000
8000
9000
10000 F-L peak (R2 = 0.27)
Inte
nsity
DOC (mg/l)
100 150 200 250 300 350 4004000
6000
8000
10000
12000
14000
16000 H-L peak (R2 = 0.90)
Inte
nsity
DOC (mg/l)100 150 200 250 300 350 400
6000
8000
10000
12000
14000 F-L peak (R2 = 0.60)
Inte
nsity
DOC (mg/l)
Figure 5.16 Comparisons of fluorescence intensities with DOC for (a) fresh waste
leachates; H-L: y = 7109 – 1.52x; F-L: y = 6554 + 0.93x and (b) composted waste
leachates; H-L: y = 22355 – 51x; F-L: y = 5125 + 18x
Results and Discussion: Anaerobic biodegradation of solid waste
124
(c)
(d)
(Newspaper waste)
(Synthetic waste)
1000 1500 2000 2500 3000 3500 40002500
3000
3500
4000
4500
5000
5500 F-L peak (R2 = 0.54)
Inte
nsity
DOC (mg/l)
1400 1600 1800 20006000
7500
9000
10500
12000 H-L peak (R2 = 0.94)
Inte
nsity
DOC (mg/l)1400 1600 1800 2000
7500
8000
8500
9000
9500 F-L peak (R2 = 0.89)
Inte
nsity
DOC (mg/l)
Figure 5.16 (contd.) Comparisons of fluorescence intensities with DOC for (c)
newspaper waste leachates; H-L: y = 34711 – 17x; F-L: y = 13180 – 2.9x and (d)
synthetic waste leachates; F-L: y = 5254 - 0.8x
Results and Discussion: Anaerobic biodegradation of solid waste
125
Regarding the protein-like compounds, the intensities of Tyr-L and Trp-L compounds in
general followed a decreasing trend over the whole period of anaerobic biodegradation for
fresh waste, composted waste and newspaper waste leachates and an increasing trend for
synthetic waste leachates. This implies the general biodegradability of the protein like
compounds in fresh waste, composted waste and newspaper waste leachates. However, for
fresh waste and newspaper waste leachates, protein-like compounds accumulated from day 50
to day 70, i.e. intensity of Trp-L compounds increased (Table 5.4). This increase might be
explained by the release of protein from the wastes through the degradation of structures and
also by the production of protein from microbial synthesis (Wu, 2002). This increase of
proteins from day 50 to 70 in fresh waste leachate could also favour the intensity decrease of
H-L and F-L compounds through the sorption by lipophilic extractives as explained
previously. The increasing trend of Trp-L compounds observed in synthetic waste leachates
might also be explained by the protein-like compounds generated by waste degradation and
microbial synthesis. From Table 5.4, it was also found that the intensities of protein-like
compounds for composted waste leachates were significantly higher than for the other
leachates. This indicates a significant protein transfer into the composted waste leachate.
However, the carbon mass balance of different wastes (section 5.4) where carbon transfer in
biomass form cannot be taken into account showed that the lowest amount of utilized total
carbon was transferred into the composted waste leachate. This result simply indicates that
majority of the total carbon which was transferred into leachate in protein form might be
transferred as bacterial biomass (Ivanova, 2007).
The ratios of H-L/F-L over time can be used as an indicator of waste biodegradation and are
presented in figure 5.17 for fresh waste, composted waste and newspaper waste leachates (the
ratio of H-L/F-L for synthetic waste leachates was not calculated as H-L compounds were not
observed in this leachate). The ratio of H-L/F-L showed the existence of three phases of
anaerobic biodegradation in the fresh waste leachates. From day 50 to day 70 (termed as
phase 1) the ratios of H-L/F-L remained constant. The ratio decreased during the second
phase (from day 70 to day 90) possibly because of the synthesis of F-L compounds.
Thereafter in the last phase (from day 100 to the end of the test) the ratios increased by the
increase of H-L compounds and decrease of F-L compounds. However, in composted waste
leachates, phase 1 was absent. Contrary to the fresh waste and composted waste leachates,
Results and Discussion: Anaerobic biodegradation of solid waste
126
only the phase 3 was present in newspaper leachates as the decrease in H-L/F-L ratio was not
observed in this leachates. The values of the H-L/F-L ratio at the end of the biodegradation in
the leachates followed the sequence: composted waste>newspaper waste>fresh waste.
(c)
(a) (b)
20 40 60 80 100 120 140 1600.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8Composted waste
Phase 3Phase 2
H-L
/F-L
ratio
Time (day)40 60 80 100 120 140 160
0.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8Fresh waste
Phase 3Phase 2
Phase 1
H-L
/F-L
ratio
Time (day)
40 60 80 100 120 140 1600.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8Newspaper waste
Phase 3
H-L
/F-L
ratio
Time (day)
Figure 5.17 H-L/F-L ratio and phases in the fresh waste, composted waste and
newspaper waste leachates
Results and Discussion: Anaerobic biodegradation of solid waste
127
The above results suggest that fluorescence spectroscopy may be a suitable tool for
investigating the recalcitrant organic compounds formed in leachates during the anaerobic
biodegradation of wastes. The intensity, positioning and shift of the fluorescence peaks of
different leachates may be explained by the variability in biodegradability of the waste
components. This could be related to the chemical and structural changes of individual
organic compounds during biodegradation.
5.5.1.3 Fluorescence analyses of humic and fulvic like materials in leachates in emission
scanning mode
In this section the structure of various compounds in leachates under anaerobic
biodegradation of waste is considered. As discussed in Chapter 4, fluorescence spectra in
emission scanning modes provide information about the aromatic structure of compounds.
Figure 5.18 (a-d) presents the fluorescence emission characteristics at specified excitation
wavelengths for leachates taken from four different wastes. The leachates from all four wastes
exhibited two characteristic peaks denoted Peak I (at a shorter wavelength of 340-350 nm)
and Peak II (at a longer wavelength of 430 nm). Kang et al. (2002) reported that the shorter
wavelength peaks (360 nm) can be interpreted as being due mainly to the presence of simple
aromatic rings in the molecule, whereas the peaks in the longer wavelength (430 nm) are due
to the condensed aromatic rings and conjugation of simple aromatic rings.
Figure 5.18 (a-d) shows that the intensity of Peak II at (300-390) nm excitation was stronger
than that of Peak I in all of the leachates. For composted waste leachates, Peak I at (230-260)
nm excitation had the highest intensity. As compared with the EEM spectrum (figures 5.11,
5.12, 5.13 and 5.14), Peak I can be assigned as Trp-L fluorophore. Peak II at (230-260) nm
excitation is located in the range of H-L and at (300-390) nm excitations in the range of F-L
fluorophores. It was found that for composted waste leachates; Peak II at (230-260) nm
excitation was in a higher wavelength region (450 nm) than for the other leachates. This may
mean that H-L compounds in composted waste leachates contained more condensed forms of
aromatic rings than the other leachates. It was also observed that for synthetic waste leachates;
the Peak II at (270-340) nm excitation was in the region around 400 nm, indicating that F-L
Results and Discussion: Anaerobic biodegradation of solid waste
128
compounds in synthetic waste leachates contained less condensed forms of aromatic rings
than the other leachates.
From figure 5.18 (a-d), the intensity of Peak II increased during 150 days of anaerobic
biodegradation for all leachates. This suggests that condensation of the aromatic structure of
H-L and F-L molecules increased over time.
Figure 5.18 (a-d) also shows that the ratio of intensities of the peaks of longer wavelength to
the shorter wavelengths for newspaper waste leachates was higher than for fresh waste and
synthetic waste leachates. This implies that newspaper waste leachates contained more
condensed forms of aromatic structures than fresh waste and synthetic waste leachates. In
addition, it appears that composted waste leachates contained more condensed forms of
aromatic H-L compounds than the other leachates. These fluorescence spectroscopic results
led to the conclusion that composted waste and newspaper waste would contribute to
leachates with higher concentrations and more condensed forms of recalcitrant H-L and F-L
compounds than fresh waste and synthetic waste.
Results and Discussion: Anaerobic biodegradation of solid waste
129
Fresh waste leachates
Composted waste leachates
300 350 400 450 500 55001234567
(ex. = 230-260)
Peak II
Peak I
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
(a) (b)
300 350 400 450 500 55001234567
(ex. = 300-390)
Peak IPeak II
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
300 350 400 450 500 55001234567
(ex. = 300-390)
Peak IPeak II
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
300 350 400 450 500 55002468
10121416
Peak II
Peak I(ex. = 230-260)
Inte
nsity
Emission Wavelength (nm)
Day 50 Day 100 Day 150
(c) (d)
Figure 5.18 Fluorescence spectra of fresh waste (a, b) and composted waste (c, d)
leachate samples in emission scanning mode at excitation wavelengths 230-260 nm and
300-390 nm (dilution factor 25 and 50 for fresh waste and composted waste respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
130
Newspaper waste leachates
Synthetic waste leachates
300 350 400 450 500 55001234567
(ex. = 230-260)
Peak I
Peak II
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
(f)
300 350 400 450 500 55001234567
(ex. = 300-390)
Peak I
Peak II
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
300 350 400 450 500 55001234567
Peak IIPeak I
(ex. = 270-340)
Inte
nsity
Emission wavelength (nm)
Day 50 Day 100 Day 150
(e)
(g)
Figure 5.18 (contd.) Fluorescence spectra of newspaper waste (e, f) and synthetic waste
(g) leachate samples in emission scanning mode at excitation wavelengths 230-260 nm
and 300-390 nm (for synthetic waste, excitation wavelength was 270-340 nm) (dilution
factor 40 and 10 for newspaper waste and synthetic waste respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
131
5.5.2 UV spectroscopic results
As discussed in Chapter 2, the specific UV absorbance (SUVA) can be used to study the
aromaticity of leachates. Figure 5.19 shows the SUVA calculated for fresh waste, composted
waste, newspaper waste and synthetic waste leachates with time during biodegradation. It was
found that the SUVA of these leachates increased with time, indicating the concentration of
aromatic molecules increased with the waste biodegradation. This result is in agreement with
the previous findings reported by Kang et al. (2002) and Fan et al. (2006). At the end of
biodegradation, the SUVA values for fresh waste and synthetic waste leachates were
approximately 0.065 and 0.06 respectively. The SUVA of newspaper waste leachates was
about twice this value, whereas the SUVA of composted waste leachates was 30 times as
great implying that the molecular characteristics of the composted waste and newspaper waste
leachates were more aromatic in nature.
0.001
0.01
0.1
1
10
5 10 40 50 60 70 80 90 100 120 140 150
Time (days)
SUVA
(l/m
g/cm
)
Fresh wasteComposted wasteNewspaper wasteSynthetic waste
Figure 5.19 The SUVA for fresh waste, composted waste, newspaper waste and synthetic
waste leachates (dilution factor 25, 50, 40 and 10 for fresh waste, composted waste,
newspaper waste and synthetic waste respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
132
The ratio of absorbance at 465 nm and 665 nm (E4/E6) can provide an indication of the
molecular size of leachates (Kang et al., 2002). Figure 5.20 shows the E4/E6 ratio of fresh
waste, composted waste, newspaper waste and synthetic waste leachates which decreased
over time. It is known that the E4/E6 ratio is inversely proportional to molecular weight, which
implies that the E4/E6 ratio decreases with increasing molecular weight. Therefore, the
decrease of the E4/E6 ratio indicates that the concentration of high molecular weight molecules
increased with waste biodegradation. This agrees with the results reported by Kang et al.
(2002).
0
0.5
1
1.5
2
2.5
3
3.5
5 10 40 60 100 120 150
Time (days)
E4/E
6
Fresh wasteComposted wasteNewspaper wasteSynthetic waste
Figure 5.20 E4/E6 ratio for fresh waste, composted waste, newspaper waste and synthetic
waste leachates (dilution factor 25, 50, 40 and 10 for fresh waste, composted waste,
newspaper waste and synthetic waste respectively)
Results and Discussion: Anaerobic biodegradation of solid waste
133
These above results can be correlated with the general trend observed in fluorescence
spectroscopic results. They suggest that waste biodegradation increased aromaticity and the
molecular weight of organic compounds in leachates. Composted waste and newspaper waste
leachates had the higher fluorescence intensities indicating a more condensed form of
aromatic structure of H-L and F-L compounds than fresh waste and synthetic waste leachates.
As aromaticity is one of the main characteristics of recalcitrant H-L and F-L compounds, the
UV and fluorescence spectroscopic results allow us to conclude that leachates produced
during the anaerobic degradation of composted waste and newspaper waste may contain a
higher proportion of condensed forms of aromatic organic compounds than fresh waste and
our synthetic waste leachates.
5.6 Summary
In this chapter, the evolution and characteristics of recalcitrant organic compounds in
leachates during anaerobic biodegradation of four types of solid waste was investigated.
Biochemical Methane Potential (BMP) tests were carried out on fresh waste, composted
waste, newspaper waste and synthetic waste over a period of 150 days and leachates produced
from the degradation of all four wastes were characterised using conventional methods as well
as by UV absorption and fluorescence spectroscopy. The increase of the fluorescence
intensities related to humic (H-L) and fulvic (F-L) compounds in the leachates formed during
anaerobic biodegradation was considered to be indicative of the increased proportion of
recalcitrant organic compounds. The recalcitrant organic compounds (H-L and F-L
compounds) in all four leachates were influenced by the nature of the wastes and also the
anaerobic biodegradability of the component solid materials. It was estimated that the total
carbon released from solid waste which was transferred to leachate and gas was low in
composted waste because of the amount of biodegradable organic compounds that had been
removed during composting. Subsequent anaerobic biodegradation of the remaining organic
compounds produced more biologically resistant compounds. Additionally, the relatively less
resistant cellulose and high lignin present in newspaper waste contributed to an increase in the
proportion of recalcitrant organic compounds in the leachate. As expected, trends in
biochemical composition of waste and gas production indicated that fresh waste and synthetic
waste had a higher biodegradation potential than composted waste and newspaper waste.
Results and Discussion: Anaerobic biodegradation of solid waste
134
Fluorescence and UV spectroscopic results led us to conclude that composted waste and
newspaper waste would contribute to the generation of higher concentrations of more
condensed aromatic structures of recalcitrant H-L and F-L compounds than fresh waste and
synthetic waste. These results indicate that in a landfill system, composted waste and
newspaper waste would generally result in a higher concentration of recalcitrant organic
compounds than fresh waste and synthetic waste.
Conclusions and future work
135
Chapter 6 Conclusions and future work
6.1 Conclusions
This research investigated the characteristics of recalcitrant organic compounds in landfill
leachates and also studied the possible use of UV absorption and fluorescence spectroscopy
for rapid characterisation of these compounds in leachates. These investigations comprised
both studies of leachates collected from two UK landfill sites (Pitsea and Rainham) and of
leachates generated from laboratory scale anaerobic biodegradation of solid waste
components. This fulfils the research objective of understanding the nature of the recalcitrant
organic compounds in leachates in course of a biological treatment processes, comparing the
biodegradation associated evolution of the recalcitrant organic compounds in two UK landfill
sites and assessing the influence of individual waste types on the recalcitrant organic
compounds. The investigation of the recalcitrant organics was carried out by established
methods of COD and DOC measurements, and by the potential new techniques, UV
absorption and fluorescence spectroscopy. The UV and fluorescence spectroscopy provided
rapid ways of monitoring some constituent compounds of leachates and the biodegradation of
leachates were studied by estimating the eventual removal of the measurable variables in
course of the treatment. UV spectroscopy was used to characterize the aromatic compounds
whereas the fluorescence spectroscopy was used for fingerprinting different organic
compounds such as humic, fulvic, proteins in leachates and their biodegradation. The
biodegradability of different leachates thus indicated by the UV and fluorescence
spectroscopy was compared with the results of conventional methods to assess the validity of
the new approaches. This investigation satisfied the research objective of the basic feasibility
study on the application of UV absorption and fluorescence spectroscopy for a rapid and an
economical characterisation of the organic compounds in landfill leachates and also provided
insight into the recalcitrant organic compounds.
Leachate samples were collected from two UK MSW landfills, Pitsea and Rainham in Essex
and a laboratory scale aerobic biological treatment was carried out in glass reactors over a
Conclusions and future work
136
period of 30 days with an air supply arrangement. The conventional methods of
characterisation demonstrated the expected gradual decrease of total/dissolved chemical
oxygen demand (COD) and total/dissolved organic carbon (DOC) in all leachates indicating
the reduction of organic compounds by the aerobic biological treatment. The observed COD
and DOC values of Pitsea (P4), Rainham (P2) and Pitsea (LTP) leachates were considerably
higher than Rainham (LTP), Rainham (FE) and Rainham (LTP Haz) leachates indicating that
Pitsea (P4), Rainham (P2) and Pitsea (LTP) leachates contained comparably higher amounts
of organic compounds than Rainham (LTP), Rainham (FE) and Rainham (LTP Haz)
leachates. In addition, no discernible difference was observed between total and dissolved
carbon contents of different leachates which indicated that the organic compounds in these
leachates were mostly in dissolved form. The biodegradability of different leachates was
estimated from the remaining percentages of DOC and COD in course of the treatment. The
results of %DOC reduction trend showed that the Rainham (LTP) and Rainham (FE)
leachates which had lower concentrations of organic compounds were more easily
biodegradable leachates whereas the Pitsea (LTP), Pitsea (P4) and Rainham (P2) leachates
which had higher concentration of organic compounds were not so easily biodegradable.
A study of constituent organic compounds during biodegradation of leachates using UV
absorption and fluorescence spectroscopy provided for a rapid investigation and insight into
the recalcitrant organic compounds in leachates. It was found that leachates with high UV254
absorbance values exhibited a low percentage reduction of UV254 absorbance and DOC in
course of the treatment. This suggested that leachates containing high aromatic organic
compounds usually result in low percentage removal of DOC in the course of a treatment
process. UV absorption allowed the study of the effect of biodegrdation on different leachates
through the reduction of aromatic compounds in course of the treatment. The study also
showed that Rainham (LTP) and Rainham (FE) leachates were more easily biodegradable
than the Pitsea (LTP), Pitsea (P4) and Rainham (P2) leachates which were in agreement with
results estimated using conventional methods. The good agreement in the degradation trends
observed in different leachates by UV254 absorption and DOC measurements probably
indicated that organic compounds contributing DOC in these leachates were mostly aromatic
in nature. Fluorescence spectroscopy was used to study the biodegradation of organic
compounds such as humic, fulvic, proteins in leachates. This study in general showed a lower
Conclusions and future work
137
reduction of the fluorescence intensities of the humic-like (H-L) and fulvic-like (F-L)
compounds in comparison to the protein-like compounds for all leachates in course of the
biodegradation. This indicated that humic-like (H-L) and fulvic-like (F-L) compounds were
the key components of recalcitrant organic compounds. It was also found that the reduction of
the fluorescence intensities of the humic-like (H-L) and fulvic-like (F-L) compounds in
Rainham (LTP) and Rainham (FE) lechates were around 60% after 30 days of aeration
whereas this value for Pitsea (LTP), Pitsea (P4) and Rainham (P2) leachates were around 4 to
30%. This result indicated the biodegradation potential of different leachates which were in
agreement with UV spectroscopy and conventional DOC measurement results. Although
%COD removal results were not an exact replica of %DOC removal and spectroscopic
results, in general this was also in fair agreement, as %COD removal was also higher for
Rainham (LTP) leachate than for Pitsea (P4) leachate. Thus the general agreement observed
between the spectroscopic method and conventional methods in assessing biodegradability of
different leachates simply pointed to the fact that a rapid characterisation of the organic
components of landfill leachate that were more resistant to aerobic degradation using
spectroscopic method was reliable.
To understand the influence of different types of waste on the generation of recalcitrant
organic compounds in leachates, an anaerobic biodegradation experiment was carried out for
four types of wastes, i.e., fresh waste, composted waste, newspaper waste and synthetic waste
in the BMP test reactors and leachates generated during biodegradation processes were
characterized. The biodegradation processes took place over a period of 150 days at
approximately 30oC. General waste biodegradation was first studied through different
measurements. The loss of cellulose, hemi-cellulose and lignin in the waste samples due to
biodegradation was estimated by the measurements of Neutral Detergent Fibre (NDF), Acid
Detergent Fibre (ADF) and Acid Digestible Lignin (ADL). As expected it was found that
composted waste contained the lowest percentages of cellulose and hemi-cellulose of all of
the wastes. The percentage of lignin in composted waste and newspaper waste was found to
be higher than in fresh waste and synthetic waste. The estimated cellulose and hemi-cellulose
degradation after 150 days of anaerobic biodegradation were 46%, 32%, 54%, 41% and
10.5%, 8.3%, 18.7% and 8% for fresh waste, composted waste, newspaper waste and
synthetic waste respectively. Thus cellulose showed a better degradation trend than hemi-
Conclusions and future work
138
cellulose in all of these wastes. The lignin fraction after 150 days of biodegradation was
barely affected in all these waste samples. This indicated the low degradation of lignin and
relatively high degradation of cellulose and hemi-cellulose over time and thereby a mild
increase in the percentage of lignin after 150 days of biodegradation. Considering cellulose
and hemi-cellulose degradation, composted waste was found to be least biodegradable among
the four wastes under consideration. The low biodegradation of composted waste was as
expected and could be explained by the fact that most of the readily biodegradable cellulose
and hemi-cellulose in the original waste would already have been degraded during
composting. Lignin was known to significantly reduce the biodegradation primarily through
sheathing of cellulose and hemi-cellulose. The (Cellulose + Hemi-cellulose) to Lignin ratio,
(C+H)/L was also used to asses the extent of biodegradation of different wastes. After 150
days of anaerobic biodegradation, the (C+H)/L values of fresh waste, newspaper waste and
synthetic waste fell from 2.3, 2.81 and 3.5 to 1.23, 1.2 and 1.58 respectively, while the
(C+H)/L ratio of the composted waste fell from 0.9 to 0.56. This result again indicated the
low biodegradation potential of composted waste. The carbon mass balance suggested that
8.5% (fresh waste), 7.0% (composted waste), 10.5% (newspaper waste) and 15% (synthetic
waste) of utilized total carbon was transferred into the leachate and 39% (fresh waste), 15%
(composted waste), 22% (newspaper waste) and 48% (synthetic waste) of utilized total carbon
was transferred into the biogas by the end of 150-day monitoring period. Thus mass balance
analysis in general allowed us to rate the observed biodegradation in these wastes as synthetic
waste > fresh waste > newspaper waste > composted waste. It was interesting to note that the
lower mass transfer of newspaper waste into gas (22%) in comparison to the synthetic waste
(48%) and fresh waste (39%) was also reflected in the cumulative gas production after 150
days. The total produced gas was found to be highest for fresh waste and synthetic waste and
lowest for composted waste. The newspaper waste exhibited an intermediate gas production.
Leachates generated in the BMP test reactors were also taken periodically and investigations
were performed to determine the composition of the leachates generated during different
stages of waste biodegradation. Conventional method of characterisation showed that TOC
values increased from initial values of 1308 mg/l to 4625 mg/l after 35 days, 1813 mg/l after 5
days, 3888 mg/l after 25 days and 5468 mg/l after 22 days and then decreased gradually to
levels of 705 mg/l, 176 mg/l, 1408 mg/l and 1160 mg/l for fresh waste, composted waste,
Conclusions and future work
139
newspaper waste and synthetic waste respectively by the end of the test. A similar trend was
observed for dissolved organic carbon (DOC) and total chemical oxygen demand (TCOD)
values of the leachate samples in course of biodegradation. This initial increase was explained
by the transfer of the solid organic matter into leachates from waste although biodegradation
of organic materials in leachates was simultaneously active. Once waste biodegradation
reached certain levels, biodegradation of leachates dominated thereby reducing DOC, COD
values. However, it was observed that total organic carbon (TOC), DOC, chemical oxygen
demand (TCOD and DCOD) values for fresh waste, newspaper waste and synthetic waste
leachates were significantly higher than for composted waste leachates. In addition, the TOC
value in composted waste leachate reached its peak value within only 5 days. These results
indicated that, as expected there was a lower organic contents in composted waste leachates in
comparison to other leachates. It was also suggested that further removal of organic
compounds might be possible through biodegradation in fresh waste, newspaper waste and
synthetic waste leachates. It is worth mentioning that unlike the landfill samples the
biodegradability of these laboratory scale waste generated leachates was not estimated from
the remaining percentages of DOC and COD due the possibility of continuous mass transfer
from the wastes.
Fluorescence spectroscopy provided an excellent way to study recalcitrant materials in this
case by estimating humic-like (H-L) and fulvic-like (F-L) compounds which were known to
be the key components of the recalcitrant organic compounds. It was found that the
fluorescence intensities of humic-like (H-L) and fulvic-like (F-L) compounds in composted
waste and newspaper waste leachates were significantly higher than in fresh waste and
synthetic waste leachates. As expected, this indicated that composted waste and newspaper
waste leachates contained relatively higher concentrations of recalcitrant organic compounds
than fresh waste and synthetic waste leachates. The presence of high proportions of
recalcitrant organic compounds in composted waste leachate was explained by the fact that
composting significantly degraded readily biodegradable organic carbon and hence a
subsequent anaerobic biodegradation of the remaining organic matter produced more
biologically resistant compounds. For newspaper waste leachates, a high percentage of
cellulose and lignin could contribute to the generation of biologically resistant compounds. In
addition, a low mass transfer of newspaper waste into gas would also suggest recalcitrant
Conclusions and future work
140
material formation in leachates through subsequent biodegradation. Fluorescence
spectroscopy also revealed that humic-like (H-L) fluorophores of composted waste leachates
were in the long emission wavelengths thereby indicated that H-L fluorophores of this
leachate was more aromatic and higher molecular weight than those of the other leachates.
The biodegradability of the laboratory scale waste generated leachates was studied as well by
estimating the evolution of aromatic compounds using the specific UV absorbance (SUVA).
At the end of biodegradation, the SUVA values for fresh waste and synthetic waste leachates
were approximately 0.065 and 0.06 respectively. The SUVA of newspaper waste leachates
was about twice this value, whereas the SUVA of composted waste leachates was 30 times as
great implying that the molecular characteristics of the composted waste and newspaper waste
leachates were more aromatic in nature. The ratio of absorbance at 465 nm and 665 nm
(E4/E6) was also used to provide an indication of the molecular size of these leachates. It was
known that the E4/E6 ratio was inversely proportional to molecular weight, which implied that
the E4/E6 ratio decreased with increased molecular weight. After 150 days of biodegradation
the lowest values of E4/E6 ratio were observed for composted and newspaper waste leachates
thereby indicating the molecular complexity and the recalcitrant nature of the leachates
generated from the biodegradation of composted and newspaper wastes.
This aforementioned research supported the application of UV and fluorescence spectroscopy
for a rapid and on-site monitoring of landfill leachates that might help scientist and engineers
to assess leachate quality, identify some groups of organic compounds and to optimise
leachate treatment processes. The analyses performed on Pitsea and Rainham leachates were a
promising step towards developing a database of representative information for leachates
from different UK landfills. While a gradual reduction of organic compounds in course of a
biodegradation is generally expected, this research made a systematic and comparative
analysis on the degradation of organic compounds through conventional methods, aromaticity
of organic compounds through UV absorption and on the evolution, structural strengths of
some groups of organic compounds through fluorescence spectroscopy for leachates collected
from different phases of Pitsea and Rainham sites. This analysis revealed that organic
compounds in Pitsea (LTP), Pitsea (P4), Rainham (P2) leachates were more recalcitrant in
nature in comparison to Rainham (LTP), Rainham (FE) leachates and hence, it can be
unambiguously decided that Pitsea (LTP), Pitsea (P4), Rainham (P2) would require more
Conclusions and future work
141
rigorous treatment than Rainham (LTP), Rainham (FE) leachates to meet effluent standard.
This research also provided an understanding of the evolution of recalcitrant materials from
different wastes and it can be conclude that deposited wastes enriched with composted and
newspaper waste would initially tend to produce leachates with higher concentrations and
more condensed aromatic structures of recalcitrant organic compounds than fresh waste in a
landfill system until all readily degradable carbon was utilised.
6.2 Future work
This study shows a way of fingerprinting and estimating the constituents of recalcitrant
organic compounds using UV and fluorescence spectroscopy. An extensive application of
these techniques to study the impact of various pre-treatment processes that would help to
reduce the recalcitrant organic compounds in leachate could be done. This would give useful
information about the necessary treatments in reducing some of the components of
recalcitrant substances and also would help to devise appropriate treatment processes to be
applied at the landfill sites.
While the nature of recalcitrant organic compounds in landfill leachates could indicate the
necessary treatments, an interesting research topic would also be the characterisation of
leachates samples collected from different landfills of varying ages, composition of wastes
and sizes to develop a database of representative information of leachates’ chemical
composition of different UK landfills. In particular the research should target the investigation
of chemical and structural composition of humic, fulvic or any other compounds that could
potentially be detected by fluorescence spectroscopy to forecast the required treatment for any
particular landfill site.
In addition to the above mentioned research topics, the research could be further strengthened
by performing following experiments
• In this work, all the fluorescence spectroscopic analyses were carried out at constant
20oC temperature. However, the work should be extended at different temperatures to
provide additional chemical structural information of leachate DOM (Baker, 2005).
Conclusions and future work
142
This could permit the use of thermal fluorescence properties for leachate quality
monitoring in the treatment processes.
• Fourier Transform Infrared (FTIR) analyses of leachate samples needs to be done.
This analysis could serve as a qualitative monitoring of chemical groups and bands of
leachate DOM and provide more information about the structural changes of leachate
DOM evolution (Kang et al., 2002; Huo et al., 2008).
• Synchronous fluorescence spectra (SFS) could also be carried out to explore additional
structural information with improved resolution for leachate DOM which sometimes
quite difficult to interpret clearly in the EEM spectra (Chen et al., 2003; Zhang et al.,
2008).
• In the BMP test, compositional analyses of solid wastes were determined at the
beginning and at the end of the experiment. Further research is needed to evaluate the
compositional characteristics of wastes at the different stages of biodegradation
processes. This would help to evaluate the relationship between the concentration of
recalcitrant organic compounds in leachates and the degradation of waste components.
Appendix A: EEM spectra of landfill leachates
143
Appendix A EEM Spectra of Landfill Leachates
2020
20
20
20
2040
20
20
204020
20
20
20
20
40
40
4020
8060
20
40
404040
20
1601401201008060
40
60
40
60
40
1601401201008060
60
4040
60
80
22020018016014012010080
60
60
60
80
6040
60
1801601401201008040
80
1201008060
100
6040
40
80
140120100806040
60
100
1201008060
100
60
60
80
1801601401201008080
80
100100
100
140120100
80
60
40
60
4060
1201008060
20
40
80
60
60
1401201008060
100
40
40
80
20
60
40
1201008060
100
401201008060
40
20
40
60
1008060
4060
402012010080
40
120
60
20
40
60
100806040
40
20
806040
20
20
20
40
806040
20
40
20
4020
20
40
20
20
604020
20
202040
40
20
20
40202020
20
20
20
2020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
(2)
(1)(4)(4)(3)
Pitsea (LTP)Day 10
20
20
20
20
2040
2020
20
2020
20
20
40
20
6040
20
20
4060
20
20
40
40
40
100806040
60
20
20
20
40
40
100806040
40
40
40
80
60
4060
1201008060
60
40
40
80
60
60
60
1201008060
40
80100
80
60
220200180160140120
80
100
80
40
60
100806060
80
20018016014012010080
100
100
40
6080
100806060
8060
100
10080
80
40
80
8060
40
806040
40
1201008060
806040
60
60
60
80
1801601401201008040
60100
40
1201008060
60
20
40
40
60
80604040
40
120
1008060
40
20
40
80604020
100806040
20
20
20
6040
20
4020
20
20
20
20
80604020
20
2040
4020
2020
2020
20
4020
20
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (LTP)Day 20
(2)
(1)(4)
(4)(3)
80604020
2020
80604020120100
20
20806040
20
20100
20
201008060
40
20
40
20
20
40
8060
40
20
120100
2040
240
240
220
220
200
200
180
180
160
160
140
140
120
120
100
100
100
80
80
80
60
60
60
4020
4040
1008060
60
60
240
240
220
220
200
200
180
180
160
160
10080
20
40
140
140
120
120
100
100
80
80
60
60
40
40
20
20
20
340320300280
280
260
260 60
140120
4060
280
280
40
80
40
340340320
320
300
300
80
60 80
60
140120 100
100
100
120
6040
120
140
8060
140
80
6080
160
806080
80
40
120
20
60
300
4020
20
80
340320
40
320
100
80
80
60
60
340
60
280
80
340320300
60
260
40
60
40
340320300280
40
320
60
340
40
220
60
6080
40
340320300280260240
20
120
40
340320300280260240220200180160140
40
40
40
340320300280260240220200180160
20
40
300280260240220200180160140
20
20260240220200180160140120
20
80
40
40
20200180160140120100
20
60
40
20
1601401201008060
1008040
20
208060
20
806040
20
60402020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (P4)Day 10
(2)
(1)(4)(4)(3)
12010080604020
2020
20120100806040160140
20
20
20120100806040
20
140
20
201401201008060
40
20
6040
20
4020
120100806040
40
20
40
180160140
12010080
40
340
340
320
320
300
300
280
280
260
260
240
240
220
220
200
200
180
180
180
160
160
160
140
140
140
120
120
120
100
100
100
80
80
80
60
60
60
4020
40
40
220200180160140
40
340
340
320
320
300
300
280
280
260
260
240
240
220
220
120100
8060
60
200
200
180
180
160
160
140
140
120
120
100
100
80
80
60
60
40
40
40
20
20
20
60
180160140
8060100
100
60
60
80
180160140120
80
100
100
120
120
10080
140
100
80
80
160
120
140120200160
140
120
40
100
20
6060
4020
20
80
80
60
80
8080
60
40
6080
60
300
40
80
340320
60
200
60
40
340320300280260240220
40
60
40
40
340320300280260240
40
340320300280260240220200
20
20
4020
34032030028026024022020040
40
20
40
320300280260240220200180160
40
40
240220200180160140120
20
80
160140120100
20
60
20
4020
1601401201008020
120100806040
20
20
20
20
10080604040
20
6020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Pitsea (P4)Day 20
(2)
(1)(4)(4)(3)
80604020
20
4020
60
40208060
60
40
2040
604020
2020
20
208060
4020
20
20
4020
1008060
40
20
160140120
40
20
20
60
60
40
40
40
40
120100 80
20
20
20
20
20018016014012010080
60
200180160140
4040
20
60
20
20
100
60
18016014012010080100
240220200180160140120
20
20
40
220200180160140120
40
80
20
20018016014012010080
60
40
160140120100
60
40
100
60
40
160140120100
80
60
60
80
60
60
220200180160140120100
40
80
80
160140120100
6040
100120
40
100
20
200180160140120240220200180160140
120
120
6040
60
80
140
40
22020018016014012010080
160
60
60
180
40
20018016014012010080
200
60
240220
6080
20018016014012010080
80100
280260
60
80
320300340
20018016014012010080
120
140
80
160
200180160140120100
180
40
120
220200
2001801601401201008060
240
4040
260
1801601401201008060
160140120
6060
100
100
18016014012010080
80
6060
80
16014012010080
60
60
4060
160
40
1801601401201008060
40
40
606040
180
60
1601401201008060401401201008060
20
40
40806020
20
20
14012010080604020
201008060402020160140120100806040
20
60
20
208020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP)Day 10
(2)
(1)(4)(4)(3) 80604020
4020
4020
10080604020140120
20
12010080604020140
2020
20
20
20
20
20
120100806040
20
14020 40
40
1601401201008060
40240220200180
40
40
40
20
60
40
40
160140120100
80
80
20
20
20
20018016014012010080
60
260240220200180
4040
60
100
18016014012010080
20
60
20
220200180160140120
40
1801601401201008060
20
406040
20
140120100
80
20
100160140120
40
340320300280260240220200180
140120
40
40
160140120100
80
60
6020
6060
40
80
20
80
8040
60
200180160140120
100
100
40
200180160140120
40
120140
6040
60
220200180160100
60
220200180160140120
40
100
100
120
260240220200180160140120
40
60
60
160
160
140
14040
40
16014012010080
60
180
20
60
200
40
16014012010080
80
260240220
60
280340320300
6060
22020018016014012010080
100
120
60
140
100
16014012010080
160
100
80
180
200180160140120100
200
120100
4040
220
1601401201008060
120
60
60
80
100
220200180160140120100804060
60
60
8060
140120100804040
160
40
1401201008060
40
60606040
180
240220200180160140120100806060
20
2001801601401201008040
60
60
40
20018016014012010080
20
80
20
200180160140120100
20
6028026024022020018016014012010080
402001801601401201008060
20
60
10080
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP)Day 20
(2)
(1)(4)
(4)(3)
(a)
(b)
(c)
Figure A1 Excitation-emission matrices (EEM) for landfill leachates at day 10 and 20
during aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like) (dilution factor 25)
Appendix A: EEM spectra of landfill leachates
144
16014012010080604020
2020
1601401201008060402022020018020
20
2018016014012010080604020020
20
20
202001801601401201008060
40
40
40
2040
20
60
18016014012010080
60
40260240220200
408060
160
160
140
140
120
120
120
100
100
100
80
80
80
60
60
60
40
40
20
20
120100140
60
180
180
180
180
160
160
140
140
120
120
100
100
100
20
18016014012010080
80 40
80
80
80
60
60
60
40
40
40
20
20
20
580560540520500480460440420400380360340
320300
300
280
280
260
260
260
240
240
240
240
220
220
220
220
200
200
200
200
20
280260240220200
40
320
20
60
20
80
600580560540520500480460440420400380360
340
80
340
60
200180160140120100
600580560540520500480460440420400380
360380440420400
6040
40
620600580560540520500480460
40
80
380400
60
600580560540520500480460440420
120
80
320
100
60
600580560540520500480460440420400380360340280
40
600580560540520500480460440420
400380360340320300420
80
40
600580560540520500480460440
80
660640620
120100
60
640620600580560540520500480680660
100
40
620600580560540520500480460660640
60
20
260
120
40
600580560540520500480460440420400380360340320
300280
280
140
400480460440420
160
6060
620600580560540520500
180
40
300
60
200
60
600580560540520500480460440420400380
360340320
80
220
380
240
60
620600580560540520500480460440420400
280260
400
300
4060
620600580560540520500480460440420
100
320
10080
360600580560540520500480460440420400380
40
240
40
60
600580560540520500480460440420400380360340320300280260
80
60
60
4060
600580560540520500480460440420400380360340320
20
40
40
60
40038036034032030028026040
40
20
420400380360340320300280260240220200180160
20
20
40038036034032030028026024022020018016014020
40
120
4020440420400380360340320300280260240220200180160140
20
340320300280260240220200180160
4020
28026024022020018016014040
20
2020
2001801601401201008060
40
2018016014012010080604040
208060
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (FE)Day 10
(2)
(1)(4)(4)(3) 604020100
2020
10080604020140120
20
80604020100
20
20
20
20
20
160140120100806040
2020
40
40
20
20
20018016014012010080
60
40280260240220
6040
100
100
80
80
60
60
60
40
40
20
20
80
60
40
100
100
80
80
60
60
22020018016014012010080
8040
40
40
40
20
20
20
340320300280260240220200180160
140
120
120
20
340320300280260240
2040
140
4060
20
80
340320300280260240220200180160
20
160
40
20
240220200180160140120100
20
340320300280260240220200180
20
140
60
340320300280260240220200180160
20
160
40
100
120100
340320300280260240220200180
140
120
80
340320300280260240220200180
160
140
20
160
40
40
40
340320300280260240220200180
180
40
340320300280260240220200
40
200220
60
340320300280260240
6040
180
40
340320300280260240220200
100
40
120
40
340320300280260240220200180160140
120
120
40
40
60
140
340320300280260240220200180160140
160
40
120
60
180
340320300280260240220200180160140
200
80
220
4060
320300280260240220200180160140120100
260240
8060
160
80
280
80
340320300280260240220200180
300
40
100
80
280260240220200180160140120
60
6040
60
40
60
3002802602402202001801601401201008040
60
40
22020018016014012020
40
6040
20
28026024022020018016014012010080
20
40
40
20
220200180160140120
20
40
60
40 20
20
1801601401201008060
202202001801601401201008080
16014012010060
1801601401201008020
20160140120100806040
20
2018016014012010080604040
80602020
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (FE)Day 20
(2)
(1)(4)(4)(3)
20
20
20
20
20
20
20
20
20
20
20
40
40
20
20
20
20
20
604080
4020
40
604080
40
60
40
40
8060
40
20
40
60
20
20
1801601401201008060
60
60
60
100
60
4040
60
200180160140120100
80
60
80
80
80120
100
60
60
180160140120100
40
80
100
100
80
200180160140120100
60
80
120
120
60
200180160140120100
120
60
80
200180160140120100
80
140
80
100
6080
180160140120100
160
140
100
80
180160140120
180
100
80
80
18016014012010040
120
20
80
2001801601401206040
20
60
20018016014012010080
80
60
6040
80
20
60
18016014012010080
40
604020
100
80
60
16014012010080
60
404020
40
18016014012010080606040
60
60
20
40
14012010080
20
40
40
40
40
20
40
1401201008060
20
20
20
10080604020
40
20
20
80604020
20
20
604020206040
20
20
20
20604060402020
20
20404020
4020204020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (P2)Day 10
(2)
(1)(4)
(4)
40202040
20
202040
20
20
20
20
20
20
2040
20
20
40204040
20
40
40
60
40
40
120
120
100
100
80
80
60
60
4020
60
240220200180160
160
140
140
140
140
120
120
100
100
80
80
60
40
40
60
60
40
40
20
20
340320300280260
60
60
80
40
160
160
20
60
40
340320300280260
240
220200
200180
180
180 80
60
240
40
80100
340320300280260
100
260
80
60
340320300280
120
120
280
60
340320300
140
80
200
100
80
120
80
340320300280260240
220
160
100
220
80
340320300280260240
180
80
260
80
300340320280
200
140
100
240
120
340320300280260
1008080
20040
60
34032030028026024022020
80
18040
100
80
34032030028026024022020020
80
60
14040
60
20340320300280260240220200180160
40
16040
80
20
100
3403203002802602402202001801004020
60
60
60
280260240220200180160140120
40
14040
60
20
40
320300280260240220200180160
40
100
40 40
20
2040
26024022020018016014012040
20
2001801601401201008060
20
20
60
20
20
1601401201008040
140120100806040
202001801601401201008060
20
806020
206040
20
4020
20
402020
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (P2)Day 20
(2)
(1)(4)
(4)
12010080604020
20
80604020
12010020100806040
20
2020
14012020180160140120100806040
2040
20
20
20
120100806040
180160140
40
40
140
140
120
120
100
100
80
80
80
60
60
60
40
40
40
20
20
20
8060
20
140
140
120
120
100
100
20
1201008060
40
2040
80
80
60
60
40
40
40
20
20
20
340320300280260240220200
200180
180
180
180
160
160160
16020
180160140
220
60
340320300280260240
40
200
200
20
20
16014012010080
60
340320300280260240
220
220
220
40
60
340320300280
260240
240
60
14012010080
40
40
200180160
60
260
80
340320300
280
60
280
60
10080
340320
300
6040
300
40
80
40
340320
60
80
80
604040
300
60
100
340320280
40
340320300240
120
2040
340320300280260180
60
340320300280260240220200180
140
340320300280260240220200
60
60
240
80100
340320300280260
60
120
40
6040
340320300280260240220200180160140
40
260240220200180160
20
80
20
340320300280260240220200180160140120100
20
20
340320300280260240220200180160140
20
20300280260240220200180160140
20
20
22020018016014012060
20
28026024022020018016014012010080
20
160140120
20
6024022020018016014012010080
2012010080604040
80604020
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
((4)
Rainham (LTP Haz)Day 10
(2)
(3)(4) 40
20
2060402080
2020
604020
20
20
604020
20
8020
201008060
40
20
160140120
20
20
40
20
20
1008060
40
40
40
806040
20
140120
40
20
60
1401201008060
40
40
20
16014012010080
4040
1008060
40
20
20
60
8060
40
40
1401201008060
1201008060
200180160
40
40
60
1401201008060
40
40
40
80
10080
120100806040
40
40
120100806060
80
8014012010080
40
160140120100
60
20
60
10080604040
60
100
1201008060
8040
20
80120
80604020
6080
80
120100806040
6040
60
60
1201008040
60
80
100806020
604020
120100806040
40
40
20
6020
40
40
100806040
20
20
40
6040
20
20
20
6040
20
20
4020
120100806040
20
20
4020
14012010080604040
80601008060402020
10080604040
20
2010080604020
20
20
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Rainham (LTP Haz)Day 20
(4)
(2)
(3)(4)
(f)
(e)
(d)
Figure A1 (Contd.) Excitation-emission matrices (EEM) for landfill leachates at day 10
and 20 during aeration (Zone 1 H-L; 2 F-L; 3 and 4 protein like) (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
145
Appendix B EEM Spectra of leachates from BMP reactors
1
1
11
2131213
213
1
32154
1
1
1
1
32
432
654
1
1
2
4321
543
1
4321
1
2
1
322
387654
3
1
2
111
321
1
1
43
2
25432
1
5432543
2
1
2
2
6543
1
1
1
3
4321
4
432
5
1
1
76
12
5432
2
8
1
1
109
432
3
1211
2
1
313
5432
4
161514
2
1432
1
2
5432
2
321
1
1
1
3
1
321
654321
15432
1
15432
1
154321
65432543211
2
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 50
1212
2121
214312
432
1
43
1
1
321
1
432
4321
1
1
321
26543
1
2
32
1
1
1
432143
22
1
1
431
1
4321
1
5432
2
2
1
1
32
1
3
32
1
4
1
5
1
1
432
6
1
2
87
32
9
1
3
10
2
432
1211
321
1
4321
2
1
1
2
1
2
2154321
3
43211
1
4321
432154324321
12
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 60
(a) (b)
112
1212
2132
1
321
31
321
32
321
1
211
2543
2
1
2
1
1
1
321
1
321
1
321321
432
2
2
2121
1
34
321
56
1
21
2
7
1
8
32
109
21
32
321
1
1
1
21
2
2114321321321
132
43213211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 80
(d)
112
1212
21432
1
321
431
3211
32
32
1
211
26543
2
1
2
1
1
1
321
1
3212
1
431
1
4321432
2
2
21
1
3
21
45
4321
6
2
7
2
21
8
1
9
32
3
1110
3214321
1
1
1
1
2
2
2143213211
132
132
43213211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 70
(c)
Figure B1 3D EEM fluorescence spectra of control (blank) reactors at day (a) 50, (b) 60
and (c) 70 and (d) 80 showing Trp-L fluorescence at 220-240 nm and 270-280 nm
excitation and 340-370 nm emission (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
146
11
1213
21
21
31
21
32
21
1
211
5432
1
21
1
2
1
1
1
321321
321432
21
2
21
1
34
321
56
21
2
7
2
8
321
9
2
21
1
1
321
1
21
1
211
2
3213213213214321321
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 90
11
112
1
21
321
21
2
21
1
11
432112
1
121
212
1
132
21
2
21
3
21
1
45
1
6
12
21
2
87
1
21
1
21
1
1
1
13212121
2132121
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 100
11
211
21
21
1
21111
211
121
2
1
2121
2
21
3
21
1
4
21
56
21
2
7
1
1
1
111
1
11
1
11
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 120
1
11
211
21
21211
1
11
211
1221
1
2121
2
2121
1
34
21
5
1
6
2
1
1
1
111
1
11
1
11
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control reactorDay 140
(e) (f)
(g) (h)
Figure B1 (contd.) 3D EEM fluorescence spectra of control (blank) reactors at day (e)
90, (f) 100 and (g) 120 and (h) 140 showing Trp-L fluorescence at 220-240 nm and 270-
280 nm excitation and 340-370 nm emission (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
147
121212 1
1
121
1
1
132
1
54
1
1
1
1
1
1
32
432
54
1
1
1
1
1
2
432
2
1
1
2
2543
1
4321
3
4322
2
3
3
87654
2
1
1
432
2
1
2
2
43
2
2
2
1
543
4
2
3
1
6543
3
876541
65432
4
1
1
2
4321
2
5
432
6
276543
3
78
1
9
3
432
10
1
4
2
432
3
321
2
1
2
432
1
2
321
12
321
1
1
1
3
43214321
1
1
321
14325432143211
2
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 80
21
1
1
3213
21
1
1 3321
1
1
1
1
32
1
654
1
1
1
2
43
65432
765
1
1
1
2
3
54322
2
2
7654
1
65431
3
5432
1
2
2
3
3
4
987654
2
1
1
2
543
2
2
2
3
65432
2
6543
2
3
4
1
8765
2
4
3
2
4
4
98765
2
2
2
2
765431
5
2 3
65432
6
15432
3
7
2
89
9876543
4
10
26543
5
3
1
3
6543212
6
54321265432
2
2
12432
12
2
1
3
432
1
12
4
1
16543212
54321
1
14321
154321
65432543211
2
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
ngth
(nm
)
200
250
300
350
400
450
500Control uncorrectedFW day 70
1
1
1
1
11
1
2
1
1
2
11
1
2
1
1
22
2
1
1
3
1
1
1
1
1
3212
2
431
1
1
1132
1
1
1
3
1
1
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 70
1
1
1
1
11
2
2
1
1
1
2
1
22
1
1
2
3
1
1
1
1
1
1
1
321243
2
121
1
1
11
432
3
1
21
11
1
1
1
11
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 80
(c) (d)
(e) (f)
1
1
21121
1
1213
1
1
1
1
1
32
1
542
1
1
43
1
1
5432
652
22
3
4321
654
1
54321
3
3
432
1
2
2
3
4
2
987654
4
2
1
4
4
1
2
432
2
1
3
543
2
2
3
2
6543
4
4 2
3
4
2
7654987652
2
3
65431
5
54321
6
432
2
7
9876543
89
2
4
10
543
3
1
2
5432
5
12
3
4321
12
2
54321
2
12
1
321
2
3
32
1
12
1
1
54321
1432
1432
1
14321
543243211
2
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedFW day 60
1
1
1
1
1
2
1
1
1
1
2
2
22
1
1
1
3
132
1
1
1
2
1
4
1
12
1
1
1
322
3
54321
1
2
21
3
11
543212
4
1
1
1
2
1
1
1
1
1
1
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 60
(a) (b)
Figure B2 3D EEM fluorescence spectra of fresh waste (FW) leachates at day 60, 70 and
80 for control uncorrected and corrected samples (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
148
1
1
1
1
1
1
1
1
2
2
2 2
1
1
1
11
3
1
321
11
2
1
1
2
1
1
1
1
322
5431
2
21
1
1
11
3
1
4321
1
2
2
1
1
4
1
2
1
1
1
1
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedFW day 90
121221
1
121
1
132
1
1
54
1
1
1
1
1
32
432
54
1
1
1
1
1
2
432
2
2
1
2543
11
4321
3
32
3
2
2
387654
2
1
1
432
2
1
2
2
43
2
4
2
2
543
2
2
3
3
1
65434
3
87651
2
5432
4
1
14321
2
2
5
432
6
2876543
3
7
1
98
3
432
10
1
4
2
432
3
321
2
1
2
432
1
2
21
12
321
1
1
1
2
43213211
132
132
43213211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 90
1121
1
12211
1
1
2
1
43
1
1
1
32
1
1
432
1
1 2
2
54
1
1
1
2
321
2543
1
432
3
1
1
3
322
2
37654
4
2
1
1
2
32
2
1
3
43
2
2
2
43
2
2
3
1
654
2
348765
1
4
3
5432
1
1
2
4321
2
5
2
32
3
2
6
876543
7
1
8
3
432
9
2
1
4
10
2
432
1211
3
2
2
3211
432
1
2
21
12
121
1
3211321
2
2132143213211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 100
11
1
1
11
1
1
1
2
1 2
1
2
1
2
1
1
3
11
32
1
4
1
12
2
2
1
1
1
3
2
1
1
2
2
1
4323654
1
3
1
1
21
2
3
11
54321
1
2
2
1
1
4
1
2
1
1
1
1
2
1
1
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 100
(g) (h)
(i) (j)
Figure B2 (contd.) 3D EEM fluorescence spectra of fresh waste (FW) leachates at day 90
and 100 for control uncorrected and corrected samples (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
149
1
1
121
1
3211
1
2
11
1
43
1 1
2
2
2
1
1
21
1
3
3
1
2
2
2
1
1
1
13121110987654
3
4
2
2
141312111098765
2
3
3
1098765
4
4
2
5
2
131211109876
3
5
2
3
14131211109876
4
1
4
3
3
14131211109876
5
4
2
5
3
4
4
2
4
17161514131211109876
3
5
2
1211109876
2
6
2
3
11109875
2
2
11109876
1
4
3
3
121110987653
2
2
10987654
2
1
3
2
2
987654
1
3
2
10987654
2
3
1
1
987654
1
21
1
4876531
17654321
4321
15432
1
4321
32
1
3211
132321
1
43211
111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedFW day 120
1
1
1221
1
321
11
1
2
11
1
431
1
2
1
3
3
2
2
1
1
21
1
1
1
2
2
2
1
1
10987654
43
4
2
2
2
1098765
3
3
1098765
4
4
2
2
5
3
3
2
109876
3
5
2
3
109876
54
1
4
4
3
109876
5
4
2
5
3
4
2
4
109876
3
6
32
10987
2
6
4
32
3
109875
22
1098764
3
3
1098765
1
3
2
10987654
2
2
1
3
2
2
987654
3
1
3
2
10987654
2
3
1
2
1
9876542
1
1
98765431
17654321
54321
1
154321
15432
1
32
1
132
1
1432321
1
43211
111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedFW day 140
21
1
122131
1
1
4321
1
1
32
11
54
2
3
3
2
2
1
1
1
1
2
2
11
1
1
1
3
3
2
2
32
1
1
15141312111098765
4
4
54
5
2
32
3
15141312111098764
3
2
654
12111098765
2
5
3
3
3
3
141312111098766
4
3
151413121110987
5
21
4
4
4
16151413121110987
65
5
2
6
3
5
2
19181716151413121110987
3
6
3
14131211109877
4
32
3
13121110986
22
131211109875
3
4
1514131211109876
1
4
222
12111098765
1
3
2
2
10987654
5
1
4121110987654
3
2
6
12
111098765
2
3
2
987654
1
29876543
2
1
6543
1
1
2
1
1
65432
2
16543
2
1
15431
11
154321
1
14321
432
1
1154323211
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedFW day 120
21
1
212131
1
1
432
1
21
1
1
32
11
54
2
3
3
2
2
2
1
1
1
2
1
1
2
111
3
3
2
2
3
2
1
1
1098765
4
4
54
5
2
2
32
3
1098764
3
2
654
109876
56
3
32
3
3
4
3
109876
4 3
10987
4
65
21
4
4
10987
65
5
5
2
6
3
10987
3
6
3
5
109877
3
4
3
1098
2
6
232
109875109876
1
4
3
22
4
2
1098765
1
3
2
10987654
5
1
4
2
10987654
3
2
6
12
1098765
2
3
2
10987654
6
1
2
5
98765432
1
76543
1
2
2
1
1
765432
176543
2
1
1
5431
1
1
154321
1
15432
1
432
1
1154323211
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedFW day 140
(k) (l)
(m) (n)
Figure B2 (contd.) 3D EEM fluorescence spectra of fresh waste (FW) leachates at day
120 and 140 for control uncorrected and corrected samples (dilution factor 25)
Appendix B: EEM spectra of leachates from BMP reactors
150
1213
1
1
1
321
1
1
1
4321
2
1
5
1
1
1
2543
3
1
876
2
3
1
765
4
4
1
432
1
1
1098
11
1
1
5
4321
9876
22
4321
6
121
322
2
765431
22
32
2
1
2
5
2
2
43
2
2
4
543
1
2
2
5
5
543
32
2
2
543
3
6
1
7
2
654321
8
2
432
9
1
3
10
4321
4
54321
5
4 34
432
6
27654312
5
5432
4
3
2
5
3
65431
3
2
32
2
1
3
2
1
432
1
2
6543211
2
1765432
2
1
65432
654329876543
1
7654321
32
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 60
12213
1
1
1
213
1
1
1
1
4321
2
1
1
43
3
2
1
765
21
3
176
5
4
4
1
321
1098
11
5
321
876
22
321
1
5
1
21
6
2
12
2
543
2
5
1
2
1
1
1
2
321
2
2
1
431
5
1
4
32
2
1
5
2
32
6
1
1
432
3
3
7
1
2
8
2
2
32
9
1
10
321
3
4
12
4321
3
32
5
12
4
654312
43
432
2
1
3
3
54321
2
52
3
2
212
1
2
32
1
432
2
1
1
1
654321
1
54321
654321
987654321
654321
32
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 80
121
1
1
1
213
1
1
1
1
321
24
1
1
4
3
2
1
7653
21
176
5
4
4
1
21
1098
11
5
21
876
22
21
1
5
1
21
6
2
12
2
543
21
2
1
1
5
2
321
2
1
4
4321
5
51
2
32
2
1
1
2
1
32
6
1
3
432
3
7
1
3
8
12
2
910
3
321
1
4
1
432
2
1
33
3
2
5
12
4
6543
43
432
2
1
3
4
4321
42
3
2
1
2
1
32
1
43211165432
1
1
1
54321
54321
98765432165432
132
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 90
1
12 1
1
1
1
22
1
32
1
14
3
3
765
1
1
4
1
1
5
11
4
1
2
3
1
211
45
2
678
1
1 109
1
3
121
132
1
111
2
22
1132121
211432321
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 60
1
12 1
1
1
1
22
1
32
1
14
3
3765
1
1
4
1
1
5
1
4
11
1
1
2
1
3
1
3
211
45
2
678
1
1109
1
3
21
132
2
1
2
1
21
11
321212
114323211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 80
1
1
2 1
1
1
22
1
32
1
154
3
3
876
1
1
1
1
4
1
1
5
21
4
1
54
2
1
4
3
211
45
2
6
2
789
1
110
3
12
1
132
2
11
2
1
21
1
1132121
211
54323211
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 90
(a) (b)
(c) (d)
(e) (f)
Figure B3 3D EEM fluorescence spectra of composted waste (CW) leachates at day 60,
80 and 90 for control uncorrected and corrected samples (dilution factor 50)
Appendix B: EEM spectra of leachates from BMP reactors
151
1
1
1
1
1
1
2123
1
1
1
32
2
1
42
1
1
3
3
1
654
4
1
987
1
5 6
2
5
11
2
6
2
2
4
1
12
5
21
3
6
3
3
3
7
2
8910
1
1121113
1
14
132
1
33
1321
1
2
2
11
21
1
132
1
1
21
32165432321
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 100
11
1
1
21
1
1
1
22
1
32
1
1
33
354
1
9876
1
4
1
1
5
2
5
11
2
2
4
2
1
1221
5
3
6
2
3
7
2
89
1
110
12
12
2
1
1
1
1
2
1
1
12
1
1321
212
11
54323211
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 120
11
1
1
21
1
1
1
22
1
32
1
1
3 33
54
1
876
1
1
1
5
2
4
1
5
1
2
21
4
1
1
2
1
1221
5
3
6
2
7
2
3
8
2
3
91 10
12
1
2
2
1
2
1
1
12
1
1321
212
11
543221
1
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedCW day 140
1
11
1
1
1
1
1
3212
1
12
2
1
3
3
1
543
21
1
11
7655
44
4
1
1098
1
5
1
876
22
1
6
1
1
1
2
2
1
2
1
5
11
2
32
3
1
1
3221
21
67
21
89
2
21
10
2
3
1
112
21
12
4
151413
1
21
1617
3 22
21
5
18
1
1
2
2543
4
32
1
1
4
3
2
321
1
3
1
2
1
1
1
22
1
1
21
1
54321
321
43216543254321
12
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 100
1
1
1
1
1
1
1
1
1
3212
1
12
2
1
3
3
1
543
21
1
1
1
654
44
1
1
10987
1
5
1
76
22
1
6
1
1
2
2
1
2
1
5
11
2
2
3
1
1
3221
1
67
1
8
2
1
9
2
3
1
10
1
4
1
321
11
5
432
4
21
3
1
21
3
2
1
11
115432
1
321
4321654324321
12
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 120
1
1
1
1
1
1
1
1
1
3212
1
12
2
1
3
3
1
543
21
1
1
1
1
655
44
4
1
10987
1
5
76
2
1
2
6
1
2
2
1
2
1
5
112
3
1
1
3221
1
6
1
7
1
1 8
2
1
9
2
3
1
10
1
1
4
1
32
1432
21
21
3
3
21
2
1
1
21
1432
132
1321
6543243211
2
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedCW day 140
(g) (h)
(i) (j)
(k) (l)
Figure B3 (contd.) 3D EEM fluorescence spectra of composted waste (CW) leachates at
day 100, 120 and 140 for control uncorrected and corrected samples (dilution factor 50)
Appendix B: EEM spectra of leachates from BMP reactors
152
1
1
112
21213
1
1
2
2
1
1
1
11
1
2
1
1
1
9876543
3
3
1
1
2
109876542
1
32
98765433
2
987654
2
1
4
2
1098765
3
1
2
2
12
10987654
3
3
3
3
121110987654
2
4
32
3
987654
2
987653
1
2
1
9876543
2
9876542
12
876543
1
2
1
76543
3
2
2
8765432
31
765432
2
6543
1
1
54321
4321
4321
1
14321132
1
1321321
2132111
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedNW day 70
1
1
112
21
1
21
13
1
2
2
1
1
1
1
1
1
1
2
2
2
1
1
1
109876543
3
3
3
2
1
2
11109876543
23
987654
2
2
4
2
2
1098765
1
4
32
111098765
4
1
3
3
12
111098765
4
4
3
33
1312111098765
2
4
2
987655
2
98764
2
2
98765
1
3
2
109876542
2
9876543
1
2765432
1
98765433
21
87654276543
1
2
1
654321
1
4321
1
432
1
1
1
432
1
1132
1
1321
1
132321321211
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedNW day 90
1
1
112
1
321
1
21
1
3
1
2
2
1
1
111
1
2
2
2
1
1
1110987654
3
3
4
2
1
121110987653
2
2
2
9876544
2
2
2
111098765
1
5
2
3
3
1211109876
4
1
3
3
12
1312111098765
4
2
4
3
151413121110987655
2
4
11109876
4
5
2
1098764
3
3
121
10987654
2
1110987653
2
2
12
2
987654
1
2876543
3
1
2
1
987654
1
3
1
1
87654
1
2
11
765431
654321
4321
4321
4321
2
1
2132121
1
321
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedNW day 100
1
1
12213211
1
1
132
11
54
1
2
2
1
1
1
1
11
2
2
3 2
1
1
11109876543
3
54
3
2
1
1
2
11109876543
26543
2
10987654
2
4
2
3
2
1110987654
2
111098765
4
1
3
3
2
12111098765
4
4
4
4
3
3
141312111098765
2
4
2
3
10987655
109876
2
4
2
1110987653
22
2
1110987654
1
3
2
9876542
2
9876543
5
1
310987654
6
3
7
2
987654
8
2
3
9
876543
1
1
2
76543226543
1
1
2
5432
1
1
2
165432
1
15432
1
4321
132
143243213211
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedNW day 70
1
1
1221321
1
321
1
4
1
2
2
1
1
1
1
1
1
2
2
32
1
1
11109876543
3
4
3
2
1
2
1211109876543
43
987654
2
4
2
2
2
3
1110987655
3
2
2
1211109876
4
1
3
3
2
12111098765
4
3
41413121110987654
2
10987655
3
3
1098764
22
10987653
2
1
2
11109876543
2
987654
1
2
2
876543
4
1
3987654
2
3
5
1
9876542
2
876543
6
1
1
5
7654321
1
5432
1
1
1
1
21
54321
54321
14321
1
14321
1
1243
1432
43213211
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedNW day 90
1
1
1221
1
3211
1
1
132
11
1
4
1
2
2
1
1
2
1
3
3
1
11
3
3
2
2
32
1
1
1
13121110987654
4
4
4
2
2
141312111098765
2
4
543
111098765
2
2
5
3
3
131211109876
3
5
2
3
14131211109876
54
1
4
3
1514131211109876
5
4
2
5
3
4
4
2
4
1817161514131211109876
3
5
2
1211109876
2
6
2
3
1211109875
22
12111098764
3
1312111098765
1
3
22
3
1110987654
1
3
2
2
987654
2
4
1
31110987654
2
3
1
12
9876542
21
98765431
3
876543226543
1
1
1
1
1
654322543
1
1
1432
1
1432
1
1
1432
1
32
1
4321211
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedNW day 100
(a)
(c) (d)
(e) (f)
(b)
Figure B4 3D EEM fluorescence spectra of newspaper waste (NW) leachates at day 70,
90 and 100 for control uncorrected and corrected samples (dilution factor 40)
Appendix B: EEM spectra of leachates from BMP reactors
153
21
1
122131
1
1
4321
1
1
32
11
54
2
3
3
2
2
1
1
1
1
2
2
11
1
1
1
3
3
2
2
32
1
1
15141312111098765
4
4
54
5
2
32
3
15141312111098764
3
2
654
12111098765
2
5
3
3
3
3
141312111098766
4
3
151413121110987
5
21
4
4
4
16151413121110987
65
5
2
6
3
5
2
19181716151413121110987
3
6
3
14131211109877
4
32
3
13121110986
22
131211109875
3
4
1514131211109876
1
4
222
12111098765
1
3
2
2
10987654
5
1
4121110987654
3
2
6
12
111098765
2
3
2
987654
1
29876543
2
1
6543
1
1
2
1
1
65432
2
16543
2
1
15431
11
154321
1
14321
432
1
1154323211
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedNW day 120
213
1
212131
1
1
1
432
1
2
1
1
1
32
11
654
2
2
3
3
2
2
1
1
2
1
4
4
111
43
3
243
2
2
1
1
109876
5
65
5
2
2
32
3
1098765
3465
3
109876
3
6
4
3
2
3
4
10987
2
6
3
3
10987
4
65
21
4
4
2
10987
65
5
5
6
3
54
10987
5
3
7
53
10987
3
4
3
10986
3
232
4
2
109875109876
1
4
3
2
32
4
2
10987654
2
1098765
5
1
4
2
1098765
3
4
2
3
1
12
2
1098765
2
3
1
109876542
5
98765432
1
76543
1
2
1
1
765432
176543
2
2
1
1
1
5431
1
154321
1
15432
1
432
1
1154323211
111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedNW day 140
1
1
11
21
11
1
1
1
21
1
1
2
1
2
2
1
1
1
3
3
1
2
2
2
1
1
1
1
1211109876544
21
2
1312111098765
2
3
3
10987654
2
5
2
2
131211109876
3
4
3
12111098765
4
1
3
3
3
1
14131211109876
54
2
5
3
4
4
2
17161514131211109876
321
5
23
111098765
2
111098764
2
111098765
2
4
3
3
12111098765
1
3
2
2
3
10987654
2
1
3
2
987654
1
3
1
1
2
2
10987654
1
2
3
1
987654
1
1
2
1
8765431
1
7654321
4321
54321
54321
2
1
121
2121
1
1
11
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedNW day 120
1
1
122111
1
3211
1
1
1
32
11
4
3
3
2
2
1
1
1
1
21
1
1
1
1
3
3
2
2
2
1
1
10987654
4
43
5
2
32
1098764
2
2
43
1098765
2
2
5
3
3
3
1098766
3
2
4
3
3
10987
4
5
1
4
4
109876
5
2
5
32
109876
3
6
3
5
10987
5
2
6
3
4
3
10987
2
5
4
232
1098765
3
109876
1
4
3
22
1098765
1
3
3
2
2
2
2
987654
2
3
1
310987654
2
4
1
2
2
1
10987653
1
1
1
9876541
187654322
5431
1
154322
1
15431
1432
1
132
1
1
14321
32
1
43211
111
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedNW day 140
(g) (h)
(i) (j)
Figure B4 (contd.) 3D EEM fluorescence spectra of newspaper waste (NW) leachates at
day 120 and 140 for control uncorrected and corrected samples (dilution factor 40)
Appendix B: EEM spectra of leachates from BMP reactors
154
1
1
1
321
1
21
12
1323
1
321
12
11
11
1
32
2654
32123
1
2
432
2323
21
765
2
1
21
1
1
2
3
2
321
1
76543
432
32
154
1
34
32
3
4
32
1
1
5
4
32
2
765
44
2
1
1
109876
2
2
2
3
32
542
1
5
32
7654 3
1
5
3
633
32
32
1
2
4
4
321
3
454
32
2
1
2
2
44
432
321
1
3
2
32
54376
1
12
321
3
1
3
1
32
2
1
1
2
1
1
2
1
22
5432
1
1
21
1
21
12
1
11
21
32
1
3
98765432
1
1
2
21
121
1
1
1
2121
1
1
1
54321
1
1
1432
11
1
21
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control correctedSW day 70
2
1 1
1
32
321
42
133
4
42143
1321
43213
4
11
1432
1
1
1
1
4323
321
1 654
544
2
2
2
3
1
1
3542
2
2
21
1
1
1
1
11109876543
3
3
543
3
1
12
111098765
4
1
4
1
98765432
324
2
111098765
1
4
3
222
12111098765
3
71098
1
876543
54
20191817161514131211
14131211109
3
6
2
3
111098765
4
21
4
1
9876543
11109876
5
2
5
5
4
1211109876
365
7
3
98
1110987354
5
1716151413121110
3
4
5
5
1211109876
6
3
654
121110987654
4
43
3
2
5
3
8765
1211109876543
4321
1
1
1110987654
3
3
2
3
431
121110987654
1
4
1
2
2
5
12111098765
1
4
1
432
1
12111098765
1
3
1
132
131211109876543
1
1
121110987654
2
4
1
121110987653
1
110987654
21
310987654
432
2
12
1876543
1
2
32
76543
1
21543
1
11432
3214321
4321
21
1
1
23
2
541
21
122
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedSW day 90
432
21
6
21
1
543
1
4321
53
244
65
65321354
154321
654321
1
4
6
121
654321
11
1
1
432
1
4
432198765
65
1
5
2
2
2
1
4
12
4
3
3
3
16532
1
2
3
3
3
2
2
2
432
12
1
1
1
10987654
4
4
765
4
221
2
10987
6
5
1
6
109876543
43 5
3 3
10987
1
21
6
254
1
3
3
10987
4
910
10987654
765
1
4 1
1
3
4
109876
5
2
21
5
21098765
4
1098
7
67
3
56
56
1098
487
4
109546
7
5
65
7
4
7
10985
65487
56
6
109876
5
54
5
3
3
6
109876
57
109876
2
4
6654
321
2
5
1098765
4
4
45
21
1
5
10987656
3
4
6
4
10987
5
21
5
65433
109876
21
4
45
1098765
1
4321
4
3
2
1
31
1098765
2
5
1
12
109876
1
5
1
2
109876
1
321
4
1
1098765
1
543
3
213
1
1
10987654
321
3
1
54
987654
12
3187654
21
1
765432
43215
1
32
1
65432 1
12
2
243
43
6522
1
1132
23
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedSW day 100
21
1
4
1
1
3214
21
11
43
21
321
14
1
432
1
3
1
65432
12
1
1 2
1
65432
2987
432
3
343
2
3
1
1
1
6543
3
2432
1
1
1
3
65432
1
1
109873
4
1
43
32
4
3
6543
2
2
2
3
1098765
543
6543
265
2
4
432
45
4
543
2
1
3
57
6 1
2
987654
98
1
6
543
2
2
10987
3
5
4
654
3
3
3
532
3654
765
4
3
5
6743
6
4
7654
32
3
4
76542
4
6
65
23
8765432
5
6
76543
7
4 321
2
6
2
6543
7
6587
1
5
8
3213
765432
9
2
10
213
6543
4
2
3
6543
4
21
18765432
2
32
1
3
32
65432
2
1
1
2
5432
1
1
33
1
3
2
5432
43
2
1
322
3
1098765433
3
12
1654
2
1
22
1
5432
1
1
1
1
11
654321
765432
1
1
1
1
98765432
1
216543
1
1
1
121
1
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedSW day 70
32
21
6
1
1
43
1
4321
4323
3
65
542134
154321
54321
1
4
6
11
54321
1
6
1
1
432
1
3
543219876
54
2
4
2
2
2
1
1
1
3
1
3
3
3
15432
1
2
3
3
2
2
2
432
12
1
1
1
13121110987654
4
765
4
2
22
1
23
14131211109876
5
1
5
1211109876543
354
3 2
14131211109876
1
321
5
2
4
11
3
2
14131211109876
3
81211109
987654
654
2019181716151413
16151413121110
4
3
4
131211109876
5
2
21
5
2
10987654
45
13121110987
6
6
54
5
151413121110987
476
91110
3
1211109845
6
201918171615141312
5
6
4
4
65
1514131211109875
654
6
55
5
161514131211109876
4
4
876955
1413121110987654
2
5
543215
1312111098765
6
3
47
21
15141312111098765
45
3
8
6
2
14131211109876
9
32
5
3
2
656
1514131211109876
10
321
3 151413121110987654
4321
4
321
15141312111098765
1
4
321
1
3
1413121110987654
12111098765
321
3
2
13121110987654
354
3
23
1
2
10987654
321
3
1
4
987654
12
3187654
1
1
18765432
3214
1
21
1
765432
2
1
2
23
43
5421
21
122
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
ngth
(nm
)
200
250
300
350
400
450
500Control uncorrectedSW day 90
4321
21
7
211
543
54321
6543
24
4
876
1
65321465
1654321
1
6543215
7
11
21
11765432
1
11
8
2
1
1
543
2
1
1
4
65432
10987
1
765
2
6
3
3
3
2
2
1
4
4
4
4
16543
1
4
4
3
3
3
5432
2
2
2
2
1
1
1
1098765
5
109876
5
321
3
10987
6
2
6
3
1
10987654
4365
11
3
10987
21 3
6
254
4
3
10987
24
31098765
765
5
2
8
4
4
10987
6
321
6
210987
65
56
109
8
78
23
75
6
109
2
5987
5
1
10576
1
3
7
4
6
8
5
56
109
8
6
776587
6
6
6
10987
6
5
5
71098
6
7
109876
7
5
3
765 4321
109876
8
4
46321
7
109876
5
9
7
7
4
10
86
1098
32
6
4877
10987
432
4
2
1098765
2
5432
5
3321
109876
1
6
2
2
423
10987
1
1
5
21
43
109876
2
4321
4
11098765
2
47653
4
2
32
1
3
3
1098765
421
3
2
65
10987654
21
3
1987654
21
29876543
43215
11
32
1
98765432
32
12
3
2
3
1
4
54
7652
31
1132
23
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedSW day 100
(a) (b)
(c)
(e)
(d)
(f)
Figure B5 3D EEM fluorescence spectra of synthetic waste (SW) leachates at day 70, 90
and 100 for control uncorrected and corrected samples (dilution factor 10)
Appendix B: EEM spectra of leachates from BMP reactors
155
54321
21
8
211
543
654321
3654
25
4
987
1
76321 465
1
1
1765432
76543215
8
21
1
21
1
11
8765432
21121
6543
2
121
5
76543212111098
76
2
6
3
3
3
2
2
1
4
2
5
4
4
4
17643
2
4
4
4
3
3
3
6543
2
2
2
2
1
1
1
19181716151413121110987
6
5
5
1110987
3
3
6
3
321
3
1918171615141312111098
7
2
7
31716151413121110987654
54376
1
1
3
1918171615141312111098
4321
7
1
65
1
4
3
1918171615141312111098
54
1216151413
1312111098765
876
20191817
2
20191817161514
5
4
5
1918171615141312111098
76
6
321
7
2
15141312111098765
2
19181716151413121110
9
8
3
3
9
76
77
2019181716151413121110
4
5109
8
12
5
151413
6
171615141312115768
2019181716
76
9
6
8
2019181716151413121110
7
7
5
86598
77
7
201918171615141312111098
65
6
4
76
1211109878
19181716151413121110987
8
5
7
87654321
3
191817161514131211109876
89
5
5
56
3217
2019181716151413121110987
6
10
8
7
4
11
9
201918171615141312111098
12
432
7
4
13
9878
201918171615141312111098
14
432
520191817161514131211109876
65432
6
4
2
321
2019181716151413121110987
5
6
2
4
1
23
2
4
2019181716151413121110987
1
1
6
211716151413121110987
4321
5
11817161514131211109876
2
47653
4
324
3
1413121110987651
4321
4
2
65
12111098765
213
4
11098765
21
2109876543
543216
11
432
1
98765432
32
12
3
214
4
54
76532
321
1
12
3 23
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control uncorrectedSW day 120
654321
21
9
321
113
654
2
654321
8765432
65
987
1
874321
1
5876
1
1
1
1765432
876543216
98
21
1221
1
1
1
98765432
1
212
12
76543
121
6
765432
21312111098
987
2
8
4
4
4
3
3
3
2
2
1
1
5
2
6
5
5
5
1987543
2
5
5
5
4
4
4
3
3
3
6543
23
3
2
2
2
1
1
1
201918171615141312111098
7
6
6
1110987
7
3
1
3
321
1
44
3
2019181716151413121110
9
8
2
9
1
201918171615141312111098765
654
87
4 4
2019181716151413121110
5321
9
1
376
4
34
2019181716151413121110
1
65
141918171615
2
15141312111098765
9879
20
2019181716
10
6
2
5
5
20191817161514131211109
87
7
3
4321
8
3
1817161514131211109876
78
20191817161514131211
10
910
4
897
89
20191817161514131211
6121110
14
6
18171615
5
7
20191817161514137698
9
2019
87
10
7
10
201918171615141312117
97611109
88
8
201918171615141312111098
98
769
7
4
9
7
1514131211108
2019181716151413121110986
8
9
2
3
987654321
1
2019181716151413121110987
9
6
6
687
321
9
201918171615141312111098
7
94
9
3
5
10
87
106
2019181716151413121110
3
11
432
8
5
12
9877
20191817161514131211109
432
6
9
9
2019181716151413121110987
7654321
7
4
2
321
201918171615141312111098
5
7
22
4
1
213
4
201918171615141312111098
1
7
3
201918171615141312111098
54321
6
12019181716151413121110987
2
48765
5
2
3254
3
1514131211109876
4321
4
2
765
1312111098765
123
4
1111098765
21
211
9876543
1
6543217
11
5432
1
9876543
2
432
132
3
2
4
1
5
654
987633
321
21
12
4 34
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500
Control uncorrectedSW day 140
432
21
7
21
1
543
1
54321
65432
44
876
65321465
11
165432
76543215
87
11
21
11765432
1
11
8
2
1
1
6543
2
1
4
6543211110987
1
765
2
6
3
3
3
2
2
2
1
4
4
4
4
176543
1
4
4
3
3
3
54 32
2
2
2
2
1
1
1
171615141312111098765
5
9876
5
321
3
1716151413121110987
6
21
6
2
1514131211109876543
54376
3
1716151413121110987
1
321
6
1
254
3
3
1716151413121110987
1
54
111615141312
121110987654
87 6
1
7
20191817
2019181716151413
5
2
8
4
4
1716151413121110987
6
321
6
2
141312111098765
17161514131211109
8
7
3
8
675
67
191817161514131211109
598
11
5
141312
16151413121110576
1
1
3
7
201918171615
4
6
8
5
8
1918171615141312111096
75987
6
66
2019181716151413121110987
6
57
5
7
5
111098127
171615141312111098765
3
7654321
171615141312111098764
2
4
2
65 21
6
20191817161514131211109876
57
2
35 7
18171615141312111098
2
7
32
6
1
476544
19181716151413121110987
321
4
6
6
1918171615141312111098765
74321
5
321
191817161514131211109876
1
32
4
6
21
2
1817161514131211109875
32
1514131211109876
4321
4171615141312111098765
76543
4
213
1
2
12111098765
321
3
1
654
1110987654
21
4198765
21
29876543
543216
11
432
1
8765432
32
12
2
214
4
543
76532
31
1
12
3 23
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedSW day 120
5432
321
8
21
1
654
1
54321
76543
25
5
9876
8764321576
1
1765432
876543216
8
2221
1
21
1
11
8765432
2112
1
1
7654
32
1
6
1165432
1110987
1
987
2
7
4
4
4
3
3
3
2
2
2
1
1
5
1876543
1
4
4
4
3
3
3
5432
213
3
2
2
2
1
1
1
20191817161514131211109876
6
5
5
5
5
9876
6
33
333
321
3
20191817161514131211109
8
7
1
21
8
1
3171615141312111098765
4
54 876
4
20191817161514131211109
321
8
43765
1
4
34
20191817161514131211109
65
131817161514
2
15141312111098765
987
2019
2019181716
10
6
5
5
20191817161514131211109
87
7
3
3214
8
2
17161514131211109876
3
2019181716151413121110
910
4
7986
88
20191817161514131211
611109
13
6
17161514
1
2019181716151413126987
1
9
201918
5
87
9
6
10
20191817161514131211107
897611109
78
8
201918171615141312111098
798
76
6
9
6
1413121110158
7
20191817161514131211109876
98765
4321
3
2019181716151413121110987
7
5
576
321
2
6
2019181716151413121110987
6
8
8
4
65
9
20191817161514131211109
2
321
7
1
87654
5
4
201918171615141312111098
321
5
766
20191817161514131211109876
54321
6
3
1
1
4321
2019181716151413121110987
2
54
7
1
2
2
1
13
1
2019181716151413121110986
1
32
19181716151413121110987
54321
5
1
191817161514131211109876
7654
4
43214
1
2
141312111098765
4321
4
21
1
765
112111098765
213
4
11098765
21
211
9876543
1
6543217
11
5432
1
98765432
432
132
3
2
4
1
5
54
987633
321
1
12
43 23
4
Emission wavelength (nm)200 250 300 350 400 450 500 550 600
Exci
tatio
n w
avel
engt
h (n
m)
200
250
300
350
400
450
500Control correctedSW day 140
(g) (h)
(i) (j)
Figure B5 (contd.) 3D EEM fluorescence spectra of synthetic waste (SW) leachates at
day 120 and 140 for control uncorrected and corrected samples (dilution factor 10)
References
156
References
Abbt-Braun, G. and Frimmel, F. H. (2002) The relevance of reference materials: isolation and
general characterization. In: Frimmel, F. H., Abbt-Braun, G., Heumann, K. G., Hock, B.,
Lüdemann, H.-D., Spiteller, M. (eds). Refractory organic substances (ROS) in the
environment. Weinheim: Wiley – VCH, 1–38.
Abbt-Braun, G., Frimmel, F. H., and Schulten, H. R. (1989) Structural investigations of
aquatic humic substances by pyrolysis-field ionization mass spectrometry and pyrolysis-gas
chromatography/mass spectrometry, Water Research, 23, 1579–1591.
Ahmad, S. R. and Reynolds, D. M. (1995) Synchronous fluorescence spectroscopy of
wastewater and some potential constituents. Water Research 29 (6), 1599–1602.
Ahmad, S. R. and Reynolds, D. M. (1999) Monitoring of water quality using fluorescence
technique: prospect of on-line process control. Water Research 33 (9), 2069–2074.
Al-Yaqout, A. F., and Hamodoa, M. F. (2003) Evaluation of landfill leachate in arid climate-
a case study. Environment International 29, 593-600.
Amokrane, A., Comel, C., and Veron, J. (1997) Landfill leachate pre-treatment by
coagulation–flocculation. Water Research 31 (11), 2775–2782.
APHA, AWWA, and WPCF (1998) Standard method for the examination of water and
wastewater. 20th ed., McGraw-Hill, Washington, DC.
Artiola-Fortuny, J., and Fullar, W. H. (1982) Humic substances in landfill leachates: I. Humic
acid extraction and identification. Journal of Environmental Quality 11, 663-668.
Bari, A., and Farooq, S. (1984) Measurement of wastewater treatment efficiency by
fluorescence and UV absorbance. Environmental Monitoring and Assessment 5, 423-434.
References
157
Baker, A. (2001) Fluorescence excitation–emission matrix characterization of some sewage-
impacted rivers. Environmental Science and Technology 35, 948–953.
Baker, A., and Lamont-Black, J. (2001) Fluorescence of dissolved organic matter as a natural
tracer of ground water. GroundWater 39, 745–750.
Baker, A. (2002) Spectrophotometric discrimination of river dissolved organic matter.
Hydrological Processes 16, 3203–3213.
Baker, A., Inverarity, R., Charlton, M., and Richmond, S. (2003) Detecting river pollution
using fluorescence spectrophotometry: Case studies from the Ouseburn, NE England.
Environmental Pollution 124 (1), 57–70.
Baker, A., and Spencer, R.G. M. (2004) Characterization of dissolved organic matter from
source to sea using fluorescence and absorbance spectroscopy. Science of the Total
Environment 333 (1–3), 217–232.
Baker, A., Ward, D., Lieten, S. H., Periera, R., Simpson, E. C., and Slater, M. (2004)
Measurement of protein-like fluorescence in river and wastewater using a handheld
spectrophotometer. Water Research 38, 2934-2938.
Baker, A., and Curry, M. (2004) Fluorescence of leachates from three contrasting landfills.
Water Research 38, 2605-2613.
Baker, A. (2005) Thermal fluorescence quenching properties of dissolved organic matter.
Water Research 39 (18), 4405–4412.
Baig, S., and Liechti, P. (2001) Ozone treatment for biorefractory COD removal. Water
Science and Technology 43 (2) 197-204.
References
158
Bila, D. M., Montalvao, A. F., Silva, A. C., and Dezotti, M. (2005) Ozonation of landfill
leachate: evaluation of toxicity removal and biodegradability improvement. Journal of
Hazardous Materials B117, 235-242.
Bilgili, M. S., Demir, A., Akkaya, E., and Ozkaya, B. (2008) COD fractions of leachate from
aerobic and anaerobic pilot scale landfill reactors. Journal of Hazardous Materials 158
(2008), 157–163.
Blaser, P., Heim, A., and Luster, J. (1999) Total luminescence spectroscopy of NOM-typing
samples and their aluminium complexes. Environment International 25 (2–3), 285–293.
Bookter, T. J., Asce, A. M., and Ham, R. K. (1982) Stabilization of solid waste in landfills.
Journal of Environmental Engineering 108 (6), 1089–1100.
Bourgeois, W., Burgess, J. E., and Stuetz, R. M. (2001) On-line monitoring of wastewater
quality: A review. Journal of Chemical Technology & Biotechnology 76 (4), 337- 348.
Brookman, S. K. E. (1997) Estimation of biochemical oxygen demand in slurry and effluent
using ultra-violet spectroscopy. Water Research 31, 372-374.
Briggs, D., and Seah, M. P. (1990) Practical Surface Analysis, Vol. 1, Auger and X-ray
Photoelectron Spectroscopy, 2nd ed. John Wiley & Sons, Chichester New York Brisbane
Toronto Singapore.
Bubert, H., Lambert, J., and Burba, P. (2000) Structural and elemental investigations of
isolated aquatic humic substances using X-ray photoelectron spectroscopy. Journal of
Analytical Chemistry, 368, 274–280.
Calace, N., and Petronio, B. M. (1997) Characterization of high molecular weight organic
compounds in landfill leachate: humic substances. Journal of Environmental Science and
Health A32 (8), 2229-2244.
References
159
Calace, N., Liberatori, A., Petronio, B.M., and Pietroletti, M. (2001) Characteristics of
different molecular weight fractions of organic matter in landfill leachate and their role in soil
sorption of heavy metals, Environmental Pollution 113, 331–339.
Cammack, W. K. L., Kalf, J., Prairie, Y. T., and Smith, E. M. (2004) Fluorescent DOM in
lakes: relationship with heterotrophic metabolism. Limnology and Oceanography 49 (6),
2034-2045.
Carley, B.N., and Mavinic, D.S. (1991) The effects of external carbon loading on nitrification
and denitrification of a high-ammonia landfill leachate. Journal of Water Pollution Control
Federation 63, 51–58.
Castagnoli, O., Musmeci, L., Zavattiero, E., and Chirio, M. (1990) Humic substances and
humification rate in a municipal refuse disposed of in a landfill. Water Air Soil Pollution 53,
1-12.
Cecen, F., and Aktas, O. (2004) Aerobic co-treatment of landfill leachate with domestic
wastewater. Environmental Engineering Science 21, 303–312.
Chae, K. J., Oh, S. E., Lee, S. T., Bae, J. W., and Kim, I. S. (2000) The optimum substrate to
biomass ratio to reduce net biomass yields and inert compounds in biological leachate
treatment under pure-oxygen conditions. Bioprocess Engineering 23, 235-243.
Chen, Y., and Schnitzer, M. (1989) Sizes and shapes of humic substances by electron
microscopy. In: Hayes, M. H. B., MacCarthy, P., Malcolm, R. L., Swift, R. S. (eds). Humic
substances II: In search of structure. New York: John Wiley & Sons, 621–638.
Chen, Y., Chefetz, B., and Hadar, Y (1996) Formation and properties of humic substances
originating from compost. European Commission International Symposium. The Science of
Composting, 382-393. Blackie Academic and Professional, England.
References
160
Chen, Y. (2003) PhD dissertation. Effect on aging on the bioavailability of tolune sorbed to
municipal solid waste component. North Carolina State University, Raleigh, NC, USA.
Chen, J., LeBoeuf, E.J., Dai, S., and Gu, B. (2003) Fluorescence spectroscopic studies of
natural organic matter fractions. Chemosphere 50, 639–647.
Chevakidagarn, P. (2005) Surrogate parameters for rapid monitoring of contaminant removal
for activated sludge treatment plants for Para rubber and seafood industries in Southern
Thailand. Journal of Songklanakarin 27 (2), 417–424.
Chevalier, L. R., Irwin, C. N., and Craddock, J. N. (2002) Evaluation of InSpectra UV
Analyzer for measuring conventional water and wastewater parameters. Advances in
Environmental Research 6 (3), 369–375.
Chian, E. S. K, and DeWalle, F. B. (1976) Sanitary landfill leachates and their treatment.
Journal of the Environmental Engineering Division 102 (EE2), 411–431.
Chian, E. S. K. (1977) Stability of organic matter in leachates. Water Research 11, 225-232.
Cho, S.P., Hong, S.C., and Hong, S. (2002) Photocatalytic degradation of the landfill leachate
containing refractory matters and nitrogen compounds, Applied Catalysis B: Environmental
39, 125–133.
Christensen, T. H., Kjeldsen, P., Bjerg, P. L., Jensen, D.L., Christensen, J.B., Baun, A.,
Albrechtsen, H-J., and Heron, G. (2001) Biogeochemistry of landfill leachate plumes. Applied
Geochemistry 16, 659–718.
Christensen, J. B., Jensen, D. L., Gron, C., Fulip, Z., and Christensen, T. H. (1998)
Characterization of the dissolved organic carbon in landfill leachate-polluted groundwater.
Water Research 32 (1), 125-135.
References
161
Christensen, T. H., Kjeldsen P., and Lindhardt, B. (1996) Gas generating processes in
landfills. In: Christensen, T. H., Cossu, R..and Stegman, R. (Eds.) Landfilling of Waste:
Biogas. E. and F. N. Spon, London.
Christensen, T. H., and Kjeldsen, P. (1995) Landfill emissions and environmental impact: An
introduction in Sardinia 1995 Landfill Symposium. Proceeding Fifth International Landfill
Symposium, CISA, Cagliari, Sardinia, Italy.
Christman, R. F., Norwood, D. L., Seo, Y., and Frimmel, F. H. (1989) Oxidative degradation
of humic substances from freshwater environments. In: Hayes, M. H. B., MacCarthy, P.,
Malcolm, R. L., Swift, R. S. (eds). Humic substances II: In search of structure. New York:
John Wiley & Sons, 33–67.
Chu, L. M., Cheung, K. C., and Wong, M. H. (1994) Variations in the chemical properties of
landfill leachate. Environmental Management 18 (1), 105-117.
Coble, P.G. (1996) Characterization of marine and terrestrial DOM in seawater using
excitation–emission matrix spectroscopy. Marine Chemistry 51, 325–346.
Cox, L., Celis, R., Hermosin, M.C., Cornejo, J., Zsolnay, A., and Zeller, A. (2000) Effect of
organic amendments on herbicide sorption as related to the nature of the dissolved organic
matter. Environmental Science and Technology 34, 4600–4605.
Cumberland, S. A. and Baker, A. (2007) The freshwater dissolved organic matter
fluorescence- total organic carbon relationship. Hydrological Processes 21, 2093-2099.
DEFRA, 2007 Mechanical biological treatment of municipal solid waste.
(http://www.defra.gov.uk/environment/statistics/wastats/bulletin.htm).
Determann, S., Lobbes, J. M., Reuter, R., and Rullkotter, J. (1998) Ultraviolet fluorescence
excitation and emission spectroscopy of marine algae and bacteria. Marine Chemistry 62 (1–
2), 137–156.
References
162
Dinel, H., Schnitzer, M., and Dumontet, S. (1996) Compost maturity: chemical characteristics
of extractable lipids. Compost Science Utility 4, 16-25.
Effland, M. J. (1977) Modified procedure to determine acid-insoluble ligninin wood and pulp.
Technical Association of Pulp and Paper Industry 60, 143-144.
Elkins, K. M., and Nelson, D. J. (2001) Fluorescence and FT-IR spectroscopic studies of
Suwannee river fulvic acid complexation with aluminium, terbium and calcium. Journal of
Inorganic Biochemistry 87 (1–2), 81–96.
Elliot, S., Lead, J. R., and Baker, A. (2006) Thermal quenching of fluorescence freshwater
planktonic bacteria. Analytica and Chemica Acta 564, 219-225.
Ellis, K. V. (1989) Surface Water Pollution and its Control, Chapter i, Macmillan, London.
Esteves da Silva, J. C. G., Machado, A. S. C., Oliveira, C. J. S., and Pinto, M. S. S. D. S.
(1998) Fluorescence quenching of anthropogenic fulvic acids by Cu(II), Fe(III) and UO22+.
Talanta 45 (6), 1155–1165.
Fan, H.-J., Shu, H.-Y., Yang, H.-S., Chen, W.-C. (2006) Characteristics of landfill leachates
in central Taiwan. Science of the Total Environment, 361, 25-37.
Florencio, L., Field, J.A., and Lettinga, G. (1995) Substrate competition between
methanogens and acetogens during the degradation of methanol in UASB reactors. Water
Research 29, 915-922.
Francou, C., Linères, M., Derenne, S., Villio-Poitrenaud, S., and Houot, S. (2008) Influence
of green waste, biowaste and paper–cardboard initial ratios on organic matter transformations
during composting. Bioresource Technology 99, 8926–8934.
References
163
Frimmel, F. H., Abbt-Braun, G., Heumann, K. G., Hock, B., Lüdemann, H. D., and Spiteller,
M. (2002) Refractory organic substances (ROS) in the environment. Weinheim: Wiley –VCH.
Fu, P., Wu, F., Liu, C., Wang, F., Li, W., Yue, L., and Guo, Q. (2007) Fluorescence
characterization of dissolved organic matter in an urban river and its complexation with
Hg(II). Applied Geochemistry 22, 1668-1679.
Galapate, R. P., Baes, A. U., Ito, K., Mukai, T., Shoto, E., and Okada, M. (1998) Detection of
domestic wastes in Kurose River using synchronous fluorescence spectroscopy. Water
Research 32(7), 2232–2239.
Germili, F. B., Orhon, D., and Artan, N. (1991) Assessment of the initial inert soluble COD in
industrial wastewaters. Water Science Technology 23, 1077-1086.
Göbbels, F., and Püttmann, W. (1996) Structural investigation of isolated aquatic fulvic and
humic acids in seepage water of waste deposits by pyrolysis-gas chromatography/mass
spectrometry. Water Research 31 (7), 1609-1618.
Gourdon, R., Comel, C., Vermande, P., and Veron, J. (1989) Fractionation of the organic
matter of a landfill leachate before and after aerobic and anaerobic biological treatment.
Water Research 23, 167-173.
Gottschalk, C., Libra, J. A., and Saupe, A. (2000) Ozonation of water and wastewater: a
practical guide to understanding ozone and its application. Wiley-VCH, London.
Ham, S.Y., Kim, Y.J., and Lee, D.H. (2008) Leaching characteristics of PCDDs/DFs and
dioxin-like PCBs from landfills containing municipal solid waste and incineration residues.
Chemosphere 70 1685–1693.
Hansen, T. L., Schmidt, J. E., Angelidaki, I., Marca, E., Jansen, J. C., Mosbaek, H., and
Christensen, T. H. (2004) Method for determination of methane potentials of solid organic
waste. Waste Management 24 (4), 393–400.
References
164
Harmsen, J. (1983) Identification of organic compounds in leachate from a waste tip. Water
Research 17, 699-705.
Harries, C. R., Cross, C. J., and Smith. R. (2001) Application of BMP testing to the study of
the MSW decomposition in a MSW lysimeter. An introduction in Sardinia 2001 Landfill
Symposium. Proceeding Eighth International Landfill Symposium, CISA, Cagliari, Sardinia,
Italy.
Hayes, M. H. B., MacCarthy, P., Malcolm, R. L., and Swift, R. S. (1989) Humic substances:
II In Search of Structure, Wiley, Chichester, 729.
He, P. J., Xue, J. F., Shao, L. M., Li, G. J., and Lee, D. J. (2006) DOM in recycled leachate of
bioreactor landfill. Water Research 40, 1465-1473.
Heerenklage, J., and Stegmann, R. (2001) Comparison of test systems for the determination of
gas potential from waste. An introduction in Sardinia 2001 Landfill Symposium. Proceeding
Eighth International Landfill Symposium, CISA, Cagliari, Sardinia, Italy.
Her, N., Amy, G., McKnight, D., Sohn, J., and Yoon, Y. (2003) Characterization of DOM as
a function of MW by fluorescence EEM and HPLC-SEC using UVA, DOC, and fluorescence
detection. Water Research 37, 4295–4303.
Hossain, M. S., Gabr, M. A., and Barlaz, M. A. (2003) Relationship of compressibility
parameters to municipal solid waste decomposition. ASCE Journal of Geotechnical and
Geoenvironmental Engineering 129 (12), 1151–1158.
Huang, S., Diyamandoglu, V., and Fillos, J. (1993) Ozonation of leachates from aged
domestic landfills. Ozone Science Engineering 15 (5), 433-444.
Hautala, K., Peuravuori, J.,and Pihlaja, K. (2000) Measurement of aquatic humus content by
spectroscopic analysis. Water Research 34 (1), 246–258.
References
165
Hudson, N., Baker, A., and Reynolds, D. (2007) Fluorescence analysis of DOM in natural,
waste and polluted waters – A review. River Research and Applications 23, 631-649.
Huo, S-l., Xi, B-d., Yu, H-c., and Liu, H-l. (2008) Dissolved organic matter in leachates from
different treatment processes. Water and Environment Journal 1-8.
Huo, S., Xi, B., Yu, H., Liansheng, H., Shilei, F., and Hongliang, L. (2008) Characteristics of
dissolved organic matter in leachates with different landfill ages. Journal of Environmental
Sciences 20, 492-498.
Ikan, R., Iosetis, P., Rubinsztain, Y., Aizenshtat, Z., Pugmire, R., Anderson, L. L., and
Ishiwatari, R. (1986) Carbohydrate Origin of Humic Substances. Naturwissenschaften 73,
150.
Imai, A., Onuma, K., Inamori, Y., and Sudo, R. (1998) Effects of pre-ozonation in refractory
leachate treatment by the biological activated carbon fluidized bed process. Environmental
Technology 19, 221-273.
Imai, A., Fukushima, T., Matsushige, K., Kim, Y-H., and Choi, K. (2002) Characterisation of
dissolved organic matter in effluents from wastewater treatment plant. Water Research 36,
859-870.
Inber, Y., Chen, Y., and Hadar, Y. (1989) Solid-state carbon-13 nuclear magnetic resonance
and infrared spectroscopy of composted organic matter. Soil Science American Journal 53,
1695-1701.
Ince, O., Germili, F. B., Kasapgil, B., and Anderson, G. K. (1998) Experimental
determination of the inert soluble COD fraction of a brewery wastewater under anaerobic
conditions. Environmental Technology 19, 437-442.
References
166
Ivanova, L. K. (2007) PhD dissertation. Quantification of factors affecting rate and magnitude
of secondary settlement of landfills. University of Southampton, Southampton, UK.
Jaffe, R., Boyer, J. N., Lu, X., Maie, N., Yang, C., Scully, N. M., and Mock, S. (2004) Source
characterization of dissolved organic matter in a subtropical mangrove-dominated estuary by
fluorescence analysis. Marine Chemistry 84(3–4), 195–210.
Jahnel, J. B., Brinkmann, T. Abbt-Braun, G., and Frimmel, F. H. (2002) Occurrence of amino
acids, carbohydrates, and low-molecular-weight organic acids in refractory organic
substances. In: Frimmel, F. H., Abbt-Braun, G., Heumann, K. G., Hock, B., Lüdemann, H.-
D., Spiteller, M. (eds). Refractory organic substances (ROS) in the environment. Weinheim:
Wiley – VCH, 264–281.
James, K. E., William, C. B., and Kevin, L. W. (1985) Surrogate parameters for monitoring
organic matter and THM precursors. Journal of American Water Works Association 77 (4),
122–132.
Ivanova, L. K., Richards, D. J., and Smallman, D. J. (2007) Assessment of the anaerobic
biodegradation potential of MSW. Waste and Resource Management 161, 167-180.
Jirka, A. M., and Carter, M. J. (1975) Micro semiautomated analysis of. surface and waste
waters for chemical oxygen demand. Analytical Chemistry 47, 1397-1402.
Johansson, L.-S., Campbell, J. M., Koljonen, K., and Stenius, P. (1999) Applied Surface
Science 144–145, 92–95.
Kang, K., Shin, H. S., and Park, H. (2002) Characterization of humic substances present in
landfill leachates with different landfill ages and its implications. Water Research 36, 4023-
4032.
Kargi, F., and Pamukoglu, M. Y. (2004) Adsorbent supplemented biological treatment of pre-
treated landfill leachate by fed-batch operation. Bioresources Technology 94, 285-291.
References
167
Kettunen, R.H., and Rintala, J.A. (1995) Sequential anaerobic-aerobic treatment of sulphur
rich phenolic leachates. Journal of Chemical Technology Biotechnology 62, 177-184.
Kettunen, R.H., and Rintala, J.A. (1998) Performance of an onsite USAB reactor treating
leachate at low temperature. Water Research 32 (3), 537-546.
Kettunen, R.H., Hoilijoki, T.H., and Rintala, J.A. (1996) Anaerobic and sequential anaerobic–
aerobic treatments of municipal landfill leachate at low temperatures. Bioresources
Technology 58, 31–40.
Khan, A.W. (1977) Anaerobic degradation of cellulose by mixed culture. Canada Journal
Microbiology 23, 1700-1705.
Kitcherside, M. A., Glen, E. F., and Webster, A. J. F. (2000) FibreCap: An improved method
for the rapid analysis of fibre in feeding stuffs. Animal Feed Science and Technology 86 (1–
2), 125–132.
Knox, K. (1983) Treatability studies on leachate from a co-disposal landfill. Environmental
Pollution (Series B) 5, 157- 174.
Komilis, D. P., and Ham, R. K. (2003) The effect of lignin and sugars to the aerobic
decomposition of solid wastes. Waste Management 23, 419-423.
Kouzeli-Katsiri, A., Bosdogianni, A., and Christoulas, D. (1999) Prediction of leachate
quality from sanitary landfills. Journal of Environmental Engineering 125 (10), 950-958.
Kulikowska, D., and Klimiuk, E. (2008) The effect of landfill age on municipal leachate
composition. Bioresource Technology 99, 5981–5985.
Lakowicz, J. R. Principles of Fluorescence Spectroscopy, Kluwer Academic/Plenum, New
York, 1999.
References
168
Larsson, T., Wedborg, M., and Turner, D. (2007) Correction of inner-filter effect in
fluorescence excitation-emission matrix spectrometry using Raman scatter. Analytica Chimica
Acta 583, 357–363.
Lema, J. M., Mendez, R., and Blazquez, R. (1988) Characteristics of landfill leachates and
alternatives for their treatment: A review. Water, Air, and Soil Pollution 40, 223–250.
Lo, I. M.-C. (1996) Characteristics and treatment of leachates from domestic landfills.
Environment International 22(4), 433-442.
Lopez, A., Pagano, M., Volpe, A., and Pinto, A. D. (2004) Fenton’s pre-treatment of mature
landfill leachate. Chemosphere 54, 1005–1010.
Loukidou, M.X., and Zouboulis, A.I. (2001) Comparison of two biological treatment
processes using attached-growth biomass for sanitary landfill leachate treatment.
Environmental Pollution 111, 273–281.
Lu, J. C. S., Eichenberger, B., and Stearns, R. J. (1985) Leachate from Municipal Landfills:
Production and Management, Noyes Publications, Park Ridge, NJ.
Maehlum, T. (1995) Treatment of landfill leachate in on-site lagoons and constructed
wetlands. Water Science and Technology 32, 129–135.
Maie, N., Scully, N.M., Pisani, O., Jaffe, R., (2007) Composition of a protein-like fluorophore
of dissolved organic mat ter in coastal wet land and estuarine ecosystems.
Water Research 41, 563–570.
Marttinen, S.K., Kettunen, R.H., Sormunen, K.M., Soimasuo, R.M., and Rintala, J.A. (2002)
Screening of physical–chemical methods for removal of organic material, nitrogen and
toxicity from low strength landfill leachates. Chemosphere 46, 851–858.
References
169
Matsche, N., and Strumworher, K. (1996) UV absorption as a control parameter for biological
treatment plants. Water Science and Technology 33, 211-218.
McBean, E. A., Rovers, F. A., and Farquhar, G. J. (1995) Solid waste landfill engineering and
design. Prentice Hall PTR, Englewood Cliffs.
MacCarthy, P., and Rice, J. A. (1985) Spectroscopic methods (other than NMR) for
determining functionality in humic substances. In Humic Substances in Soil, Sediment and
Water, Eds G. R. Aiken, D. M. McKnight, R. L. Wershaw and P. MacCarthy, pp. 527± 559.
John Wiley & Sons, New York.
McKnight, D.M., Boyer, E.W., Westerhoff, P.K., Doran, P.T., Kulbe, T., and Andersen, D.T.
(2001) Spectrofluorometric characterization of dissolved organic matter for indication of
precursor organic material and aromaticity. Limnology Oceanography 46, 38–48.
Micales, J. A., and Skog, K. E. (1997) The decomposition of forest products in landfills.
International Biodeterioration and Biodegradation 39, 145-158.
Monje-Ramirez, I., and Orte de Velasquez, M. T. (2004) Removal and transformation of
recalcitrant organic matter from stabilized saline landfill leachates by coagulation-ozonation
coupling processes. Water Research 38, 2359-2367.
Morawe, B., Ramteke, D.S., and Vogelpohl, A. (1995) Activated carbon column performance
studies of biologically treated landfill leachate. Chemical Engineering Processes 34, 299–303.
Mostofa, K. M. G., Yoshioka, T., Konohira, E., Tanoue, E., and Takahashi, K. H. M. (2005)
Three dimensional fluorescence as a tool for investigating the dynamics of dissolved organic
matter in the Lake Biwa watershed. Limnology 6, 101-115.
Musikavong, C., and Wattanachira, S. (2007) Reduction of dissolved organic matter in terms
of DOC, UV254, SUVA and THMFP in industrial estate wastewater treated by stabilization
ponds. Environmental Monitoring and Assessment 134, 489-497.
References
170
Nanny, M. A., and Ratasul, N. (2002) Characterization and comparison of hydrophobic
neutral and hydrophobic acid dissolved organic carbon isolated from three municipal landfill
leachates. Water Research 36, 1572-1584.
Nilsum, H. I. (1998) Light-enhanced chemical oxidation for tertiary treatment of municipal
landfill leachate. Water Environment Research 70 (6), 1161-1169.
Ntampou, X., Zouboulis, A. I., and Samaras, P. (2005) Appropriate combination of physico-
chemical methods (coagulation/flocculation and ozonation) for the efficient treatment of
landfill leachates. Chemosphere 62 (5), 722-730.
Ogawa, H., Amagai, Y., Koike, I., Kaiser, K., and Benner, R. (2001) Production of refractory
dissolved organic matter by bacteria. Science 292, 917–920.
Ohno, T., and Bro, R. (2007) Dissolved organic matter characterization using multi way
spectral decomposition of fluorescence landscapes. Soil Science Society American Journal 70,
2028–2037.
Ozkaya, B., Demir, A., and Bilgili, M. S. (2006) Soluble substrate concentrations in leachate
from field scale MSW test cells. Journal of Hazardous Materials A134, 19-26.
Parlanti, E., Worz, K., Geoffroy, L., and Lamotte, M. (2000) Dissolved organic matter
fluorescence spectroscopy as a tool to estimate biological activity in a coastal zone submitted
to anthropogenic inputs. Organic Geochemistry 31(12), 1765–1781.
Patel-Sorrentino, N., Mounier, S., and Benaim, J. Y. (2002) Excitation-emission fluorescence
matrix to study pH influence on organic matter fluorescence in the Amazon basin rivers.
Water Research 36(10), 2571–2581.
Pelaez, A. I., Sanchez, J., and Almendros, G. (2009) Bioreactor treatment of municipal solid
waste landfill leachates: Characterization of organic fractions. Waste Management 29, 70-77.
References
171
Perdue, E. T. (1998) Chemical composition, structure and metal binding properties. In:
Hessen, D. O., Tranvik, L. J. (eds). Aquatic humic substances. Berlin: Springer-Verlag, 41–
61.
Perminova, I. V., Frimmel, F. H., Kovalevskii, D. V., Abbt-Braun, G., Kudryavtsev, A. V.,
and Hesse, S. (1998) Development of predictive model for calculation of molecular weight of
humic substances, Water Research, 32, 872–881.
Pichler, M., and Kogel-Knabner, I. (2000) Chemolytic analysis of organic matter during
aerobic and anaerobic treatment of municipal solid waste. Journal of Environmental Quality
29, 1337-1344.
Pohland, F. G., Engelbrecht, R. S. (1976) Impact of sanitary landfills: An overview of
environmental factors and control alternatives. Report to the American Paper Institute, 12–22.
Pohland, F. G., and Harper, S. R. (1985) Critical review and summary of leachate and gas
production from landfills, EPA/600/2-86/073, US Environmental Protection Agency,
Cincinnati.
Qasim, S. R., and Burchinal, J. C. (1970) Leaching from sanitary landfills. Journal of the
Water Pollution Control Federation 42 (3), 371–379.
Ragle, N., Kissel, J., Ongerth, J. E., and DeWalle, F. B. (1995) Composition and variability of
leachate from recent and aged areas within a municipal landfill. Water Environment Research
67 (2), 238–243.
Renou, S., Givaudan, J. G., Poulian, S., Dirassouyan, F., and Moulin, P. (2008) Landfill
leachate treatment: Review and opportunity. Journal of Hazardous Material 150, 468-493.
Reynolds, D., and Ahmad, S. R. (1995). Effect of metal ions on the fluorescence of sewage
wastewater. Water Research 29(9), 2214–2216.
References
172
Reynolds, D. M., and Ahmad, S. R. (1997) Rapid and direct determination of wastewater
BOD values using a fluorescence technique. Water Research 31(8), 2012–2018.
Reynolds, D. M. (2002) The differentiation of biodegradable and non-biodegradable dissolved
organic matter in wastewaters using fluorescence spectroscopy. Journal of Chemical
Technology and Biotechnology 77 (8), 965–972.
Reynolds, D. M. (2003) Rapid and direct determination of tryptophan in water using
synchronous fluorescence spectroscopy. Water Research 37 (13), 3055–3060.
Rees, J. F. (1980) The fate of carbon compounds in the landfill disposal of organic matter.
Journal of Chemical Technology and Biotechnology 30, 161-175.
Rivas, F.J., Beltran, F., Gimeno, O., Acedo, B., and Carvalho, F. (2003) Stabilized leachates:
ozone-activated carbon treatment and kinetics. Water Research 37, 4823–4834.
Robinson, H. D., and Grantham, G. (1988) The treatment of landfill leachates in on-site
aerated lagoon plants: Experience in Britain and Ireland. Water Research 22, 733-747.
Robinson, H. D., and Knox, K. (2003) Environment Agency Technical Report PI-496/TR(2)
Updating the landfill leachate pollution inventory reporting tool.
Saadi, I., Borisover, M., Armon, R., and Laor, Y. (2006) Monitoring of effluent DOM
biodegradation using fluorescence, UV and DOC measurements. Chemosphere 63, 530-539.
Schnitzer, M., and Khan, S. U. (1978) Humic substances in the environment, Dekker, New
York.
Schulten, H. R., Leinweber, P., and Jandl, G. (2002) Analytical pyrolysis of humic substances
and dissolved organic matter in water. In: Frimmel, F. H., Abbt-Braun, G., Heumann, K. G.,
References
173
Hock, B., Lüdemann, H.-D., Spiteller, M. (eds). Refractory organic substances (ROS) in the
environment. Weinheim: Wiley – VCH, 163–187.
Schulten, P., Bayer, A., Kuhn, F., Luy, T., and Volkmer, M. (1995) H2O2/O3, H2O2/UV, and
H2O2/Fe+2 processes for the oxidation of hazardous wastes. Ozone Science Engineering. 17,
119-134.
Schmitt-Kopplin, P., Garrison, A. W., Perdue, E. M., Freitag, D., and Kettrup, A. (1998)
Capillary electrophoresis of humic substances. Facts and artefacts. Journal of
Chromatography A 807, 101–109.
Senesi, N. (1990) Molecular and quantitative aspects of the chemistry of fulvic acid and its
interactions with metal ions and organic chemicals. Part 2. The fluorescence spectroscopy
approach. Analytica Chemica Acta 232, 77–106.
Senesi, N., Miano, T. M., Provenzano, M. R., and Brunetti, G. (1991) Characterization,
differentiating, and classification of humic substances by fluorescence spectroscopy. Soil
Science 152, 259–271.
Sharpless, C. M., and McGown, L. B. (1999) Effects of aluminium-induced aggregation on
the fluorescence of humic substances. Environmental Science and Technology 33 (18), 3264–
3270.
Sierra, M. M. D., Giovanela, M., Parlanti, E., and Soriano-Sierra, E. J. (2005) Fluorescence
fingerprint of fulvic and humic acids from varied origins as viewed by single-scan and
excitation/emission matrix techniques. Chemosphere 58, 715-733.
Silva, A. C., Dezotti, M., Sant’Anna, G.L. (2004) Treatment and detoxication of a sanitary
landfill leachate, Chemosphere 55, 207–214.
Smith, B., Higgo, J. J. W., Moody, P., Davis, J. R., Williams, G. M., and Warwick, P. (1991)
Comparative study of humic substances in groundwater: 1. The extraction of humic material
References
174
from Drigg groundwater and a study of its ability to form complexes with cobalt and nickel.
British Geology Survey. Nottingham, Technical report WE/90/43.
Stedmon, C.A., Thomas, D.N., Granskog, M., Kaartokallio, H., Papadimitriou, S., and
Kuosa, H., (2007) Characteristics of dissolved organic matter in Baltic coastal sea ice.
Allochthonous or autochthonous origins? Environmental Science and Technology 41, 7273–
7279.
Stevenson, F. J. (1982) Humus Chemistry: Genesis, Composition, Reactions. Wiley, New
York.
Stevensen, L. (1994) Humus chemistry: Genesis, composition, reactions, second ed. Wiley,
New York, USA.
Tatsi, A. A., Zouboulis, I., Matis, K. A., and Samaras, P. (2003) Coagulation–flocculation
pretreatment of sanitary landfill leachates, Chemosphere 53, 737–744.
Tchbanoglous, G., Theisen, H., and Vigil, S. A. (1993) Integrated solid waste management.
The McGraw-Hill Companies, Inc.
Thurman, E. M., and Malcolm, R. L. (1981) Preparative isolation of aquatic humic
substances. Environmental Science and Technology 15, 463-466.
Timur, H., and Ozturk, I. (1999) Anaerobic sequencing batch reactor treatment of landfill
leachate, Water Research 33, 3225–3230.
Tong, X. G., Smith, L. H., and McCarty, P.L. (1990) Methane fermentation of selected
lignocellulosic materials. Biomass 21, 239-255.
Tuomela, M., Vikman, M., Hatakka, A., and Itavaara, M. (2000) Biodegradation of lignin in a
compost environment: a review. Bioresource Technology 72, 169-183.
References
175
Tucker, S.A., Amszi, V.L., and Acree, W.E. (1992) Primary and Secondary Inner Filtering
Effect of K2Cr207 on Fluorescence Emission Intensities of Quinine Sulfate. Journal of
Chemical Education 69.
Uyguner, C.S., and Bekbolet, M. (2005) Evaluation of humic acid photocatalytic degradation
by UV–vis and fluorescence spectroscopy. Catal.Today 101, 267–274.
Uygur, A., and Kargi, F. (2004) Biological nutrient removal from pre-treated landfill leachate
in a sequencing batch reactor. Journal of Environmental Management 71, 9-14.
Varadachari, V., and Ghosh, K. (1984) On humus formation. Plant Soil, 77: 305-313.
Vansoest, P. J., and Robertson, J. B. (1980) Systems of analysis for evaluating fibrous feeds.
In Standardization of Analytical Methodology for Feeds (Pigden, W. J., Balch, C., and
Graham, M. (eds)). International Research Development Center, Ottawa, Canada, 49–60.
Vodacek, A., and Philpot, W. D. (1987) Environmental effects on laser-induced fluorescence
spectra of natural waters. Remote Sensing of Environment 21(1), 83–95.
Wang, Z. P., Zhang, Z., Lin, Y., Deng, N., Tao, T., and Zhuo, K. (2002) Landfill leachate
treatment by a coagulation-photooxidation process. Journal of Hazardous Materials 2887, 1–
7.
Wang, Y. S., Caroline, S. B., and. Barlaz, M. A. (1994) Anaerobic biodegradability of
cellulose and hemicellulose in excavated refuse samples using a biochemical methane
potential assay. Journal of Industrial Microbiology 13, 147-153.
Waste Management Paper 26B, Landfill Design, Construction and Operational Practice
(1995) Department of the Environment, HMSO, London.
Weishaar, J. L., Aiken, G. R., Bergamaschi, B. A., Fram, M. S., Fujii, R., and Mopper, K.
(2003) Evaluation of specific ultra violet absorbance as an indicator of the chemical
References
176
composition and reactivity of dissolved organic carbon. Environmental Science and
Technology 37, 4702-4708.
Weis, M., Abbt-Braun, G., and Frimmel, F. H. (1989) Humic like substances from landfill
leachates- characterization and comparison with terrestrial and aquatic humic substances. The
Science of the Total Environment 81/82, 343-352.
Welander, U., and Henrysson, T. (1998) Physical and chemical treatment of a nitrified
leachate from a municipal landfill. Environmental Technology 19 (6), 591-599.
Westerhoff, P., Chen, W., and Esparza, M. (2001) Fluorescence analysis of a standard fulvic
acid and tertiary treated wastewater. Journal of Environmental Quality 30, 2037–2046.
Williams, G. M., Trick, J. K., Noy, D. J., and Ogilvy, R.D. (2002) The impact of disused
landfills on groundwater in the UK Chalk. In: Proceedings of the Waste 2002 Conference,
Stratford-Upon-Avon, England 24–26 September, 521–530.
Wu, B. Y. (2002) Effect on aging on the bioavailability of tolune sorbed to municipal solid
waste component. PhD thesis. North Carolina State University, Raleigh, NC, USA.
Yamashita, Y., and Tanoue, E. (2003) Chemical characterization of protein-like fluorophores
in DOM in relation to aromatic amino acids. Marine Chemistry 82, 255–271.
Yang, Y., and Zhang, D. (1995) Concentration effect on the fluorescence spectra of humic
substances. Communications in Soil Science and Plant Analysis 26 (15–16), 2333–2349.
Yuan, G., Soma, M., Seyama, H., Theng, B. K. G., Lavkulich, L. M., and Takamatsu, T.
(1998) Geoderma 86, 169–181.
Zehnder, A. J. B. (1978) Ecology of methane formation. Water Pollution Microbiology 2,
349-376.
References
177
Zhang, T., Lu, J., Ma, J., and Qiang, Z. (2008) Fluorescence spectroscopic characterization of
DOM fractions isolated from a filtered river water after ozonation and catalytic ozonation.
Chemosphere 71, 911–921.
Zheng, Z., He, P.J., Shao, L.M., and Lee, D.J. (2007) Phthalic acid esters in dissolved
fractions of landfill leachates. Water Research 41, 4696–4702.
Zheng, B., Richards, D. J., Smallman, D. J., and Beaven, R. P. (2007) Assessing MSW
degradation by BMP and fibre analysis. Proceeding of the Institution of Civil Engineers,
Waste and Resource Management 160 (4), 133–139.
Zoubouli, A. I., Loukidou, M. X., and Christodoulou, K. (2001) Enzymatic treatment of
sanitary landfill leachate. Chemosphere 44, 1103-1108.
Zsolnay, A., Baigar, E., Jimenez, M., Steinweg, B., and Saccomandi, F. (1999) Differentiating
with fluorescence spectroscopy of dissolved organic matter in soils subjected to drying.
Chemosphere 38 (1), 45-50.