Leachate Quality Analysis and Passive
Treatment Options
Pema Choden
Master of Philosophy (Environmental Science)
A Thesis submitted for the fulfilment of M Phil. Degree,
The University of Newcastle, N.S.W.
July, 2011
“I hereby certify that the work embodied in this thesis is the result of original research
and has not been submitted for a higher degree to any other University or Institution.”
Signed
……………………………
Acknowledgements
First of all, I would like to gratefully and sincerely thank Associate Proffessor Phillip
Geary for his continuous guidance, support and patience. He has contributed
significantly in gaining new experiences and immense knowledge in my life. I would
also like to thank Mr Joe Whitehead for his assistance as a co-supervisor and for
providing valuable discussions and accessibility. I have deep appreciations to both of
them for being very kind, co-operative and understanding throughout the study period.
Secondly, I would like to thank Dr Kim Edmunds for her immense support in editing.
Dr Kim worked hard and spent much time reviewing the draft chapters of the thesis. She
also gave me necessary inspiration to have faith in myself while writing the thesis.
I am grateful to Dr Steven Lucas, Mr Richard Bale, Mr Christopher Dever and Mr
Matthew Davies for all the laboratory assistance whenever needed. To all my friends in
Geology 109, they have provided me some much needed company, entertainment and
humour in what could have otherwise been a stressful study environment. I always felt
happy to have them around and talk to them whenever depressed.
Finally, and especially, for my parents (Dad, Mom and Sister, Pema Wangmo), I thank
you for the unwavering support, prayers and strong faith in me and for being tolerant
while being away from home, over a long period. Without them, this would not have
been possible.
Table of Contents
Acknowledgements .................................................................................................................... i
Table of Contents ..................................................................................................................... ii
List of Tables ........................................................................................................................... iv
List of Figures ........................................................................................................................... v
List of Equations ..................................................................................................................... vi
Abbreviations, Acronym and Symbols ................................................................................. vii
Abstract .................................................................................................................................... ix
Chapter 1 Introduction ............................................................................................................ 1
1.1
Waste Management in Bhutan.................................................................................... 4
1.2
Research Aims .............................................................................................................. 5
1.3
Thesis Approach and Structure .................................................................................. 6
Chapter 2 Background ............................................................................................................ 8
2.1
Solid Waste Management Techniques ....................................................................... 8
2.1.1
2.1.2
2.1.3
2.2
Composting ..................................................................................................................................... 8
Incineration.................................................................................................................................... 12
Landfills ........................................................................................................................................ 13
Review of Leachate Treatment Options ................................................................... 19
2.2.1
2.2.2
2.2.3
2.2.4
Aerobic biological treatment (Aeration) ........................................................................................ 19
Flocculation ................................................................................................................................... 22
Membrane separation .................................................................................................................... 24
Activated Carbon Adsorption (ACA) ............................................................................................ 24
Chapter 3 Solid Waste Management .................................................................................... 26
3.1
Waste Management in Bhutan.................................................................................. 26
3.2
Summerhill Waste Management Centre (SWMC), Newcastle .............................. 28
3.3
Leachate Characterization and Statistical Evaluation ........................................... 30
3.3.1
3.4
Temporal variation in leachate quality .......................................................................................... 32
Outcomes of Leachate Characterization and Evaluation ...................................... 34
Chapter 4 Materials and Methods ........................................................................................ 41
4.1
Aerobic Biological Treatment ................................................................................... 41
4.1.1
4.1.2
4.1.3
4.1.4
4.2
Site selection ................................................................................................................................. 41
Physical and chemical parameters of the leachate (sample) .......................................................... 41
Characterisation of leachate sample .............................................................................................. 44
Set-up of the aeration experiment .................................................................................................. 45
Column Experiments ................................................................................................. 45
4.2.1
4.2.2
4.2.3
Setup of the column experiments .................................................................................................. 45
Characterisation of leachate sample .............................................................................................. 47
Determination of physical and chemical properties of the filter materials .................................... 47
4.3
Materials for Leachate Characterisation and Evaluation ...................................... 51
Chapter 5 Results ................................................................................................................... 52
5.1
Aerobic Biological Treatment ................................................................................... 52
5.1.1
5.1.2
5.1.3
5.1.4
5.1.5
5.2
Change in the nitrogen compounds ............................................................................................... 54
Turbidity ....................................................................................................................................... 56
Electrical conductivity (EC) .......................................................................................................... 58
pH and total alkalinity .................................................................................................................. 59
Phosphorus .................................................................................................................................... 60
Column Experiments ................................................................................................. 61
5.2.1
5.2.2
5.2.3
5.2.4
5.2.5
Solube reactive phosphorus removal ............................................................................................ 66
Turbidity and Colour ..................................................................................................................... 69
Electrical conductivity ................................................................................................................... 71
Total alkalinity and pH .................................................................................................................. 72
Ammonia-nitrogen, Nitrite-nitrogen and Nitrate-nitrogen ............................................................ 74
Chapter 6 Discussion ............................................................................................................. 76
6.1
Aerobic Biological Treatment ................................................................................... 76
6.1.1
6.1.2
6.1.3
6.1.4
6.1.5
6.1.6
6.1.7
6.1.8
6.2
Ammonia-nitrogen removal .......................................................................................................... 77
Turbidity and electrical conductivity ............................................................................................. 79
Colour and odour ........................................................................................................................... 80
Phosphorus .................................................................................................................................... 80
Temperature .................................................................................................................................. 81
Retention time ............................................................................................................................... 81
Application .................................................................................................................................... 83
Drawbacks of the experiment ........................................................................................................ 84
Column Experiments ................................................................................................. 85
6.2.1
6.2.2
6.2.3
6.2.4
6.2.5
6.2.6
P-removal ...................................................................................................................................... 85
The effect of chemical composition .............................................................................................. 86
Contact time & hydraulic conductivity ......................................................................................... 87
Longevity ...................................................................................................................................... 88
Particle size distribution, porosity percentage (%) and bulk density ............................................. 88
Drawbacks of the experiment ........................................................................................................ 89
Chapter 7 Conclusions and Recommendations ................................................................... 91
7.1
Laboratory Aeration Experiments ........................................................................... 91
7.2
Laboratory Column Experiments ............................................................................ 91
7.3
Leachate Characterisation ........................................................................................ 92
7.4
Solid Waste Management Options for Bhutan........................................................ 93
7.5
Recommendations and Further Research ............................................................... 94
References ............................................................................................................................... 95
Appendix A ................................................................................................................................ I
Appendix B ............................................................................................................................. VI
List of Tables
Table 2.1 Classification of Landfill Leachate According to the Composition Changes. ...... 16
Table 2.2 Landfill Gas Composition. .................................................................................... 17
Table 3.1 Composition of Landfill Leachate (1995-2009) from SWMC, Newcastle,
Australia. ............................................................................................................... 31
Table 3.2
Correlations Between Rainfall in mm and Leachate Volume in KL: i) monthly
(2001-2009) and ii) annual (1995 -2009). ............................................................. 39
Table 3.3 Spatial and Altitudinal Variation in Annual Rainfall by Region and Percentage of
Total Area in Different Altitudinal Zones in Bhutan. ........................................... 40
Table 4.1 Initial Physio-Chemical Characteristics of the Raw Leachate Sample (11th
February 2010). ..................................................................................................... 45
Table 4.2 Initial Physio-Chemical Characteristics of the Raw Leachate Sample and
DeIonised Water Used in Column Experiments (18th May 2010). ....................... 48
Table 4.3 Typical Chemical Composition of BFS (amorphous-0.25-4 mm). ....................... 50
Table 5.1 The Change in Characteristics of Landfill Leachate in the two Tanks operated at
21ºC for 20 days. Tank A was the control while Tank B was aerated. ................. 53
Table 5.2 Physical Properties of Three Different Filter Materials: GAC, BFS and sand. .... 63
Table 5.3 Operating Parameters for the Column Experiments. ............................................ 63
Table 5.4 Mean Results Obtained from Column Experiments using three Different Filter
Materials. ............................................................................................................... 67
List of Figures
Figure 2.1 Typical temperature curve observed during different compost phases.. ............... 10
Figure 2.2 A sketch of a MSW landfill construction.. ............................................................ 14
Figure 2.3 A cross section of a leachate collection system.... ................................................ 17
Figure 2.4 A landfill gas collection and recovery system. ..................................................... 18
Figure 2.5 An aerated lagoon plant......................................................................................... 22
Figure 3.1 Average percentage composition of MSW in Bhutan.. ......................................... 27
Figure 3.2 Frequency distribution curve of ammonia between 1995 and 2009. .................... 35
Figure 3.3 An illustration of six month moving average – leachate percolation versus rainfall
(January 2001- December 2009). .......................................................................... 36
Figure 3.4
An illustration of the annual leachate percolation versus annual rainfall received
at the landfill site (1995-2009). ............................................................................. 37
Figure 3.5 A linear relationship between rainfall (mm) and leachate volume (KL): a) monthly
(R2= 0.238) and b) annual (R2= 0.367). ................................................................ 38
Figure 3.6 A plot of variation in rainfall with changes in season and altitude in Bhutan. ..... 40
Figure 4.1 SWMC landfill and methane collection facility. ................................................... 42
Figure 4.2 Set up of the aeration experiment using glass-sided tanks and aerators showing
Tank A (left) and Tank B (right). .......................................................................... 46
Figure 4.3 Setup of the column experiment showing: a) different columns and b) constant
head apparatus. ...................................................................................................... 47
Figure 5.1 The changing pattern of ammonia concentration in Tank A and Tank B over a
period of 20 days. .................................................................................................. 54
Figure 5.2 The changing pattern of nitrite concentration in Tank A and Tank B over a period
of 20 days. ............................................................................................................. 55
Figure 5.3 The changing pattern of nitrate concentration in Tank A and Tank B over a period
of 20 days. ............................................................................................................. 56
Figure 5.4 The changing pattern of turbidity in Tank A and Tank B over a period of 20 days.
57
Figure 5.5 The changing pattern of EC in Tank A and Tank B over a period of 20 days. ..... 58
Figure 5.6 The difference in colour of the leachate samples from Tank A (left) and Tank B
(right). .................................................................................................................... 59
Figure 5.7 The changing pattern of pH in Tank A and Tank B over a period of 20 days. ..... 59
Figure 5.8 The changing pattern of alkalinity in Tank A and Tank B over a period of 20 days.
60
Figure 5.9 The changing pattern of phosphorus in Tank A and Tank B over a period of 20
days........................................................................................................................ 61
Figure 5.10 Plot of % of media finer by weight as a function of particle size (mm). .............. 64
Figure 5.11 The changing level of phosphorus in the leachate samples following different runs
throughout the experiment. ................................................................................... 66
Figure 5.12 P-sorption isotherms for three different filter materials developed using five
different P-solutions in the batch scale experiment (Lanfax Laboratories). ......... 69
Figure 5.13 The changing level of turbidity in the leachate samples following different runs
throughout the experiment. ................................................................................... 70
Figure 5.14 The difference in the colour of the leachate samples for three different runs along
with their respective controls. a) GAC, b) BFS and c) sand. ................................ 71
Figure 5.15 The changing level of EC in the leachate samples following different runs
throughout the experiment. ................................................................................... 72
Figure 5.16 The changing level of alkalinity in the leachate samples following different runs
throughout the experiment. ................................................................................... 73
Figure 5.17 The change in pH values of the leachate samples following different runs
throughout the experiment. ................................................................................... 74
Figure 5.18 The changing level of ammonia in the leachate samples following different runs
throughout the experiment. ................................................................................... 74
Figure 5.19The changing level of nitrate in the leachate samples following different runs
throughout the experiment. ................................................................................... 75
List of Equations
Equation 4.1 Calculation of filter material % retained on sieve.. ............................................ 49
Equation 4.2 Calculation of % finer by weight. ...................................................................... 49
Equation 4.3 Calculation of hydraulic conductivity of the filter material.. ............................ 49
Abbreviations, Acronym and Symbols
Abbreviation
Term
SWM
Solid Waste Management
MSW
Municipal Solid Waste
SWMC
Summerhill Waste Management Centre
AWQGFMW
Australian Water Quality Guidelines for Fresh and
Marine Waters.
BOM
Bureau of Meteorology
RGoB
Royal Government of Bhutan
BFS
Blast Furnace Slag
GAC
Granular Activated Carbon
PAC
Powdered Activated Carbon
DO
Dissolved Oxygen
BOD
Biochemical Oxygen Demand
COD
Chemical Oxygen Demand
TOC
Total Organic Carbon
XOCs
Xenobiotic Organic Compounds
OM
Organic Matter
NCM
Non-compostable Materials
TKN
Total Kjeldahl Nitrogen
SRP
Soluble Reactive Phosphorus
Ksat
Saturated hydraulic conductivity (permeability)
D10
Effective size of filter material finer by 10%
LEL
Lower Explosive Limit
kL
Kilolitre
EC
Electrical Conductivity
pH
measure of H+ activity
UMR
Under Measuring Range
NA
Not Applicable
P
Phosphorus
CH4
Methane
CO2
Carbon dioxide
NH3
Ammonia
CaCO3
Calcium Carbonate
Abstract
More than 90% of municipal solid waste (MSW) in developing countries is disposed of
in landfills. In Bhutan, about 90% of solid waste is disposed of in landfills. One of the
significant problems associated with landfills is the generation of leachate. Landfill
leachates are highly contaminated waste waters containing high concentrations of
organic matter (OM) measured as biochemical oxygen demand (BOD) and chemical
oxygen demand (COD), ammonia, halogenated hydrocarbons and trace elements. The
direct release of leachate into the environment may pose potential risks and hazards to
public health and ecosystems. As a result, cost effective and environmentally acceptable
treatments of leachate are sought. This research aims to examine leachate characteristics
and two low cost passive treatment options.
The characteristics of a typical leachate generated at a modern sanitary landfill were
investigated by analysis of long-term monitoring data collected at Summerhill Waste
Management Centre (SWMC), Newcastle, NSW. Leachate production from the SWMC
landfill was clearly related to rainfall events at the landfill site. Rainfall has a direct
impact on the volume of leachate produced and consequently on its chemical
characteristics. Thus, leachate treatment systems must have provisions for the variation
in concentrations.
The primary goal of this study was to investigate the suitability of two low cost passive
leachate treatment options, which are viable and suitable for adoption in Bhutan. Two
laboratory bench scale experiments were undertaken. The first experiment involved
surface aeration of raw leachate over a period of 20 days, while the second investigated
the treatment performance of three low cost filter media (Granular Activated Carbon
(GAC), Blast Furnace Slag (BFS) and sand) by examining their sorption efficiencies in
a series of column experiments. Leachate samples were collected from the landfill at
SWMC.
The results of the research showed that a medium strength landfill leachate can be
treated by both methods to reduce the concentrations of certain parameters. Aerobic
treatment enhanced the leachate quality mainly through removal of ammonia and OM
(>95%). It resulted in significant pollutant reductions as opposed to no aeration, which
resulted in anoxic conditions. Column experiments provided leachate treatment
essentially by lowering soluble reactive phosphorus (SRP) concentration. BFS and GAC
have performed comparatively better with P-removal efficiencies of 92% and 67%,
respectively, than sand (40%) in the laboratory work undertaken.
Finally, the research results also suggested landfills in Bhutan do not have appropriate
leachate and gas collection facilities. Due to a lack of proper waste segregation, the
leachates produced in landfills could be chemically complex. Composting is suggested
as a sustainable alternative for SWM in Bhutan to reduce the 50-60% of organic waste
disposed of in the landfills if leachate collection and treatment cannot be afforded.
Chapter 1 Introduction
A large number of reports on Solid Waste Management (SWM) from many countries
indicate that sanitary landfill is still the most affordable disposal option for solid waste,
especially in the developing countries. About 90% of solid wastes generated in the
world are disposed in landfills, for reasons of convenience and affordability. Despite the
availability of many alternatives, land filling (variously called a garbage dump or
rubbish dump) of Municipal Solid Waste (MSW) is still the main method of disposing
solid waste in most countries across Asia, Africa, and Europe. In Asia alone, about ten
countries dispose more than 70% of their MSW in landfills ( Autret et al., 2007; Henry
et al., 2006; Lopez et al., 2010).
Landfills mostly vary in structure, design and management based on the social,
technical and economic development status of that country. While some landfills remain
simple and traditional, and pose an adverse impact on the environment, others
developed into modern sanitary landfills ensure minimal environmental impact.
However, a common and major drawback associated with all the landfills is the
generation (formation) of heavily polluted leachates (Christensen et al., 2001;
Mehmood et al., 2009; Renou et al., 2008).
The formation of leachate begins in the landfill where the waste gets mixed up with
water from the air and moisture from waste and ground water. Immediately, waste starts
to undergo physical, biological and chemical decomposition under aerobic and
anaerobic - conditions, producing a large volume of liquid called leachate. Thus,
leachate consists of two components: one, a liquid generated by the breakdown of waste
in the landfill; and two, the infiltration of precipitation that has percolated through the
landfill (Duggan, 2005).
Leachates are generally characterised by dissolved contaminants, volatile organic acids,
trace elements, and high concentrations of organic matter (high biochemical oxygen
demand (BOD) and chemical oxygen demand (COD)) and ammonia (Tyrrel et al.,
2002). However, the compositions can vary depending on the nature of waste, the active
1
microbial flora and the characteristics of the soil, the rainfall pattern and the age of the
landfill (Cecen et al., 2003; Kargi & Pamukoglu, 2004; Mehmood et al., 2009).
When a leachate containing high concentrations of dissolved contaminants is directly
released into the environment, it may contaminate soil and water bodies, and may pose
potential risks and hazards to the environment and to public health. For instance,
numerous cases of contamination of streams, creeks and ground water have been
associated with leachate contamination from landfill sites (Christensen et al., 2001;
Duggan, 2005; Lim et al., 2009; Pivato & Gaspari, 2006; Robinson et al., 2005; Yang et
al., 2008). The high organic compounds, heavy metals and ammonia present in the
leachate are identified as potential sources of ground and surface water contamination.
Furthermore, if leachates are discharged directly to municipal wastewater treatment
plants, they may affect chlorine disinfection efficiency, cause corrosion of the pump
station, and sludge build up and settling problems. As a result, the treatment of leachate
is necessary before discharging to the environment.
Amongst many other factors, the successful treatment or appropriate remedial action for
leachate will depend on how its characteristics are understood. In fact, leachate
characterization is pivotal in evaluating the environmental risks associated with it. In
this regard, a part of this research involved the study of leachate formation and its
characteristics at a modern sanitary landfill in Australia, where best management
practices are used.
Located on an old mine site, the sanitary landfill (Summerhill Waste Management
Centre) is located near Newcastle, NSW. It receives four different types of wastes:
inert, solid, industrial and hazardous waste. Accordingly, the landfill is divided into
different cells, where different waste can be processed and deposited in these respective
cells. The main feature of this landfill comprises a series of discrete sections called
“cells” lined with an impermeable leachate barrier, which captures the leachate for
appropriate disposal. A landfill compactor is used to compact the waste and it is covered
daily with approximately 150 mm of cover material (soil). Other important features
include provisions for recycling, sorting and gas and leachate collection facilities.
2
In addition, recent studies indicate that apart from the type of waste that goes into the
landfill, the volume of water that enters the landfill has an important bearing on the
characteristics of the leachate (Tatsi & Zoubolis, 2002). A low strength and dilute
leachate is produced when the water percolation is high, whereas, lesser volumes of
water percolating through the landfill resulted in high strength leachates with
concentrated dissolved contaminants. This phenomenon indicates that water is the most
important element in leachate formation. Although, surface drainage and irrigation
water are considered to be some of the sources of water, precipitation (rainfall) is the
most dominant factor in the volume of leachate produced at a landfill site (Tatsi &
Zoubolis, 2002).
To enable further understanding of this concept, an overview of the long term leachate
monitoring data from a modern sanitary landfill (SWMC) is presented in this study. The
study took into account both temporal and long term changes of the leachate
characteristics. Further discussions were made regarding the change in leachate
characteristics, the volume percolated and the climatic factors (rainfall) at the site to
assess the likely correlations existing between them.
Generally, there are various treatment methods or remedial actions that have been
developed in recent decades aimed at minimising the concentration of the contaminants.
The remedial methods in landfills rely on a range of physical, chemical,
physiochemical, or biological processes depending on the method used. Physical
processes in landfills include dilution (Christensen et al., 2001) whereas physiochemical
and chemical processes involve processes such as sorption (Foo & Hameed, 2009),
filtration, ion exchange, precipitation and flocculation (Amokrane et al., 1997;
Nehrenheim et al., 2008). Biological remedial actions mostly deal with aeration and
microbiological (degradation) processes.
While many of these methods are now known to be highly successful and efficient,
most of them involve huge capital investment and require extensive use of resources
and technology, which makes them less than or not feasible in many places, particularly
in developing countries where financial constraints and resource limitations are already
a problem.
3
1.1
Waste Management in Bhutan
Bhutan is a small land-locked country situated between India and China and bordered
by Nepal and Bangladesh in the west. It has a total land area of 38,394 km2 and a
population of approximately 873,700 persons. Virtually all of Bhutan is mountainous
with an elevation of 100 m above sea level in the south to over 7500 m in the north.
Three major landform features are evident in Bhutan: the southern foothills; the inner
Himalayas; and the high Himalayas featuring snow capped mountains (Uddin et al.,
2007). Water and forests are Bhutan’s renewable energy resources and the country
depends highly on subsistence farming and hydroelectricity. The major land uses in
Bhutan are agriculture (>80% of land use) and forestry.
One of the most important threats to Bhutan’s environment is the problem of ever
increasing solid waste (Penjor, 2008; Phuntsho et al., 2008; Uddin et al., 2007). While
the magnitude of the problem is relatively small and manageable in rural areas, it is
growing significantly in urban areas in recent time subsequently posing threats to the
environment. The increase in waste generation is primarily attributed to factors such as
rapid rates of urbanisation, rural urban migration (6.7%), changing consumption pattern
and high population growth rate (11%). As per the national survey of solid waste
conducted between 2007 and 2008, mean waste generation has been estimated at
approximately 0.5 kg/person/day (Phuntsho et al., 2008). By comparison, the per capita
waste generation is still smaller than other countries - North America (2.0
kg/day/person), China (1.59 kg/day/person), Australia (just over 1.0 kg/day/person) and
Europe (0.6 kg/day/person) (EPA, 2008; cited in Farrell & Jones, 2009).
For instance, Thimphu is the largest city and the capital of Bhutan. The population in
Thimphu rapidly increased from 25,000 in 1992 to 80,000 in 2005. Consequently daily
waste generation increased by four fold from 17.5 tonnes in 1992 to 64.5 tonnes in 2007
(Penjor, 2008). Similar situations were observed in the three other urban centres Phuentsholing, Paro and Samtse.
In the absence of incineration and combustion, land filling of solid waste is the most
common method of solid waste disposal in Bhutan. The majority of the country
4
disposes of solid waste in the landfill-like structures, which includes open dumps in
narrow valleys, streams and rivers. Due to the lack of vacant land, finance, technology
and labour, the landfills do not have proper design, facilities and leachate treatment
systems. Hence, leachate contamination at landfill sites is a significant environmental
issue. For example, the biggest landfill in the capital (Memelakha Waste Disposal Site)
is already overflowing (Allison, 2008; Penjor, 2008).Without proper leachate and gas
collection facilities, the issues of free flow of leachate in the environment and release of
toxic gases (e.g. methane) in the atmosphere are of concern. In the past few years,
several forest fires have been reported to have occurred near the landfill site and
methane combustion has been suggested as one of the possible causes of each incident.
Given the situation in Bhutan, which is similar in many developing countries, this
research has been undertaken to examine a number of simple, effective and affordable
treatment measures to reduce leachate contaminant concentrations, which are an
important environmental issue.
1.2
Research Aims
Although aeration is used for treating wastewater, so far very little is known about its
application in the field of leachate treatment (Mehmood et al., 2009; Robinson &
Grantham, 1988). Few studies have shown the efficacy of the aeration process in
treating leachate, and those that did have used complex processes that are less efficient
in terms of cost, resource and application for large scale treatment (Berge et al., 2006;
Mehmood et al., 2009; Shao et al., 2008).
Similarly, a group of researchers indicated the potential ability of filter materials, both
natural and manmade products, for treatment of water, wastewaters and leachates
through bench scale column studies (Johansson, 1999; Mohan & Gandhimathi, 2009;
Nehrenheim et al., 2008). According to their studies, limestone, blast furnace slag, sand,
coal, peat and pine bark are some of the potential media capable of adsorbing and
reducing contaminant concentrations in wastewater. However, not much is known with
regard to their application in leachate treatment. Previous studies suggest further
research on the use of these materials in leachate treatment would not only assist in
5
treating leachate, but also enable the use of industrial by-products like slag and coal
(Johansson, 1999; Mohan & Gandhimathi, 2009; Nehrenheim et al., 2008).
The main part of this research has evaluated the suitability of two low cost passive
treatment methods for treatment of landfill leachates. The first experiment involved
aerobic biological treatment (aeration) of leachate in two aeration tanks (25 L) over a
period of 20 days, while the second, a column experiment, investigated the suitability of
three low cost filter media in treating leachate, by examining their sorption efficiencies
and therefore, their capacity to reduce contaminant concentrations.
The results from this research will enable recommendations to be made regarding SWM
in Bhutan and provide low cost and simple solutions for leachate treatment. To achieve
this, the following work was undertaken as part of this research:
To review different SWM techniques prevalent in other countries and suggest an
improved system of disposing solid waste for Bhutan; and
To be able to determine two low cost passive leachate treatment options, which
are suitable, environmentally friendly and economically viable for Bhutan.
1.3
Thesis Approach and Structure
This thesis is divided into seven different chapters. Chapter One is the introduction to
the thesis and covers the main topics to be discussed, the aims of the research and the
structure of the thesis. Chapter Two is the background of the thesis and is divided into
two parts. The first half of the chapter reviews the prevailing SWM techniques in
different countries, their advantages and disadvantages. The second part introduces
leachate, its impact on the environment and low cost passive treatment options that have
been so far adopted in treating leachate.
Chapter Three compares in brief the waste management system in Bhutan to that of
Summerhill Waste Management Centre (SWMC) in Newcastle. In addition, this chapter
presents an overview of the main physio-chemical characteristics of leachate formed at
the SWMC sanitary landfill during the past 14 years. The outcome of the statistical
evaluation of long term leachate monitoring data set is discussed in relation to leachate
6
formation at the landfill. Further correlations between the volume of leachate formed
and the amount of rainfall received at the site is applied in estimating leachate formation
in Bhutan.
Chapter Four essentially describes the methodology and materials used in the two bench
scale experiments: a) the effect of aerobic biological treatment on leachate quality
improvement; and b) column experiments using three different filter media to determine
the suitability of low cost filter materials in leachate treatment methods. One approach
of the study includes an analysis of long term leachate monitoring data obtained from
the SMWC landfill along with climate data obtained from Bhutan and Australia.
Chapter Five comprises two main sections dealing with results obtained from the
respective bench scale laboratory experiments in relation to their impact on leachate
quality improvement. Results of both experiments are based on nine important
parameters of leachate. Both experiments evaluated the possibility of treating leachate
quality in a simple, affordable and effective way.
Chapter Six is divided into two sections. The changes in the quality of the leachate from
both bench scale leachate treatment experiments are discussed in detail. The application
and limitations of this treatment study including concerns with respect to their
application as an onsite leachate treatment plant in Bhutan is outlined.
Lastly, Chapter Seven summarizes and concludes this thesis by highlighting the
outcomes of the study in relation to its application in treating leachate in Bhutan. Some
future research directions are also discussed towards the end of the chapter.
7
Chapter 2 Background
This chapter is divided into two parts. The first part presents a review of common SWM
techniques widely practised in different countries, their advantages and disadvantages.
The second part reviews low cost leachate treatment methods used for treating landfill
leachates.
2.1
Solid Waste Management Techniques
With a steady increase in solid waste generation, SWM has become a necessity
throughout the world. Essentially, SWM deals with collection, storage and disposal of
waste in an environmentally sustainable way, with minimum hazard to human health
and the environment. There is no single correct method to achieve proper waste
management, because management of waste has several aspects: political, social,
environmental, economic and technical (Rushbrook & Finnecy, 1988). Depending on
these factors, a wide range of SWM practices are prevalent in different countries.
Nevertheless, this study will look at three common methods that are efficient, affordable
and require minimal use of technology.
2.1.1
Composting
Similar to landfills and incineration, the method of composting solid waste is nothing
new. The practice of composting solid waste was prevalent over 100 years ago (Ernst,
1990; Wei et al., 2007). However, as with landfills, scientific understanding of the
biological, and particularly the microbiological processes involved, has only come about
recently. Until then, traditional methods of composting were largely dominant. The OM
of the waste was decomposed and degraded only on a small scale under simple and
natural processes and the product formed was used for agricultural land as a substitute
for fertilizers. Today, composting has become an essential component of MSW
management in various countries; the United Arab Emirates (UAE) – more than 50%;
the USA - 33%; the UK - 15% (Farrell & Jones, 2009; Hamoda et al., 1998; Slater &
Frederickson, 2001).
8
Generally, composting is described as the biological decomposition of OM under
controlled aerobic conditions to form a stable, humus-like end product (Adani et al.,
1995; cited in Wei et al., 2000; Farell & Jones, 2009). When MSW is involved, it is
defined as a biological decomposition of the biodegradable organic fraction of
municipal waste under controlled conditions, to a state sufficiently stable for nuisancefree storage, handling and for safe use in agricultural lands (Tchobanoglous & Kreith,
2002). However, in all cases, the process is facilitated by a diverse population of
microbes, whose population dynamics vary greatly, both temporally and spatially, and
generally involve the development of thermophilic temperatures as a result of
biologically produced heat (Swan et al., 2002; cited in Farrell & Jones, 2009).
Composting Procedures
According to Ernst (1990) and Hamoda et al. (1998), modern composting operations
consist of four basic steps: (i) preliminary treatment of feedstock, with separation of
undesirable content called processing; (ii) fermentation with aeration and turning
(decomposition); (iii) preparation of final product, which involves grinding, screening
and final separation of undesirable items from the product; and finally (iv) marketing.
The first step is necessary because the quality of compost produced will depend on the
starting material. MSW is a heterogenous material containing non-compostable
materials (NCM) such as heavy metals or persistent organic compounds (Lopez et al.,
2010). As a result, processing of waste is required in places where there is no source
segregation and sorting of waste.
In the second step, fermentation/decomposition of waste begins with the establishment
of composting conditions. Active involvement of microorganisms like bacteria,
actinomycetes, fungi, protozoa, worms and some larvae, decompose the organic
fraction, utilising it as a source of carbon. This eventually introduces an ecological
succession of microbial pollutants where resident or indigenous microbes, capable of
utilizing nutrients in the raw waste, immediately begin to proliferate. Due to the activity
of this group, the composting mass becomes favourable for other indigenous
populations to proliferate. Therefore, by plotting the effect of succession of total
bacterial content of the mass would result in a curve (Figure 2.1), showing three
different stages or phases in composting (Tchobanoglous & Kreith, 2002, p.12.5).
9
Finally, the third and fourth step is an essential step in determining the success of the
composting and deals with the post processing and marketing of the final product.
Except for the second step, the rest is common to all composting plants or industries.
Types of Fermentation/Decomposition Methods
To achieve fermentation or decomposition, especially for MSW composting on a
commercial scale, various composting methods are employed that use systems of
varying complexity. There are essentially two main types: turned or forced aeration
systems. Turned systems are commonly based upon the windrow system, which entails
the feedstock being piled in elongated heaps up to 2 m high and 50 m in length.
Figure 2.1 Typical temperature curve observed during different compost
phases (Source: Tchobanoglous & Kreith, 2002, p.12.5).
These piles are turned with decreasing frequency throughout the period of active
composting to maintain oxygen (O2) and moisture levels, and to release spent air. MSW
windrowing is often done indoors within large commercial premises to minimize
leachate production, improve odour control and reduce visual impact (Farrell & Jones,
2009).
10
A turned or windrow system sometimes consists of rotating drums responsible for
receiving the waste stream. After a certain number of days, the material exits the drum
and is screened. Residuals such as plastics, glass and metals are compacted and sent to
landfill and organics are transferred to storage buildings for composting. Windrows are
turned about two or three times per week for a total of eight weeks, during which
temperatures reach 15O oC. Finished compost is screened again, prior to being sold or
given away to farmers and landscapers.
In contrast to turned systems, forced aeration systems are often more complex with
computer controlled aeration regimes, offering a greater control over the process
conditions. The Bedminster Composting Technology used by Port Stephens Council in
Newcastle, Australia, is one such facility. It utilises a combination of ‘accelerated
aerobic composting’ process (including a patented rotating digester system) and product
screening systems to convert the organic fraction of the MSW stream into a range of
high quality compost products, which are subsequently marketed. Apart from
maximising the recovery of recyclable content and removing contaminants, the facility
boasts achieving approximately 45% reduction in the amount of putrescible waste
(Schmidt, 2005).
Compost Application
Compost has numerous agronomic, horticultural, and forestry uses. The largest potential
user of MSW compost is the agricultural industry (Shiralipour et al., 1992). Compost
that does not meet minimum environmental standards for food crop production may be
used for growing nursery stock, forest seedlings, field and container grown ornamental
plants (Wei et al., 2000). Low-grade composts are suitable for establishment and
maintenance of public gardens and landscapes; and for reclamation of disturbed lands.
Compost with excessive levels of heavy metals may be used as landfill cover or for
other uses on land dedicated to disposal of waste materials.
Drawbacks
Even though composting is less expensive than incineration and has few negative
effects on the environment, the process is a capital intensive method when compared to
landfills. In addition, compost products have uncertainties attributed to pathogens,
11
heavy metals, phytotoxic compounds, and foreign objects (Rynk, 1992; cited in Wei et
al., 2000).
2.1.2
Incineration
Although, sanitary landfills provide a simple and affordable solution for SWM
throughout the world, the need for other waste management technology does not end
there. In fact, various technologies were developed in response to some fundamental
flaws associated with landfills. Some of these problems were highlighted as scarcity of
land, prolonged stabilisation of land filled waste and high pollution impact on the
environment. The other motivation was to develop an alternative treatment tool that
would allow waste reduction, material and energy recovery simultaneously within a
short period of time. As a result, the period between 1970 and 2000 saw rapid
development of incineration technologies (Autret et al., 2007; Joos et al., 1999; Larsen
& Borrild, 1991; Yamamura, 1983).
Incineration for SWM was first developed between the 1970s and 1980s in European
countries, when a number of waste-energy plants were constructed.
Germany,
Switzerland, France, Denmark and the Netherlands chose the technology as an
alternative to landfills (Autret et al., 2007; Joos et al., 1999; Larsen & Borrild, 1991).
By the mid 1990s, most Asian countries (Korea, India, Taiwan and China) started to
adopt incineration as an alternative to landfills (Hunsicker et al., 1996). Over time, the
technology also underwent a dramatic improvement. Old incinerators with less capacity
were gradually replaced with new high capacity plants, equipped with energy recovery
and ash treatment facilities (Autret et al., 2007). Today, incineration is the primary
means of waste disposal in several countries, including Japan (59%) and France (93%).
Essentially, incineration is defined as an oxidation by controlled burning of materials to
simple, mineralised products such as carbon dioxide and water. It is also described as a
thermal treatment measure before the disposal of waste (Johnke, 1992). The actual
combustion of waste takes place in the gas phase and simultaneously releases energy.
This leads to a thermal chain reaction and self-supporting combustion, which explains
why there is no need for the addition of other fuels.
12
Depending on the type of waste and its hazardous properties, incinerators are of
different types with some being more elaborate and complex. The three most common
incineration types are grate, rotary kiln and fluidised bed incinerators. Grate incinerators
are widely used for incineration of municipal wastes while rotary and fluidised
incinerators are used for hazardous wastes and pre treated wastes (sewage sludge),
respectively. Grate incinerators form almost 84% of the total incinerators in France
(Autret et al., 2007).
Compared with landfills, incineration has several attractive features. First of all, it
provides an efficient means of recovery of energy from much of the waste. For instance,
the amount of heat produced from combustion is converted to steam and electricity
(Baird, 1999). Secondly, it involves a large reduction in the volume to 6% and
composition of the waste (Larsen & Borrild, 1991). Another major advantage is that it
destroys some or all of the hazardous constituents of the solid waste and eliminates
problems of methane generation and leachate production at the landfill site.
Drawbacks
Incineration is highly expensive and requires extensive use of technology for operation,
making it less effective in terms of cost and resources. This undermines its
implementation in developing countries where financial and technical resources are
already a major problem. In addition, incineration of waste generates a lot of pollutants
(dioxins, furans, acid gases, dust and heavy metals) along with a large amount of solid
slag, called “bottom ash” (Autret et al., 2007; Mendes et al., 2004).
2.1.3
Landfills
One of the main methods of disposing of municipal solid waste is to place it in a
landfill, variously called a garbage dump or a rubbish dump. Initially, landfills were
simply holes or a levelled piece of ground usually used for dumping municipal wastes
(Al-Yaquout et al., 2002; Henry et al., 2006; Ngoc & Schnitzer, 2009; Read et al.,
2001). Such landfills lacked proper design, structure and management, which led to
numerous problems: pollution of surface and underground waters, unpleasant odours,
pest infestation, gas explosions, public health deterioration and flooding (Ayomoh et al.,
2008).
13
However, over time, landfills have undergone an immense change in structure and
design due to social, technical and economic development with the purpose of ensuring
minimal environmental impact (Christensen et al., 1992; Ngoc & Schnitzer, 2009).
Thereby, simple and traditional landfills eventually developed into modern sanitary
landfills that are now the foremost method of solid waste disposal in many countries
across Asia, Africa and Europe (Renou et al., 2008). Together, traditional and modern
sanitary landfills serve as the ultimate disposal destination for more than 90% of the
world’s solid waste.
Design and Structure
The design and structure is the most important aspect of the modern sanitary landfill.
Typically, the structure consists of an excavated site that has numerous “cells”, which
are separate, yet connected so as to function as a single system. Figure 2.2 illustrates the
desired components of a well-controlled landfill disposal facility.
Figure 2.2 A sketch of a MSW landfill construction (Source:
http://www.sfb477.tu-bs.de/english/tp_d1/tpd1.html).
14
Each cell is lined with a lining material used in place of, or in addition to, low
permeability soils as a base or cover liner. Base liners are placed below waste to prevent
leachate from making its way into the surrounding earth and ground water system, while
cover liners are placed above the final waste configuration to keep water, usually rain or
snow melt, from entering the waste. Although, most traditional landfills use clay or
gravel for the landfill base and cover lining systems, materials such as polyethylene
(PE) geo-membranes, poly vinyl chloride (PVC) and chlorinated poly ethylene (CPE)
are also often used.
Once the cells are ready, solid waste is deposited in them on a daily basis and
periodically covered with a layer of earth or dirt, thereby establishing pockets of solid
waste in the landfill (Read et al., 2001). After the landfill is filled to a predetermined
amount, the site is then covered with a suitable covering material such as a layer of
earth, clay or a synthetic liner to minimize leachate formation. Usually, a landfill cover
with a sloping surface is preferred as it enhances surface runoff and reduces the volume
of water inflow (El-Fadel et al., 1997).
Leachate and Leachate Collection System
As soon as the waste is disposed of in the landfill, it immediately undergoes
decomposition under different conditions - aerobic and anaerobic - where it gets mixed
with water, air and moisture from the waste, producing a large volume of leachate.
Leachate is a highly contaminated wastewater characterised by dissolved organic and
inorganic contaminants, capable of contaminating soil and water (Table 2.1).
In order to reduce the potential risk to the environment and public health, leachate must
be consistently monitored and controlled (Christensen et al., 2001; Duggan, 2005; Lim
et al., 2009; Pivato & Gaspari, 2006). Thus, sanitary landfills are required to have
systems to pump and collect leachate for further treatment. Landfills are also designed
with various hydraulic barriers such as trenches and extraction and gradient control
wells to make them more environmentally sustainable (Harris & Gaspar, 1989; cited in
El-Fadel et al., 1997).
15
Table 2.1 Classification of Landfill Leachate According to the Composition
Changes.
Intermediate
Stabilisation
Age (Years)
pH
COD
Type of leachate
<5
<6.5
>10,000
5-10
6.5-7.5
4,00010,000
>10
>7.5
>4,000
BOD5 / COD
0.5-1.0
0.1-0.5
<0.1
Organic
Compounds
80%
Volatile
Fatty
Acids
<400
<0.3
0.1-0.2
5-30% VFA
+ Humic
and Fulvic
Acids
NA*
0.3-0.5
NA
Humic and
Fulvic
Acids
Low medium
Low
Low
Important
Medium
Low
Ammonia
Nitrogen
TOC/COD
Total Kjeldahl
Nitrogen
Heavy Metal
Biodegradability
Young
>400
>0.5
NA
Note: All units in mg/L unless otherwise indicated. * Not Applicable.
(Source: Foo & Hameed, 2009).
Figure 2.3 shows a section of a typical leachate collection system. A leachate collection
system usually consists of a network of perforated pipes located under the cover and/or
above the bottom line. Each pipe is connected to a different cell, which later ends in the
main system. The leachate drains are gravity fed from the landfill cells to lined leachate
holding ponds (SWMC, 2002). The collected leachate is either piped to an onsite
leachate storage tank for treatment or transported to an approved offsite wastewater
treatment plant for disposal. In some landfills, it is recirculated in the landfill itself as a
method of easier treatment and to provide water for microbial processes.
Landfill Gases and Gas Collection System
Apart from leachate formation, landfills pose high pollution potential owing to their
ability to generate toxic gases. According to several researchers, methane and carbon
dioxide are quantitatively by far the two principal components of landfill gas and form
more than 90% of the total gas generated (Christensen et al., 2001, El-Fadel et al., 1997;
Foo & Hameed, 2009). Nitrogen and oxygen are normally present in small quantities.
Table 2.2 summarizes the composition of a typical landfill gas.
16
Figure 2.3 A cross section of a leachate collection system (Source:
http://www.cintec.ca/english/03technologies/01landfills.htm).
Table 2.2 Landfill Gas Composition.
Concentration Range Percent
Component
(Dry Volume Basis)
Methane
40-70
Carbon Dioxide
30-60
Carbon Monoxide
0-3
Nitrogen
3-5
Oxygen
0-3
Hydrogen
0-5
Hydrogen Sulfide
0-2
Trace Compounds
0-1
Source: El-Fadel et al. 1997, p.4.
Various gases are produced due to different decomposition patterns established in the
landfill. During the initial stage, waste decomposition is mainly aerobic and does not
contribute to methane gas generation. OM reacts quickly with oxygen to form carbon
dioxide, water and other by-products (e.g. bacterial cells). However, as oxygen is
depleted within the landfill, the onset of a dominant anaerobic phase decomposition
starts, which is more significant in the formation of methane gas (El-Fadel et al., 1997),
as well as vapour-phase volatile organic compounds (Read et al., 2001). Apart from
carbon dioxide, methane is an explosive greenhouse gas with a global warming
potential estimated to be 23 times greater than that of the same volume of carbon
17
dioxide. Hence, release of methane and carbon dioxide from the landfill surfaces gives
rise to greenhouse gases (El-Fadel et al., 1997; Themelis & Ulloa, 2007). Based on this
determination, it is apparent that landfill gas control and recovery measures are essential
in eliminating or minimizing environmental impacts.
To efficiently control gas and avoid odour problems, gas extraction systems may require
installation of larger pipes, blowers and related equipment early in the landfill’s
operational life. Horizontal trenches, vertical wells, near surface collectors, or a hybrid
system may be used for gas extraction. Typically, landfills should have one well per
acre (0.4 ha) but greater gas flows can also be readily accommodated by increased pipe
diameter. An example of a landfill gas collection and recovery system is shown in
Figure 2.4. With the help of an extractor fan to pull the gas from the collection wells,
landfill gases are then piped to a main collection header where they are treated or flared.
Figure 2.4 A landfill gas collection and recovery system (Source:
Tchobanoglous & Kreith, 2002, p.14.7).
18
2.2
Review of Leachate Treatment Options
Historically, most landfills rely on the slow uncontrolled release of leachate into the
surrounding geology and ground water or surface water (Duggan, 2005). While many
such cases have been superseded by modern sanitary landfills and better technologies,
much still needs to be done in order to reduce contamination of the environment. As a
result, treatment of leachate is still a priority for the waste industry. A treatment method
is referred to as any method capable of reducing the concentration of the pollutants in
the leachate to a level considered safe for the environment and human health. Currently,
a number of options are available for leachate treatment depending upon the nature and
strength of leachate and the degree of treatment required. These treatment options are
based on physical, chemical, biological or a combination of one or two of these
processes.
2.2.1
Aerobic biological treatment (Aeration)
Definition
One of the simplest forms of on-site treatment of landfill leachate is by means of
aeration. A typical aerated lagoon or pond refers to basins constructed in, or on the
ground surface, using earthen dikes to retain wastewater, within which natural
stabilization processes occur with the help of natural or artificial aeration. Because,
most landfill leachates are characterised by high concentrations of OM and ammoniacal
nitrogen, this method is chosen particularly to solve this problem with minimum cost
incurred. It is therefore important to remember that most often, the method may not
achieve the best effluent quality technically possible, but will still provide the effluent
to meet discharge consent conditions, while minimising the total cost of onsite
treatment.
The Concept of Aeration
The principle of aeration is that it can bring about changes in the leachate via chemical
and biological oxidation, thereby allowing its quality to improve (Mehmood et al.,
2009; Nivala et al., 2007; Robinson & Grantham, 1988). The concept of supplying
oxygen to the leachate is to bring about aerobic degradation of organic material and
subsequently transform the nitrogen content of the leachate. As aeration continues, the
19
heterogeneous mass of new cells synthesized is further destroyed due to endogenous
respiration/self oxidation in order to yield the required energy. This happens in all the
aerobic treatment systems such as the aerobic stabilisation pond, oxidation ditches and
aerated lagoons. Once the carbonaceous demand is satisfied, oxidation of nitrogenous
material occurs via nitrification.
Nitrification occurs because leachates usually contain nitrogen compounds in the form
of ammonia (NH3), nitrite (NO2-), nitrate (NO3-), amines and other nitrogenated
compounds. Nitrification occurs in a two-step aerobic biological process. Firstly,
ammonia is converted to nitrite by Nitrosomonas and thereafter, nitrite is quickly
converted to nitrate through the action of Nitrobacter. The process of nitrification is
influenced by several factors: dissolved oxygen (DO); temperature; pH; total alkalinity;
and detention time.
Classification of Aerobic Treatment of Leachate
There are three types of aerobic treatment based on the way total solids in the system
are handled:
i)
Facultative type,
ii)
Aerobic flow-through type,
iii)
Extended aeration type.
Facultative Lagoons
Facultative lagoons are akin to algal ponds used for waste stabilisation except that the
O2 is derived from mechanical aeration instead of algal photosynthesis. In this kind of
pond, the power input is only sufficient to diffuse enough oxygen into the liquid and not
to maintain all the solids in suspension. Consequently, suspended solids in the raw
sewage entering the lagoon tend to settle down and undergo anaerobic decomposition at
the bottom. The activity in such a lagoon is therefore partly aerobic and partly
anaerobic, which contributes to the name “facultative” (Arceivala, 1973). Facultative
ponds in the form of onsite-aerated lagoons are so far the most common method of
treating wastewater in developing countries. Researchers have identified the method as
a promising technology among others because of its extreme simplicity, low cost and
minimum usage of technology (Mehmood et al., 2009; Robinson & Grantham, 1988).
20
Such lagoons can achieve good BOD removals and are mostly used in the treatment of
sewage, industrial wastes and leachates.
Frascari et al. (2004) and Mehmood et al. (2009) reported on the successful facultative
lagoon treatment used for treating landfill leachate in Italy and the United Kingdom
(UK), respectively. The facultative lagoon achieved significant removal efficiency of
64%, 40%, 77%, 77% and 63% for BOD, COD, ammonia, Total Kjeldahl Nitrogen
(TKN), and nitrate, respectively. In the UK, facultative aerobic systems involving
sequential aerobic and anaerobic microbial oxidations achieved 75% COD and 81%
ammonia removal, respectively, at a higher hydraulic retention time of 56 days.
Aerobic Flow-through Type
Similar to the facultative lagoons, the aerobic flow-through uses oxygen or air and
microbial action to bio-treat the pollutants in wastewaters. However, it makes extensive
use of energy to keep all the solids in suspension, ensuring complete mixing conditions.
Another different feature is that this method does not hold back or retain solids. Instead,
solids are allowed to pass out along with the effluent. Consequently, the BOD removal
efficiency is not very high. An example of the efficacy of a flow-through lagoon was
assessed by Robinson & Maris (1983) in a bench scale experiment. Apart from
demonstrating a good removal of BOD (97.5%) and COD (92%) at mean solid retention
time (SRT) of 10 days, the study exhibited the possibility of metal removal via the
process of aeration. Removal efficiency for various metals was recorded as: iron (Fe)
(>98%); manganese (Mn) (>92%); and zinc (Zn) (94%).
Extended Aeration Lagoons
This type of lagoon is similar to the aerobic flow through type in the sense that it also
maintains solids in suspension and complete mixing conditions are attained. However
solids are not allowed to flow out with the effluent, rather, they are deliberately made to
accumulate or build-up in the system, either by providing a separate settling tank or
recirculation, or by incorporation of a settling compartment in the lagoon itself. Similar
methods of biological processes for treatment of wastewaters are also prevalent in
various hybrids but all the methods have in common the use of air and microbial action
to bio-treat the pollutants in wastewater.
21
Design and Operation
Aerated lagoons are generally rectangular in shape and are three to five metres in depth
and built in earthwork with slopes partly or fully pitched in stone or concrete. Figure 2.5
below shows an example of an aerated lagoon built in an concrete structure. The inlet
and outlet are located on opposite banks. To avoid percolation, a lagoon has to be
located in a relatively impervious soil or suitably lined or constructed in masonry and
concrete and in most cases lined with high-density chlorinated polyethylene materials.
Leachate is collected by differing methods including pneumatic pumps, hydraulic
pumps and standard electric pumps. In situations where the landfill is raised above the
land level, it is also possible to use gravity to collect and move leachate at a lesser cost.
One very important feature of an aeration plant is the provision of efficient and proper
diffusion of air. Usually, a range of methods are available - surface aerators, venturi
injectors and air diffusers - but a spiral flow system, which uses an air diffuser is
considered more convenient and advantageous because a single air header can serve two
tanks with diffusers located along one wall of the tank. It also allows fine bubble
diffused aeration, which helps in providing a high level of oxygen transfer and is very
gentle with bacteria. Another less expensive method of aeration is to use floating
aerators and air from atmospheric diffusion and photosynthetic sources.
Figure 2.5 An aerated lagoon plant (Source: http://www.adadget.com/domestic.html).
22
2.2.2
Flocculation
Coagulation/flocculation is a relatively simple technique that may be employed
successfully for the treatment of stabilised or older landfill leachates and wastewater. It
involves physio-chemical processes considered effective for removing metals, turbidity
and refractory humic substances (Tatsi et al., 2003; Li et al., 2010).
Quite often, flocculation is combined with precipitation and sedimentation in removing
metals from leachate. Precipitation involves the addition of chemicals to the leachate to
transform dissolved contaminants into insoluble precipitates, followed by flocculation
and then finally, removed by sedimentation or filtration (McArdle et al., 1988). The
precipitation method is necessary in flocculation because it allows metals to precipitate
from leachates as hydroxides, sulphides, or carbonates by the addition of an appropriate
chemical precipitant and the adjustment of its pH to favour solubility. Although better
removal efficiencies are possible with sulphide precipitation, hydroxide precipitation
with lime or caustic as the precipitant is practised widely because of the materialshandling and cost advantages (Canter & Knox 1986; cited in McArdle et al., 1988).
Principle of Flocculation
In the wastewater treatment industry, the first step in the coagulation-flocculation
method is to destabilise the dispersion and coagulate the contaminants. This is generally
done via the addition of positively charged species, called as coagulants, in appropriate
quantities to neutralise the charge on the impurities. A flocculation step is then used to
bring together the small flocs formed by coagulation so that larger flocs that will
precipitate are produced.
Coagulants form an essential part of the coagulation-flocculation treatment, thus it is
important to know the various coagulants used. The most widely used coagulants are
salts of aluminium (Al) and iron (Fe) such as aluminium sulphate, ferrous sulphate and
chlorides and polyaluminium chloride (Amokrane et al., 1997; Aziz et al., 2007; Li et
al., 2010; Ntampou et al., 2006). While Al and Fe salts have been widely used for
removing humic substances from water, ferric chloride is suggested as a viable
coagulant in managing colour problems in leachate treatment. Polyaluminium silicate
chloride is a new coagulant reagent, which has enhanced aggregating power and the
23
ability to form bigger and denser floc formation (Tzoupanos et al., 2008). Calcium
hydroxide (lime) is used most of the time for precipitation.
Drawbacks
Flocculation firstly allows only moderate removal of COD (or TOC) content. Secondly,
the process results in production of excessive sludge and increased aluminium (Al) or
iron (Fe) concentrations in the resulting effluent (Maranon et al., 2008). As a result,
coagulation-flocculation has been proposed mainly as a pre-treatment method for fresh
leachates prior to biological, physical or other chemical techniques or as a posttreatment technique for partially stabilized leachate (Tzoupanos et al., 2008).
2.2.3
Membrane separation
Membrane separation is based on the principle that semi-permeable membranes allow
only water and certain solutes to pass through them. Essentially, two processes are
involved: ultrafiltration and reverse osmosis.
Based on a similar concept, granular-media filtration uses a bed of granular material to
remove the suspended solids from leachate by forcing the fluid through this porous
medium. The granular media filtration system may be classified by: 1) direction of flow,
2) type of filter beds, 3) waste water driving force, and the method of flow rate control.
For high turbidity leachate, dual or tri-media filter beds are more desirable because they
have greater solids storage capacity.
2.2.4
Activated Carbon Adsorption (ACA)
A notable trend in the development of activated adsorption was seen over the last few
years in wastewater treatment industries. Activated Carbon has a superior ability for
removal of a wide variety of organic and inorganic pollutants dissolved in aqueous
media (Foo & Hameed, 2009) due to its large porous surface area, controllable pore
structure, thermo stability and low acid or base reactivity. Gradually, activated carbon is
used as an adsorbent for separation of marginally biodegradable organic and highly
toxic inorganic micro pollutants (heavy metals) (Christensen et al., 1992). Thus, it has
greater applications in leachates containing non biodegradable substrates, which are not
24
removed by biological treatment alone (Cecen & Cakiroglu, 2001; cited in Cecen et al.,
2003).
Apart from granular activated carbon (GAC), powdered activated carbon (PAC) is also
used in water treatment. Cecen & Aktas (2001) found that PAC is suitable in leachate
treatment based on the ability to enhance biological treatment efficiency and
nitrification by completely preventing nitrification inhibition. PAC, when combined
with activated sludge, improves sludge dewater ability and increases the removal
efficiency by adsorbing non-biodegradable, toxic or inhibitory organics and also some
metals (Metcalf & Eddy, 2004 ; cited in Aghamohammadi et al., 2007).
The advantages of using activated carbon adsorption in the passive leachate treatment
industry over the last 15 years are reflected in several studies (Foo & Hameed, 2009;
Kargi & Pamukoglu, 2004) The studies illustrated the use of GAC and PAC as a
successful treatment option with 91% and 95% COD removal, respectively. Recently,
GAC, granular activated alumina and ferric chloride were utilised for the treatment of
heavy metals (Cd, Cu, Cr, Mn, Pb and Zn). Foo & Hameed (2009) studied the use of
GAC and limestone in removal of ammonium nitrogen.
25
Chapter 3 Solid Waste Management
This chapter compares in brief the waste management system in Bhutan to that of
Summerhill Waste Management Centre (SWMC) in Newcastle. An overview of the
main physio-chemical characteristics of leachate formed at the SWMC sanitary landfill
during the past 14 years is presented. In addition, the outcomes of the statistical
evaluation of long term leachate monitoring data set are discussed in relation to leachate
formation at SWMC sanitary landfill. The correlations between the volume of leachate
generated and the rainfall received at the site are applied in estimating leachate
formation at SWMC sanitary landfill and Bhutan.
3.1
Waste Management in Bhutan
As in other countries, waste management in Bhutan consists of two main steps:
In the first step, collection and transportation of waste is carried from house-tohouse at regular intervals by employing trucks/tippers. The recent trend in urban
centres is towards using concrete receptacles and bins placed at strategic points
from where the garbage is lifted for removal by trucks or tractors (State of the
Environment, Bhutan, 2001). However, all the three components of waste
management - collection, transportation and disposal - lack in terms of
infrastructure, maintenance and up gradation.
The second step deals with disposal of waste in the landfills. In the absence of
alternative waste management methods, 90% of the MSW is disposed of in the
landfills. Generally, each urban centre has only one or no landfill. These
landfills receive the entire MSW. Where there is no landfill site, MSW is
disposed of either in rivers/streams, valleys or in low lying areas. About 50-60%
of the MSW disposed in the landfill is organic in nature, which is an indication
that MSW contains very little glass, electronics and metals (Phuntsho et al.,
2008). The overall composition of the MSW waste in Bhutan is shown in Figure
3.1.
26
Figure 3.1 Average percentage composition of MSW in Bhutan (Source:
Phuntsho et al., 2008).
3.1.1
Problems of SWM in Bhutan
In many regions, the solid waste problems are becoming acute. New sites that are both
accessible and technically suitable for landfills are almost impossible to obtain, due to
lack of vacant land, finance and labour. Currently, Memelakha Waste Disposal Site in
Thimphu is the largest landfill in the country. It is located 12 km away from the city and
receives about 64.5 tonnes of MSW daily. Collection of solid wastes, transportation
and disposal at the landfill site was initiated in 1993. However, the landfill has exceeded
its volumetric capacity of 70,000-80,000 m3 for years and is already overflowing with
wastes. Yet no alternative site has been identified (Allison, 2008; Penjor, 2008).
Furthermore, the landfill does not have adequate leachate and gas collection and
treatment systems.
Likewise, there are already serious concerns about the sustainability and management of
landfills in several other cities. Concerns centre on three main features: (i) increasing
population and waste output (ii) lack of land to identify a landfill and (iii) pollution
potential of the landfill leachate and gases.
Apart from the problem of disposal, the worst problem of waste management in Bhutan
is the lack of proper segregation, sorting and recycling of waste (Phuntsho et al., 2008).
27
Proper segregation of waste into different components and their separate collection do
not occur at source. Thus, the entire organic and inorganic wastes are disposed of in the
landfills.
3.2
Summerhill Waste Management Centre (SWMC),
Newcastle
Newcastle is the second most populated area in the Australian state of New South
Wales. It is situated 162 kilometres northeast of Sydney on the southern bank of the
Hunter River. According to the 2006 census the city has a population of 288,732
persons, which is equivalent to a population of several cities in Bhutan and one third of
population of Bhutan.
SWMC in Newcastle is a high-tech, high capital cost facility with many state-of-the-art
aspects of waste management (Whitehead et al., 2000). Located on the site of a former
open-cut and underground coal mine, which includes tracts of remnant vegetation and
areas disturbed by disused workings, the site is close to Wentworth Creek and Flaggy
Creek catchments (SWMC, 2002). It is a class one solid waste landfill managed by the
City of Newcastle and has a total approved capacity of 3.5 million m3 (Whitehead et al.,
2000). Since 1995, some 150,000 tonnes of waste has been received at the site annually.
The waste received by the landfill includes both putrescible and non-putrescible waste.
However, unlike in Bhutan, where MSW mainly originates from residential sources
(69%) (Penjor, 2008), the majority of the waste received by the Summerhill landfill
originates primarily from commercial and industrial sites in Newcastle. A smaller
portion of the waste comes from public or council operations. For example, in 2001 and
2002, 55% of its waste received by the landfill was produced by commercial and
industrial sites; 29% household and 16% by the public or from council operations. Solid
waste containing large proportions of putrescibles waste constitutes the major portion
(52%) of the waste stream (SWMC, 2002). This indicates that about 50% of the waste
received at the centre undergoes compaction, decomposition and thus, is liable for
leachate production in the long run.
28
Unlike landfills in Bhutan, the special feature of the SWMC landfill is that it comprises
a series of discrete sections called “cells” where different wastes can be processed and
deposited. The cells are constructed with a foundation layer of between 0.5 and 2.0
metres of onsite clay over a prepared subgrade. It is then lined by manufactured
geocomposite clay and a high density polyethylene (HDPE) synthetic liner. The HDPE
liner functions as an impermeable barrier to leachate and helps in collection and
reduction of leachate generation, thereby minimising the contamination of groundwater
and downstream surface water (Whitehead et al., 2000).
In contrast to disposal of waste in Bhutan, SWMC has a special method of disposing
waste in the cells. Waste is deposited and compacted in successive layers called “lifts”
and covered daily with approximately 150 mm of soil in order to minimise and control
water infiltration, odour, flies, litter, fire and vermin. Finally, the cells are covered with
600 mm of compacted low permeability clay cover, 1250 mm of bulk capping, topsoil
and vegetation (Whitehead et al., 2000).
On top of that, the landfill site has been designed to carry out various functions
necessary for proper management of waste. The design consists of a computerised
weigh bridge system, waste segregation and recycling facilities. The centre also has an
administrative, exhibition and educational facilities on waste minimisation and
recycling activities. Other provisions of the centre include comprehensive monitoring
of leachate, surface and groundwater, noise, dust, odour and litter (SWMC, 2002;
Whitehead et al., 2000). The landfill sites in Bhutan do not have such facilities.
While Bhutan does not have a pricing mechanism imposed on waste disposal, SWMC
imposes a levy of $65.30 per tonne on all waste disposed of at the centre. The
imposition of landfill levy is in accordance with New South Wale Government’s Waste
Minimisation and Management Act 1995, which set an ambitious target of a 60%
reduction in waste to landfill by 2000 (Oakes, 2009).
29
3.3
Leachate Characterization and Statistical Evaluation
The characteristics of a typical leachate generated at the SWMC landfill have been
investigated by analysis of long-term leachate monitoring data collected from the centre.
The main purpose of this analysis was to understand various aspects of a landfill
(sanitary), particularly, the pollution potential associated with leachate formation. Apart
from this, the investigation was expected to establish, if any, the correlation of leachate
characteristics to some important climatic factors (rainfall), which might be useful in
considering leachate characteristics and leachate generation in Bhutan.
Table 3.1 presents a summary of the main physio-chemical characteristics of the
leachate monitored during the past 14 years by SWMC. The different water quality
parameters are divided into three different groups and the values are calculated by
various preliminary statistical treatments, i.e., calculation of average (mean) and range
(minimum to maximum).
A wide variation is observed in the quality of leachate. Overall, the organic content of
the leachate exhibited very high variations followed by inorganic macro components.
Trace elements were found only in traces and showed much lesser variation. Physical
pollution parameters such as pH, EC, total alkalinity, and suspended solids also showed
frequent variations. The organic content of the leachate are indicated by various
parameters (BOD, COD and TOC) in the range (min-max) 2-2750 mg/L, 80-3850
mg/L, 1-1840 mg/L, respectively.
In terms of inorganic macro components, ammonia-nitrogen is the most significant
pollutant with concentration in the range 0.01-970 mg/L. With the exception of nitritenitrogen (0.01-0.50 mg/L), nitrate-nitrogen (0.01-0.2 mg/L), phosphate (0.10 mg/L) and
Manganese (Mn) (0.2-3.0 mg/L), metal salts such as sodium (Na) and potassium (K)
were also present in high concentrations (50-2160 mg/L and 5-900 mg/L, respectively).
Chloride and sulphate varied from 30-2450 mg/L and 1-400 mg/L, respectively.
30
Table 3.1 Composition of Landfill Leachate (1995-2009) from SWMC,
Newcastle, Australia.
pH
No. of
sample
188
Range
6-8
Mean
7.8
Electrical Conductivity
(µs/cm)
157
105016500
6250
59
170
76-6950
10-7220
1600
337
187
2-2750
306
180
178
80-3850
1-1850
944
309
189
37
0.01-970
0.01-0.50
200
0.10
78
1
78
84
78
78
77
78
56
46
0.01-2.00
0.10
30-2450
1-400
50-2160
5- 900
2-300
30-250
0.01-430
0.2-3.0
0.10
0.10
728
85
597
217
86
118
19
1.0
42
79
9
0.00-0.05
0.00-0.20
0.01-0.07
0.01
0.05
0.03
Parameter
Total Alkalinity (as
mg/L as CaCO3)
Suspended Solids (NFR)
Organic Matter
Biochemical Oxygen
Demand (BOD5)
Chemical Oxygen
Demand (COD)
Total Organic Carbon
(TOC)
Inorganic
macro components
Ammonia-nitrogen
Nitrite-nitrogen
Nitrate-nitrogen
Phosphate
Chloride
Sulphate
Sodium
Potassium
Calcium
Magnesium
Iron
Manganese
Inorganic trace elements
Cadmium
Chromium
Cobalt
Copper
Lead
Mercury
Nickel
Zinc
0.0076
18.00
0.50
43
0.00-0.10
0.04
17
0.00-0.08
0.01
74
0.01-0.4
0.10
45
0.01-0.1
0.20
Note: All units in mg/L unless otherwise indicated.
31
Compared to Na and K, the concentrations of other metal salts are comparatively low
and lay in the range 2-280 mg/L for calcium (Ca), 30-250 mg/L for magnesium (Mg)
and 0.01-430 mg/L for iron (Fe). The high concentration of inorganic components might
have led to the high values of conductivity, total alkalinity and suspended solids.
On the other hand, a large number of xenobiotic organic compounds (XOCs) (aromatic
hydrocarbon, phenols, chlorinated aliphatic compounds and pesticides) and heavy
metals (cadmium (Cd), chromium (Cr), cobalt (Co), lead (Pb), mercury (Hg), nickel
(Ni) and zinc (Zn)) were present only in traces with most of their concentrations
remaining below 0.5 mg/L. The Australian Water Quality Guidelines for Fresh and
Marine Waters (AWQGFMW) (1992), considered 0.5 mg/L or lower as a limit value
safe for fresh and marine waters. Further detail of some of this important parameter with
increasing age of the landfill is discussed in the section below.
3.3.1
Temporal variation in leachate quality
The analysis of the leachate composition data for the past 14 years indicated that
temporal variation in leachate quality is evident at the SWMC landfill. Figures a-g in
Appendix B-1 illustrate the temporal variations in some of the main pollutant
parameters of the leachate as a function of time (landfill age) that extended over a
period of 14 years (1995-2009). The parameters are ammonia-nitrogen, pH, EC,
suspended solids, BOD, TOC, COD, metals and salts. The respective curves clearly
show that the landfill underwent two distinct decomposition phases with increasing age.
The first phase (1995-1999) resulted in production of weaker or less toxic leachate,
while the latter (2000-2009) recorded the production of higher strength leachate.
The First Phase - from 1995 - 1999
The first phase of the landfill included an initial stabilisation phase that lasted for
approximately four to five years. Leachate collected during this phase was a result of
first decomposition activity taking place within the landfill. Mean concentrations of the
resulting leachate shows that the leachate produced during the initial period is less toxic
with a fairly neutral pH (mean value ~7) and low concentrations of ammonia-nitrogen
(6 mg/L), nitrite-nitrogen (0.01 mg/L), nitrate-nitrogen (0.03 mg/L), Na (120 mg/L), K
(3 mg/L) and Fe (1 mg/L). Both chloride and sulphate are approximately estimated as
32
130 mg/L. Trace elements have even lower concentration of less than 1 mg/L. Total
alkalinity measured 580 mg/L asCaCO3 and EC value was 850 µS/cm. The organic
compounds in the leachate were detected only towards the end of the first year as
indicated by high BOD (900 mg/L) and COD (1300 mg/L).
The Second Phase - from 2000-2009
Unlike the first phase, the second phase of landfill was marked by a significant change
in the quality of leachate, indicating a gradual change in its decomposition pattern. The
phase was marked by an increasing concentration of ammonia-nitrogen, chloride,
sulphate and Na and K. The leachate became slightly alkaline along with an increase in
EC and alkalinity. This assumption is in agreement with Chen (1996), who reported that
an increasing age of landfill led to an increase in pH value up to a certain steady state.
In the case of high EC, the author attributes this to the increasingly dissolved solids
present in the leachate.
The changing pattern of the BOD and COD are in agreement with the changes observed
in other landfills i.e., BOD decreased with time while COD kept increasing and that
COD is higher than BOD and TOC. Of the nitrogenous compounds, ammonia-nitrogen
increased rapidly owing to the deamination of amino acids present in the organic waste
(Tatsi & Zoubolis, 2002). Since the concentrations of nitrite-nitrogen and nitratenitrogen remained constant without much change, ammonia nitrogen is the major
portion of total nitrogen present during this phase, hence an anaerobic state.
However, more changes were observed with increasing age. Leachates became more
concentrated and hence more toxic. In addition to ammonia, the landfill also produced
methane gas. A routine surface gas monitoring in 2002 detected three locations where
methane gas levels were greater than 1% Lower Explosive Limit (LEL) (SWMC, 2002).
In addition to the normal decomposition changes that are described above, highly
concentrated leachates were produced between 2001 and 2003. During this period,
leachates showed an unusual increase in some pollutant parameters, particularly with
respect to ammonia, OM and salinity. Therefore, a pollutant peak can be observed in
this period in the Figures a-g in Appendix B-1. This is related to a reduction in dilution
33
caused by the comparatively lower amount of rainfall (900 mm/year) received in this
period compared to the earlier period (1000 mm/year). Christensen et al. (2001)
observed similar changes in other landfills as well.
3.4
Outcomes of Leachate Characterization and Evaluation
The characterization of long term leachate data set resulted in three significant
outcomes:
i)
Formation of SWMC landfill leachate and its toxicity,
ii)
The relation between climatic conditions (rainfall) and leachate formation at the
SWMC landfill site,
iii)
Prediction of leachate formation in Bhutan.
Firstly, the analysis indicated the possible time of leachate formation and the main
pollutant parameters of the SWMC landfill leachate. That is leachate formation at this
landfill site occurred as early as the first year of landfill operation in 1995. Hence,
researchers have suggested leachate formation as one principal cause of environmental
risks associated with landfills (Christensen et al., 2001; Duggan, 2005; Lim et al.,
2009; Pivato & Gaspari, 2006; Tatsi & Zoubolis, 2002).
Ammonia-nitrogen was the main pollution potential of the SWMC landfill leachate
between 1995 and 2009. Although Figure 3.2 revealed that high concentrations of
ammonia-nitrogen
occurred
only
rarely,
its
mean
concentration
remained
approximately >200 mg/L, which is much higher than the 0.1-2.5 mg/L limit
recommended by AWQGFMW (1992). Thus, it is clear that ammonia can have
negative environmental impact through toxicity. Leachates of this nature, if released
directly into the surrounding environment, can elevate ammonia concentrations in fresh
waters by almost 100 times making it unsafe for fresh water aquatic organisms. Above
0.1-2.5 mg/L, ammonia can cause acute and chronic toxicity to fresh water fish
(AWQGFMW (1992).
Apart from ammonia-nitrogen, the high concentration of Fe (mean 19 mg/L or 19000
ug/L), is likely to make the ammonia in the leachate more toxic. Warnock & Bell (1969)
34
in AWQGFMW (1992) reported that Fe concentration between 320 µg/l and 16,000
µg/l can be toxic to aquatic insects. Fresh water bodies are considered safe only at Fe
concentration of 1000 µg/l or 1 mg/L.
Figure 3.2 Frequency distribution curve of ammonia between 1995 and
2009.
Secondly, the study of leachate formation at the SWMC site under varying rainfall
conditions [ i) monthly leachate production (2001-2009) against monthly rainfall (mm)
(Figure 3.3) and ii) annual leachate production (1995-2009) against annual rainfall
(mm) (Figure 3.4)] revealed that leachate percolation is closely correlated with rainfall
during the monitoring period. The seasonal variation in leachate formation is evident
from the fact that the summer season with higher rainfall resulted in higher volumes of
leachate. Climatic conditions (mean annual) during summer are generally characterized
by higher rainfall, temperatures and relative humidity compared to winter. Similarly, the
annual leachate production was roughly proportional to the amount of annual rainfall
received in each year of the monitoring record. Further confirmations of this
relationship have been derived using linear (Figure 3.5a, b) and bivariate correlations
(Table 3.2) between the rainfall and leachate formation. The results supported the
hypothesis that a positive relationship exists between the two variables and indicates
that higher volumes of leachate generated are associated with higher levels of rainfall
(r=0.488 and 0.606, p<0.05).
35
Figure 3.3 An illustration of six month moving average – leachate percolation versus rainfall (January 2001- December 2009).
36
Figure 3.4
An illustration of the annual leachate percolation versus annual rainfall received at the landfill site (19952009).
37
a
b
Figure 3.5 A linear relationship between rainfall (mm) and leachate volume (KL): a) monthly (R2= 0.238) and b) annual (R2=
0.367).
38
Table 3.2 Correlations Between Rainfall in mm and
Leachate Volume in KL: i) monthly (2001-2009) and ii)
annual (1995 -2009).
Monthly Leachate
Monthly
Volume (KL)
Rainfall (mm)
Monthly
Leachate
Volume
(KL)
Pearson
Correlation
Sig. (1tailed)
N
1
0.488**
108
Annual
Leachate
Volume (KL)
0.000
108
Annual
Rainfall
(mm)
Pearson
1
0.606**
Correlation
Sig. (1Annual Leachate
0.008
tailed)
Volume (KL)
N
15
15
Note: ** Correlation is significant at the 0.01 level (1-tailed).
Similar relations between climatic conditions and the amount of leachate production in
different climatic zones have been reported by several researchers (Frascari et al., 2004;
Nora, 2007; Robinson & Grantham, 1988; Tatsi & Zoubolis, 2002). While Frascari et
al. (2004) confirmed close relations between rainfall and leachate production at an
active Italian landfill site in north Italy, Tatsi & Zoubolis (2002) reported that
Mediterranean climates with dry, hot seasons resulted in reduced percolation of leachate
at a landfill in Thessaloniki, Greece. Further evidence of the effect of rainfall on
leachate production is observed in countries like Thailand and Malaysia, which have
extreme climates characterized by high intensity rainfall (up to 80 mm/day and above)
in the rainy season and no rainfall at all in the dry season (Nora, 2007).
Finally, considering the fact that rainfall is one of the most significant factors
responsible for leachate formation at any landfill site, landfills in Bhutan are likely to
generate very high volumes of leachate as compared to other countries because the
majority of the urban settlement areas are located in areas that receive more rainfall
(500-5000 mm per year) (Table 3.3).
39
Table 3.3 Spatial and Altitudinal Variation in Annual Rainfall by Region
and Percentage of Total Area in Different Altitudinal Zones in Bhutan.
Altitude(m)
%
Area
Region
Below 600
5.3
600-1800
22.4
1800-4000
51.8
Above 4000m
elevation
20.5
Southern
border area
Southern
foothills
Inner central
valleys
/Himalayas
Greater
Himalayas
Annual
rainfall
(mm)
30005000mm
12002000mm
5001000mm
Less
than
500mm
Source: Country Report on the Forest Resources of Bhutan, 2000.
Unlike Australia, the monsoon or summer season in Bhutan falls between June and
August; therefore, higher leachate percolation could occur during this period. The
other implication is that landfill sites situated in different altitudinal zones will have
different leachate percolation because the amount of rainfall received decreases with
the increase in altitude, as shown in Figure 3.6.
Figure 3.6 A plot of variation in rainfall with changes in season and altitude
in Bhutan.
40
Chapter 4 Materials and Methods
This chapter describes the two elements of the research undertaken. It includes
procedures undertaken to study two bench scale laboratory experiments and leachate
characterisation from long term monitoring of data collected at the SWMC, Newcastle.
The objective of this element of the research was to evaluate the suitability of two
passive leachate treatment options, which are simple, effective and affordable. In order
to achieve this, two independent bench scale experiments were conducted in the
laboratory - (1) a passive aeration of leachate in 25 litre tanks and (2) column
experiments using three different filter materials (Granular Activated Carbon (GAC),
Blast Furnace Slag (BFS) and sand). The laboratory scale experiments were undertaken
to determine whether the results can be applied to or assist with the design of larger
scale treatment methods suitable for treatment of landfill leachate in Bhutan. The
procedures associated with each of the experiments are described below.
4.1
Aerobic Biological Treatment
4.1.1
Site selection
Figure 4.1 shows the modern sanitary landfill at SWMC, Newcastle, NSW. The
background of the landfill is described earlier in Section 3.2.. General climatic
conditions between 1995 and 2009 include mean rainfall and temperature of 1165
mm/year and 24 oC, respectively, 70% humidity and minimum mean sunshine of 7
hours. The climate data were obtained from the Williamtown meteorological station
(site number 061078), which is located approximately 16 km from the landfill site at an
elevation of 9 m above sea level (Bureau of Meteorology (BOM), 2010).
4.1.2
Physical and chemical parameters of the leachate (sample)
Nine different water quality parameters were examined as part of this experiment:
turbidity and colour; pH; total alkalinity; electrical conductivity; ammonia-nitrogen;
nitrite-nitrogen; nitrate-nitrogen;and SRP (SRP). Additional factors such as DO and
41
temperature were also examined in this experiment. All the methods are based on
Standard Methods for the Examination of Water and Wastewater (Eaton et al., 1995).
The solid waste disposal area.
Figure 4.1 SWMC landfill and methane collection facility.
Turbidity and colour
Turbidity of the waste water indicates the “clarity” of the sample and is an important
determinant of its condition and productivity. Also defined as the expression of the
optical property that causes light to be scattered and absorbed, turbidity is caused by
suspended and colloidal matter such as clay, silt, finely divided organic and inorganic
matter, plankton and other microscopic organisms. The higher the intensity of scattered
light, higher the turbidity. In this study, turbidity was determined by the Absorptometric
method using a HACH turbidity meter 2100P. In addition to turbidity, colour was
qualitatively assessed.
42
Acidity (pH) and Total Alkalinity
While pH indicates the acidity of water or a solution, total alkalinity of water is its acid
neutralising capacity. Total alkalinity indicates the buffer capacity of the water, that is,
its ability to resist small additions of acids to change pH. It is expressed as mg/L as
calcium carbonate (CaCO3) and indicates the concentration of carbonate, bicarbonate
and hydroxide contents of the fresh water. The measured values may also include
contributions from borates, phosphates, silicates or other bases, if present. Alkalinity is
primarily used in interpretation and control of water and wastewater treatment
processes. In this study, total alkalinity of the leachate samples was determined by
HACH Alkalinity test kit digital titrator-16900 (model-AL-DT) for alkalinity values in
the range 10 - 4000 mg/L as CaCO3. pH was determined by a pH meter connected to a
HQ 40d DO probe.
Electrical conductivity (EC)
EC is a measure of the ability of an aqueous solution to carry an electric current. This
ability depends on the presence of ions; on their total concentration, mobility, and
valence; and on the temperature of measurement. Solutions of most inorganic
compounds are relatively good conductors. EC was determined by the conductivity
meter attached to a HQ 40d DO probe. The EC electrode was first calibrated to EC of
1412 µS/cm before determining the EC.
Inorganic Constituents (Nitrogen compounds and Phosphorus)
Ammonia-nitrogen (NH3-N) is produced largely by deamination of organic nitrogen
and by hydrolysis of urea. It is present naturally in surface and waste waters. Ammonia
concentrations encountered in water vary from >10 ug/L in some natural surface and
ground waters to more than 30 mg/L in some wastewaters. In the laboratory, NH3-N is
determined by the Phenate Method, which is based on the principle that an intensely
blue compound, indophenol, is formed by the reaction of NH3, hypochlorite, and phenol
catalysed by sodium nitroprusside. The method used HACH TNT 832 test reagent
capable of determining high range NH3-N in the range 2-47 mg/L.
Nitrite-nitrogen (NO2--N) is an intermediate oxidation state of nitrogen both in the
oxidation of ammonia to nitrate and in the reduction of nitrate. Such oxidation and
43
reductions may occur in waste water treatment plants, water distribution systems and
natural waters. NO2--N
2
-
is the actual etiologic agent of methemoglobinemia. NO2--N
is determined by HACH TNT 839 test reagent, which can determine NO2--N
-
concentration in the range 0.015-0.60 mg/L and 0.5-2.00 mg/L. The test is based on a
principle that nitrites react with primary aromatic amines in acidic solution to form
diazonium salts. These combine with aromatic compounds that contain an amino group
or hydroxyl group to form intensively coloured azo dyes.
Nitrate-nitrogen (NO3--N) forms part of the nitrogen cycle and generally occurs in trace
quantities in surface water but may attain high levels in some groundwater. In excessive
amounts, it is considered to contribute to methemoglobinemia in infants. A limit of 10
mg/L NO3--N has been thus imposed on drinking water to prevent this condition. NO3-N is determined using high range (5-35 mg/L and 22-155 mg/L) HACH TNT 836 test
reagent based on the principle that NO3--N ions in solute ions containing sulphuric acid
and phosphoric acids react with 2, 6-dimmethylphenol to form 4-nitro-2, 6dimethylphenol.
Soluble Reactive Phosphorus (SRP) occurs in natural waters and in waste waters
almost solely as phosphates in the form of particles or detritus, or in the bodies of
aquatic organisms. These are classified as SRP, condensed phosphates and organically
bound phosphates. They arise from a variety of sources such as surface run off or
drainage water. The main concern related with phosphorus is the eutrophication of
surface waters. Phosphorus determination in this experiment used the principle that
phosphate ions react with molybdate and antimony ions in an acidic solution to form an
antimonyl phosphor molybdate complex, which is reduced by ascorbic acid to phospho
molybdenum blue. HACH TNT 843 test reagent for determining low range SRP (0.051.50 and 0.15-4.50 mg/L phosphorus was used for the test. All the inorganic
constituents were measured by HACH DR2800 Spectrophotometer.
4.1.3
Characterisation of leachate sample
Raw landfill leachate was collected from the leachate pond at SWMC landfill in
February 2010 and was maintained in a condition with minimum exposure to oxygen.
To avoid contamination, clean polyethylene containers were used to collect the samples.
44
All the parameters were determined within an hour and the remaining sample was
stored in the refrigerator at 4°C until needed. The results are shown in Table 4.1.
Table 4.1 Initial Physio-Chemical Characteristics of the Raw Leachate
Sample (11th February 2010).
Parameters
Initial condition
(before aeration)
21.6
Temperature (oC)
Dark Brown
Colour
110
Turbidity (NTU)
8.2
pH
1510
Total Alkalinity (as mg/L as CaCO3)
6.64
Electrical Conductivity (dS/m)
0.5
Dissolved Oxygen
600
Ammonia-nitrogen
8.4
Nitrate-nitrogen
0.2
Nitrite-nitrogen
13
Soluble reactive phosphorus
Note:* All units in mg/L unless otherwise indicated.
4.1.4
Set-up of the aeration experiment
Two tanks (A and B) of similar volume (223×10-6 m3) were filled with approximately
25,000 mL of leachate. Tank A was subjected only to natural aeration (without artificial
aeration), while two aerators were fixed on Tank B to allow aeration. The setup of the
experiment is shown in Figure 4.2. The flow rate of each aerator was approximately 250
L/hour. The laboratory was air conditioned at a constant temperature of 22±1oC.
Approximately 100 mL of leachate from both tanks were sampled and analysed at
intervals of approximately 48 hours for nine different parameters. In addition, DO and
temperature were also monitored. The procedure was repeated until ten different
readings were obtained on the 2nd, 4th, 6th, 8th, 10th, 12th, 14th, 16th, 18th and 20th days of
aeration. The results are recorded in Table 5.1.
4.2
Column Experiments
4.2.1
Setup of the column experiments
The experiment consisted of three columns (Column 1, 2 & 3) (Figure 4.3a), each
having an overall height of 0.28 m and a cross sectional area of 0.0019 m2. These
45
Control
Experiment
Figure 4.2 Set up of the aeration experiment using glass-sided tanks and
aerators showing Tank A (left) and Tank B (right).
columns were designed to ensure uniform flow through the entire cross-section of the
column. A polythene mesh was placed at the bottom of each column to prevent loss of
the filter materials and clogging in the opening of the column. The columns were then
filled with the filter material up to a height of approximately 0.2 m. They were lightly
tapped to achieve some compaction and in order to allow uniform distribution of the
filter material. The volume of media amounted to approximately 377×10-6 m3. The
columns were mounted on a stand as shown in Figure 4.3b. Column 1 was a control
while columns 2 and 3 functioned as the experiment and were designed to replicate one
another.
Once the columns were setup, 500 mL of de-ionized water was added to Column 1
under a constant head loading condition of 0.02 m. The same volume of leachate sample
was added to columns 2 and 3. About 100 mL of sample was taken from each column,
when the respective samples infiltrated through the column for the first time (1st run).
The sample was then analyzed for nine different parameters as shown in Table 5.4. In
the same manner, the remaining samples from each column were collected and
recirculated ten times. This was done with the aim of obtaining another batch of
samples (2nd run and 3rd run), each corresponding to different contact time between the
sample and the filter materials.
46
a
b
Figure 4.3 Setup of the column experiment showing: a) different columns
and b) constant head apparatus.
The assumption was that with an increase in the number of filtrations or runs, greater
contact times between the sample and filter materials resulted. However, only two
batches of samples were taken, i.e., one at the end of the fifth run and another one at the
end of the 10th run. The approximate contact times were estimated and presented in
Table 5.3 in Chapter Five. The samples were finally analyzed and results recorded. The
same procedure was repeated for BFS and sand. In summary, a matrix of 27 (3 - 1st run,
2nd run and 3rd run) × (3 columns) × (3 different filter materials) was obtained. All
experiments were carried out at an approximate temperature of 22±1oC, inside an air
conditioned laboratory.
4.2.2 Characterisation of leachate sample
Although the leachate sample was collected from the same site as in the aerobic
treatment, it was necessary to characterize the sample again before the experiment
because samples were collected at different times of the year and hence, differ in their
chemical characteristics. Thus, the initial analyses of the leachate sample and de-ionized
water used for the column experiment are shown in Table 4.2.
4.2.3
Determination of physical and chemical properties of the filter materials
Physical Properties of the Filter Materials
Determination of pH, EC, bulk density, porosity percent and void ratio
47
The pH and EC of the filter materials were obtained by dissolving the respective
materials in de-ionized water at a ratio of 1:5 (weight: volume). The solution was then
agitated for 30 minutes at a speed of 300 OSC and the respective readings were
recorded using pH and EC meter attached to HQ 40d DO probe.
Table 4.2 Initial Physio-Chemical Characteristics of the Raw Leachate
Sample and DeIonised Water Used in Column Experiments (18th May 2010).
Parameters
DeIonized
Leachate
Water
Colourless
Dark
Colour
Brown
0.3
59
Turbidity (NTU)
6.8
8.4
pH
30
4700
Total Alkalinity ( as mg/L as
CaCO3)
15
10480
Electrical Conductivity (µS/cm)
o
NA
NA**
Temperature ( C)
9
2.2
Dissolved oxygen
UMR*
975
Ammonia-Nitrogen
UMR
0.2
Nitrite-Nitrogen
UMR
10
Nitrate-Nitrogen
0.02
25
Soluble reactive phosphorus
Note: All units in mg/L unless otherwise indicated. * Under Measuring Range.
**Not Applicable.
Bulk density, percent porosity and void ratio of each medium were obtained using the
formulae below;
Bulk Density (g/cm3) = Mass of known volume (g)/ Volume (cm3)
Porosity (n) = Volume of voids (mL)/ Total Volume (mL)
Void ratio (e) = Volume of voids (mL)/ Volume of Solids (mL) (Brady & Weil, 1996, p.114).
Particle Size Distribution
The sieve analysis undertaken in the laboratory involved a nested column of sieves with
wire mesh (screen) of varying sizes from 5 mm to less than 0.063 mm. The sieves were
placed in order of decreasing size, from top to bottom. A representative weighed sample
(approximately 400 g) was poured into the top sieve, which has the largest mesh
openings. Each lower sieve in the column has smaller openings than the one above and
finally the sample is received at the base in a round pan. The column was then placed in
a mechanical shaker. The shaker shakes the column for about 15 minutes to allow the
48
material whose diameter is smaller than the mesh opening pass through the sieves. After
the aggregate reaches the final pan, the amount of material retained in each sieve was
then weighed. The weight of the sample of each sieve was then divided by the total
weight to give a percentage retained on each sieve (Equation 4.1).
Equation 4.1
Calculation of filter material % retained on sieve (Eaton
et al., 1995).
where WSieve is the weight of material in the sieve and WTotal is the total weight of the
material. The % retained was then subtracted from 100% to find the % passing (% finer
by weight) through the sieve as shown in equation 4.2.
Equation 4.2
Calculation of % finer by weight (Eaton et al., 1995).
The values were plotted on a graph with percent passing on the Y axis and sieve size
on the X axis (Figure 5.10 in Chapter Five).
Determination of Hydraulic conductivity or permeability (Ksat-mm/hr)
Based on the volume of influent, the surface area of the columns (0.00189 m 2), and
using 1 mm =1 L/m2 and the time taken for the influent to infiltrate through the media,
the hydraulic conductivity or permeability (Ksat-mm/hr) was calculated as shown in
equation 4.3.
Equation 4.3
Calculation of hydraulic conductivity of the filter
material (Lucas, 2007, p.73).
The summary of all the physical properties determined are presented in Table 5.2 in
Chapter Five.
49
Chemical Properties
Since the chemical properties of the filter materials were not determined in this study,
the chemical properties determined by other researchers were referred as a source of
information on the likely composition of the three filter materials used in this
experiment.
BFS is an industrial by-product resulting from steel-iron making processes (Kietlinska
& Renman, 2005; Oguz, 2004). Due to high percentages of alumina and silica (Table
4.3), BFS is considered a good material for use as an economic adsorbent for large-scale
use. The other reason is that it is easily available and affordable (Nehrenheim et al.,
2008; Oguz, 2004) making it reasonable for use in the low cost treatment of waste water
and leachates.
Table 4.3 Typical Chemical Composition of BFS (amorphous-0.25-4 mm).
Parameter
% by weight.
SiO2
35.5
CaO
35
Al2O3
9.6
FeO
0.3
MgO
13.7
MnO
0.4
S
1.4
V2O5
0.1
TiO2
1.7
Others
2.3
Source:Johansson (1999).
Similar to BFS, the main components of the sand were SiO2 (69.3%), Al2O3 (13.4%),
K2O (3.4%) and Fe2O3 (3.1%) (Kietlinska & Renman, 2005). In addition, it also
contained traces of sodium (Na), magnesium (Mg) and calcium oxides. It is the finest of
the three filter materials considered for this experiment.
On the other hand, GAC is an entirely different filter material made of carbon. It is a
good adsorbent medium due to its high surface area to volume ratio. One gram of a
typical commercial activated carbon is reported to have a surface area equivalent to
50
1,000 m2 (Carbtrol Corporation, 1992). This high surface area permits the adsorptionof
a large number of contaminant molecules.
4.3
Materials for Leachate Characterisation and Evaluation
Three different data sets were obtained from different sources for the characterisation of
leachate:
Long term leachate monitoring data from the SWMC including leachate
volume and rainfall,
Climate data from the BOM, Australia website (www.bom.gov.au)
retrieved November, 2010),
Climate data of different regions in Bhutan obtained from Meteorology
Section, Hydro-met Services Division, Department of Energy, MTI,
Thimphu, Bhutan (1996-2009). In addition, secondary information
about SWM systems in Bhutan were collected using previous research
journals, documents and websites of different non government
organisations, government departments and media.
The data were analysed using various features of PASW Statistics 18 and Microsoft
Word Excel 2007. T test, correlation and descriptive analysis were performed
accordingly.
51
Chapter 5 Results
This chapter presents the results obtained from two bench scale laboratory experiments.
The results from the aerobic treatment of leachate are reported first followed by the
column experiments using three different filter materials to ascertain their viability and
suitability in leachate treatment. Both experiments of the research evaluated the
possibility of treating leachate quality in a simple, affordable and effective way.
5.1
Aerobic Biological Treatment
The results of the aerobic treatment of leachate are discussed in this section. The change
in the characteristics of the leachate samples in the two treatment units (Tank A and
Tank B) were monitored for a period of 20 days, using two glass sided tanks of 25 L
capacity. Tank A was not aerated and functioned as a control for the experiment,
whereas Tank B was aerated with two mechanical aerators. At an interval of every 48
hours, about 50-100 mL of leachate were obtained from the respective tanks (Tank A
and Tank B) and analysed for a total of nine different physico-chemical characteristics
(Appendix A-1). The overall outcomes of the study are presented in Table 5.1.
As observed, the comparison of effluent quality between Tank A and B revealed that the
aerobic treatment had a significant role in leachate treatment. A complete removal of
ammonia-nitrogen (100%) and a reduction in the majority of turbidity (96%) were
observed due to aeration in Tank B. Total alkalinity was also reduced by 40%.
However, nitrogen compounds such as nitrite-nitrogen and nitrate-nitrogen rose due to
aeration from an initial of 0.2 and 8.4 mg/L to 205 and 98 mg/L, respectively.
These results were different to those in Tank A, which was not aerated and, which
became anaerobic over the monitoring period. Ammonia-nitrogen and turbidity removal
reduced to 29% and 46%, respectively. Total alkalinity increased from 1510 mg/L to
2520 mg/L as CaCO3. There was basically no change in nitrite-nitrogen and nitratenitrogen
contents
of
the
leachate
in
Tank
A.
52
Table 5.1 The Change in Characteristics of Landfill Leachate in the two Tanks operated at 21ºC for 20 days. Tank A was the
control while Tank B was aerated.
Parameters
Tank A
Tank B
Initial
Effluent
Removal %
Effluent
Removal%
Colour
pH
Electrical Conductivity
(dS/m)
Turbidity (NTU)
Total alkalinity (as mg/L as
CaCO3)
Dark
Brown
8.2
6.64
Dark
Brown
8.8
6.11
Increase
Decrease
Light
Orange
8.8
5.78
Increase
Decrease
110
1510
60
2520
46
Increase
5
910
96
40
Dissolved Oxygen
Ammonia-Nitrogen
Nitrite-Nitrogen
Nitrate-Nitrogen
0.5
600
0.2
8.4
0.0
430
0.2
7.9
100
29
No Change
Slight
Decrease
4
UMR*
205
98
Increase
100
Increase
Increase
Soluble reactive phosphorus
13
13
No change
11
15
Temperature
21.6
2.5
No change
21.9
Slight
Increase
Note:* Under Measuring Range and all units in mg/L unless otherwise indicated.
53
One important observation was that a steady change in the leachate quality for all these
parameters in both the tanks occurred roughly between the 4th and 6th day, indicating
that even in a small volume tank a minimum of six-day detention time is required to
significantly reduce the concentration of a number of the contaminants.
Further details of the variations for individual parameters as a function of aeration time
are shown in Figures 5.1, 5.2, 5.3, 5.4, 5.5, 5.7, 5.8 and 5.9.
5.1.1
Change in the nitrogen compounds
One of the significant changes in the quality of a leachate is demonstrated by the
reduction in ammonia content. Figure 5.1 demonstrates that the concentration of
ammonia-nitrogen in both the tanks varied significantly over a period of 20 days. For
the first four days, there was an abrupt increase in ammonia-nitrogen from an initial
concentration of 600 mg/L to 1180 mg/L in Tank A and 1385 mg/L in Tank B.
Figure 5.1 The changing pattern of ammonia concentration in Tank A and
Tank B over a period of 20 days.
However, with an increasing aeration time, the ammonia-nitrogen concentration began
to decline rapidly and by the 16th day, a complete removal of ammonia-nitrogen was
achieved (0 mg/L) in Tank B. Azevedo (1993) reported the same findings. In Tank A,
there was no further decrease and the concentration stabilised at 430 mg/L. This
corresponds to an overall ammonia removal of approximately 29% and 100% for Tank
A and Tank B, respectively.
54
The changes could be attributed to the conversion of ammonia to other forms of
nitrogenous compounds, particularly nitrate in the presence of oxygen (Berge et al.,
2006; Cook & Foree, 1974). The two small aerators together had a flow rate of 500
L/hour (equivalent to approximately 8 L/Min); therefore aeration is the principal factor
behind the steady concentration (2-4 mg/L) of DO maintained in Tank B. In Tank A, the
DO was reduced to 0 mg/L within six days. The difference indicates that there was no
available oxygen for the oxidation of ammonia in Tank A.
If nitrification is considered as the main mechanism for ammonia reduction, a
substantial decrease in ammonia would simultaneously result in an increase in nitrite
and nitrate contents of the leachate. This is reflected in Figures 5.2 and 5.3, which
demonstrate the changes in nitrite and nitrate nitrogen concentration in the respective
tanks.
Figure 5.2 The changing pattern of nitrite concentration in Tank A and
Tank B over a period of 20 days.
In Tank B, there was a gradual increase in both nitrite-nitrogen and nitrate-nitrogen,
from an initial concentration of 0.2 mg/L and 8.4 mg/L, respectively to 205 mg/L and
98 mg/L particularly from the 8th day onwards. Interestingly, this time period coincided
with the time at which ammonia-nitrogen (6-8 days) began to decrease, suggesting that
55
the decrease in ammonia and the increase in nitrite and nitrate were associated with the
development of nitrification.
Figure 5.3 The changing pattern of nitrate concentration in Tank A and
Tank B over a period of 20 days.
However, the observation is not the same in Tank A. The slight decrease in ammonianitrogen (29%) did not influence the nitrite and nitrate nitrogen concentrations. Instead,
it was observed that their concentrations remained constant at 0.2 and 7 mg/L,
respectively, as the only oxygen to enter the tank was through entrainment from the
atmosphere above. In this case there was not sufficient oxygen in the leachate for
nitrification to occur. Earlier studies reported similar observations stating that below 2
mg/L, the process of nitrification is inhibited resulting in a nitrite build up. Further, at a
lower DO of 0.5 mg/L, both nitrite accumulation and ammonia consumption decreased
(Bickes & Van Oostron; cited in Shao et al., 2008).
5.1.2
Turbidity
Turbidity is caused by suspended and colloidal matter such as clay, slit, finely divided
organic and inorganic matter, plankton and other microscopic organisms. Therefore, a
decrease in turbidity of the leachate may indicate decrease in these contents. In this
experiment, there was a substantial decrease in turbidity. This was most notable in Tank
B (Figure 5.4).
56
Figure 5.4 The changing pattern of turbidity in Tank A and Tank B over a
period of 20 days.
Firstly, in both tanks the turbidity of the leachate declined rapidly within the first four
days. In Tank A, turbidity decreased from an initial of 110 NTU to 50 NTU while in
Tank B with aeration it was reduced to 25 NTU. While further increase in detention
time did not have an effect on turbidity in Tank A, the turbidity of the leachate
continued to decrease in Tank B. As a result, at the end of the 20th day, approximately
96% of turbidity was removed in Tank B, which is much higher than 46% removal in
Tank A. The results demonstrate that aeration has a significant effect on turbidity and
can improve the leachate quality. It should also be noted that in Tank B, a small portion
of the solids may have been removed by the filtering material in the aerators.
Turbidity removal in both the tanks could also be associated with the reduction in
ammonia earlier, given that a large majority of organics initially present in the leachate
can be very well degraded under both aerobic and anaerobic conditions (Gourdon et al.,
1989). However, the level of organic removal is much higher in an aerobic condition
(Tank B) because a large part of the OM not removed by the anaerobic treatment was
readily biodegradable in aerobic conditions. A comparision of figures 5.1 and 5.4 show
this occurrence.
57
5.1.3
Electrical conductivity (EC)
The EC of the leachate in Tank A decreased from an initial of 6.64 dS/m to 6.11 dS/m
while in Tank B the final EC was lower at 5.78 dS/m (Figure 5.5). It is therefore
possible the ionic solutes might have been removed because they are generally
suspended in the solution along with other organic materials.
Figure 5.5 The changing pattern of EC in Tank A and Tank B over a period
of 20 days.
5.1.3
Colour and odour
A visual assessment of the difference in leachate quality from each tank at the end of the
test period is shown in Figure 5.6. The improved clarity of the aerated sample (right) is
shown. Similarly, sensory records showed an improvement in odour of the leachate after
aeration.
58
Figure 5.6 The difference in colour of the leachate samples from Tank A (left)
and Tank B (right).
5.1.4
pH and total alkalinity
Although, the fluctuations in pH were much less in Tank A as with the case of total
alkalinity, both the parameters fluctuated to a greater extent in Tank B. That is, pH
became increasingly alkaline over the period (8.2-8.8) in both the tanks; however, the
fluctuation was greater in Tank B (Figure 5.7).
Figure 5.7 The changing pattern of pH in Tank A and Tank B over a period
of 20 days.
59
Similarly, the total alkalinity of the leachate changed over the monitoring period as
shown in Figure 5.8. For the first six days, there was a two-fold increase in total
alkalinity values in both the tanks (1510 to 2860 in Tank A and to 2670 mg/L as CaCO3
in Tank B). However beyond the sixth day, the changes in the respective tanks differed.
That is in Tank A, no further reduction was observed whereas in Tank B, a 40%
reduction was observed.
Figure 5.8 The changing pattern of alkalinity in Tank A and Tank B over a
period of 20 days.
5.1.5
Phosphorus
Phosphorus in landfill leachate includes organic and inorganic phosphorus. This nutrient
can cause detrimental eutrophication in aquatic environments (Wang et al., 2010). In
leachates, phosphorus can be measured as SRP, which is a chemically active dissolved
form of phosphorus readily available for biological uptake. A particular concern with
this pollutant is that it is readily available to algae and under certain conditions can
stimulate excess algae growth leading to subsequent depletion of DO. Apart from
leachates, SRP is found in wastewater treatment plants, feedlot runoff, and failing septic
systems.
To examine the effect of aerobic treatment on this pollutant, the changing pattern of Pconcentrations in the leachate in each tank was examined as shown in Figure 5.9.
60
Figure 5.9 The changing pattern of phosphorus in Tank A and Tank B over
a period of 20 days.
The leachate sample initially contained approximately 13 mg/L of phosphorus.
Although a slight increase (11.8-16.6 mg/L in Tank A and from 11 to 17.6 mg/L in
Tank B) was observed between day 6 and day 10 of aeration, the final P-concentration
in Tank A remained the same. In Tank B, the P-concentration was reduced to 11 mg/L,
approximately equivalent to a 15% reduction. However, the final change was minimal,
demonstrating that both aerobic and anaerobic conditions do not have great influence on
the P-concentration in the leachate.
5.2
Column Experiments
The laboratory method associated with these column experiments has previously been
described. Similar to the first experiment, nine different parameters of a typical leachate
from the SWMC were considered for the investigation. The parameters analysed were
colour, turbidity, pH, electrical conductivity, total alkalinity, ammonia-nitrogen, nitritenitrogen, nitrate-nitrogen and SRP(Appendix A-2)
61
Operating conditions
Several factors considered in operating the column experiment were:
the mass, bulk density and particle size distribution of the filter materials,
the contact time and hydraulic conductivity (permeability),
the volume of sample absorbed due to the differences in porosity.
Mass, Bulk Density and Particle Size Distribution
The purpose of determining mass and more importantly the bulk density of the filter
material was to investigate if specific filter materials were more effective in the
treatment of leachate than others. The mass of filter materials required in filling up a
column volume of 377×10-6 m3 was determined as 0.09 kg of GAC, 0.45 kg of BFS and
0.485 kg of sand (Table 5.3). Bulk densities further illustrate their mass per volume
(kg/m3) distribution, which was determined by the method discussed in Section 4.2.3.
The bulk density values for GAC, BFS and sand are determined as 300 kg/m3, 1500
kg/m3 and 1600 kg/m3, respectively (Table 5.2). With 300 kg/m3, some particles of the
GAC floated in the leachate sample.
The particle size distribution graph (Figure 5.10) was obtained from Appendix A-3 and
used to estimate the effective size (D10) of each filter material. The D10 values are also
presented in Table 5.2.
62
Table 5.2 Physical Properties of Three Different Filter Materials: GAC, BFS and sand.
Parameter
Type of media
GAC
BFS
Sand
Black
Grey
Pale Yellow
Colour
9.4
10.9
6.5
pH
526
930
10.6
Electrical Conductivity (µS/cm)
2.8
0.5
0.25-0.125
Effective Particle size (mm) (D10)
68
40
36
Porosity (%)
3
300
1500
1600
Bulk Density (kg/m )
2.1
0.7
0.6
Void Ratio
192
29
32
Hydraulic Conductivity (m/day)
Note:* Effective size D10 was calculated from Figure 5.10.
Table 5.3
Materials
GAC
BFS
Sand
Mass of
materials
(kg)
1st
Run
0.090
0.450
0.485
3
14
12
Operating Parameters for the Column Experiments.
Contact time (mins)
Volume of sample permeated (mL)
nd
rd
2
3
1st
2nd
3rd
Run
Run
Run
Run
Run
15
70
60
30
140
120
500
500
500
350
370
350
260
290
260
63
Figure 5.10
Plot of % of media finer by weight as a function of
particle size (mm).
Contact Time and Hydraulic Conductivity
A total of three runs was considered for the whole column experiment - 1st run, 2nd run
and 3rd run. Basically, each run was a flow of leachate sample at a rate determined by
the filter materials (volume = 377×10-6 m3) under a constant head loading (0.02 m)
condition. However, the 2nd and 3rd runs were obtained by recirculating the leachate
samples for five and ten times, respectively and hence were comprised of numerous
single runs.
The important aspect of each run is that it is basically determined by the filter
material’s saturated hydraulic conductivity (Ksat). The purpose of determining Ksat was
to gain insight into the rate at which the leachate samples flowed through the filter
medium in the columns. The Ksat values for the three different filter materials are also
presented earlier in Table 5.2. GAC exhibits a comparatively higher hydraulic
conductivity (192 m/day) than BFS (29 m/day) and sand (32 m/day). The other
important point is that hydraulic conductivity plays an important role in determining the
contact time period between the material and the influent. That is, a material with lower
hydraulic conductivity induces higher contact time and vice versa.
64
The actual contact time in this experiment was estimated as the total time taken by the
500 mL of sample to permeate through the respective filter material columns under a
constant head loading condition. Therefore, the contact time determined from 1st run
was multiplied by five and ten times, respectively to obtain a total of three different
contact times in an increasing order. GAC provided the least contact time period in all
the runs (1st, 2nd and 3rd) with three, 15 and 30 mins, respectively, whereas BFS
materials ensured a significantly higher contact time of 14, 70 and 140 mins,
respectively. Finally, sand also had similar contact time period to that of BFS with the
first contact time period calculated as 12 mins and the remaining as 60 mins and 120
mins, respectively (refer Table 5.3 before).
Volume of Sample Absorbed
Another important operating parameter is the ability of filter materials to absorb a
certain volume of a sample. This volume was calculated by subtracting the effluent
sample volume from the influent.
In the first run, GAC and sand columns
demonstrated that the filter materials roughly absorbed 100 mL of sample whereas BFS
could absorb only 80 mL. In the 2nd run, the absorption reduced to 40 mL for GAC and
sand and 30 mL for BFS. The difference in the ability of the filter material to absorb
sample is apparently due to differential % porosities of the filter materials (GAC (68%),
BFS (40%), and sand (36%)) (Table 5.2 before). Unlike the 1st and 2nd runs, the 3rd run
did not absorb the samples, indicating that all the filter materials were already saturated
and thus no absorption occurred. In addition, about 50 mL of the effluent sample from
each run was collected for analysis.
The consequence of the absorption phenomenon is that it varied the volume of
samples permeated in the 2nd and 3rd runs. For example, although 500 mL of sample
were permeated in all the filter columns in the 1st run, the volumes of samples
permeated in the 2nd and 3rd runs differed from one column to the other. For example,
in the 2nd run, only 350 mL of sample was permeated in the GAC and sand columns
and 370 mL in the BFS columns. Similarly, there was a substantial decrease in the
volume of sample permeated in the 3rd run with approximately 260 mL for GAC and
sand and 290 mL for BFS.
65
Column Experiment Results
The outcomes of the leachate treatment from the filter materials are presented in Table
5.4. The table presents both range and mean values of the nine parameters examined to
determine the overall impact of the filter materials on leachate treatment.
5.2.1
Solube reactive phosphorus removal
The percentage (%) removal results in Table 5.4 show the filter materials have most
significant impact on leachate sample with respect to P-removal. The final effluent Pconcentrations demonstrated a significant decrease with an estimated P-removal of
approximately 68%, 92% and 40% for GAC, BFS and sand, respectively. Earlier,
Johansson (1999) investigated BFS in a laboratory scale column experiment. The
results indicated that this material had a high P-sorption potential to remove more
than 90% of phosphorus from a P-solution at concentration 10 mg/L. Figure 5.11
further illustrates the pattern of decrease in P-concentrations as a function of three
different contact times.
Figure 5.11
The changing level of phosphorus in the leachate samples
following different runs throughout the experiment.
66
Parameter
Colour
pH
Electrical
Conductivity
(dS/m)
Turbidity (NTU)
Alkalinity (as mg/L
as CaCO3)
Ammonia-Nitrogen
Nitrite-Nitrogen
Nitrate-Nitrogen
Soluble Reactive
Phosphorus
Table 5.4 Mean Results Obtained from Column Experiments using three Different Filter Materials.
GAC
BFS
Sand
%
Effluent
Effluent
Effluent
Effluent
%
Effluent
Effluent
Removal (range)
Influent (range)
(mean)
(mean)
Removal (range)
(mean)
%
Removal
Dark
8.4
Brown
NA
8.7-8.7
Black
8.7
NA
NA
NA
9.4-9.6
Light
9.5
Brown
NA
NA
NA
8.7-8.8
Dark
8.8
Brown
NA
NA
10.48
10.1210.18
10.15
NA
9.83-9.9
9.865
NA
9.97-10.06
10.015
NA
59
202-209
205
NA
45-46
46
22
42-43
43
25
4700
4420-4780
4600
2
1450-1780
1615
66
4280-4440
4360
7
975
0.2
10
1060-1080
0.3-0.3
4.7-8.3
1070
0.3
7
NA
NA
30
1010-1265
0.1-0.1
20-22
1138
0.1
21
NA
NA
NA
1060-1260
0.2-0.2
12.8-12.8
1160
0.2
13
NA
NA
NA
25
7-9
8
68
2-2
2
92
14-16
15
40
Note: All the units in mg/L unless otherwise indicated. NA indicates Not Applicable for % removal estimation.
67
For all the materials, the maximum reduction in phosphorus (app. 64% and 76% and
28%, respectively) occurred during the 1st run (initial phosphorus~25 mg/L)
corresponding to a contact time period of 3, 12 and 14 mins, respectively. Further
reductions in P-concentration were observed in the subsequent runs essentially due to
the fivefold increase in contact time estimated as 15, 60 and 70 mins, respectively.
Beyond the 2nd run, there was no major decrease in P-concentration possibly indicating
the saturation of the materials.
This trend signified that a longer contact time resulted in a greater P-removal. However,
beyond the 2nd run, the 3rd run did not lead to greater P-removal, which is an indication
that an additional increase in contact time would not result in adsorption. The results
illustrate that although the same volume (377×10-6 m3) of filter material and the same
volume of leachate (500 mL) were considered, each filter material has a different
response towards phosphorus, thus resulting in various P-removal efficiencies at
different contact times.
P-sorption capacity
The objective of this part of the analysis was to evaluate sorption and removal of
phosphorus from model solutions by various filter materials in batch experiments. This
was carried out by developing P-sorption isotherm (curves) and could possibly verify
the assumption that P-sorption by filter materials is the main mechanism behind Premoval. For this purpose, batch tests results for P-sorption were used. The batch test
was conducted by Lanfax Laboratories according to the Five Point P-sorption Curve
method used for determining P-sorption. The details of the method are as outlined in
Patterson & Jones (2001):
Mass of sample:
4.0g
Phosphorus concentration:
25, 50, 75, 100 and 150 mg P/L
Sample/Solution ratio:
1:10
Equilibrating solution:
0.01M CaCl2
Tumbling period:
17h at 25oC.
The resulting P-sorption isotherms or curves are shown in Figure 5.12.
68
Figure 5.12 P-sorption isotherms for three different filter materials developed
using five different P-solutions in the batch scale experiment (Lanfax
Laboratories).
According to the P-sorption curves, P-sorption increases proportionally with the
increase in initial phosphorus concentration. Nevertheless, P-sorption capacity from 25
mg/L phosphorus solution was selected to resemble the P-concentration of the leachate
sample used in this experiment. When fed with 25 mg/L of phosphorus, the three filter
materials-GAC, BFS and sand- had a capacity to adsorb 249, 257 and 27 mg of
phosphorus for every one kilogram of the material, respectively.
5.2.2
Turbidity and colour
Apart from a significant P-removal, there was a considerable decrease in the turbidity of
the effluent leachates treated with BFS (22%) and sand (25%) as shown in Table 5.4.
This might be an indication that the filter materials removed a considerable amount of
suspended and colloidal matter possibly containing clay, slit, and finely divided organic
and inorganic matters during the filtration. Aziz et al. (2007) suggested that high levels
of OM (measured as COD) are associated with turbidity and suspended solids. However
the drastic increase in turbidity from 59 NTU to 205 NTU in the effluents from the
GAC
column
was
not
anticipated
(Figure
5.13).
69
Figure 5.13 The changing level of turbidity in the leachate samples following
different runs throughout the experiment.
In addition to turbidity reduction, there was an improvement in colour of the leachate
before and after the treatment, particularly effluent from the BFS and sand columns.
Hence, the overall improvement in the turbidity of the leachate is confirmed by the
changes in the colour of the leachate effluents (Figure 5.14 b, c) based on a visual
assessment.
GAC (Control)
GAC (Experiment)
a)
70
BFS (Control)
BFS (Experiment)
b)
Sand (Control)
Sand (Experiment)
c)
Figure 5.14 The difference in the colour of the leachate samples for three
different runs along with their respective controls. a) GAC, b) BFS and c)
sand.
5.2.3
Electrical conductivity
Further, adsorption of other forms of dissolved solids (particularly inorganic solutes)
present in the leachate is evident from the fact that all the filter materials were able to
reduce the EC of the leachate sample. This is shown in Figure 5.15. The EC values
decreased from an initial 10.480 dS/m to a final of 10.15 dS/m, 9.865 dS/m and 10.015
dS/m for GAC, BFS and sand, respectively.
Although, all the materials resulted in a reduction of EC, effluent samples from BFS
demonstrated extreme EC variations at different runs. For example, there was a sudden
initial increase in EC from 10.48 to a very high 10.795 dS/m in the 1st run. The values,
however, decreased eventually to 10.155 dS/m and finally to 9.865 dS/m in the 2nd and
3rd runs, respectively. The results reflect that BFS might have initially released certain
dissolved solids into the effluent and that with repeated circulations it showed an ability
to re-adsorb them. However, this did not happen to the effluent from GAC and sand.
71
Rather a substantial decrease was observed with each increasing run as shown in Figure
5.15.
Figure 5.15
5.2.4
The changing level of EC in the leachate samples
following different runs throughout the experiment.
Total alkalinity and pH
The total alkalinity removal ability for different filter materials is shown in Figure 5.16.
Over all, a significant reduction in total alkalinity (66%) was observed in the effluent
sample treated with BFS. The reduction was, however, less in the case of GAC and sand
with 2% and 7%, respectively. In the BFS columns, a significant reduction in P-removal
occurred at the 2nd run. Hence, the final total alkalinity value of 1615 mg/L as CaCO3
corresponded to a 66% decrease. The only possible explanation could be that BFS
materials might have released acids that can react with bicarbonates present in the
leachate and possibly converted them reducing the total alkalinity. Therefore, low total
alkalinity values indicated that the leachate no longer had ability to buffer against any
pH changes in the environment, thus becoming more prone to fluctuations in its quality,
especially with respect to its acidity and alkalinity.
72
Figure 5.16
The changing level of alkalinity in the leachate samples
following different runs throughout the experiment.
On the other hand, the changing pattern of the effluents treated with GAC and sand are
almost consistent and consequently the final values are only slightly lower than the
initial - 4600 and 4360 mg/L of CaCO3, respectively.
Figure 5.17 demonstrated that with each run, the effluent leachate samples became more
alkaline. This change was common to all the filter materials although it was
significantly high in the case of the effluent treated with BFS. pH values for BFS
effluent increased from 8.4 to 9.5 and to a little over 8.7 and 8.8 for GAC and sand
respectively.
73
Figure 5.17 The change in pH values of the leachate samples following
different runs throughout the experiment.
5.2.5
Ammonia-Nitrogen, Nitrite-Nitrogen and Nitrate-Nitrogen
Figure 5.18 demonstrates that ammonia concentrations in the leachate in all the three
columns reached a maximum during the experimental period. For example, ammonianitrogen concentration increased from an initial 975 mg/L to 1178 mg/L in the 1 st run
for GAC effluent, to 1108 mg/L for BFS and to 1243 mg/L for sand both in the 2nd run.
Figure 5.18
The changing level of ammonia in the leachate samples
following different runs throughout the experiment.
74
In terms of nitrite-nitrogen, effluent values for nitrite from all three columns were
similar and fluctuated only by ±0.1 as compared to the initial concentration (0.2 mg/L).
Thus, it can be assumed that there was no change in nitrite; instead, it remained
remarkably constant throughout the experiment.
The effluent samples treated with BFS showed a significant two fold increase in nitratenitrogen from an initial of 10 mg/L to 27 mg/L in the 2nd run and then to 21 mg/L in the
3rd run (Figure 5.19). This was not the case with the effluents from GAC and sand
columns because the nitrate-nitrogen concentration only fluctuated by negligible
amounts and as such there were no significant change throughout the entire
experimental period. Therefore, it may be concluded that GAC and sand neither
contributed to an increase nor to its removal from the effluent.
Figure 5.19 The changing level of nitrate in the leachate samples following
different runs throughout the experiment.
One relation from the Figures 5.18 and 5.19 suggests that when nitrate-nitrogen
concentrations
for
BFS
effluents
reached
maximum,
the
ammonia-nitrogen
concentration was comparatively low. Conversely, GAC and sand with no or little
change in nitrate concentration showed more fluctuations in ammonia-nitrogen.
75
Chapter 6 Discussion
The chapter is divided into two sections. Both sections discuss the overall outcomes of
the two bench scale laboratory experiments and their relevance and suitability as a low
cost passive leachate treatment option in Bhutan. The changes in the quality of the
leachate from both experiments are discussed in detail. The application and limitations
of this treatment study including concerns with respect to their application as an onsite
leachate treatment plant in Bhutan are outlined.
6.1
Aerobic Biological Treatment
This experiment followed earlier bench and full scale treat-ability studies, which
showed that raw leachate could be effectively treated by aerobic biological processes
and that complete nitrification of ammonia could be achieved under suitable conditions
(Boyle & Ham, 1974; Cook & Foree, 1974; Knox, 1983 & 1985; Robinson &
Grantham, 1988; Robinson & Maris, 1983) The bench scale treatment process was
designed to further investigate the treatment of leachate in onsite landfill facilities and
to particularly resemble the facultative treatment lagoons.
The experiment was carried out in two glass-sided tanks over a period of 20 days and
involved 25 L of medium strength leachate obtained from the SWMC landfill. A mean
temperature of approximately 21oC was maintained over the period of test to simulate
conditions at onsite leachate treatment plants. There was no pH manipulation or nutrient
addition. In Tank B, the two aerators together have a capacity to aerate 500 L of
leachate per hour that enables a minimum oxygen concentration of 2 mg/L – the
required limit for most aerobic treatment systems. Tank A was not aerated and thus
served as a control for the experiment.
Further the experiment demonstrated that even in a small volume tank a minimum of
six-day detention time is required to significantly reduce the concentration of a number
of the contaminants. This could be the minimum period of time required to facilitate
biodegradation in the leachate in both tanks. The four to six day time period has a close
similarity to those reported by Yahmed et al. (2009) and Robinson & Maris (1983).
76
Although certain operating conditions differed from this experiment, the aerobic
treatments of leachate were all based on smaller scale pilot experiments in the
laboratory. Yahmed et al. (2009) demonstrated that a significant OM reduction
(indicated by TOC) was achieved from an undiluted young leachate in bioreactors at a
detention time of 108 hours (~ four and one half days). Similarly, a well clarified
effluent and a high removal of BOD (>98%) and COD (>92%) was obtained from a
medium strength leachate at a slightly higher detention time of 10 days (Robinson &
Maris, 1983).
6.1.1
Ammonia-nitrogen removal
Initially, during the first four days, an abrupt increase in ammonia was observed in both
the tanks. Azevedo (1993; cited in Shiskowski & Mavinic, 1997) described this
phenomenon as a sort of ammonia “spike” believed to be occurring in both anoxic and
aerobic reactors when ammonia load is initially high. This could be attributed to the
rapid biodegradation of organic nitrogenous compounds in the leachate. Although, the
biological degradation is lesser in Tank A due to anaerobic conditions, ammonianitrogen accumulation still occurred due to its stability under anaerobic conditions. In
addition, anaerobic conditions do not have ammonia elimination process (Vigneron et
al., 2007; cited in Shou-liang et al., 2008).
However, a 100% ammonia-nitrogen removal in Tank B and 29% in Tank A at the end
of the experimental period suggested that ammonia removal was achievable, especially
in Tank B. This could be attributed to complete nitrification of ammonia brought about
by an intensive aeration at a higher pH of 8.2-8.8 optimum for nitrification. In Tank A,
the slight reduction in ammonia content could be due to the volatisation of ammonia to
the atmosphere (Robinson & Maris, 1983) or through the assimilation by anaerobic
bacteria, which utilises ammonia for cellular growth over the period of the test
(Kettunen et al., 1996; cited in Shou-liang et al., 2008).
The mechanism for the conversion of ammonia to nitrate takes place by the presence of
bacteria known as “nitrifiers”, which are strict “aerobes,” meaning they must have free
dissolved oxygen to perform their work. Therefore, nitrification occurs only under
aerobic conditions at DO concentrations of 1.0 mg/L or more. Nitrosomonas first
77
convert ammonia and ammonium to nitrite and then Nitrobacter finish the conversion of
nitrite to nitrate. The reactions are generally coupled and proceed rapidly to the nitrate
form; therefore, nitrite concentrations at any given time are usually low. At DO
concentrations less than 0.5 mg/L, the growth rate is minimal; hence nitrification is
inhibited in anaerobic conditions.
The 100% ammonia-nitrogen removal in Tank B could also mean the main constituent
of concern in the raw leachate, which can cause a range of serious problems such as
eutrophication, increase in BOD and stimulation of algal growth in water systems, was
successfully treated (Berge et al., 2006; Boyle & Ham, 1974; Cook & Foree, 1974;
Knox, 1985; Pivato & Gaspari, 2006; Wang et al., 2006; Wang et al., 2008). The treated
effluent is free of ammonia (0 mg/L) and therefore suitable to be discharged into the
environment according to this parameter. For instance, in Australia, the AWQGFMW
(1992) allows 0.5 mg/L of ammonia in the fresh and sea water systems. In case the
effluent does not meet this discharge limit, it will still be suitable for discharging into
the municipal waste water treatment or sewerage systems.
An important observation in Tank B is that the pH and total alkalinity of the leachate
sample fluctuated more significantly than in Tank A. The changes might be due to the
process of nitrification that can produce acid. Acid formation might have caused
fluctuations in pH, in most cases lowering its value by a few units. However, it did not
affect the nitrification process because pH values (8.2-8.8) still fell in the optimum
range for the nitrifiers given that most treatment plants are able to effectively nitrify
above a pH of 6.5. The assumption compared favourably with Tank A where fewer
changes in pH occurred due to the absence of nitrification.
In addition, the nitrification reaction (that is, the conversion of ammonia to nitrate) is
reported to consume 7.1 mg/L of total alkalinity as CaCO3 for each mg/L of ammonia
nitrogen oxidized. This occurrence might have resulted in a reduction of total alkalinity
in Tank B. Further observations are that reduction in ammonia-nitrogen, total alkalinity
and increase in nitrite-nitrogen and nitrate-nitrogen all occurred simultaneously between
the 6th and 10th days of aeration. So, it may be correct to assume that about 40% of total
alkalinity is consumed for 100% ammonia-nitrogen reduction. However, the reduction
78
did not affect the buffering ability of the leachate because 900 mg/L as CaCO3 is above
the minimum total alkalinity requirement (50-100 mg/L as CaCO3) for a treatment tank.
Conversely, in Tank A, where nitrification was absent, both ammonia-nitrogen and total
alkalinity reductions were insignificant or minimal. Nitrite and nitrate nitrogen
concentrations also remained consistent.
6.1.2
Turbidity and electrical conductivity
Apart from the nitrogenous compounds, leachates contain significant amounts of
organic and inorganic solids indicated variously by colour, turbidity and EC. The results
of the change in these parameters in the presence and absence of aeration show a
thorough understanding of how aerobic treatment can bring about significant
improvement in the quality of the final effluent.
The first is a decrease in turbidity that partly reflects the OM content of the leachate.
Therefore, although the BOD, COD and TOC indicators were not monitored in this
study to model biodegradation of leachate, it is noticeable from the turbidity that
substantial reduction in OM occurred from the 4th day onwards in both the tanks. By
comparison, the OM was almost completely removed (96% reduction in turbidity) in
Tank B and to a much lesser extent in Tank A (46%). The results are consistent with the
findings of Boyle & Ham (1974), Cook & Foree (1974) and Robinson & Marris (1983),
where a substantial decrease in the OM was observed at the same retention time for
leachates containing high amounts of OM. That is, when leachates with high BOD/COD
values (1550/2700 mg/L, 3000/5000 mg/L and 7100/15800 mg/L) were aerobically
treated at a bench scale, it resulted in more than 80% BOD and COD removal at a five
day retention time.
The second important parameter of the leachate is the EC. It reflects the total
concentration of ionic solutes present in a solution (Jun et al., 2009). The slight decrease
in EC in both the tanks can be associated with the removal of ionic solutes. However, as
compared to OM, ionic solute removal is less. This is because the leachate sample is
likely to contain relatively high amounts of elements such as Na, K, Ca, Mg and so on.
Since none of this can be lost from the leachate simply by aeration, there was only a
slight influence on EC.
79
6.1.3
Colour and odour
The combined effect of reduction in turbidity and EC improved the clarity (colour) of
the aerated sample, particularly in Tank B where the colour of the leachate changed
from dark brown to clear light orange after the aerobic treatment (Figure 5.6 ). In
addition, aeration had a postivive impact on removing the foul odour of the leachate
possibly due to the degradation in OM in Tank B. Conversely, in Tank A, foul odour
was developed over the period of time.
6.1.4
Phosphorus
The aerobic treatment experiment exhibited that aeration didnot have influence on the
phosphorus concentration in the leachate. With the exception of Cook & Foree (1974),
who asserted that aeration resulted in a removal of Total-phosphorus and soluble
reactive phosphorus from leachates in 10 days, it is possible there is no literature of the
positive effect of aerobic treatment for phosphorus. Instead, P-removal was mostly
based on three factors such as precipitation; adsorption or the assimilation by certain
micro-organisms present in both anaerobic and aerobic conditions. The same
assumption might possibly explain the difference in fluctuations of P-concentrations in
the respective tanks.
Although all three factors might have contributed to the fluctuations in P-concentration
in both the tanks, the slight reduction in phosphorus in Tank B must be due to the
greater possibility of removing phosphorus by micro-organism assimilation in aerobic
conditions (Sedlak, 1991). According to this author, all biological phosphorus removal
systems utilise the following two-step process description:
“Certain micro-organisms when subjected to anaerobic condition assimilate
and store fermentation products produced by other facultative bacteria. These
micro-organisms
derive
energy
for
this
assimilation
from
stored
polyphosphates, which are hydrolysed to release energy. The resulting
phosphorus is released to the mixed liquor.
These same micro-organisms, when subsequently exposed to aerobic
conditions consume both phosphorus and Oxygen to metabolize the
previously stored substrate for energy production and cell synthesis” (Sedlak,
1991, p.167).
80
6.1.5
Temperature
Although aerobic treatments are sensitive to temperature and changes in temperature are
likely to bring certain changes in the treatment efficiency, at normal temperatures, it
may not pose much effect on treatment efficiency. Earlier, Robinson & Maris (1983)
studied the effect of cold temperature (10oC) on aerobic treatment and at 20-25oC by
Cook & Foree (1974). The experiments confirmed that an efficient removal of OM
occurred at 10 days in both conditions. This indicated that within reasonable
temperature changes, the effect should not be as significant. Hence, unless the
temperature is completely extreme, treatment of leachate is still feasible.
In the present study, temperature remained almost constant throughout the experimental
period with values ranging between 21.3-21.9oC in Tank A and 21.4-22oC in Tank B.
Both resembled onsite temperatures and are basically suitable for OM degradation and
nitrification. In fact the slightly higher temperature in Tank B might be due to the rapid
activity of microorganisms. However, the heat produced was possibly offset by the
cooling effect of aeration.
With regard to the effect of temperature on nitrification, nitrification ceases only at
temperatures higher than or equivalent to 40oC and below 10oC and thus, between this
range, temperature does not have much effect on the treatment process. Berge et al.
(2007) reported that ammonia removal by the process of nitrification in a bioreactor
landfill leachate can readily occur over a range of temperatures (22oC, 35oC and 45oC.).
However, it was observed that ammonia removal was most efficient at 35oC and slower
at 22oC and 45oC.
6.1.6
Retention time
The most important factor to consider in an aerobic treatment is the retention time.
Graphical illustrations of various pollutant parameters as a function of time in Section
5.1 showed the influence of retention time on the treatment efficiency in both tanks.
Firstly, a careful assessment of the change in leachate characteristics as a function of
time reveals that both aerobic and anaerobic treatment requires a minimum six day
retention time to establish a steady state or stabilize. The first four to six days did not
81
bring about any steady change in the leachate quality. Instead, this period was
characterized by abrupt changes that were not consistent with the remaining
experimental period. As an illustration, the rapid increase in ammonia-nitrogen, total
alkalinity and P and the decline in turbidity in both the tanks, occurred between the 4th
and 6th day (see Figures 5.1, 5.8, 5.9 and 5.4, respectively). Hence, the first four to six
days is considered the minimum retention time required in facilitating biodegradation of
leachate, irrespective of whether the condition is aerobic or not.
The fact that the above changes were common in both tanks in the initial period implies
that leachates also have a natural tendency to undergo stabilization in aerobic as well as
anaerobic conditions. Rapid biodegradation of OM might have led to the rapid increase
in ammonia and decrease in turbidity during that period. In an anaerobic condition
(Tank A), this might have been possible due to the presence of dissolved oxygen
initially present in the leachate. Consumption of oxygen was evident from the fact that
dissolved oxygen was completely reduced to 0 mg/L within four days in Tank A.
Secondly, it is evident that longer retention time beyond six days brought significant
changes in the quality of leachate. Around the 16th day, a final steady change was
observed with the exception of P. At this time, in Tank B, the leachate quality improved
significantly with complete removal of ammonia and turbidity. Total alkalinity and EC
were reduced by considerable units, while nitrite and nitrate nitrogen attained maximum
concentrations and remained remarkably constant thereafter. In Tank A, a considerable
decrease in the concentrations of the pollutants occurred. Beyond the 16th day, there was
no significant change in the leachate quality in both the tanks, hence the 16 days is
considered an approximate retention time for both aerobic and anaerobic treatment in
this study.
A 16 day retention time compared favourably with other bench scale studies that
considered a 10 day retention time more effective than five day at normal operating
temperatures of 20-25oC. In addition, BOD and COD removal efficiency of 80-90% at
a 10 day retention time declined to about 40-50% when the retention time is lowered by
five days (Boyle & Ham, 1974; Cook & Foree, 1974; Robinson & Maris, 1983).
However, a longer retention time would be more appropriate and desirable when large
82
scale leachate treatment is involved. In onsite treatment systems, a higher retention time
of 32 days (Frascari et al., 2004) and 56 days (Mehmood et al., 2008) were required to
achieve similar removal efficiencies. Therefore, the retention time is an important
controlling factor for both aerobic and anaerobic treatment of leachates and should
ideally be 10 days and not less than six days.
6.1.7
Application
The suitability of aerobic treatment of leachate was suggested by Boyle & Ham (1974)
and Cook & Foree (1974). The concept has, however, mostly been used as a pretreatment step prior to a biological treatment in a municipal facility (Frascari et al.,
2004). The results obtained from this study along with Matthews et al. (2009) suggest
that the method is highly suitable as the principal leachate treatment option given that it
is highly resilient and little affected by the fluctuations in temperature, flow rate and
loading associated with effluents. The advantage further lies with the ability of the
system to remove a wide range of critical pollutants at a minimum possible investment
and operating cost, making it feasible in places where finance and technological
constraints are a consideration. The only operating cost is associated with electricity
because sufficient power input is required to diffuse enough oxygen into the leachate.
In terms of design and facility, this experiment resembles facultative lagoons that are
simple enough to provide a sufficient space for leachate retention, aeration and settling
solids. Several lagoons or basins resembling this study (facultative type) were
developed in Italy, Britain, Ireland and the UK (Frascari et al., 2004). Generally, they
are located on a clay declivity and also surrounded by a 0-2m thick layer of more
permeable clay filling. They have several common features: high density poly-ethylene
lining and few aerators; automatic timers, sensors, switches; and a provision for leachate
collection. While the capacity of the lagoon varies from place to place, for a landfill
receiving small waste input (50 tonnes per day) at a place receiving an average annual
rainfall of 1200 mm, a capacity of 1000 m3 is suitable. This capacity allows a required
retention time of 10 days at a maximum design flow rate of about 100 m3 per day
(Robinson & Grantham, 1988).
83
Given Bhutan’s SWM background, that is a similar waste input rate (64.5 tonnes per
day in urban Bhutan), less per capita waste generation (0.5 kg/day (rural/urban), the
organic nature of the waste (60% organic fraction) and similar climatic conditions, an
aerobic treatment system is highly recommendable and feasible for Bhutan.
6.1.8
Drawbacks of the experiment
Although both bench scale and onsite aerobic treatment studies came to a general
consensus that the system is little affected by the fluctuations in temperature, most
results have been reported from experiments performed at room temperature, generally
stated to be about 20-30oC. Treatment systems carried out at 3oC, 10oC and 15oC also
achieved significant outcomes, but what is not known is the effect of extreme low
temperatures (below 0oC) on treatment efficiency. Most urban places in Bhutan have an
extreme winter season where the mean temperatures fall to ¯8oC in winter. Robinson &
Grantham (1988) reported a case in mid-Wales, UK, where the lagoons froze forming
ice several inches thick at 2-3oC. Though, it was stated that there was no noticeable
effect on the effluent quality when the situation occurred, such cases might result in
other unforeseen operation problems.
Likewise, simple aeration operation style is possibly associated with sludge bulking and
scum formation. Boyle & Ham (1974) reported that a treatment unit with one hour
settling gave rise to sludge problems resulting in high effluent solid concentration. This
requires a separate collection and disposal of settled sludge in the landfill. In the present
experiment, sludge build up was negligible and scum formation was not evident by
visual assessment of the two units, apparently because bench scale experiments
involved lesser amounts of leachate.
Since the treatment units in the above study were not “fill and draw” systems, there is a
lack of information on the influent and effluent flow rates. This might give rise to
overflow of leachate, especially in the monsoon season when the rate of leachate
generation in the landfill rises. In practice, leachates have to be treated in continuous
flow rather than the batch mode system (Matthews et al., 2009).
84
6.2
Column Experiments
In total nine different characteristics of a typical leachate were discussed to determine
the suitability of filter materials in a leachate treatment. The results varied from each
filter material to the other; nevertheless a common observation is that all the materials
were highly recommendable for removal of SRP from leachate. Apart from that, the
materials have a considerable effect in reducing suspended and dissolved solids as
indicated by the slight decrease in turbidity and EC.
6.2.1
P-removal
Results of P-removal efficiency obtained from the column experiment were then
compared with the batch experiment results. The P-removal efficiencies in the columns
were highest for BFS (92%) followed by GAC (68%) and sand (40%). As anticipated,
the batch experiments also confirmed that GAC and BFS could remove phosphorus
more significantly than sand. That is BFS has the highest P-sorption capacity with an
ability to adsorb 257 mg of phosphorus in every one kg of filter material. Equally
efficient is the GAC with a P-sorption capacity estimated at 249 mg of phosphorus in
every one kg. Sand exhibited a capacity to remove 27 mg of phosphorus in every one
kg and thus will result in a lower P-removal efficiency.
The fact that both the experiments demonstrated BFS can remove phosphorus much
better than the GAC and sand, was supported by Johansson (1999), who reported that
BFS can remove more than 90% phosphorus from a P-solution graded 10 mg/L. With
respect to GAC, the material is known for its effectiveness in treating contaminated
wastewater and leachate, particularly the removal of phosphorus. This could be
attributed to the fact that GAC has a high surface area to volume/weight ratio than
most filter materials. Figure 5.10 exhibits that GAC has an approximate effective
particle size (D10) equivalent to 2.8 mm, which is much higher than BFS (0.5 mm) and
sand (0.25 mm).
In addition, the batch experiment results confirm P-sorption as one main factor behind
P-removal from the leachate. However, P-removal efficiencies obtained from column
85
filtration will be less than those obtained in batch experiments because some P-leaching
may always occur (Kuryer et al, 1995; cited in Kim, 1999).
Similar to P-removal, the variation in overall effect of the filter materials on the leachate
quality can be attributed to numerous controlling factors. Firstly, treatment of leachate
by the column filtration used three materials with large differences in their physical and
chemical properties, the details of which are discussed in the following sections.
Secondly, leachate is a complex mixture of organic and inorganic substances capable of
undergoing various physical and chemical reactions.
6.2.2
The effect of chemical composition
The need to understand a filter material by its chemical composition is fundamental
because each filter material is different on its own. BFS is essentially composed of
silicates and alumino-silicates of Ca. Sand is roughly similar to BFS and contains SiO2
as the main component followed by oxides of Al, K and Fe (Kietlinska & Renman,
2005). On the other hand, GAC is an entirely different material containing carbon as the
main component. Thus, when a leachate is permeated through them, complex physical
and bio-chemical processes are anticipated eventually resulting in a change in effluent
quality.
Since removal of phosphorus predominantly takes place by several mechanisms: rapid
removal or adsorption; chemical precipitation; and ion exchange (Aulenbach &
Meisheng, 1988; Johansson, 1999; Oguz, 2004), the chemical composition of the filter
material will have a significant impact on P-removal from the leachate. BFS mainly
contains metals like Ca, Al, K and Fe in the form of their oxides. Their presence favours
physical sorption of SRP or ortho-phosphate ions on the surface of the material
particles. In addition, the respective metal ions can help remove phosphorus by
precipitation, which occurs by transforming the soluble phosphate in the influent to
relatively insoluble metal phosphates (Ca, Al, K and Fe), particularly when the effluent
is alkaline.
86
Similarly, partial dissolution and subsequent hydrolysis of Ca and alumina silicates
from the slag allows adsorption of other metal ions (Copper (Cu), Ni and Zn) on its
surface either by exchange or replacement when releasing Ca ions into the solution.
(Dimitrova & Mehanjiev, 1999). Most researchers have in fact proposed precipitation of
any of Ca- phosphates as one mechanism for P-removal.
6.2.3
Contact time & hydraulic conductivity
Another explanation for observed differences between the P-removal efficiencies
among different filter materials is the variation in contact times, which in turn is
induced by varying hydraulic conductivity values. It is observed that the materials with
lower hydraulic conductivity ensure greater contact time and thus promote P-sorption
mechanisms to a greater extent than with shorter contact times and higher hydraulic
conductivity. As a result, hydraulic conductivity is suggested as the most conceivable
explanation for the observed differences in P-removal between GAC, BFS and sand.
The BFS column, which achieved the highest P-removal efficiency (92%), had the
lowest hydraulic conductivity (29 m/day). That is, the lower hydraulic conductivity
allows for more opportunities for leachate to come into contact with a significant
volume of filter materials. Inversely, although GAC has a better ability to adsorb
phosphorus like BFS (Figure 5.11), it resulted in a comparatively lower P-removal
efficiency (68%), mainly due to the shorter contact times induced by its high hydraulic
conductivity (192 m/day). However, in the case of sand, low hydraulic conductivity (32
m/day) and high contact time did not improve its ability to remove phosphorus, possibly
owing to its chemical composition.
Hydraulic conductivity (m/day) = GAC (192) > sand (32) > BFS (29)
Approximate contact time (minutes) for 500 mL of influent in column
configuration used = BFS (14) > sand (12) > GAC (3)
Similar conclusions were also observed by Nilsson (1990; cited in Johansson, 1999).
Therefore the effectiveness of P-removal is higher for filter materials with lower
hydraulic conductivity and better contact times.
87
6.2.4
Longevity
According to Johansson (1999), longevity of the filter material is an important practical
factor to take into consideration prior to its use in onsite treatment systems, which use
filter materials such as sand or soil. Longevity is limited by factors such as saturation,
hence, once the material attains a saturation point; it is no longer effective for further
treatment. Considering all other factors constant, longevity may be defined as a point of
time, where the material no longer significantly influences the quality of the effluent. In
this experiment, the saturation point of all the filter materials is achieved around the 2nd
run, corresponding to a contact time of 15 minutes for GAC, and approximately an hour
for both BFS and sand. This is assumed from that fact that after the 2nd run, the effluent
leachate quality remained steady for most pollutant parameters, especially phosphorus,
turbidity, EC and total alkalinity (See Figures 5.11, 5.13, 5.15 and 5.16, respectively).
This implies that different filter materials will have different durabilities in the treatment
industry.
6.2.5
Particle size distribution, porosity percentage (%) and bulk density
The particle size distribution of a filter material is yet another critical factor in defining
the varying performance of a filter material because adsorption is one main mechanism
by which the filter materials adsorb or collect molecules of dissolved compounds on the
surface of an adsorbent solid. Adsorption is essentially a surface phenomenon (Brady &
Weil, 1996) occurring as a result of weak physical interaction between the surface of
adsorbent and the metallic salts of phosphate (Oguz, 2004).
Considering the above assumptions, filter materials with smaller particle size will have
greater advantage over the ones that are larger. This is because smaller particle size will
have larger surface area available for adsorption in a given volume and ensures greater
interaction with the contaminants present in leachate. While GAC has larger particle
size ( effective size (D10) of approximately 2.8 mm) and smaller surface area available
for greater interactions, each particle has a honeycomb fabric with very high surface
area-hence its effectiveness in capturing contaminants. This leaves the material with
comparatively higher ability to remove phosphorus even at a very high hydraulic
conductivity. Better results may be achievable by mixing peat with GAC in the filter
matrix to reduce the hydraulic conductivity (Johansson, 1999). BFS has similar
88
properties with a D10 value of 0.5 mm. Although sand has a smaller particle size (D10
<0.25 mm), it do not have internal surfaces.Thus it has low effectiveness in capturing
contaminants.
Again, particle size may be related to two other factors, porosity and bulk density. That
is the bigger the particle size, the higher the proportion of pore space to solid per unit
volume (Brady & Weil, 1996) and the lower the bulk density and vice versa. The
inverse relation between estimated porosity percent (%) and bulk density is shown
below:
Porosity % = GAC (68%)>BFS (40%)>Sand (36%)
Bulk Density = (kg/m3) GAC (300) <BFS (1500) <Sand (1600)
The significance of low porosity is that it allows reduced infiltration of the influent
sample or water and thus allows very low interaction between the filter materials and
the influent. This might be one possible reason behind the low performance of sand in
treating leachate.
In addition to particle size, porosity, and bulk density, the structure of the filter material
is likely to result in differential removal of phosphorus. For example, Johansson (1999)
reported that P-removal efficiencies of BFS differed even when the structure of the filter
changed. That is, two different surface structures (amorphous and crystalline) of the
same filter material (BFS) having the same particle size (0.25 mm - 4 mm) exhibited
differential P-removal efficiency. Amorphous slag absorbed phosphorus to a greater
extent than crystalline slag owing to its vitreous surface structure that allows large
amounts of phosphate ions from the solution to penetrate between the filter particles in a
shorter period of contact time than those with fine crystalline particles. This
phenomenon partially explains why sand does not have an ability to remove P like GAC
and BFS, even though it has a similar hydraulic conductivity like BFS.
6.2.6
Drawbacks of the experiment
Since BFS contains a high percentage (%) of Ca in the form of CaO (35%), it is likely
that Ca can react with sulphate commonly found in leachate to create gypsum
(Kietlinska & Renman, 2005). Therefore, there is a need for an addition of organic
89
component (peat) to the filter matrix to decrease this effect. Peat is also known to
decrease the hydraulic conductivity of the coarse BFS materials and also act as an
adsorbent for heavy metals.
Filter materials might result in leaching several elements. Kietlinska & Renman (2005)
earlier demonstrated that BFS and sand release about 200 mg/L and 500 mg/L of Al into
the leachate. The authors attributed such occurrence to different chemical reactions such
as ion exchange and mineralisation of organic materials with strong leachate.
Finally, the experimental results with landfill leachate are more difficult to interpret than
those performed with aqueous solutions. The differences in the results obtained by the
investigation of reactive filter materials arise from the fact that leachate is a chemically
complex liquid.
90
Chapter 7 Conclusions and Recommendations
7.1
Laboratory Aeration Experiments
The major conclusions from the laboratory experiments are that the aerobic treatments
were successful in effectively treating raw landfill leachate and in reducing
concentrations of certain pollutants. Leachate was effectively stabilised at a room
temperature (21oC) without any pH adjustment or nutrient addition. The quality
improved significantly with 100% ammonia-nitrogen and 96% turbidity reduction at a
retention time of approximately 16 days. Likewise, the leachate quality improved
aesthetically both in terms of colour and odour. Such changes demonstrate that a
medium strength leachate can be treated by a simple aerobic process without involving
complex processes and technology. On the other hand, in the absence of aeration, there
was no significant improvement in leachate quality, although ammonia and turbidity
reduced by 29% and 46%, respectively. The slight reduction in ammonia-nitrogen and
turbidity was attributed to the biodegradation of organics present in the leachate in the
presence of available dissolved oxygen and volatilisation of ammonia to the
atmosphere. However, over this time, there was no improvement in the quality. Instead,
the leachate samples began to impart a foul odour with an increasing retention time
reflecting a trend toward anoxic conditions.
7.2
Laboratory Column Experiments
With respect to column experiments, out of the nine different characteristics discussed,
the filter materials were most effective at removing phosphorus from the leachate. There
was a significant decrease in the P-concentration of the leachate samples after being
filtered through the respective materials. That is P-concentration decreased by
approximately 68%, 92% and 40%, (initial~25 mg/L phosphorus) for GAC, BFS and
sand, respectively. This occurred at a corresponding contact time of approximately 30
minutes for GAC, two and half hours for BFS and two hours for sand. By comparison,
GAC and BFS were more suitable for P-removal than sand. Apart from significant Premoval, the filter materials appeared to have absorbed other forms of both organic and
91
inorganic pollutants as reflected by a considerable decrease in turbidity (with an
exception for GAC) and electrical conductivity.
Overall, an important observation is that the two treatment experiments resulted in
significant differences based on their ability to treat landfill leachate. While the aerobic
treatment is most effective and suitable for treating OM, particularly ammonia, the filter
materials are suitable for removing SRP. On this note, it may be suggested that the two
treatment experiments can together make a unique or effective combination to form an
independent leachate treatment train. In the train system, the main concept, aerobic
treatment, will function better as a primary clarifier system to remove organic materials
and ammonia first and then allow the resulting pollutants (phosphorus, metals and
refractory solids) to be removed by passing the leachate through a filter bed.
Implementation of this system would help in accomplishing complete or better effluent
quality that may not be possible using either of the processes separately.
7.3
Leachate Characterisation
Firstly, the statistical evaluation of the 14 year long term leachate monitoring data
revealed that the pattern of waste decomposition in SWMC landfill is roughly similar to
other landfills. Leachate generated from this landfill has high concentrations of OM,
ammonia-nitrogen, sodium and chlorides. These contaminants increase the pollution
potential of the leachate making it unsafe for direct release into the nearby surface water
or environment.
Secondly, a closer investigation of the change in leachate characteristics illustrated the
temporal variations of the landfill. In 14 years, the landfill showed approximately two
distinct phases. The first phase lasted for a period of first four years (1995-1999) and
included the initial stabilisation period (first year), which is the time taken by the waste
to start decomposition and for leachate to be generated. Leachates formed during this
period were essentially pH neutral, of low toxicity and characterised by very low
concentrations such as ammonia-nitrogen, electrical conductivity, suspended solids and
metals (Na, K and Fe). The concentrations of, BOD, COD and TOC were however high.
With the onset of the second phase (2000-2009), leachates became slightly alkaline and
92
more toxic. Very high concentrations of the above pollutants were then recorded in the
leachate samples.
Finally, the relation between rainfall and leachate production from the landfill for the
past 14 years helped in estimating the approximate volume of leachate generation. The
prediction is that for a mean annual rainfall of 1165 mm/year, the leachate produced
was up to 400 mm/year. This roughly indicates leachate percolation was up to 35% of
the rainfall and falls within the range (15-50%) reported by Canziani & Cossu (1989;
cited in Tatsi & Zoubolis, 2002).
7.4
Solid Waste Management Options for Bhutan
Currently, land filling of MSW is the primary method of solid waste disposal in Bhutan.
Therefore, low cost leachate treatment methods are necessary to reduce the
contamination of landfill leachate on the environment. When leachate treatment is still
not feasible, composting may be adopted as an alternative waste management technique.
After all the waste generation trend indicates a fourfold increase in waste generation for
the urban centres of Bhutan by 2020. Therefore, a sustainable and cost effective option
is also necessary to overcome what may prove to be a significant environmental
management issue for the country.
Likewise, this thesis has indentified the main problems of SWM in Bhutan based on
secondary evaluation of the performance of SWM systems in the country. The
evaluation, which is based on earlier research carried out in Bhutan, suggested that
unlike other countries, there is a lack of proper sorting and segregation of waste at the
source. Different types of waste were all collected together and transported to the
landfills. It is also clear that most landfills in Bhutan are mere open dumps or low lying
depressions on the ground and thus do not have a proper design and structure nor a
provision for leachate and gas collection facilities.
93
7.5
Recommendations and Further Research
The following recommendations are based on the results of the laboratory experiments,
a review of the literature on SWM in various countries and prevailing low cost leachate
treatment options.
Although the laboratory experiments were highly convincing in demonstrating that
leachate treatment is possible both through aerobic treatment and media filtration, it is
very important to remember that the results obtained from this study are achieved in the
laboratory and at a bench scale. The conditions prevailing here might differ from those
occurring in the field or on-site. With this qualification, the laboratory results indicate
what could be expected in the field. Further research is therefore suggested at the field
scale to increase the knowledge of using the filter materials or aeration in order to
implement it successfully.
The analysis of various research on SWM in Bhutan points out that SWM systems in
Bhutan have existed only for a shorter period of time as compared to Australia and
some other countries in the region. Even though collection, transportation and disposal
of waste are prevalent in major urban cities in Bhutan, all the three components of an
integrated waste management system such as infrastructure and maintenance are lacking
there.
94
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Appendix A
1) Aerobic Treatment of Leachate (Control and Experiment) over a period of 20 days.
Number of Days
Parameters
Initial
2
Dark
Brown
Dark Brown
with greenish
layer
formation at
the top.
Dark
Brown
Turbidity (NTU)
PH
Electrical
conductivity (dS/m)
Alkalinity ( mg/L as
CaCO3 )
DO
Ammonia-nitrogen
Nitrite-nitrogen
Nitrate-nitrogen
Soluble reactive
phosphorus
110
8.2
100
8.8
6.64
Temperature (oC)
Colour
4
6
12
14
16
18
20
Dark
Brown
8
10
Control
Dark
Brown
with oily
Dark
layer and
Brown foul smell
Turbid
Brown
and
foul
odour
Turbid
Brown
and
foul
odour
Turbid
Brown
and
foul
odour
Turbid
brown
and foul
odour
Turbid
Brown
50
8.7
50
8.7
70
8.7
60
8.7
115
8.6
65
8.8
65
8.8
55
8.8
60
8.8
6.08
6.12
6.10
6.17
6.17
6.24
6.11
6.15
6.12
6.11
1510
0.5
600
0.2
8.4
\
13
2720
0.6
410
0.3
11.2
2570
0.1
1180
0.2
8.4
2860
0
435
0.2
7.2
2610
0
460
0.2
7
2630
0
460
0.2
7.2
2520
0
480
0.3
7.9
2520
0
415
0.2
8.1
2590
0
530
0.3
7.5
2410
0
400
0.2
7.9
2520
0
430
0.2
7.9
13.3
11.8
16.6
16.3
15
14.3
13.8
14.7
13.2
13
21.6
21.9
21.6
21.8
21.6
21.4
21.1
21.5
21.3
21.7
21.5
Note: All units in mg/L unless otherwise indicated. * UMR is under measuring range.
I
1) Aerobic Treatment of Leachate (Control and Experiment) over a period of 20 days (continued).
Colour
Turbidity (NTU)
PH
Electrical
conductivity
(dS/m)
Alkalinity ( mg/L
as CaCO3 )
DO
Ammonianitrogen
Nitrite-nitrogen
Nitrate-nitrogen
Soluble reactive
phosphorus
o
Temperature ( C)
Dark
Brown
Experiment
Light
Light
Brown Brown
Dark Brown
Light
Brown
Light
Brown
Light
orange
Light
orange
Clear
orange
Clear
orange
Clear
Orange
110
8.2
70
8.6
25
8.7
15
8.2
15
8.7
10
8.6
5
8.4
5
8.3
5
8.2
5
8.7
5
8.8
6.64
6.38
6.27
6.14
6.04
6.00
5.91
5.78
5.72
5.72
5.78
1510
0.5
2630
6
2130
3
2670
3
2470
2
2210
2
1700
2
1350
3
980
2
990
4
910
4
600
0.2
8.4
450
0.2
8.7
1385
0.2
8.3
390
1.3
7.6
405
7.2
9.4
285
38
20
185
85
35
110
130
35
15
195
90
*UMR
200
87
UMR
205
100
13
11.6
12.2
17.6
17.2
15.2
14.6
13
13
12
11.3
21.6
22
21.8
22
21.9
21.4
20.9
21.8
21.6
21.9
21.9
II
2) Column Experiment Results.
Initial
Control
Parameter/ Runs
Colour
Turbidity (NTU)
pH
Electrical Conductivity
(µS/cm)
Alkalinity ( mg/L as
CaCO3)
Dissolved Oxygen (DO)
Ammonia-nitrogen
Nitrite-nitrogen
Nitrate-nitrogen
Soluble reactive
phosphorusExperiment
Colour
Turbidity (NTU)
pH
Electrical Conductivity
(dS/m)
Alkalinity (as mg/L as
CaCO3)
Dissolved Oxygen (DO)
Ammonia-nitrogen
Nitrite-nitrogen
Nitrate-nitrogen
Soluble reactive
phosphorous
Granular Activated Carbon (GAC)
Transparent
0.29
6.8
1st
Light Black
22
8.5
5th
Clear transparent
10.7
9.3
Blast Furnace Slag (BFS)
Deionised Water
10th
1st
5th
10th
Clear transparent Clear
clear
Clear
2.89
12.3
0.85
1.58
9.19
11.75
11.83
11.82
Sand
1.48
143.4
418
532
3150
3560
3680
67.5
56.7
61.8
30
320
160
230
70
130
150
40
40
60
8.96
0.156
0.002
UMR
0.023
9.54
0.29
0.06
0.78
9.66
0.33
0.06
1.14
9.64
1.26
0.03
0.68
13.14
UMR
0.1
20.9
12.12
UMR
0.09
25.3
11.44
UMR
0.017
17.3
12.71
1.12
0.009
5.76
13.03
1.08
0.01
4.69
12.92
1.26
0.014
4.79
0.657
0.80
2.27
0.439
0.63
0.793
0.057
0.052
0.027
Dark Brown
58.8
8.38
89.5
Dark Brown
8.55
Dark Brown
121.5
8.69
Dark Brown
205.1
8.72
Leachate
Dark Brown
315.5
8.9
Orange
52.75
9.33
Orange
45.9
9.53
Light Brown
51.05
8.705
Light Brown
45.1
8.775
Light Brown
42.55
8.795
10.48
10.47
10.22
10.15
10.795
10.155
9.865
9.840
9.960
10.015
4700
4420
4920
4600
3415
1915
1615
3990
4230
4360
2.19
975
0.20
9.46
7.76
1177.5
0.24
11.40
7.60
1115
0.28
7.60
7.70
1070
0.28
6.49
9.755
1090
0.156
12.3
9.205
1107.5
0.142
26.85
5.825
1137.5
0.121
20.8
9.305
1095
0.2095
13.35
10.205
1242.5
0.244
13.05
9.19
1160
0.223
12.8
25.3
8.95
9.41
7.94
6.235
2.42
1.5
17.9
14.2
15.25
1st
Clear
1.72
7.37
Sand
5th
Clear
2.39
7.44
10
Clear
2.29
7.54
th
Note: All units in mg/L unless otherwise indicated. * UMR is under measuring range.
III
3) Particle Size Distribution Analysis.
Media
GAC
BFS
Sieve Size (mm)
5
4
2.8
2.4
2
1
0.5
0.25
0.125
0.063
<0.063
5
4
2.8
2.4
2
1
0.5
0.25
0.125
0.063
<0.063
Mass retained on Sieve (g)
0
160
126
8
3
0
0
0
0
0
0
0
0
0
0
186.64
92.45
61.47
34.45
15.05
5.42
0.47
% of total mass retained on sieve
0
53.0
42.0
2.7
1.0
0.0
0.0
0.0
0.0
0.0
0.0
0
0
0
0
47.14
23.35
15.52
8.70
3.80
1.37
0.12
% of total mass passing sieve
100
47.00
5.00
2.30
1.30
0.00
0.00
0.00
0.00
0.00
0.00
100
100
100
100
52.86
29.51
13.99
5.29
1.49
0.12
0.00
IV
3)
Media
Sand
Particle Size Distribution Analysis (continued).
Sieve Size (mm)
5
4
2.8
2.4
2
1
0.5
0.25
0.125
0.063
<0.063
Mass retained on Sieve (g)
0
0
0
0
0.14
1.69
97.01
284.65
12.54
0.19
0.19
% of total mass retained on sieve
0
0
0
0
0.04
0.43
24.47
71.81
3.16
0.05
0.05
% of total mass passing sieve
100
100
100
100
99.96
99.54
75.07
3.26
0.10
0.05
0
V
Appendix B
1)
Temporal variations in Leachate Quality (1995-2009)-Figures 1 a-g represents leachate
quality relative to time from SWMC landfill, Newcastle, Australia (1995-2009).
Figure 1-a) Temporal variations in ammonia relative to time in leachate
from SWMC landfill, Newcastle, Australia (1995-2009).
Figure 1-b) Temporal variations in pH relative to time in leachate from
SWMC landfill, Newcastle, Australia (1995-2009).
VI
Figure 1-c) Temporal variations in electrical conductivity relative to time in
leachate from SWMC landfill, Newcastle, Australia (1995-2009).
Figure 1-d) Temporal variations suspended solids relative to time in leachate
from SWMC landfill, Newcastle, Australia (1995-2009).
VII
Figure 1-e) Temporal variations in TOC/BOD/COD relative to time in
leachate from SWMC landfill, Newcastle, Australia (1995-2009).
Figure 1-f) Temporal variations in Sodium and Potassium relative to time in
leachate from SWMC landfill, Newcastle, Australia (1995-2009)
VIII
Figure 1-g) Temporal variations in chloride and sulphate relative to time in
leachate from SWMC landfill, Newcastle, Australia (1995-2009)
IX
2) Rainfall Data (1996-2009), Bhutan (Hydro-met Services Division, Department of Energy, MTI, Thimphu, Bhutan
(1996-2009)).
Station: 220m
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
70
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
25
0
15
0
0
17
24
11
0
3
5
0
0
6
8
86
2
0
0
0
10.3
0
24
23.4
2
0
39
0
0
0
2
0
0
0
0
0
0
0.2
0
0
0
19
10
0
0.2
0
0
12
0
0
0
0
0
0
0
30
11
0
0
15.5
22
94
22
29
0.5
0
6.1
3.5
6
57
0.2
0
11
0
74
41
0
2.1
20
6
0.6
6
19
9
1
4.5
24
19
2.5
7
2
0.5
71
29
20
20
50
42
25
27
32.7
110
4.3
2.8
2
3
6.5
0
13
0
0
7
0
15
0
1
0
0
44
3
0
79
66
8
15
29
86
29
100
90
7
62
26
46
32
18
7.6
18
2.2
0
95
9
7.5
0
70
100
21
25.2
6.2
39
34
30
44
2.5
8
0
5
23
0.7
0
47
0
8
10.2
14.6
2.2
0
0
12
10.2
2.5
0
0
0
0
2.5
28
39
10.5
2.5
0
14
0
0
0
0
0
22.5
0
0
0
0
0
0
0
22.9
120.8
0
0
22.5
25
0
0
0
0
0
0
10
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
5.6
0
0
0
1
0
0
0
0
0
0
0
0
0
0
0
4
0
0
0
0
0
0
0
0
0
5
4
0
Note: Rainfall in mm
X
2) Rainfall Data (1996-2009), Bhutan (Hydro-met Services Division, Department of Energy, MTI, Thimphu,
Bhutan (1996-2009)). (continued).
Station: 375 m
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
32.8
0
0
0
0
11.8
0
0
0
0
0
0
0
0
0
0
0
17.6
0
0
0
0
0
16
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
6.2
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3.6
0
0
13.2
0
10.2
0
0
12.2
3.6
4.6
0
0
12.2
7.2
0
6.2
0
0
7.2
2.8
0
0
0
0
0
30.2
8.2
0
7.2
0
0
0
0
0
0
0
0
0
57.2
64.4
3.8
0
0
31.6
0.6
4.2
0
21
9
0
4.8
1.8
0
4.2
0
3.6
17.6
0
0
0
6.4
0
0
10.8
55.8
20
25
0
0
24.6
1.2
40.2
1
10.2
0
39
0
57.4
40.8
6.8
75.2
7.2
1.2
7
10.4
54.4
0
0
1
32.2
16.2
107.4
46.8
31.8
0
134.6
168.6
29
191.2
268.8
160
69
205.8
89
99.8
30.2
1.4
65.8
2.4
7.6
0
0
0
0
4.6
26.6
24
0
5.6
9.6
0
1.4
11.8
3.6
257.2
34.8
19.4
0
163
22.4
85.8
24.4
8
0
14.2
10.6
27
21
1.2
3.2
2
6.2
90.8
42.4
10.8
9.2
113
40.6
29.2
189
64
21.8
168.8
30.6
7.2
13
0
2.8
3.6
1.2
76.2
19.6
0
0
0
0
0
11.8
0
1.4
0
13.6
35.4
7
0
30.8
7.6
9
77.4
256.6
30.2
0
19
0
29.4
28.2
4.6
0
0
0
0
0
0
0
0
0
0
5.8
57.6
204.8
34.4
0
16.2
0
0
0
0
0
7.6
4.4
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3.2
0
0
0
0
0
0
0
0
0
0
17.6
0
0
0
0
0
0
0
0
0
0
0
1
0
Note: Rainfall in mm.
XI
2) Rainfall Data (1996-2009), Bhutan (Hydro-met Services Division, Department of Energy, MTI, Thimphu,
Bhutan (1996-2009)). (continued).
Jan
Station: 1930m
0
1
0
2
0
3
0
4
0
5
0
6
0
7
0
8
0
9
0
10
0
11
0
12
0
13
0
14
0
15
0
16
0
17
0
18
0
19
0
20
0
21
0
22
0
23
0
24
0
25
0
26
0
27
0
28
0
29
0
30
0
31
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3.8
29.8
11.8
0
8.8
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.6
5.6
0
0
0
0
23.8
0
0
2
12.2
4.8
0
2.6
1.2
0
25.6
12
0
11.6
30.8
0
4.4
0
0
3.6
5.4
0
0
0
0
1.4
1.2
0
0
0
0
0
0
2.2
0
0
0
2
15.4
0
0
0
0
0
0
3.8
2.8
5.6
0
0
0
3.6
0
0
0
0
0
0
3
71.8
34
8.8
0
0.6
0
0
0
0
0
9.6
0
0.4
0
0
0
1
0
0
1.2
0
0
0
0
0
0
0
4
0
0
0
0
4.2
0
12.4
28.4
18.2
8.4
42.2
15.4
1.6
0
0
3.8
0
4.2
0
3.8
0
1.4
0.6
1.8
0
0
39.8
2.4
0
9.2
0
0
6.4
8
15.2
2
33.6
14
9.2
0.8
0
0
7.4
2.8
10.6
5.2
10.6
0
0
0
1.4
11.6
2.4
0.8
6.2
1.4
3
34.2
18
2
0.8
0.6
29.6
16.8
10
11.8
7.6
0.8
1
3.8
0.6
0
0
0
2.6
3.8
14.6
15.2
3.2
9.6
3.4
0
0
0
0
0
0
0
0.8
28
0
0
0
0
0
0
5.6
0
14.6
0
0
0
0
2.2
3.8
1
25.2
38.8
24.8
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.8
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0.4
1.2
XII
2) Rainfall Data (1996-2009), Bhutan (Hydro-met Services Division, Department of Energy, MTI,
Thimphu, Bhutan (1996-2009)).(continued).
Jan
Station: 2380m
0
1
0
2
0
3
0
4
0
5
0
6
0
7
0
8
0
9
0
10
0
11
0
12
0
13
0
14
0
15
0
16
0
17
0
18
0
19
0
20
0
21
0
22
0
23
0
24
0
25
0
26
0
27
0
28
0
29
0
30
0
31
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3
8.8
3
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3.3
0
0
0
0.6
5.6
0
0
0
0
0
0
0
9.7
6.7
0
2.6
0
0
0
0
0
1.9
0
0
0
0
0
0
0
0
0
0
0
0.5
0
0
0
1.8
6.3
0.9
0
0
0
0
1.6
3.5
0.2
7.1
0.2
0
0
1.9
0
0
0
0
0
0
0
0
4.9
72.5
47.2
0
2.3
0
0.7
0
1.1
9
0.4
15
1.4
3.1
0
0.6
0
0
0
0
0.8
0
0.3
0
0
1
0.8
0
0
0
0
0
0
0
0
0
4.8
0
2.8
2.9
0.9
0
0
10.6
0.6
0.2
0.6
2.6
0.6
0
0
0
0
0
3.2
1.9
0
1.9
0
1.5
0
2.2
14
1.4
3.3
13.4
0
3.2
0
0
0
20.5
18.7
26.4
9.2
0.6
0
12
6.2
13
0
0
2.8
0
0.5
8.1
3.7
0
0
15.6
8.7
4.2
24.4
1.7
0.1
1.4
7.1
0.2
1
1
6.9
3.4
0
0
6.3
1.4
3.1
7.1
0
4.6
0
1.2
0
0
0
0
0.2
0.1
1.8
0
0
0
0
1.4
1.6
0.3
0
0
0
0
0
0
0
0.3
27.7
81.4
0.1
0
0.1
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
Note: Rainfall in mm
XIII
2) Rainfall Data (1996-2009), Bhutan (Hydro-met Services Division, Department of Energy, MTI, Thimphu,
Bhutan (1996-2009)). (continued).
Station:2760m
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
3
3.1
4
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
2
0
0
0
0
0
2
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
2
3
3
2
4
2
0
1
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
10
0
0
6
0
0
5
15
5
0
0
13
0
0
13
27
0
27
0
0
0
4
1
2
6
11
8
2
4
14
1
1
0
0
0
2
0
0
2
3
0
2
3
0
3
0
2
0
2
4
3
0
0
3
2
0
0
3
3
4
3
3
3
5
18
5
3
5
4
0
2
3
0
3
1
1.2
2.4
10
1
4
3
1
0
17.2
5.2
1
1.4
1.2
12
14
31.4
6.4
2
8
18
5
8
17.4
12.4
13.4
17
6.4
14
9
9.4
3.4
26.4
6.4
11
6
21.4
13
11.6
4
1.2
9.4
14.4
16.4
43
6
1
0
30.4
23
7.4
0
9.4
1.5
10
13
22
24
2.4
10.4
28.8
12.4
12
18
20.4
12.4
40
17.4
1
1.4
11
4
7.4
24
22
7.4
13.4
7
11
7.4
4.4
14.4
2.2
3.4
6
0
23.2
23.2
14.4
20
26.4
11.4
3.2
17
10.4
4
14.2
0
2
2
4
0
4.4
12
35.4
2.4
8
8.4
0
1
2
2.4
0
0
0
0
0
0
2
2
0
0
15
0
4
0
7.2
9
0
2.2
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
3.2
4.3
0
0
0
0
0
0
0
0
0
0
0
-
Note: Rainfall in mm
XIV
3) Monthly Rainfall Data (1995-2009), Williamtown Meteorological Station
(site number 061078), Newcastle, Australia.
Year
Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
Annual
1995
105.4
54.4
194.8
11.6
188.4
101.6
1.8
0
115.2
30
146.8
87.8
1037.8
1996
117.8
94.6
78.6
4.6
112
156.6
47.4
73.4
62
48
87
60.4
942.4
1997
99.2
152.2
101.8
42.2
135.4
128.2
190.4
40.8
94.6
38.8
11.2
44.4
1079.2
1998
67
95.6
27.6
104.8
340
156.8
102.8
169
94.4
69.2
241.4
91.6
1560.2
1999
72.2
252
46.6
332.2
27.6
212
141.8
96.2
58.8
129.2
111.8
61
1541.4
2000
60.8
35.4
359.8
104.6
59.2
91.6
43.8
66.8
25.8
35.8
79
33
995.6
2001
62.2
113.2
154.2
103.8
410.2
14.6
74.2
41.2
54
38
105.4
56
1227
2002
17.2
270.4
126
68
129
82.8
36.6
61.2
45.2
7
38.4
171.8
1053.6
2003
9.6
74.6
58
167.8
178.8
29.8
94.6
50
0.4
74.4
115
44.8
897.8
2004
140.8
178.6
90.8
56.6
20.6
25.6
78.6
103.6
91.2
186.4
74.6
68.4
1115.8
2005
68.4
143
220.8
34.6
222.4
88.4
16.6
0.8
49.8
55.8
82
14.4
997
2006
41.2
99.8
78
90.4
79.4
105
101.8
100.2
162.2
12.8
126.8
68.6
1066.2
2007
14.8
59.2
137.6
250.4
42.6
414.2
28
92.8
50.8
49.4
104
85.2
1329
2008
181.4
222.6
64
256.6
35.8
142.6
92.8
34
179.2
76.2
88.2
90.2
1463.6
2009
20.6
229.8
49.2
196.2
153
112.6
50.2
1
20.2
61.4
Note: Rainfall in mm.
XV
4) Monthly Temperature (mean maximum) Data (1995-2009), Williamtown Meteorological
Station (site number 061078), Newcastle, Australia.
Year
Jan
Feb
Mar
Apr
May
Jun
Jul
Aug
Sep
Oct
Nov
Dec
Annual
1995
26.3
26.9
25.8
23.2
20.6
17.5
17.2
21.9
19.9
24.2
26.3
25.1
22.9
1996
26.7
25.9
25.2
24.2
21.2
18.8
17.2
19.3
22.4
23.4
25.2
26.6
23
1997
26.2
27.6
26.4
24.8
20.5
18.2
16.6
19.1
20.1
24.6
28.1
30.5
23.6
1998
30.2
30.4
28.3
24.7
20.5
17.2
16.4
18.6
22.1
23.7
22.9
26.8
23.5
1999
28.8
26.9
27.2
21.9
21.6
17.7
17.5
18.9
21.8
24.3
23.3
25.8
23
2000
26.3
29.6
25.9
23.5
20
17.2
17.8
18.4
23.4
24.2
24.6
29.7
23.4
2001
30.9
29.4
26
24.3
19.2
19
17.6
19.3
22.3
24.5
24.6
28.2
23.8
2002
29
26.7
26.5
24.3
20.2
18
18.4
19.8
22.8
27
28.3
27.8
24.1
2003
29.5
28.2
25.1
22.5
20.5
19.3
17.3
18.9
23.3
22.5
25.2
28
23.4
2004
30
29.5
26.3
24.2
20.7
19
17.6
19.5
21.8
23.7
26.9
26.6
23.8
2005
28.6
28.2
25.2
24.9
20.4
18.5
18.3
20
21.2
25
26.2
30.7
23.9
2006
29.6
29.9
27.3
24.4
20.1
17.3
17.6
19.5
23.2
25.1
26.5
26.1
23.9
2007
29.5
28.6
27.6
23.2
22.3
16.7
17
20.3
21.9
26.9
25.1
26.2
23.8
2008
27.5
25.3
25.8
21.9
20.5
18.5
17.2
17.6
22.1
24.4
24.7
27.9
22.8
2009
30.2
28.4
27
23.3
20.6
18.1
17.7
21.6
23.7
23.1
28.5
27.5
24.1
o
Note: Temperature in C.
XVI
5) Leachate Monitoring Data (1995-2009) from SWMC, Newcastle, NSW, Australia.
Date
5.07.1995
31.07.1995
3.08.1995
10.08.1995
17.08.1995
25.08.1995
7.09.1995
14.09.1995
21.09.1995
28.09.1995
5.10.1995
13.10.1995
20.10.1995
27.10.1995
3.11.1995
10.11.1995
16.11.1995
20.11.1995
21.11.1995
22.11.1995
23.11.1995
24.11.1995
1.12.1995
31.01.1996
15.02.1996
8.03.1996
12.04.1996
10.05.1996
12.06.1996
26.07.1996
23.08.1996
26.09.1996
24.10.1996
1.11.1996
21.11.1996
20.12.1996
Ammonia
0.05
0.01
0.01
0.01
0.01
0.01
0.01
0.01
6.07
29.5
9.62
9.37
8.62
7.66
5.53
4.03
3.69
5.15
3.99
2.5
10.2
10.7
16.6
15.8
12.01
14.1
20.4
37.3
25.5
35.5
30.1
49.7
52.6
41.4
33.2
Nitrite
Nitrate
EC
(us/cm)
pH
Alkalinity
(mg/L of
CaCO3
BOD
COD
Fl
Cl
Sulphate
Ca
Mg
K
Na
0.76
161
103
140
133
156
27
51
80
232
126
138
179
150
170
172
160
390
280
150
98
110
60
59
66
54
52
5
5
133
103
112
117
124
88
131
280
122
139
111
110
117
93
117
175
105
117
8
9
6
7
19
63
27
29
145
155
48
69
100
209
116
128
171
147
131
110
107
1
31
70
75
73
78
53
146
119
99
83
97
96
138
110
90
75
86
90
33
27
22
25
25
26
157
132
109
96
102
106
104
92
60
183
89
57
190
101
71
264
Fe
Pb
5
0.01
0.12
0.01
0.01
0.06
0.01
0.02
0.01
0.01
0.02
0.01
0.01
0.01
0.01
270
336
623
723
730
522
460
569
714
686
632
840
676
580
418
528
478
962
0.01
2368
0.01
0.08
2219
2360
2500
7.8
7.2
7.3
7.3
6.4
6.6
6.6
8.1
7.9
8
7.9
7.2
7.8
7.7
8
6.5
7.5
0.4
0.2
2710
2660
7.5
7.5
0.01
0.04
0.07
0.03
0.03
0.03
0.03
7.7
7.6
7.8
7.3
6.8
6.6
7.6
6.6
6.3
7.3
7.4
1850
0.04
2250
800
820
3
0.65
31
57
42
694
2750
857
1060
1060
1170
1060
1160
1070
1080
983
663
1080
1260
1860
783
265
70
40
351
101
41
25
482
271
26
365
102
81
985
3850
1280
1420
1540
1160
1410
1380
1130
1150
1070
818
1510
2270
1260
545
231
171
609
205
242
174
855
465
274
174
0.46
0.22
0.82
0.22
0.11
0.51
0.57
0.54
2.7
2.1
0
0
0
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0
0.1
0.1
0.1
0.1
0.1
0.1
6
0.72
280
95
0.49
235
28.6
0.59
318
39
86
426
0.1
0.1
16
0.1
XVII
Date
31.01.1997
20.02.1997
20.03.1997
16.04.1997
12.05.1997
14.05.1997
16.05.1997
4.06.1997
17.07.1997
14.08.1997
25.09.1997
31.10.1997
19.11.1997
20.11.1997
18.12.1997
21.01.1998
19.02.1998
26.03.1998
23.04.1998
5.05.1998
7.05.1998
30.06.1998
23.07.1998
27.08.1998
30.09.1998
22.10.1998
30.11.1998
24.12.1998
21.01.1999
28.01.1999
4.02.1999
9.02.1999
24.03.1999
22.04.1999
24.05.1999
1.07.1999
21.07.1999
11.08.1999
Ammonia
11.4
53.5
57.3
25.7
6.99
16.2
17.4
15.2
86.7
Nitrite
Nitrate
0.01
0.12
0.43
0.03
23.2
57.3
42
9.73
28.3
63.5
35.7
62.9
95.9
158
164
145
167
266
84.5
46.2
74.9
35.9
22.4
17.8
37.2
44.7
64.4
105
51.5
96.3
202
0.46
0.01
0.02
0.07
0.01
0.01
0.02
0.03
EC
(us/cm)
2612
1378
1261
1051
1300
2045
3934
3840
1613
3590
3760
3500
4360
4840
1860
4476
3860
3222
3685
4567
0.01
0.04
0.03
0.01
0.01
2720
3277
3366
4137
3360
1653
1731
2.09
1883
2250
4137
0.08
3120
5410
pH
7.3
7.3
7.4
8.1
6.4
6.8
6.8
7.2
7.1
7.6
7.2
7.8
8.3
7.3
8.9
8.7
8
7.7
7.8
6.9
6.9
7.3
7.7
7.6
7.3
7.7
8.7
8.6
7.2
7.4
8.3
6.8
7.4
7
7.7
6.8
7.13
7.6
Alkalinity
(mg/L of
CaCO3
998
190
329
1470
1560
972
1200
2180
1160
76.2
675
877
1610
2560
BOD
COD
26
99
285
22
239
361
395
421
1560
530
140
164
57
404
322
474
167
26
47
68
23
96
781
930
69
48
62
74
17
14
40
17
28
89
107
19
23
92
500
828
533
375
314
495
404
454
1080
1510
414
438
424
534
149
344
531
293
128
243
320
168
230
789
451
1440
1110
Fl
Cl
Sulphate
Ca
Mg
K
Na
0.55
279
47
82
94
76
254
0.1
0.46
0.54
82
118
48
49
26
29
12
86
0.1
0.1
408
1
84
145
98
352
0.1
1.3
681
13
8.4
211
621
0.1
0.82
650
5
36
192
549
705
2310
973
Fe
Pb
564
372
528
1.7
0.1
262
314
20
27
0.1
0.1
0.64
0.42
333
353
34
1
116
1.6
94
108
#
83
99
0.8
537
14
221
155
##
474
0.1
1.08
537
38
34
123
##
421
0.1
0.32
0.36
0.36
204
211
202
121
77
39
29
48
65
53
57
60
49
38
51
198
203
196
0.1
0.1
0.1
0.91
547
35
89
138
##
434
0.1
0.54
580
1
154
166
##
530
26
XVIII
Date
6.09.1999
7.10.1999
10.11.1999
2.12.1999
13.01.2000
9.02.2000
6.03.2000
5.04.2000
25.05.2000
15.06.2000
12.07.2000
10.08.2000
24.08.2000
7.09.2000
11.10.2000
7.11.2000
5.12.2000
21.12.2000
12.01.2001
7.02.2001
7.03.2001
11.04.2001
10.05.2001
5.07.2001
16.08.2001
6.09.2001
14.11.2001
11.12.2001
7.01.2002
6.02.2002
21.03.2002
24.05.2002
20.06.2002
7.08.2002
11.09.2002
2.10.2002
22.10.2002
22.11.2002
Ammonia
Nitrite
165
88.6
66.5
137
136
110
200
136
227
223
223
268
430
173
168
192
128
62.8
68.4
79.3
73.6
483
206
874
970
146
180
94.3
970
685
479
564
514
842
458
429
Nitrate
0.1
5090
4910
3520
9200
7910
6490
10800
3690
8390
6910
7590
7010
0.01
1200
9010
9200
0.04
0.16
0.09
0.03
EC
(us/cm)
0.02
0.01
0.01
0.01
0.01
0.01
0.04
0.01
0.05
0.01
6290
8740
3360
3320
2320
2330
11400
9600
16300
16400
7250
9800
2450
16500
12100
9870
10900
11100
1530
13000
pH
7.56
7.15
7.8
8.12
7.44
7.53
7.66
6.93
7.54
8.26
8.42
8.46
8.1
7.91
8.59
8.64
7.3
8.17
8.23
6.94
7.89
6.55
6.9
7.71
8.2
7.51
7.24
8.76
8.69
7.27
7.37
8.07
7.93
7.61
7.91
7.93
8.49
8.73
Alkalinity
(mg/L of
CaCO3
1320
1960
3370
3350
2730
BOD
COD
249
121
175
55
151
22
92
191
308
430
460
644
117
378
208
148
134
368
352
664
610
461
429
625
297
927
388
784
588
744
952
2080
734
739
1140
1010
614
1380
444
390
940
748
795
3720
6950
771
4710
3940
4230
4370
204
524
162
181
143
266
213
179
113
93
504
435
582
80
75
37
562
1032
211
964
1610
1550
1320
798
1010
667
1730
1340
1430
1160
1430
2020
1830
2660
Fl
Cl
Sulphate
Ca
Mg
K
Na
Fe
0.53
388
8
97
93
##
466
4.1
0.6
###
20
60
211
##
748
9.6
1.1
###
96
122
241
##
855
99
0.9
###
79
25
76
216
##
910
12
1.2
###
106
16
55
250
##
###
9.6
0.5
463
65
107
90
##
312
15
0.4
0.3
220
44
10
63
68
59
50
86
77
178
166
8.5
8.1
1
###
90
87
199
##
###
20
1.2
###
55
117
225
##
###
20
0.4
262
81
44
40
92
196
4.7
1.1
###
95
101
190
##
###
12
1.2
###
45
99
184
##
###
26
1.2
1.4
###
###
86
111
103
78
196
201
##
##
###
###
8.9
12
Pb
0
0
XIX
Date
23.01.2003
5.02.2003
3.03.2003
7.04.2003
21.05.2003
3.06.2003
15.07.2003
18.08.2003
1.09.2003
2.10.2003
30.10.2003
1.12.2003
6.01.2004
18.02.2004
2.03.2004
6.04.2004
6.05.2004
21.06.2004
1.07.2004
5.08.2004
1.09.2004
30.09.2004
3.11.2004
1.12.2004
11.01.2005
1.02.2005
2.03.2005
12.04.2005
24.05.2005
2.06.2005
4.07.2005
3.08.2005
1.09.2005
5.10.2005
14.11.2005
5.12.2005
10.01.2006
1.02.2006
Ammonia
423
880
183
502
241
264
486
197
241
262
356
74.2
256
83.4
108
122
334
646
490
445
342
170
128
156
193
100
197
240
330
275
46
328
296
311
553
92.9
26.8
122
Nitrite
Nitrate
0.06
0.61
0.03
0.08
0.01
0.36
0.01
0.06
0.04
0.01
0.203
0.458
0.077
0.013
0.067
0.018
0.045
0.01
EC
(us/cm)
pH
11600
15300
8030
9500
5660
6270
9640
7380
8300
11300
10200
4650
7950
7900
4920
4750
9170
11200
11200
10100
6520
5170
5220
6900
7090
7310
7600
6960
5290
6050
1610
5150
8190
9100
9870
3510
1550
7930
7.98
8.18
8.16
8.26
8.06
8.09
8.13
8.8
8.9
8.79
8.45
8.9
8.41
8.63
8.61
8.62
8.71
8.09
8.35
8
8.58
8.26
8.36
8.4
8.49
8.63
8.32
8.53
7.88
8.05
8.37
8.43
8.76
8.5
7.97
7.48
7.24
8.35
Alkalinity
(mg/L of
CaCO3
5550
1460
2050
1150
2070
2750
2200
1140
1780
2140
BOD
COD
92
66
582
318
164
6
14
298
76
20
124
111
165
112
40
88
44
63
206
143
51
59
107
96
96
135
53
59
10
50
31
17
252
822
171
24
2
356
1730
664
1510
855
363
574
887
1400
936
1310
1420
646
1370
891
431
555
921
1230
1040
1200
896
607
730
1190
1440
1310
890
1000
441
358
79
466
1980
1680
1620
403
166
1800
Fl
Cl
Sulphate
Ca
Mg
K
Na
Fe
1.3
###
76
53
222
##
###
16
0.8
649
369
160
83
##
432
2
1.3
###
125
62
129
##
862
8.4
0.6
908
132
26
79
##
655
1.3
1.2
###
85
29
150
##
###
1.9
1
###
140
40
146
##
###
4.6
1
###
103
54
108
##
811
4.7
0.6
810
125
34
57
##
623
4
1
###
108
45
110
##
###
2.8
0.6
768
121
61
54
##
429
3.7
0.8
844
164
94
78
##
607
9.3
1.5
###
126
68
128
##
###
12
1.5
146
137
63
116
##
###
8.4
Pb
0
0
0
XX
Date
1.03.2006
11.04.2006
3.05.2006
31.05.2006
3.07.2006
2.08.2006
5.09.2006
11.09.2006
15.09.2006
3.10.2006
1.11.2006
4.12.2006
22.01.2007
5.02.2007
5.03.2007
2.04.2007
2.05.2007
1.06.2007
13.06.2007
22.06.2007
29.06.2007
5.07.2007
2.08.2007
11.09.2007
2.10.2007
7.11.2007
3.12.2007
8.01.2008
11.02.2008
26.03.2008
2.04.2008
7.05.2008
11.06.2008
3.07.2008
5.08.2008
4.09.2008
1.10.2008
17.11.2008
Ammonia
89.3
402
502
218
420
193
535
42
53.8
590
437
488
363
251
108
252
191
252
49.9
86.5
150
123
72.4
255
437
294
399
471
154
353
468
308
209
105
91.6
148
85.8
248
Nitrite
Nitrate
0.083
0.048
0.051
0.01
0.01
0.015
0.011
0.01
0.039
0.386
0.104
0.027
0.013
0.019
0.01
0.01
0.01
0.012
0.01
0.01
0.01
0.044
0.073
0.01
0.021
0.149
0.039
0.038
0.555
0.02
0.06
EC
(us/cm)
pH
2300
9710
5780
6210
7170
4570
10800
1600
1900
10500
15200
11200
11200
15900
3500
7780
3490
5130
2120
2800
3990
3480
2930
6530
8490
10600
8190
11900
3780
7900
10100
7270
4700
3310
2980
5890
2720
5450
7.35
8.17
8.55
8.39
8.14
8.22
7.16
6.01
6.31
7.78
8.02
7.93
7.72
7.85
7.13
7.32
7.23
7.64
6.84
6.68
7.14
7.27
7.78
7.72
7.42
8.14
7.63
7.6
7.23
7.98
7.88
7.97
7.97
7.58
7.58
8.07
7.9
7.76
Alkalinity
(mg/L of
CaCO3
BOD
COD
72
112
89
242
286
395
343
438
398
306
215
129
89
256
21
103
1230
301
1320
1200
721
949
43
984
118
218
468
170
120
106
70
65
41
28
45
40
41
46
270
1830
1800
1350
1730
1380
1600
850
840
1940
1810
1230
1660
1710
522
768
2030
1260
1990
2080
285
1880
1280
1600
1490
1760
495
1070
1040
611
376
230
175
428
310
573
Fl
Cl
Sulphate
Ca
Mg
K
Na
Fe
Pb
1.5
###
86
64
111
##
###
7.9
0
0.7
773
84
97
80
##
641
4.9
0.3
0.5
192
248
81
104
52
86
29
36
43
64
161
240
7.1
12
1.1
###
92
126
202
##
###
9.9
1.4
###
88
95
223
##
###
12
0.5
375
53
81
73
##
394
16
0.4
0.5
268
389
40
45
104
106
69
78
84
##
257
401
22
16
0.5
363
164
65
46
56
482
5.5
1.1
###
73
109
183
##
###
6.5
0.6
583
20
81
65
##
418
11
0.7
890
48
79
105
##
727
9.9
0.6
402
196
62
59
71
422
20
0.7
763
74
51
74
##
598
8.3
0
0
XXI
Date
1.12.2008
19.01.2009
4.02.2009
30.03.2009
28.04.2009
12.05.2009
Ammonia
275
375
418
245
479
735
Nitrite
Nitrate
0.5
0.5
0.1
0.1
EC
(us/cm)
pH
5140
9790
10800
9560
8760
9880
7.72
8.28
8.49
8.56
7.97
8.06
Alkalinity
(mg/L of
CaCO3
BOD
COD
46
649
96
116
52
71
443
1650
1580
1330
746
1010
Fl
Cl
Sulphate
Ca
Mg
K
Na
Fe
1.2
###
10
119
185
##
###
8.5
1.1
###
71
114
162
##
970
7.2
Pb
0
Note: All units in mg/L unless otherwise indicated. Blank cells indicate “No Sample”.
XXII