innovative applications of natural zeolite bertholini_thesis.pdfgrease are held in these filters....
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INNOVATIVE APPLICATIONS OF NATURAL
ZEOLITE
Marita Guarino Bertholini
Master of Engineering Management
Bachelor of Chemical Engineering
Submitted in fulfilment of the requirements for the degree of
Master of Engineering (Research)
School of Physics, Chemistry and Mechanical Engineering
Science and Engineering Faculty
Queensland University of Technology
2016
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Keywords
Ammonium, filter media, hydraulic conductivity, isotherm, landfill leachate, laterite,
natural zeolites, resin, stormwater runoff
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Abstract
Water is the most precious resource for life as we know it. Ever increasing global population
represents increased pressure to our limited freshwater supplies, not only due to higher
consumption rates but also because of accelerated instances of pollution of water bodies.
Likewise, we are also observing an increase in waste production and consequently in
activities such as landfilling. Common to problems with water and wastewater, is the
presence of ammoniacal nitrogen.
It was our hypothesis that ammoniacal nitrogen should be regarded as a nutrient source and
not as a waste material to be disposed of. Hence, we proposed that what is required are
new materials which can exchange ammonium species not only in higher amounts but also
selectively in the presence of competing cations. Natural zeolites were of particular interest
as they were naturally available and abundant, low cost and stable.
The use of natural zeolites generally involves a filtration system. However, methods of
determination of key performance parameters, such as the hydraulic conductivity, for the
design of these types of equipment were found to be lacking. Hence, in this research a
stormwater filter design comprising zeolite and laterite ore was used to study the hydraulic
conductivity. Standard test methods were not representative of the system and hydraulic
conductivity was found to have been measured with two different approaches in relation to
the Δh reference point. The calculation of hydraulic conductivity by the tailwater Δh
reference point demonstrated insensitivity to changes in filter depth, while hydraulic
conductivity calculated at different depths using the bed top Δh reference showed
significantly different results.
Natural zeolites were demonstrated to be able to remove ammonium ions from solution.
Sodium forms of zeolite were more effective (12.24 g NH4/kg zeolite) than calcium
exchanged natural zeolite (9.18 g NH4/kg zeolite) and acid forms (6.12 g NH4/kg zeolite). In
contrast, synthetic exchange resins equilibrated with ammonium relatively fast but
exhibited minimal selectivity for ammonium in the presence of other cations. The zeolite
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performance was limited by very slow diffusion and showed a complex relationship
between sorbing and desorbing species which was not stoichiometric as would be expected
with an ion exchange process.
Treatment of ammoniacal solutions with modified zeolites of a core-shell structure was
positive in that ammonium uptake was promoted. However, issues with increased pH of the
solution were discovered which may require modification of the Si/Al ratio of the zeolite to
minimize this latter issue.
Synthesis of core-shell zeolites was investigated in more detail to ascertain the variables
important during the zeolite modification process. Application of non-hydrothermal
conditions was found to be inefficient at producing a shell comprising of zeolite N and/or W.
Instead, hydrothermal reactions at 175 oC were recommended. The zeolite synthesis
process was shown to result in a shell of nano-crystalline zeolites. The success in creating
the latter material depended upon the scale of the synthesis reaction and reaction
conditions such as agitation method.
Overall, the concept of using a core shell zeolite based upon a natural zeolite core and an
outer layer of zeolite N and/or W has been demonstrated. Future work is required to
optimise the preparation of the core shell structure and to improve its performance in
relation to ammonium removal and recovery from solution.
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Table of contents
1.1 Background ................................................................................................................... 12
1.2 Research Need .............................................................................................................. 12
1.3 Research Goals .............................................................................................................. 13
1.4 Significance and Innovation .......................................................................................... 14
1.5 Thesis Outline................................................................................................................ 15
2.1 WATER RESOURCES ...................................................................................................... 18
2.2 AMMONIA ..................................................................................................................... 18
2.2.1 Effects of ammonia contamination ....................................................................... 18
2.2.2 Sources of ammonia in the environment .............................................................. 19
2.2.2.1 Stormwater runoff ................................................................................................................................ 19
2.2.2.2 Landfill leachate ..................................................................................................................................... 20
2.3 TREATMENT .................................................................................................................. 22
2.3.1 Biological treatment .............................................................................................. 23
2.3.2 Ion exchange .......................................................................................................... 24
2.3.3 Air Stripping ........................................................................................................... 27
2.3.4 Ammonia adsorption on charcoal ......................................................................... 27
2.3.5 Operational considerations for filters ................................................................... 29
2.4 ZEOLITES ........................................................................................................................ 31
2.4.1 Crystal structure and classification ........................................................................ 31
2.4.2 Exchange mechanism ............................................................................................ 33
2.4.3 Natural zeolite ....................................................................................................... 34
2.4.4 Synthetic zeolite .................................................................................................... 34
2.4.5 Modification of natural zeolites ............................................................................ 36
2.5 SUMMARY AND IMPLICATIONS .................................................................................... 37
2.6 REFERENCES .................................................................................................................. 38
3.1 INTRODUCTION ............................................................................................................. 46
3.2 MATERIALS AND METHODS .......................................................................................... 48
3.2.1 Hydraulic Conductivity ........................................................................................... 48
3.2.1.1 Experimental set-up .............................................................................................................................. 48
3.2.1.2 Testing protocol ..................................................................................................................................... 49
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3.2.1.3 Calculations ............................................................................................................................................. 54
3.2.2 Characterisation of Media Materials ..................................................................... 56
3.2.2.1 X-ray diffraction (XRD) ......................................................................................................................... 56
3.2.2.2 Particle size distribution ................................................................................................................... 56
3.2.2.3 Bulk density ............................................................................................................................................. 57
3.3 RESULTS AND DISCUSSION ........................................................................................... 59
3.3.1 Zeolite and Laterite Ore Characterisation ............................................................. 59
3.3.1.1 Particle size distribution ...................................................................................................................... 59
3.3.1.2 Dry bulk density and apparent specific gravity ........................................................................... 61
3.3.2 Hydraulic Conductivity ........................................................................................... 62
3.3.2.1 Media density ......................................................................................................................................... 67
3.3.2.2 Media particle size ................................................................................................................................ 68
3.3.2.3 Bed height ............................................................................................................................................... 69
3.4 CONCLUSIONS ............................................................................................................... 72
3.5. ACKNOWLEDGEMENTS ................................................................................................. 73
3.6 REFERENCES .................................................................................................................. 74
SUPPLEMENTARY INFORMATION ............................................................................................ 78
4.1 INTRODUCTION ............................................................................................................. 81
4.2 MATERIALS AND METHODS .......................................................................................... 84
4.2.1 Materials ................................................................................................................ 84
4.2.1.1 Zeolite ........................................................................................................................................................ 84
4.2.1.2 Resin .......................................................................................................................................................... 84
4.2.1.3 Test Solutions ......................................................................................................................................... 84
4.2.2 Equilibrium Exchange Isotherms ........................................................................... 85
4.2.3 Column Trials ......................................................................................................... 86
4.2.4 Analysis .................................................................................................................. 86
4.3 RESULTS AND DISCUSSION ........................................................................................... 88
4.3.1 Ammonium exchange equilibria from NH4Cl solutions – Natural Zeolite ............. 88
4.3.1.1 “As Received” Natural Zeolite ............................................................................................................ 88
4.3.1.2 Sodium Natural Zeolite........................................................................................................................ 91
4.3.1.3 Acid Pre-Treated Natural Zeolite ...................................................................................................... 93
4.3.2 Ammonium exchange equilibria from NH4Cl solutions – Resin .......................... 95
4.3.2.1 H+ - Resin ................................................................................................................................................. 95
4.3.2.2 Na+ - Resin................................................................................................................................................. 96
4.3.3 Landfill Leachate Equilibria - Pre-RO (field sample) .............................................. 98
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4.3.3.1 “As Received” Natural Zeolite ........................................................................................................ 98
4.3.3.2 Sodium Exchanged Natural Zeolite ............................................................................................... 101
4.3.3.3 Sodium Exchanged SAC Resin ......................................................................................................... 104
4.3.4 Column Trials ....................................................................................................... 105
4.3.4.1 Sodium Modified Natural Zeolite ................................................................................................ 106
4.3.4.2 Sodium Exchanged Strong Acid Cation Resin .................................................. 107
4.4 CONCLUSIONS ............................................................................................................. 110
4.5. ACKNOWLEDGMENTS ................................................................................................. 112
4.6. REFERENCES ................................................................................................................ 112
5.1 INTRODUCTION ........................................................................................................... 119
5.2 Materials and Methods ............................................................................................... 122
5.2.1 Natural Zeolite ..................................................................................................... 122
5.2.2 Core-Shell Zeolite Synthesis ................................................................................ 122
5.2.3 Zeolite Scale up .................................................................................................... 124
5.2.4 Process Optimization ........................................................................................... 125
5.2.5 Core Shell Zeolite Performance ........................................................................... 126
5.2.6 Analysis ............................................................................................................ 127
5.2.6.1 Inductively Coupled Plasma – Optical Emission Spectroscopy (ICP-OES) ....................... 127
5.2.6.2 X-ray Diffraction ........................................................................................................................... 127
5.2.6.3 Cation Exchange Capacity ......................................................................................................... 127
5.2.6.4 Particle Size Distribution ........................................................................................................... 127
5.2.6.5 Optical Microscopy .......................................................................................................................... 128
5.2.6.6 Laser Ablation Inductively Coupled Mass Spectroscopy (LA-ICP-MS) ............................. 128
5.2.6.7 Scanning Electron Microscopy (SEM) ......................................................................................... 128
5.2.6.8 Surface Area Analysis ................................................................................................................. 128
5.3 Results and Discussion ................................................................................................ 129
5.3.1 Core-shell Zeolite Production .............................................................................. 129
5.3.1.1 Impact of Synthesis Temperature .......................................................................................... 133
5.3.1.2 Agitation ......................................................................................................................................... 135
5.3.1.3 Process Optimization .................................................................................................................. 139
5.3.2 Modified Zeolite Performance ............................................................................ 143
5.3.2.1 Ammonium Chloride solution – 250 ppm NH4+ ...................................................................... 143
5.3.2.2 Pre-RO landfill leachate – 50 ppm NH4+ ............................................................................... 144
5.4 CONCLUSIONS ............................................................................................................. 145
5.5 ACKNOWLEDGEMENTS ............................................................................................... 146
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5.6 REFERENCES ................................................................................................................ 146
6.1 Conclusions ................................................................................................................. 151
6.3 Future research and recommendations ..................................................................... 152
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Statement of Original Authorship
The work contained in this thesis has not been previously submitted to meet
requirements for an award at this or any other higher education institution. To the best of
my knowledge and belief, the thesis contains no material previously published or written by
another person except where due reference is made.
Signature:
Date:
QUT Verified Signature
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Acknowledgements
I would like to thank some people for their direct or indirect participation in the
achievement of this research and the presentation of the work:
My supervisors, Dr. Sara Couperthwaite and Professor Graeme Millar, for their mentoring
and guidance and for their immense support throughout this time, from which I emerge as a
better person and a more experienced and capable professional. Thank you.
Mr. Gregory Stephen and Zeolite Australia Pty. Ltd. for the opportunity of tapping into so
many contemporary environmental problems and hopefully making a contribution to
advancing their solutions.
Queensland University of Technology and its staff for all the infrastructure, student support,
and opportunities.
My colleagues for their always available advice, sharing and listening; in particular Vishakya,
John, Amy, Kenny, and Mitch for all the great hands-on help.
My parents for being my role models, keeping me on a progressing path and always
believing in me.
My husband, for the endless support and infinite patience.
Thank you all, this realization is ours.
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Chapter 1: Introduction
The overall goal of this research program was to advance the understanding of natural
zeolite applications for environmental protection and remediation, as well as to examine
novel means of modifying the zeolite structure and determine the feasibility of scaling up
production of this new material. More specifically, the use of natural zeolites for
stormwater filtration and landfill leachate treatment was evaluated.
1.1 Background
Zeolites have been successfully used as sorbents, ion exchangers, and catalysts for many
decades since their discovery in the mid 1700’s. These materials are crystalline, highly
porous, aluminosilicates with remarkable ion exchange and sorption capacities. The various
crystalline combinations of the zeolite’s building units generate a great number of distinct
framework types. Reversible ion exchange and dehydration are the main mechanisms
behind the applications of these materials. In addition, natural zeolites are inexpensive and
readily available, as zeolite-rich rock deposits are widespread throughout the globe. The
properties and characteristics of zeolitic tuffs vary with their origin, including structure type
and zeolite content in the mineral.
1.2 Research Need
Environmental problems and pollution mitigation are of central importance to our quality of
life. There is a continuing demand to improve technologies to meet ever stricter
environmental regulations. Specifically of interest in this project were the cases of
stormwater runoff filters and landfill leachate treatment.
Stormwater runoff is usually contaminated with heavy metals, oils and grease,
hydrocarbons, nutrients and suspended solids. Filters placed at strategic locations, like
parking lots, provide a preliminary treatment before the run-off enters the drainage system
or is directed to ground infiltration areas. Generally, litter, suspended solids and oils and
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grease are held in these filters. There is a need to provide up to tertiary treatment using
layered stormwater filters. The most advanced step is intended to remove ammonium and
heavy metals by means of a media such as natural zeolite. Consequently, there is a
requirement to design zeolite beds which are efficient for these contaminants including the
critical aspect of hydraulic properties.
In the case of landfill leachate, the challenge is the removal of ammonia from this
wastewater to the extent required by discharge limits imposed to the landfill. This leachate
originates from liquids that permeate the waste in a landfill, carrying along suspended and
dissolved contaminants and becoming more concentrated as it flows through the layers of
waste. At the bottom of the landfill, a collection system directs the leachate to treatment,
which usually consists of a combination of techniques (or a treatment train). Even with
systems using reverse osmosis, the ammoniacal nitrogen content is still an issue with this
effluent, especially considering that the discharge limits for this contaminant are relatively
low (<1 mg/L).
The fundamental need of this project is the development of new materials. There is a
growing demand for media such as zeolites with enhanced efficiency and innovative
configuration. Critical aspects include: increased ammonium capacity; enhanced
ammonium selectivity; structural strength; reduced diffusion limitations; ease of synthesis;
and cost of manufacture.
1.3 Research Goals
Zeolite materials are normally used in filter columns when purifying water and wastewater.
One of the most important parameters influencing the filter design is the hydraulic
conductivity of the filter media. The hydraulic conductivity parameter relates to the rate at
which water will flow through the media and it is influenced by a range of parameters such
as type of media and particle size. Applications such as filters for stormwater run-off
require a precise knowledge of the hydraulic conductivity value for the materials used in the
filter. Surprisingly, current standards for measuring the hydraulic conductivity of materials
relate mainly to soil science and not to natural zeolites. Thus, the determination of the
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hydraulic conductivity of various zeolite media, and elucidation of the key factors which
influence the measurement, was the first research goal for this project.
Ammoniacal nitrogen removal remains a challenge for many industries, such as the example
of landfill leachates. The current goal is to create effective materials which can remove and
recover ammonia from solution, thus allowing for the possibility of reusing the nitrogen as a
fertilizer. Therefore, media such as zeolites and resins which can exchange ammonium ions
are of interest. However, we require to understand how efficient is the ion exchange
treatment of for example landfill leachate, using not only natural zeolites and resins but also
modified zeolites which are designed to improve performance.
As outlined above, the development of new zeolite forms is a critical path for this research.
As such, we need to be able to characterize modified zeolite materials and determine if they
can be scaled up in production. Consequently, a goal of this research was to examine core-
shell zeolite materials wherein the core was natural zeolite and the shell was Zeolite N
which is known to be very selective to ammonium species.
1.4 Significance and Innovation
Modified zeolites based on a natural zeolite core with a synthetic zeolite shell have not been
tested previously for ammonium removal and recovery from solution. Natural zeolites are
abundant and comparatively inexpensive. These materials exhibit some capacity for
ammonium species and good selectivity. However, natural zeolites suffer from problems of
very slow exchange kinetics (typically up to 72 hours to reach equilibrium) and cation
capacity values of <120 meq/100 g. In contrast, zeolite N has been demonstrated to possess
very high capacity for ammonium ions (500 meq/100 g) and excellent selectivity in the
presence of common competing cations such as calcium and magnesium. However, zeolite
N is more expensive than natural zeolite and requires binding into pellets which are
acceptable for practical use. Zeolite N also has an inherently high pH in solution due to the
use of 1:1 Si:Al ratios in the zeolite framework. The innovation of making a core-shell zeolite
structure potentially has the benefits of: reducing the amount of zeolite N present;
significantly lessening the diffusion pathlength as zeolite N is nano-crystalline and only
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present at the edge of the natural zeolite particle; may not require binder as natural zeolite
grains are already of the correct size and strength; could potentially have lower pH in
solution due to a change in the Si:Al ratio.
The significance of the hydraulic conductivity parameter and its influencing factors relates to
a gap in the literature regarding zeolite application. Hydraulic conductivity is a critical
aspect in the design of filters and permeable barriers, and the study of this parameter has
until now primarily been dedicated to soils or sand filters. This study looked at innovative
stormwater runoff filters which used layers of natural zeolite and other media from Zeolite
Australia, immediately before the zeolite layer. Significantly, this study intended to create
new testing protocol which could be applied to determination of hydraulic conductivity of
relatively coarse zeolite media.
Landfill leachate represents a significant environmental liability and as such a technical
solution is of great significance. In this study, the aim was to determine if either natural
zeolites or resins were applicable to landfill leachate, especially in relation to situations
where a reverse osmosis system was available to remove the main contaminants apart from
ammonium species. Additionally, the performance of modified zeolites of the core-shell
structure for leachate treatment was examined to determine whether it had potential for
use or required further development.
1.5 Thesis Outline
Each of the three main topics studied in this project, although interrelated, generated an
independent document with the goal of publishing the results at an appropriate time in a
quality journal. As such, each topic was made into one chapter of this thesis and each has
their own introduction, consisting of a literature review directly related to the study of that
topic. Figure 1.1 indicates the structure of this thesis.
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Figure 1.1: Mind Map of Thesis Structure
Chapter 2 is a brief literature review, which sets the context for the introduction of each
chapter and the connection of the topics. Chapter 2 describes zeolites as a material and
their uses, including: the issue of ammonia contamination, stormwater runoff, landfill
leachate and finally, zeolite modifications and manufacture.
Chapter 3 concerns the topic of stormwater runoff filtration, mainly relating to the
determination of hydraulic conductivity of filters using zeolite media. The configuration of
the filter is considered innovative and the zeolite particle size presented has not previously
been studied in this context.
Chapter 4 is focused on the treatment of landfill leachate by ion exchange, within the
context of a landfill facility which has the ability to pre-treat the leachate with reverse
osmosis if required to meet ammonia discharge limits. In this chapter natural zeolite,
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modified zeolite and commercial resins were tested and compared (i.e. natural zeolite as
received, sodium exchanged natural zeolite, acid exchanged natural zeolite, H+ SAC resin,
and sodium exchanged SAC resin).
Chapter 5 is dedicated to the development of an innovative modification of natural zeolite
and its production. A core shell zeolite product was synthesised with the aim of improving
ammonium selectivity and capacity, as well as minimizing diffusion pathlengths. This
chapter investigated the process of scaling up the production and the performance of the
modified zeolite for treatment of landfill leachate.
Lastly in Chapter 6, the research project was summarised as a whole and the conclusions
drawn from each topic connected under the overarching objective. Future works based on
the outcomes of this study are discussed, ending with recommendations and implications of
the findings for the future of environmental engineering and the zeolite industry.
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Chapter 2: Literature Review
2.1 WATER RESOURCES
With half the world’s population living in cities, urban water usage is substantial; in Australia
between the years of 2011 and 2012 urban water use was 1,530 GL [1]. According to the
United Nations water statistics [2], by 2050 global water withdrawals should increase by
50%. However, depletion of fresh water resources is not the only problem. Degradation of
water bodies is a major global issue; aquatic ecosystems are the most impacted, with 80% of
the world’s developing countries sewage being dumped in water bodies without treatment
[2]. In Australia, a study of approximately 30% of the country’s river basins, by Australian
Water Resources Assessment, showed a number of basins with reduced water quality, as
measured by the aquatic biota index [3]. A river was deemed impaired when 20 to 100% of
the aquatic invertebrate species that should be present were lost; a third of the river length
assessed was found to be within that spectrum. The worst case reported was in New South
Wales which was characterized by an impaired biota level in about 50% of its assessed
waters. The changes in water quality were associated with levels of turbidity (suspended
solids), salinity and nutrients (phosphorous and nitrogen).
2.2 AMMONIA
2.2.1 Effects of ammonia contamination
The issue of ammoniacal nitrogen contamination is a well-known challenge. If excessively
discharged to the aquatic environment, it causes eutrophication. The degradation of
nutrients is an oxygen consuming process, thus excessive nutrient loads promote enhanced
oxygen consumption and proliferation of specific plant species which feed on the nutrients.
Blue-green algae (or cyanobacteria) are one such species which can proliferate such that
they cover the surface of the water body, thus preventing the penetration of sunlight [4, 5].
That effect in itself is damaging to the aquatic diversity whose photosynthesis processes
become hampered. In addition, blue-green algae also produce toxins that are noxious to
fauna and flora in and out of the water body, e.g. fish and birds [5, 6]. The elevated nutrient
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load alone is also harmful to fish and human health [4, 6, 7]. In light of these issues,
regulations are strict for nitrogen ammonium content in discharge streams, with limits
below 1 mg/L commonly found [8]; discharge limits in the United States of America are less
than 0.02 mg/L [9], while in Europe discharge limits are less than 0.5 mg/L.
The terms ammonia, ammoniacal nitrogen, ammonia-nitrogen and ammonium are in many
cases used interchangeably. Although ammonium refers to the aqueous ionic form NH4+
and ammonia to the free gaseous form NH3, these species exist in equilibrium in solution,
therefore the terms, most of the time refer to the two species together [5]. The
concentration of one or the other is dependent upon temperature, salinity and pH, where
elevation of these parameters favours the ammonia side of the equilibrium [5].
2.2.2 Sources of ammonia in the environment
Some examples of wastewater with serious ammonium contamination are landfill leachates,
sewage effluent, industrial (e.g. oil refineries, pharmaceutical, paper) and agricultural
sources such as slaughterhouses and dairy farms [9].
2.2.2.1 Stormwater runoff
Stormwater runoff treatment is a relatively modern problem, essentially caused by
extensive urbanisation [10, 11]. The covering of the ground with construction and
pavements prevents the drainage of rainwater into the soil and holds pollutants which are
swept up by the rainwater, turning it into a polluted runoff [10, 12]. Stormwater quantities
and the concentration of pollutants are proportional to urbanisation and vary with activities
developed in a location, weather, climate and permeability of the soil covers [10, 11, 13].
Typically, stormwater runoff comprises heavy metals, hydrocarbons, suspended and
dissolved solids, nitrogen, phosphorous and microorganisms [14-16].
Stormwater contamination is now considered during urban wastewater management (or
water cycle) and urban planning and design (Water Sensitive Urban Design (WSUD)) [10, 13,
20
17]). A range of treatment options have been described by Melbourne Water [17] including
bio-retention basins, sand filters and constructed wetlands. Meanwhile Reddy et al. [18]
reported the use of emerging, physico-chemical methods such as electrodialysis, reverse
osmosis and ion exchange. Reddy et al. [18] highlighted that interest in the development of
more advanced filters for stormwater runoff derived from the challenge of using less space
and money to provide the treatment. Indeed, several studies considered alternative
materials for stormwater filters including tyre crumbs, sawdust, coconut fibre, wood chips
and the more commonly used activated carbon, zeolite, laterite and sands [19-21].
However, Bratières et al. [13] stated that most filtration systems need to improve their
performance, in particular nitrogen removal.
2.2.2.2 Landfill leachate
The disposal of urban solid waste which is not recyclable into landfills is currently accepted
as the best way to deal with the large amounts of waste produced by an increasing
population [22-24]. The degradation of waste produces liquid, which when associated with
rainfall, captures contaminants as it runs over the landfill before pooling at the lowest point
on the site [4]. The catchment area of landfill facilities follows a layered approach in which
the bottom layer is impervious, preventing the contamination of the soil by the leachate.
The other layers are arranged and designed to retain solids suspended in the leachate and
filtrate, along with its mild treatment of cations depending on the material used at the
leachate facility [Figure 2.1] [4]. The collected leachate is either pre-treated and directed to
a treatment plant or treated in the landfill facility to mitigate the contamination by toxic
compounds, like heavy metals and pathogens, before being discharged to the environment
[4, 22].
The composition of landfill leachate varies significantly from one landfill to another and as
such, a range of treatments are applied in each situation. According to Delkash et al. [24],
the main characteristics of landfill leachate are pH, chemical and biological oxygen demand
(COD and BOD), ammoniacal nitrogen, heavy metals and total suspended solids (TSS). The
concentrations vary widely, for example Davis and Cornwell [4] reported typically pH 6, and
concentrations of 18,000 mg/L COD, 10,000 mg/L BOD, 200 mg/L NH3-N and 500 mg/L TSS.
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Meanwhile, Abbas et al. [22] and Mukherjee et al. [7] reported for a medium age leachate,
pH 6.5-7.5, COD 3.0-15 g/L, BOD5/COD ratio 0.1-0.5, NH3-N 400 mg/L and heavy metals <2.0
mg/L. As a landfill ages, its leachate composition and characteristics change dramatically on
account of the biological degradation of the waste. A young landfill will have high chemical
and biological oxygen demand (COD and BOD) as it has more organic content, however,
over time the bacterial degradation of the waste increases the ammonia nitrogen content
altering the leachate characteristics [4, 22]. Halim et al. [25], point out that bacterial activity
is hampered by the reduced BOD/COD ratio and significantly elevated ammoniacal nitrogen
concentrations. In addition, increase of compounds which are not biodegradable (or
refractory), such as humic and fulvic acids, are also an inhibitor of bacterial efficiency [22].
After the first ten years the landfill is considered mature and is stabilised in its internal
processes. Overall, landfill leachates are subject to a high level of complexity since their
composition varies with waste characteristics (initial source and type of waste), landfill
location, local weather, landfill age and various other factors [7, 25].
Figure 2.1: Landfill layers scheme and leachate formation
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It is not unusual to find ammonium in leachates in the order of thousands of mg/L,
especially in mature landfills with stabilised leachates [23, 25]. Bashir et al. [26] have
recently published a study comprising six commonly employed technologies (biological, ion
exchange, coagulation-flocculation, carbon-zeolite mixture adsorption, advanced oxidation
processes (AOPs) and flotation) for landfill leachate treatment using leachate from the same
source. In this study, where the leachate complexity was circumvented, the main
conclusion was that no single technology could effectively treat landfill leachate to
discharge standards and a combination of treatments was required. The success of each
technique was based on the removal (%) of colour, COD and NH3-N, where all performed
well for one or two contaminants but not all. Ion exchange achieved good results, albeit
only when performed with a combination of cationic and anionic resins in sequence, which
had high efficiency for ammonia and for colour and COD, respectively. Still, even with the
best performing technique (resin ion exchange) the final concentration of ammonia (125
mg/L) was significantly in excess of the discharge limit (5 mg/L). Hence, the authors
concluded that a combination strategy, with multiple treatment stages would be the best
approach and suggested ion exchange as the final step, to remove the ammonia [26]. Other
studies have pointed in that same direction with ion exchange as a final step for ammonium
and ammonia removal in leachate treatment trains, even when they include stages as
advanced as reverse osmosis [22, 27]. Bashir et al. [26] also tested the ion exchange
treatment upfront, using raw leachate directly, and found that a cation-anion resin
sequence performed well but required a large amount of material, making costs too high.
2.3 TREATMENT
A range of treatment options for ammonia removal is available, as well as a range of
treatments applied to specific wastewaters such as the case of landfill leachate, discussed
previously. Coagulation-flocculation, for instance was tested for leachate treatment by
Bashir et al. [26] and Abbas et al. [22] and both studies found that it was reasonable for
removal of non-biodegradable organic matter, but ineffective for ammonia. Increases in
sludge production and in aluminium concentration (the main coagulant) in the liquid phase
were emphasised as the major drawbacks [22, 26].
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Another method cited was advanced oxidation processes (AOPs), in which ammonia and
organic-N are converted into nitrogen gas or nitrate, as follows [7]:
The chemical oxidation in AOPs, is achieved with a range of oxidants, which perform
differently for various contaminants in the leachate. Generally, oxidation is based on a
combination of oxidants such as ozone, persulphate and UV. Mukherjee et al. [7] state that
ozone is effective in removing ammonia however ozone alone is not efficient. Bashir et al.
[26] reported 94.4% removal of COD and 96.9% of colour using electro-Fenton (Fe2+/H2O2)
oxidation, while 76% removal of NH3-N was achieved using an ozone/persulfate
combination which also removed 96% colour and 72% COD. Other oxidants and oxidant
combinations used in AOPs include: persulfate (as Na2S2O8) [26]; ultraviolet (UV)/O3,
H2O2/UV, O3/H2O2 [22, 28], and photochemical iron mediated aeration (PIMA) [7]. AOPs
may require a significant amount of electricity for UV lights or ozonizers for instance, and
this is reported as a major drawback of this method [7, 28]. However, it is not the only
negative point associated with AOP: the advanced oxidation may not completely degrade
the contaminants unless a large dosage of oxidant is added, which would increase the costs
substantially [28]. Another issue for some oxidants is the formation of other pollutants as
intermediates during the reactions [22, 28].
Many other ammonia removal methods are known: precipitation as struvite, breakpoint
chlorination, chemical reduction of nitrate, chloramine removal by selected activated
carbon; the most relevant techniques for this study are discussed in the sections below.
2.3.1 Biological treatment
Traditionally, nutrients (i.e. phosphorous and nitrogen) are removed by a biological
treatment process whereby noxious forms of these elements, such as ammonium, are
converted into gaseous harmless forms (such as N2) by bacteria in a nitrification-
denitrification process [5, 8, 29]. Nitrification is an oxygen- intense process which results in
the biological transformation of ammonium/ammonia into nitrate. Denitrification in turn, is
24
an anaerobic process, which uses the oxygen from nitrates to biologically transform them
into nitrogen gas (N2). However, biological treatment drawbacks can be significant,
including slow performance, demand for large treatment area, high dissolved oxygen
requirements and weather vulnerability [9]. In addition, high loads (100’s of mg/L) of total
ammonia actually inhibit the nitrification process [9], a problem seen frequently in landfills
when leachate is recirculated into the waste [7].
2.3.2 Ion exchange
A range of ion exchange media have been employed in ammonia removal, such as zeolites
and resins [25]. According to Johnson et al. in 2014, [30] zeolites had a global market of 1.8
million tonnes per annum; the highest volume application for zeolite was use in detergents
as ion exchangers to remove Ca2+ and Mg2+ from water [31, 32]. However, zeolites as
catalysts represented the highest value application for processes such as fluid catalytic
cracking (FCC). A prominent area of use is in environmental remediation, in particular, heavy
metal and nutrient (nitrates and phosphates) removal due to the size of their cavities and
selectivity for metals (Cd2+, Cu2+, Ni2+, Zn2+, Fe3+, Pb2+, As3+) and ammonium [33]. Ion
exchange with zeolites is known for being environmentally friendly, cost effective
(compared to other sorbents) and resilient to temperature and weather changes [34, 35].
The ability of zeolite processes to sustainably remove and recover ammonia in the form of a
fertilizer is of particular interest.
Although ion exchange is quite effective and has many advantages, limitations have also
been identified: the flow to be treated by ion exchange most of the time requires pre-
treatment (otherwise, efficiency may be heavily impaired and clogging problems are likely);
the regeneration of the media may require high volumes of regenerant, which may incur
significant cost; adequate disposal or re-purpose of the spent regenerant (brines) carrying
high concentrations of nutrients is required [9, 36]. Evidently, these issues may be resolved
with the use of the brine (or the nutrient loaded zeolite) as a soil additive in the case of
ammonia removal, however, if used for removal of heavy metals or other contaminants,
other solutions need to be considered [35].
25
Zeolites have been the subject of substantial studies in relation to the treatment of
ammonium [25], with performance relating at least in part to initial ammonium
concentrations [37]. The natural zeolite clinoptilolite can be relatively selective for
ammonium (Cs+ > Rb+ > K+ > NH4+ > Ba2+ > Sr2+ > Na+ > Ca2+ > Fe3+ > Al3+ > Mg2+ > Li+) and is
typically used in wastewater treatment processes [38]. The selectivity sequence may be
slightly different for different deposits due to variation in composition. The selectivity
sequence has also often been presented in other formats, e.g. alkaline and alkaline earth
metals are presented on separate sequences [33, 35]. The consensus, however, is that the
natural zeolite clinoptilolite is highly selective for ammonia as most wastewaters do not
contain Cs+ and Rb+.
The ion exchange process between zeolite and NH4+ happens spontaneously [39], but it is
dependent on pH, initial concentration and time of contact between solution and zeolite
[33]. Ye et al. [23] have proposed the following reaction:
where represents the zeolite surface, M represents the exchangeable cations on the
zeolite surface, while n is the cation charge. There are numerous applications of zeolites in
water treatment, such as the treatment of industrial and domestic wastewaters,
contaminated leachates from agricultural fertilisers and landfills, and acid mining drainage
[24]. Zeolites are popular in wastewater treatment due to their pronounced advantages
over other ion exchangers, such as high selectivity, low cost, availability, thermal and
mechanical stability, moderate pH control, regeneration capacity and the fact that they do
not add any unwanted ions to the environment where they are employed [40, 41]. The
treatment of contaminated streams with zeolite will generally consist of a column-type filter
or a packed bed [33, 42]. In municipal wastewater, ammonia is one of the main
contaminants [35] and its removal is achieved using zeolites in cases where the biological
process (nitrification-denitrification) is not sufficient, which may be due to low BOD/N ratio,
low temperatures or the presence of inhibitors affecting the bacteria [35]. In addition, the
recovery of ammonia as a valuable product may drive the use of physico-chemical processes
alongside biological treatment [5]. The costs of biological nitrification-denitrification are
26
centred in aeration energy and pumping and the nitrogen cannot be reused, as it is released
to the atmosphere as dinitrogen gas (N2) [41]. The removal of nutrients by means of ion
exchange is classified as an advanced treatment within the conventional wastewater
treatment plant processes (typically primary, secondary and tertiary treatments)
Resin ion exchangers are polymers which have functional groups covalently attached to
their cross-linked matrix [43]. In fact, as Bashir et al. [44] elucidated, the material is a co-
polymer where each polymer has a specific function: the main structure of the chain
(styrene) and the cross linker (divinyl benzene) in the case of a strong acid cation (SAC)
resin; the functionality of acid cation resins is provided by the sulphonic group (-SO3-H+),
where (-SO3-) is fixed and H+ is the exchangeable ion (mobile). SAC resins are also often
used on their sodium form, whereby the resin is in contact with NaCl solution, usually
flushed through a resin column, until all mobile hydrogen cations are replaced by sodium
cations. Strong acid cation (SAC) resins have high capacity for ammonia uptake but do not
have the selectivity for ammonia that is characteristic of zeolites [41, 45]. Bashir et al. [46]
achieved good results when treating raw landfill leachate using a sequence of cationic and
anionic resins (removal of 96.8% colour, 87.9% COD and 93.8% NH3-N). However, for raw
leachate, this technique represented a high cost, as it required large quantities of the resins
and consequently of regenerant. Malovanyy et al. [41] tested zeolite and resins (strong and
weak acid cation) side-by-side to evaluate their efficiency in treating municipal wastewater
(simulated and real) and the treatment of the spent regenerant. The selectivity for
ammonium of the natural zeolite (clinoptilolite) proved to be superior to all other materials,
removing calcium and magnesium completely in addition to ammonium. In equilibrium
tests of a series of resins, Vignoli et al. [27] reported that a SAC resin had the best results
with an 80% removal of ammonium from evaporated landfill leachate (initially containing
approximately 950 mg/L NH4+) and subsequent recovery of 70% of that ammonium as
ammonium sulphate, after treating the spent regenerant with H2SO4. With regards to the
drawbacks of resins, their lesser selectivity may prevent or hamper the recovery of
ammonia from the spent regenerant, due to the presence of other contaminants (such as
heavy metals). As pointed by Malovanyy et al. [41], for instance, the presence of Ca2+ and
Mg2+ in the spent regenerant (for low selectivity materials) may be a problem in regenerant
treatment, since they will most likely precipitate as hydroxides or carbonates.
27
2.3.3 Air Stripping
Nitrogen removal by air stripping is based on the adjustment of the equilibrium between
ammonium and ammonia in the solution:
The principle is to remove nitrogen in the gaseous form by passing a flow of air in counter-
current to the contaminated water (sprayed), in a packed tower, thereby degasifying the
solution [5]. To achieve this, the pH of the wastewater is raised above 11, at which point
over 99% of the nitrogen is in the NH3 form [5]. Lime is commonly used to reach the latter
pH levels. The product of the spray tower is a contaminated gas, which then is treated with
H2SO4 or with HCl [22, 28]; nitrogen can be recovered as ammonium sulphate. Air stripping
was reported by Abbas et al. [22] and Renou et al. [28] in their reviews of landfill leachate
treatment techniques, as the process that is used most commonly for removal of ammonia
nitrogen. Although a popular option, air stripping has disadvantages, the most serious one
is the emission of ammonia to the atmosphere, in cases where the acid absorption stage is
not perfect [5, 22, 28]. Because of the high volume of lime employed to raise the pH, air
stripping towers often suffer from calcium carbonate scaling issues [4, 22, 28]; a decrease in
ammonia removal from 98 to 80 % was reported by Viotti and Gavasci [47] owing to scaling
of the packing material. According to Rumana Riffat [5], the method can be used
concomitantly for phosphorous precipitation with lime, in which case it is an economical
option for advanced wastewater treatment. In addition, Booker et al. [29] considered the
cost of air stripping excessive for the treatment of streams with low ammonia concentration
(<100 mg/L).
2.3.4 Ammonia adsorption on charcoal
In adsorption processes, the contaminants are removed from the water by being trapped on
the interface between solid (adsorbent) and fluid; it is a physical process which happens on
the surface of the sorbent, hence the use of carbon, a highly porous material after thermal
or oxygen activation [4]. Activated carbon adsorption is frequently used in wastewater
28
decontamination, either as a powder or granules. Powdered activated carbon (PAC) can be
added to the water directly, in the aeration tank, raw water or secondary effluent [4, 5].
Granular activated carbon (GAC) is used in contact columns or filters [5, 20]. In general,
activated carbon is used to remove odours and tastes, and is known for its capacity for
removing organic contaminants from liquid or gas phases [25]. Pawluk and Fronczyk [20]
employed granular activated carbon for the removal of heavy metals (Cd, Cu, Ni, Pb, Zn)
from aqueous solutions and found good results, although GAC was exhausted faster than
the other media tested (zeolite and silica spongolite). However, AC is not very effective for
ammonia since the surface of activated carbon is non-polar [25]. As remarked by Mojiri et
al. [43] the effectiveness of activated carbon is excellent for the removal of non-
biodegradable pollutants. Treatment of landfill leachate with activated carbon, zeolite and
a composite containing both materials by Mojiri et al. [43] found that AC performed poorly
for ammonia removal (not quite reaching 30%), while the composite performed nearly as
well as zeolite alone (both achieving over 70%). Composite, zeolite and activated carbon
each had an ammonia adsorption capacity (determined by isotherms) of 32.89, 17.45 and
6.08 mg/L, respectively [43]. In agreement with these results, a composite containing
zeolite, limestone, activated carbon and rice husk carbon (45, 15.31, 4.38 and 4.38 %,
respectively) showed 43.67 % removal of colour, 22.99 % removal of COD and 24.3 %
removal of ammonia in the study by Bashir et al. [26] of landfill leachate. The zeolite and
activated carbon alone did not perform well (ammonia removal of 17.45 and 6.08 %,
respectively) [26]. The lower performance of zeolite in relation to the composite was
probably due to competing cations in the solution which may be held back by the other
components of the composite.
Overall, it is clear from this discussion that treatments for ammonia removal are varied and
each shows advantages and disadvantages depending on context. Generally, nitrogen
recovery, cost-effectiveness and resilience can be appointed as advantages of physical or
physico-chemical processes over biological. The formation of by-products and other
pollutants is of higher concern regarding AOPs and air stripping, but may be extended to ion
exchange if one considers the spent regenerants management, albeit to a lesser scale. It was
also apparent from the discussion that charcoal adsorption may not be considered effective
29
at all unless combinations of media, such as charcoal and resin or zeolite, are employed; in
which case the primary ammonia removal mechanism turns back to ion exchange.
Drawbacks of ion exchange in turn, include the requirement of pre-treatment of the
contaminated stream, regenerants costs and lifecycle, and selectivity/efficiency
relationships for different media.
2.3.5 Operational considerations for filters
Many of the ammonia removal methods involve some form of filter or packed media.
Therefore, an understanding of physical parameters involved in the flow of fluids through
solid media is essential. The rate of permeation and flow through a filter will determine its
flow capacity, holding time and life cycle. As stated by Rumana Riffat [5], with activated
carbon columns, design parameters typically considered for this equipment include
hydraulic loading rate, the media particle size, bed depth and contact time. Lang et al. [48]
list media geometry, size distribution, effective size, media shape, bed porosity and ratio of
the diameters of the filter and the media effective size. Furthermore, the solution
characteristics and its interaction with the media may also impact the filter performance.
Kandra et al. [49, 50] used simulated stormwater (tap water with sediment from a
stormwater pond) to test the clogging effects of filters packed with zeolite, scoria and glass
beads. They found that the shape and smoothness of the media did not have a significant
effect on the clogging rate [49], while the filter depth affected the life span of the filters
whereby the deeper configurations took longer to clog. Shallower filters retained most
solids on the top layer, thus clogged faster; meanwhile a ponding depth was associated with
the better performance of higher beds by forcibly disturbing the surface layer [50]. With
zeolite filters, Kandra et al. [50] reported that using mixed particle sizes in the bed, either
packed in layers or mixed together, showed better sediment removal and longer operational
life than single sized media beds. Of all the parameters tested in these studies, the flow rate
through the media and particle size were identified as critical factors for design and
performance.
Evidently, the flow rate controls the contact time between fluid and media determining the
rate of removal of dissolved pollutants; similarly, the particle size determines permeability
30
through the particles and available surface area for pollutant removal, in addition to
efficiency of solids retention and clogging. The media’s mechanical resistance was also an
important factor to consider, as breakage of particles may alter the performance during
operation [49]. In another study, Hatt et al. [10] reported that a soil-based media actually
leached nutrients into the solution and although the pollutants in the inflow stream were
retained in the top 20 % of the filter, the discharge was loaded with nutrients (phosphorous
and nitrogen) and TOC (total organic carbon). Another problem observed in this study was
the compaction of the media under the flow pressure, which in turn reduced hydraulic
conductivity across the filter [10].
In the case of landfill leachate, hydraulics affects not only the filters at treatment stage, but
also the quality and volume of leachate produced. As the waste settles there is less space
for the liquid (either from rainfall or biodegradation) to flow through, i.e. the void ratio is
reduced [51]. The landfill liners performance also strongly depends upon the hydraulic
characteristics of the materials, whether they are meant to simply filter suspended solids or
to react with dissolved and colloidal matter in the leachate, i.e. permeable reactive barriers
(PRBs). Permeable reactive barriers incorporating zeolites are reported to be quite
effective, as liners and zeolite/bentonite mixtures as caps for the landfill (reducing the
rainwater infiltration through the waste) [24, 52]. Analogous to filters, to achieve better
barrier efficiency, mixtures of materials are often used. Delkash et al. [24] reported a higher
cation exchange capacity and lower hydraulic conductivity for zeolite/bentonite liner.
A combination of treatments is often employed when treating landfill leachate due to its
complexity; generally any single treatment will be effective for one range of contaminants
but perform poorly for another. An example is the case of natural zeolite and activated
carbon, the former with good efficiency for ammonia and poor for COD, and the latter with
the inverse capacity and limitation [53]. Delkash et al. [24] described in their review, studies
which incorporated zeolite and zeolite mixtures with other adsorbents (bentonite, activated
carbon) in landfill liners and caps. Advantages cited in doing so included the possibility of
having thinner liner layers, increased adsorption efficiency and an increase in removal of
organic matter and heavy metals.
31
2.4 ZEOLITES
The volume of research on the subject of zeolites is substantial and only the relevant
fundamental aspects will be discussed here. In 1973, Breck [54] reported in his book
(Zeolite Molecular Sieves) the existence of over 7,000 papers and 2,000 US patents on
zeolites between 1948 and 1972. More recently, in 2013, Margeta et al. [33] reported an
additional 2,000 papers and 410 patents on zeolites between 2003 and 2013 for
clinoptilolite alone. The marked interest in these materials is owed to their particular
crystalline construction and the numerous applications that derive from their properties and
characteristics (cation exchange capacity, selective ion exchange, reversible dehydration,
adsorption).
2.4.1 Crystal structure and classification
Zeolites are based on aluminosilicate (aluminium (Al), silicon (Si) and oxygen (O)) minerals,
containing H2O and metals of the alkaline and alkaline earth groups; their empirical formula
is M2/nO·Al2O3·ySiO2·wH2O, in which n = cation valence, y = 2 to infinity, and w = the water
molecules within the framework’s channels [30, 54]. The crystalline structure of zeolites is
based on AlO4 and SiO4 tetrahedra that connect to one another through oxygen bonds [39,
55]. The Al and Si tetrahedra are the primary building units (PBU) of the zeolite structure,
and the connections between these form the secondary building units (SBU) that then
repeat infinitely in a three-dimensional structure [Figures 2.2 and 2.3] [33, 56]. As displayed
in Figure 2.3, the aluminosilicate tetrahedra form rings that connect with one another
forming channels, cages or cavities. The negatively charged framework attracts positively
charged cations. These cations reside in the cages along with water molecules that are
loosely bound to the structure and are therefore mobile [23, 54].
The different combinations in which the rings and channels interconnect are responsible for
the vast number of zeolite frameworks that have been identified. Framework types (or
groups) are organised according to “corner-sharing network of tetrahedrally coordinated
atoms”, meaning that the frameworks formed by different combinations designate each
group, regardless of the extra-framework chemistry [57]. As such, the framework type
defines most characteristics of a given zeolite, it is the framework type which determines
32
Figure 2.2: Primary and secondary building units for natural zeolites
Figure 2.3: Crystal structure of Clinoptilolite-Na (|Na3K
3 (H
2O)
24| [Al
6Si
30O
72])
Channels: A, B and C (perpendicular to A and B)
the dimensions of the channels (and pores or cavities) and consequently the ion affinity.
The ion selectivity of an ion exchange process involving zeolite is based at least in part on
the size of the ions in the fluid, i.e. the zeolite will most likely retain those which are not too
33
large that they cannot enter the channels and pores, nor too small that they will flush
through the framework, hence the term “molecular sieve” defining zeolites [52]. Over 230
framework type codes (including 60 natural zeolites) can be found in the International
Zeolite Association (IZA) Database classified by a three-letter code, generated mostly from
the name of the material type (e.g. heulandite (HEU), edingtonite (EDI), faujasite (FAU), etc.)
[24, 57-59].
2.4.2 Exchange mechanism
The channel system in the zeolite’s structure allows free cations to be exchanged for cations
in a solution. The ion exchange process is reversible, i.e. the cations captured in the zeolite
structure can be released back into another solution when that solution contains an excess
of a more favourable cation. For example, clinoptilolite is commonly used for ammonium
removal and can be recovered after the zeolite is exhausted (loading is at capacity) by
flushing a highly concentrated sodium chloride solution through the media. Even though
NH4+ is preferred the high concentration of Na+ induces a concentration flux reaction,
meaning that the difference in concentration dislocates the equilibrium and the sodium
cations dislodge the ammonium cations from the framework taking their place. This is
known as regeneration, the sodium cations replace the ammonium in the zeolite and the
media can be used again; other regenerant used includes NaOH, HCl and H2SO4. The zeolite
does not necessarily lose its cation exchange capacity because of the regeneration process;
in fact it may be increased as reported by Ye et al. [23], who observed better ammonium
removal capacity on natural zeolite after three regeneration cycles (6.02 mg/g capacity on
natural zeolite and up to 6.44 mg/g on regenerated zeolite). This regeneration ability is an
important feature of zeolitic materials, which in addition to extending their life cycle allows
the recovery of cations of interest which would otherwise be eliminated from the system.
The spent NaCl regenerant is loaded with ammonium (ammonium chloride (NH4+Cl-)) and it
is possible to use this latter solution as a fertiliser. A number of methods are available to
recover the ammonium, including ammonia stripping (with previous pH elevation),
precipitation as struvite (NH4MgPO4·6H2O), electrolysis and biological conversion [41, 60].
34
2.4.3 Natural zeolite
Zeolites form in hydrothermal conditions found in nature and occur in geological deposits
such as volcanic and sedimentary rocks and tuffs [33, 61]. The origin of the zeolite-rich (or
zeolitic) rock influences the mineral composition of zeolite [62]. The discovery of zeolites
dates back to 1756, when Crönstedt first noticed the peculiar behaviour of the mineral
appearing to boil upon being heated; indicating a porous nature [61]. Due to their unique
porous properties, they are now used in a variety of applications that consume an estimated
2.5 to 3 million tonnes globally [63].
Clinoptilolite ((Ca0.5,Na,K, Sr0.5,Ba0.5,Mg0.5)6(H2O)20|[Al6 Si30 O72]) is one of the most common
natural zeolites and presents in various forms, namely: Clinoptilolite-K (potassium
dominated), Clinoptilolite-Na (sodium dominated) and Clinoptilolite-Ca (calcium dominated)
[58]. Na exchanged clinoptilolite has been reported to have the highest ammonia removal
capacity [33]. The crystal structure of clinoptilolite is classified as part of the heulandite
(HEU) group (Ca0.5,Sr0.5,Ba0.5,Mg0.5,Na,K)9(H2O)24|[Al9 Si27 O72]. Heulandite and clinoptilolite
are natural zeolites with the same crystal structure (monoclinic, space group C2/m),
however they have different water and cation distributions in the framework. According to
the International Zeolite Association (IZA) [58] there has been (and still is) some discussion
around the nomenclature of these minerals, since both have the same crystal structure and
morphology. The difference between the two is set on the silicon to aluminium ratio (Si/Al),
whereby a Si/Al smaller than 4 defines heulandite and Si/Al greater than 4, clinoptilolite
[64].
2.4.4 Synthetic zeolite
When compared to synthetic materials, natural zeolites do not possess the highest capacity
for ion uptake from solution. Resins and synthetic zeolites have higher capacity and
kinetics, albeit at times, at the cost of lower selectivity. Natural zeolites, as previously
described, are found in nature embedded in tuffs and rocks and as such, their use is
inherently impacted by other components of the rock, and their composition may vary even
within the same deposit [65]. Interest in synthesising zeolites goes as far back as the 1800s
when St Claire Deville allegedly produced levynite in laboratory in 1862 [66]. Nearly a
35
century later, in 1948, the first report of a zeolite created entirely from non-zeolitic
materials was made by Richard Barrer [61]. This achievement was a result of years of
pioneering work from Barrer and Robert Milton, who developed the hydrothermal synthesis
of zeolites which is used to date. Their work began in the late 1940s and by 1953 the
synthesis of 14 completely new zeolites had been reported by Milton et al. [66]. Research
on synthetic zeolites continues to grow and new materials are developed frequently.
Zeolites with framework cations other than aluminium and silicon have been discovered and
there are entire families of these materials, e.g. aluminophosphates (AlPOs), organic,
silicophosphates (SAPO-n), etc. [61, 66].
In relation to ammonium ion uptake, synthetic zeolites such as zeolite N (or zeolite K-F(Cl)),
can be 11 times more efficient than natural zeolites in the presence of competing cations [8,
67]. Zeolite N has a framework of the EDI type [Figure 2.4], with channels through which a
sphere with a maximum diameter of 3.20 Å (channels a and b) and 3.44 Å (channel c) can
diffuse; the maximum diameter that can be added to the framework is 5.72 Å [56].
Figure 2.4: [001] view of the EDI framework [56].
However, synthetic zeolites are relatively expensive; as remarked by Majano et al. [68], their
production is not only costly but resource and time intensive. Moreover, often there is
excessive liquor left after the synthesis which needs to be treated. Most synthetic zeolites
are formed under hydrothermal conditions [54]. Even though they are considerably more
36
expensive than natural zeolites, Perego et al. [35] reported a market of approximately 2
million ton/annum worldwide for synthetic zeolites.
The synthesis of zeolite N has greatly advanced since this compound was first identified by
Barrer et al. [67, 69]; it was described in detail by Christensen and Fjellvag [69] who made
zeolite crystals in a 170 h long process at a temperature of 300°C. More recently,
Mackinnon et al. [67, 70] discussed milder conditions for the synthesis of zeolite N, at
temperatures lower than 200°C and reaction times of only a few hours.
Cation exchange capacities for ammonium ions on the order of 500 meq/100g were
observed for pure zeolite N [70], while for natural zeolites a capacity of 120 meq/100g, as
noted by Wang and Peng [71] can be considered high.
2.4.5 Modification of natural zeolites
To achieve better capacity and efficiency with natural zeolites, several modifications have
been reported, such as microwave, exchange with solutions of specific cations or calcination
[38]. In the case of clinoptilolite, treating the mineral with specific solutions (brines, acids,
bases and surfactants) is known to enhance its efficiency for specific pollutants [38]. As
Margeta et al. [33] stated, the treatments modify the zeolite’s properties and characteristics
as a result of cation relocation within the structure; the cation migration impacts the
openings and pore sizes in the framework, altering the zeolite’s selectivity, exchange
capacities and other features. For instance, the modification of clinoptilolite with Fe(II) has
been reported by Lv et al. [72], to increase the material’s capacity to remove Cr(VI) while
maintaining the physical characteristics (particle size and hydraulic conductivity). Margeta
et al. [33] reported that clinoptilolite treated with NaCl shows an increase of 34 % in Pb2+
and 33 % in NH4+ removal, whereas when treated with NaCl + NaOH the increase in NH4
+
uptake increased to 45 %. Similarly, Cheng and Ding [34] compared their modifications of
natural zeolite by NaCl and by NaCl + calcination, and observed a 1.5 fold increase in
ammonium removal rate with both modifications. In addition to the improvement in
kinetics, Cheng and Ding [34] reported a 40 % higher capacity for ammonium adsorption
after the NaCl modification and 50 % after the NaCl + calcination process.
37
2.5 SUMMARY AND IMPLICATIONS
The problem of ammonium contamination in wastewaters has been studied at length and
yet a definitive solution has not been found. Sorption systems appear to have promise as
they can be operated sustainably. The concept of viewing wastewater as a resource for
nutrient recovery and not as a cost centre is one which represents the future of the
wastewater treatment industry. However, to achieve the latter outcome requires
development of improved materials which are not only selective to ammonium ions but also
of high capacity (>200 meq NH4/100 g) and regenerable.
Consequently, this research project had the main aim of developing improved zeolite
materials which exhibited enhanced ammonium ion capacity, acceptable hydraulic
properties and could be scaled up to practical amounts.
To achieve the outlined aim required completion of the following objectives:
Investigate the methodology to determine hydraulic properties of natural zeolites
o Experimental protocol which are relevant to zeolite samples
o New procedures for hydraulic conductivity measurements
o Understanding of important factors which influence results
Develop innovative means of modifying natural zeolites that enhance its ammonium
removal capacity
o Creation of core/shell zeolite types wherein a more selective zeolite for
ammonium ions comprises the shell structure
o Reduction of the diffusion path for ammonium ions which aids more rapid
uptake of ammonium species and reduces the volume of chemicals for
regeneration
Investigate modified zeolite synthesis methods
o Bench and floor scale zeolite manufacture
o Elucidation of variables such as reaction time, temperature and agitation
mode
38
2.6 REFERENCES
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4. Davis, M.L. and D.A. Cornwell, Introduction to environmental engineering. Vol. 4th.
2008, Dubuque, IA: McGraw-Hill Companies.
5. Riffat, R., Fundamentals of wastewater treatment and engineering. 2013, Boca
Raton, Fla: CRC Press/Taylor & Francis.
6. Jones, J., N.B. Chang, and M.P. Wanielista, Reliability analysis of nutrient removal
from stormwater runoff with green sorption media under varying influent conditions.
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7. Mukherjee, S., et al., Contemporary environmental issues of landfill leachate:
Assessment and remedies. Critical Reviews in Environmental Science and
Technology, 2015. 45(5): p. 472-590.
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44
Chapter 3: Hydraulic Conductivity of Coarse Media Filters for Stormwater Run-Off Applications
Hydraulic Conductivity of Coarse Media Filters for Stormwater Run-Off Applications
Marita Guarino Bertholini1, Sara J. Couperthwaite, Graeme J. Millar* and Ian D.R.
Mackinnon
Institute for Future Environments & 1School of Chemistry, Physics and
Mechanical Engineering, Science and Engineering Faculty, Queensland University
of Technology (QUT), Brisbane, Queensland 4000, Australia.
Zeolites and laterite ores are of interest for application in stormwater filters due to a
combination of their availability, effectiveness, and low cost. However, there exists a lack of
information regarding key performance parameters such as the hydraulic conductivity.
Tests were conducted upon a series of different size ranges of granular zeolite and laterite
ore in a column arrangement. It was found that Standard Test Methodology issued by
ASTM International were not representative of the system studied. Consequently, a
number of methods were procured. Two were deemed as sufficiently representative of the
field conditions. In relation to the hydraulic conductivity measurements, it was found that
two approaches have been previously used which differed in terms of the Δh reference
point. Hence, data was reported for both calculation methods. The results showed that
larger media grain sizes produced higher hydraulic conductivity values and that settling and
saturation periods impacted the results. The calculation of hydraulic conductivity by the
tailwater Δh reference point demonstrated insensitivity to bed height changes. Conversely,
when calculated by the media top Δh reference, the height of the media bed proved to have
a significant influence upon hydraulic conductivity values. The type of media employed was
found to have minimal influence on the hydraulic conductivity, wherein the much denser
laterite samples had the same hydraulic conductivity as their zeolite counterparts.
45
KEYWORDS: hydraulic conductivity, filter media, natural zeolite, laterite, stormwater
*Corresponding author:
Professor Graeme J. Millar
Science and Engineering Faculty, Queensland University of Technology, P Block, 7th
Floor, Room 706, Gardens Point Campus, Brisbane, Queensland 4000, Australia
ph (+61) 7 3138 2377 : email [email protected]
46
3.1 INTRODUCTION
Stormwater has historically been dealt with by infiltration into soil whereby it replenishes
groundwater resources [1]. However, with the continuous progress in urbanisation,
stormwater runoff carries pollutants at increasing concentrations and normally requires
treatment before being allowed to enter the ecosystem [2]. Another issue is that pavement
and construction significantly reduce the available soil area as well as the infiltration
capacity, either by covering it or excessively compacting the soil matrix [1]. Stormwater
management systems typically involve a filter which acts as a physical barrier to a liquid flow
removing pollutants or particulates [3].
Zeolites have been reported to be useful as a stormwater filter material due to a
combination of their availability, relatively low cost, performance, and robustness [1, 4, 5].
For example, Reddy et al. [6] examined several media filters including natural zeolites and
found that a range of heavy metal ions could be removed from solution. Doping the zeolite
media with copper ions has also been shown to allow bacteria reduction in stormwater
samples [7]. Sorption of species such as nitrate and phosphate was demonstrated with
zeolite materials for urban stormwater applications [8] as was the control of polycyclic
aromatic hydrocarbons (PAHs) such as napthalene and phenanthrenes [9]. Laterite ores are
similarly found in abundance and have known adsorption properties for a number of anionic
pollutants [10, 11]. Craig et al. [12] reported that laterite from Ghana could remove fluoride
ions from groundwater, albeit fine grain sizes were required to enhance the degree of
uptake. Maiti et al. [13] determined that modification of laterite ore with acid then base
substantially increased the capacity for arsenic species. Improvement in both surface area
and sample porosity were considered to have promoted the arsenic loading performance of
the laterite.
The removal rate of pollutants is dependent upon properties of the media such as affinity
for specific species, particle size, density, and hydraulic conductivity. Hydraulic conductivity
(k) is a property of aggregate materials and soils that can be defined as “the ease with which
water moves through an aquifer” [14]. The hydraulic conductivity of soils and streambeds is
an important parameter with respect to control of water permeation into the soil [15, 16].
This latter parameter is frequently considered when dealing with problems derived from
47
infiltration events, such as landfill leachate, agricultural fertiliser run-off, pesticides
contamination of groundwater and soil erosion [16]. Hydraulic conductivity values are
related to factors such as the grain size and saturation [14, 17]. Typically, larger particle size
promotes greater hydraulic conductivity values.
Stormwater filters are subject to a surprising level of complexity and often, studies have not
encompassed the entire situation. The impact of media particle size is one such example
where knowledge gaps are present, as the majority of hydraulic conductivity studies have
been related to soils and relatively fine materials (on the scale of µm) such as clays, silt and
loam. Kandra et al. [5] studied the clogging phenomena in stormwater filters specifically for
particle sizes between 1 and 5 mm, as previous literature focused on either relatively large
or very fine particles. Other factors inducing variability in hydraulic conductivity have been
identified, especially in soil testing. Deb and Shukla [18] pointed out that measurement
methods, sample support, and even decisions made by the tester can impact results. In
addition, due care has to be taken in relation to filter design conditions, including: (1) filter is
open to atmospheric pressure; (2) the flow is intermittent; (3) the media is fully saturated;
and (4) the tailwater pressure is constant (the water flows out of the filter and away,
without any backpressure). Moreover, pre-treatment of the inflow should remove oil and
grease, solids and particulates, and thus mainly dissolved pollutants and colloids should
reach the zeolite and laterite ore barriers, minimizing overloading and clogging issues in
these tertiary treatment layers [19].
Hydraulic conductivity is an essential design parameter in filters and yet there is minimal
literature regarding the influence of testing methods, especially for coarse grain media. The
bulk of literature and standard test methods is focused on the hydraulic conductivity of soils
and sand filters. Therefore, this study focussed on the determination of the applicability of
certified standard test methodology for coarser grained materials such as zeolites and
laterite ore. The primary aim was to determine the relevance of existing models to media
tests in a column environment which approximated practical situations.
48
3.2 MATERIALS AND METHODS
3.2.1 Hydraulic Conductivity
3.2.1.1 Experimental set-up
A column made of u-PVC with a diameter of 5.1 cm and total length of 1.0 m was used to
conduct the hydraulic conductivity tests. The bottom and top ends of the column were
equipped with a porous PTFE disc which prevented the media from escaping the column. A
glass volumetric cylinder was used to collect the outlet flow, which was directed from the
bottom of the column through a 0.95 cm PTFE tube. Between the porous end piece and the
tube, a fitting with the same diameter as the column and curved edges, smoothed the
transition from 5.1 to 0.95 cm, minimising any significant pressure variation [Figure 3.1].
The column was transparent and filled to at least half of its length with specimen, and to the
top with permeant liquid (tap water). This latter arrangement allowed visualisation and
measurement of the head loss above the sample at any point in time [20, 21]. Samples
were tested as-received from the supplier, as well as pre-treated by means of oven-drying
and sieving. Each column was filled with media (either zeolite or laterite ore) via a funnel to
until the required bed height was achieved. ASTM-D5084−10 [20] and ASTM-D5856-95 [21]
state a minimum specimen (bed) height and diameter (25 mm), but not a maximum limit.
The final bed height varied between 50 and 65 cm. To ensure homogeneity of particle size,
the media was thoroughly mixed prior to being sampled from different parts of the storage
container according to ASTM-C702/C702M−11 [22].
Figure 3.1: Column configuration for hydraulic conductivity tests
49
The columns were filled with tap water from the bottom entry port in order to saturate the
sample and remove as many air bubbles as possible. The water was pumped into the
column with a peristaltic pump (Masterflex II) at 80 – 100 L/h, which not only lifted the bed
to aid settling of the particles, but also removed air bubbles and fine particles. The water
flow through the column at this stage was continuous and the settling and air removal was
aided by tapping the column (vibration). The upward flow was maintained until most of the
air bubbles and fines were expelled from the column, generally taking between 10 and 20
minutes, depending on the sample. After this first loading, the water flow was stopped and
tapping of the column was used in the final settling stage, which ceased when air bubbles
were no longer released from the bed. Once the active settling was complete, the column
was left to settle and saturate overnight undisturbed.
In practical applications, the zeolite/laterite media filters are typically preceded by various
treatment stages which are designed specifically for retaining solids, grease and oils,
particulates and hydrocarbons. Therefore, tap water was considered representative of such
a pre-treated water sample and as such was selected as the permeant liquid for the testing
of zeolite and laterite. Tap water also conformed with the guidelines for permeant liquid
requirements described in ASTM-D5856-95 [21] and ASTM-D5084−10 [20]. The top of the
column was kept open and the flow downwards through the sample was due to the force of
gravity alone. The time elapsed during the flow from one point to the next (Δt) was
measured with a laboratory stopwatch with 0.01 s precision. The volume of water (V)
discharged during the test was measured with a 1 L glass measuring cylinder with a
precision of 10 ml. The head losses were measured using a tape measure with a precision of
0.01 m.
3.2.1.2 Testing protocol
According to ASTM-E2396−11 [23], measurements of the permeability of coarse materials
are impacted by the head conditions employed. That is to say that the pressure above the
filter media due to the water column (hydraulic head) during the test must be pre-defined.
Factors such as pressure, head loss measurement, field or laboratory applications and
testing systems have all been investigated [18, 24-26]. The permeability of granular
50
materials is measured with either a falling head or a constant head method, referring to the
conditions of the hydraulic head over the media [18, 24]. With respect to the constant head
method, the hydraulic head is kept constant; conversely on the falling head method, the
hydraulic head decreases as the permeant flows through the media. Thus, different
equations are employed to each test type to correct for the head loss measurements.
Darcy’s law is the basis to hydraulic conductivity and permeability calculations with many
derivations described which accommodate different systems and conditions [27, 28]. For
the constant head equation, the head loss features as an average value since the flow rate is
constant; the flow measurement is made based on the outflow of the system, for instance
by mass [20, 24]. Meanwhile, in the falling head equation (a derivation of Darcy’s law), head
loss appears as the ratio between initial and final head losses, in this case measured by the
difference in the height of liquid above the media bed [20, 24, 28]. Furthermore, the falling
head method is mostly applied to fine media (like silts and clays), with low hydraulic
conductivities [28, 29]; for instance, within ASTM International’s Test Methods, the falling
head technique is used for hydraulic conductivities lower than 10-4 m/s [30]. Generally, the
system for the falling head test consists of a permeameter cell (where the media is
contained) with a stand pipe connected directly above it (where the permeant is stored).
The measurement of the head loss is made by the height of the liquid on the stand pipe,
which normally is much smaller in diameter than the permeameter cell. For materials with
higher hydraulic conductivity, the falling head test can still be used but the higher flow rate
requires adjustment of the stand pipe size so that the measurement is possible [20, 28].
For the intended filter application, although the media is relatively coarse (up to 1.0 to 3.2
mm) and the hydraulic conductivity is expected to be in the order of 1 x 10-2 m/s, the
constant head test is not realistic, given that the filter is subject to intermittent flows.
Therefore, in order to achieve adequate measurements in this study, the stand pipe was
virtually scaled up to the same size as the permeameter. In practice, the stand pipe was
eliminated by the use of a long column, whereby media cell and permeant liquid pipe were
one and the same. Similarly, other methods for hydraulic conductivity measurement are
available to cover other scenarios and situations such as the constant volume and the
constant rate of flow [30]. The falling head permeameter is a common method and
variations for it are not unusual, such as the case reported by Johnson et al. [31] who
51
developed an automated falling head test using pressure transducers to measure hydraulic
conductivity, or Wilson et al. [24] who employed infrared emitters to measure the flow rate.
Within the falling head method, there are other variations to consider such as the
permeameter or system type and the pressure on the system outlet, namely the tailwater
elevation. Different equations apply to each tailwater condition. On ASTM’s methods
procured for this study, the tailwater is at either constant or rising elevation, that is to say
that the outlet flow is directed to a collection vessel, thus exerting backpressure on the
system outlet [20, 21, 32]. However, the methods allow for adaptations such as reported in
an automated system by Johnson et al. [31] where the outlet flow was discarded to a gutter
as it exited the testing system. In this latter study, the researchers applied a constant
tailwater pressure equation, since the variation in pressure at the outlet of the columns, if
any, was minimal.
Upon considering the test methods to employ, it was found that no standard methodology
applied directly to the stormwater filter cartridge of interest. International standard
methods (ASTM) provided several approaches for different ranges of k, materials and
pressure conditions. The filter system was designed to operate under specific criteria, which
the hydraulic conductivity test must satisfactorily reflect. Therefore, the following set of
necessary conditions defined the testing configuration and methodology: (1) predicted
hydraulic conductivity range; (2) intermittent flow operation; (3) filter cartridge open to
atmospheric pressure; (4) the media was fully saturated; and, (5) the tailwater pressure was
constant. The constant tailwater test with the outflow freely exiting the system, although
not particularly described by ASTM, is commonly found in publications on soil testing and
development of hydraulic conductivity permeameter systems [24, 28]. Given the open
character of the filter and the intermittent flow presented to it, a constant-head test was
possibly not truly representative of the real situation; whereas a falling head test appeared
more realistic. The expected hydraulic conductivity was actually a design value, under
which the filter was expected to perform in order to cope with the predicted runoff flows it
would be exposed to, which was is in the order of 1 x 10-2 m/s. Table 3.1 compares the
conditions of the filter and those covered by ASTM methods, and highlights the target test
conditions.
52
Table 3.1: Test methods for hydraulic conductivity (highlighted cells show conditions related
to the target)
Test
Method
Hydraulic
Conductivity
order (m/s)
Pressure
head
Tailwater
elevation Material type Testing System
Last updated
(year)
Target 1 x 10-2
Falling
head Constant
Porous
Saturated Open column n/a
D5084 ≤ 1 x 10-6
Falling
Head Constant
Saturated
Porous
Flexible-wall
permeameter 2010
D5856 ≤ 1 x 10-5
Falling
Head Constant Porous
Rigid-wall
compaction
permeameter
2015
D7100 ≤ 1 x 10-8
Falling
Head Constant Soils
Flexible-wall
permeameter 2011
D2434 > 1 x 10-5
Constant
Head n/a Granular soils
Method not
active
2006
Withdrawn 2015
D7664
From
saturated k to
10-11
Very low
flow
Very low
flow
Unsaturated
soils
Hydraulic
conductivity
function
2010
E2396 n/a
Low head,
Falling
head
n/a
100% ≥ 2.25
mm,
green roof
Nested
cylinders 2015
D6527 n/a Advective
flow
Advective
flow
Saturated
Porous Centrifuging 2000
D5567
≤ 5
Hydraulic
conductivity
ratio
n/a n/a
Saturated
porous, Soil-
geotextile
systems
Flexible-wall
permeameter 2011
The selected test methods for this study were D5084 and D5856, k ≤ 1 x 10-6 and ≤ 1 x 10-5
m/s respectively, which were the closest representation of the filter system design. Both of
these test methods used the same equation to determine k for Falling head and Constant
Tailwater Elevation conditions. The requirements and criteria for the successful application
of the two test methods were fulfilled in accord with best available practice, in such a
manner that the test methods were complementary to one another. It was noted that
hydraulic conductivity for granular soils was reported in relatively recent papers to be
53
evaluated using Test Method ASTM-D2434-06 [33] [5, 34]. However, this latter test method
was in fact withdrawn from ASTM’s database in January, 2015 and not superseded by
another method [30].
The entire test was carried out in a temperature-stable laboratory, at 20°C, therefore, no
temperature adjustment was required (kT = k20). The test consisted of allowing the flow of
water to initiate while simultaneously starting the stopwatch (Δt) and water collection (V)
once the water level in the column reached point A (Δh1) [Figure 3.2]. The collection of
water and the stopwatch were stopped at the same time when the water level reached
point B (Δh2). After each flow through to a certain point (e.g. point B, C or D), the water was
pumped into the column upwards with a peristaltic pump at 40 L/h to refill the column for
the next measurement. The controlled upward flow avoided excessive consolidation of the
media, which could impact on hydraulic conductivity [35]. Note that a 20L/h flow rate was
applied for samples with smaller particle sizes (such as the 0.5 - 1.2 mm), to avoid disturbing
the bed with a faster flow. Once the water reached the top four centimetres of the column,
the pump was switched off. This process was repeated for three different head-loss (Δh2)
values (points A–B, A–C and A–D); three times each day for three days providing replicates.
As soon as the water reached the end point (B, C or D), the bottom valve was closed,
keeping the water level above the specimen at all times. This latter procedure prevented
the introduction of air into the specimen bed and guaranteed the full saturation of the
media during the test, thus preventing errors in the readings caused by the changes in these
conditions. In addition, the distance between the top of the column and point A, where the
sample collection and running time started, allowed for the flow to be steady through the
bed before measuring. The initial flow also guaranteed the absence of air in the collection
tube [20].
A control run was performed to assess the hydraulic conductivity of the porous end piece of
the column by itself. The test was conducted in the exact same manner, with the same
head losses (A-B, A-C and A-D). The flow went through the column three times for each
head loss, with a total of nine measurements. The rate of flow of the system with no media,
under the same test conditions, should ideally be at least ten times greater than that
measured with the media [20, 21]. In the case where this latter condition is less than 10
54
times greater than the measured hydraulic conductivity then the obtained data can be
interpreted as being conservative due to impedance of the support plates.
Figure 3.2: Column configuration and representation of start and stop points
(pumps used to fill the columns prior to the test - idle during the measurements)
3.2.1.3 Calculations
Hydraulic conductivity was calculated according to ASTM-D5084−10 [20] and ASTM-D5856-
95 [21], with Equation 3.1 derived from Darcy’s law [28]:
Equation 3.1
Where: k = hydraulic conductivity (m/s); a = cross sectional area of the reservoir containing
the influent liquid (m2); L = length of specimen (bed height) (m); A = cross sectional area of
specimen (m2); Δt = interval of time over which the flow occurs (s); ln = natural logarithm
55
(base e=2.71828); Δh1 = head loss across the specimen at t1 (m of water); Δh2 = head loss
across the specimen at t2 (m of water).
The flow velocity, or flow rate, was calculated using time and volume data as shown in
Equation 3.2:
Equation 3.2
Where: Q = flow rate (L/h); V = volume (mL); Δt = interval of time over which the flow occurs
(s); 3.6 = conversion factor for Q in L/h
The measurement of Δh1 and Δh2 can be performed in two different ways, as pictured in
Figure 3.3. According to ASTM-D5084−10 [20] the head loss is defined as “the change in
total head of water across a given distance”. Many studies which present illustrations of
their permeameter schemes have marked the tailwater level as the reference for Δh
measurements, i.e. the end of the effluent line, as seen on the left hand side of Figure 3.3.
However, the top of the media bed as a reference point is also found in literature, such as is
the case of Wilson et al. [24]. Notwithstanding, the experimental procedure and the data
collected remained the same for both cases. Therefore in the present study both
expressions of Δh were used to calculate hydraulic conductivity and the differences and
implications were discussed.
Figure 3.3: Δh reference points: tailwater (left hand side) and media top (right hand side)
56
3.2.2 Characterisation of Media Materials
Natural zeolite (predominantly clinoptilolite) and laterite ore were both supplied by Zeolite
Australia, and tested in various grain-size ranges.
3.2.2.1 X-ray diffraction (XRD)
Diffraction patterns were collected using a PANalytical X'pert wide angle X-Ray
diffractometer, operating in step scan mode, with Co K radiation (1.7903 Å). Patterns
were collected in the range 5 to 90° 2 with a step size of 0.02° and a rate of 30s per step.
Samples were prepared as Vaseline thin films on silica wafers, which were then placed onto
aluminium sample holders. The XRD patterns were matched with ICSD reference patterns
using the software package HighScore Plus. The profile fitting option of the software used a
model that employed twelve intrinsic parameters to describe the profile, the instrumental
aberration, and wavelength dependent contributions to the profile.
3.2.2.2 Particle size distribution
For particle size distribution measurements, test method ASTM-C136/C136M-14 [36] was
followed. The filter media was supplied pre-sieved to within a specific size range. Some of
the media samples were dry but a fraction were moist and thus these materials were oven-
dried at 105 ± 5 °C overnight and cooled down prior to sieving. The analysis was carried out
using a mechanical sieve shaker, equipped with wire mesh sieves of size: 500 μm; 1.0 mm;
2.0 mm; 2.8 mm; and 4.0 mm. The sieves were weighed before each test (empty mass) and
re-weighed after the 20 minute shaking period. Each individual sieve was then hand sieved
to ensure that no more than 1 % of the retained material was passing [36]. The mass
retained on each sieve was determined using an analytical balance with a precision of ±
0.005 g as:
A minimum of one kilogram of each media was sieved [36] in portions that were within the
limit of covering the sieve surface. The percentage retained on each sieve size was
calculated from the combined masses of the portions sieved, as opposed to each portion, on
the basis of the total mass of the initial dry sample [36].
57
3.2.2.3 Bulk density
Bulk density was tested according to test method ASTM-D5057–10 [37], as well as the
apparent specific gravity (ASG) of the media. Each sample was tested on: (1) "loose" pour
conditions where the sample was gently poured into the receiving vessel without tapping or
shaking; and (2) “tapped” conditions, where the vessel was slightly tapped after pouring the
sample in order to encourage the material to settle and fill large voids. An analytical
balance was used to record the masses with a precision of ± 0.005 g. An aluminium straight
edge was used to level the materials in the weighing bottles. In addition, two different
vessels were used as weighing bottles in order to account for shape-influenced errors [98]:
plastic 1L Nalgene® bottle; and plastic 250 mL straight cylindrical container [Figure 3.4]. The
test consisted of weighing the mass of a container filled with water (R), the mass of the
same container filled with media (S), and the mass of the container filled with both media
and water (Q). In the last measurement the water is filling the voids present in the
container when it is filled with media, allowing the calculation of the real volume of sample
contained in the bottle.
Figure 3.4: Different neck shapes on containers with illustration of void space potential area
Every time the container was filled, it was filled to the point of overflow to ensure that it
was at maximum capacity each time. The excess material was levelled with the top of the
container with a straight edge, thus ensuring the repeatability of the test. The water was
also required to be levelled due to its surface tension producing an elevation of the water in
relation to the container.
Bulk density (BD) was calculated in g/mL as illustrated in Equation 3.3 [37]:
Equation 3.3
58
Where: W = weight of the empty container (with lid), R = weight of water-filled container
(with lid), S = weight of the sample-filled container (with lid), and Y = 1 g/mL, the conversion
of mass /volume at 4°C
The apparent specific gravity (ASG) is dimensionless; calculated as in Equation 3.4 [37]:
Equation 3.4
Where Q = weight of sample and water-filled container
ASTM-D5057–10 [37] recommends the use of bulk density (BD) for materials such as
granules and powders (identified as “Group B”) and ASG for materials such as gravel, paper
and wood (identified as “Group C”). These denominations and material types are somewhat
ambiguous in the case of the materials used in this study. Therefore, although BD is the
dimension in focus, ASG is calculated here as a confirmation of the values found for BD.
59
3.3 RESULTS AND DISCUSSION
3.3.1 Zeolite and Laterite Ore Characterisation
Natural zeolite and laterite ore proposed for the stormwater filter were characterised in
order to understand their physical properties prior to hydraulic conductivity testing. The
natural zeolite used in this study had been pre-treated by the supplier in order to replace a
fraction of the calcium ions which dominated the exchange sites of the “as mined” material
by sodium ions; this material is known as Na+ modified natural zeolite. Hydraulic
conductivity testing was performed on zeolite and laterite ore samples as received, dried
and re-sieved. Mineralogical composition analysis was performed on dried, micronized
zeolite and laterite ore samples.
Table 3.2 presents the phase composition of the Na+ modified zeolite and laterite ore,
determined by quantitative X-ray diffraction (XRD) using a 10% corundum internal standard.
Dominant mineralogical phases of the zeolite material were found to be clinoptilolite
(34.7%) and quartz (26.2%). A high amorphous content of 26.25% may be associated with
the micronizing method used prior to analysis. Previous work on zeolite characterisation
has shown the susceptibility of zeolites to become amorphous during crushing [99]. The
laterite ore was comprised primarily of a mixture of hematite (62.9%) and magnetite
(20.0%). X-ray fluorescence showed the clinoptilolite sample had 70.5% silicon (SiO2) and
12.15% aluminium (Al2O3). XRF also found the laterite ore contained 93.0% iron (reported
as Fe2O3) (Supplementary Information), indicating that a portion of the amorphous phase
detected in XRD was an iron mineral. BET surface areas of the zeolite and laterite ore
samples were found to be 12.8 ± 0.2 and 5.5 ± 0.1 m²/g, respectively.
3.3.1.1 Particle size distribution
Zeolite and laterite ore testing was performed using a variety of sizes, which are
represented in the particle size distribution curves in Figure 3.5. It was noted that the
zeolite samples exhibited size ranges consistent with the stated size fractions supplied;
however variations in finer and coarser particles within the size range can potentially
influence the hydraulic conductivity measurements. For example, for the two similar size
ranges, “1.0 - 3.0mm” and “1.2 - 3.0mm”, the size distributions within each range were
60
significantly different; “1.0 - 3.0mm” was comprised of larger particles well distributed (80%
passing at 1.4 mm), whereas “1.2 - 3.0mm” had a larger amount of smaller particles (80%
passing at 0.7 mm). The 80% passing (P80) of each size fraction can be found in Table 3.3.
The quantity of fines identified (< 1 mm) was typically about 5% of the zeolite sample, albeit
the 1 to 2 mm size fraction was characterised by a significantly larger quantity of fines (13.7
%).
Table 3.2: Quantitative X-ray diffraction of Na+ modified natural zeolite and laterite ore with
a 10% corundum internal standard
Zeolite Laterite ore
Phase Formula wt% Phase Formula wt%
Clinoptilolite [(Ca0.5,Na,K, Sr0.5,Ba0.5,Mg0.5)6(H2O)20]
[Al6 Si30 O72] 34.7 Hematite Fe2O3 62.9
Quartz SiO2 26.2 Magnetite Fe3O4 20.0
Sanidine K(AlSi3O8) 1.45 Goethite FeO(OH) 4.1
Albite NaAlSi3O8 6.1 Quartz SiO2 1.3
Mordenite |(Na2,Ca,K2)4(H2O)28|[Al8Si40O96] 5.3 Amorphous n/a 11.7
Amorphous n/a 26.25
Figure 3.5: Particle size distribution for zeolite and laterite samples
61
3.3.1.2 Dry bulk density and apparent specific gravity
The results of the bulk density measurements are presented in Table 3.3 where it is
indicated whether the sample received any pre-treatment, such as sieving or drying, prior to
analysis. Given that some of the samples were received moist and tested for hydraulic
conductivity as-received, one of the moist samples was also tested for bulk density in this
condition. Clinoptilolite 0.5 - 1.2 mm, showed a 0.08 and 0.21 difference between dry and
moist samples for bulk density and specific gravity, respectively. Moisture content in
material samples is known to be a source of errors in the measurement of both bulk density
and specific gravity [17, 37]. Thus, careful consideration of sample preparation methods is
required.
Table 3.3: Particle size, dry bulk density, and apparent specific gravity results
Sample P80 Pre-treatment Average bulk
density (mg/L)
Average apparent
specific gravity
Zeolite
1.2 to 3.0 mm
1.4 mm oven dried 1.12 2.24
Zeolite
1.0 - 2.8mm
1.4 mm Sieved (originally 1.0
- 3.0 mm)
1.06 2.20
Zeolite
1.2 - 2.2mm
0.7 mm as-received 1.16 2.22
Zeolite
1.0 - 2.0mm
0.7 mm sieved 1.17 2.25
Zeolite
0.5 - 1.2 mm
0.7 mm oven dried 1.03 2.08
Zeolite
0.5 - 1.2 mm
0.7 mm as-received (moist) 0.95 1.87
Average zeolite N/A 1.08 2.14
Laterite ore
1.0 - 3.2 mm
1.8 mm as-received 2.45 4.76
Laterite ore
1.0 - 2.8mm
1.8 mm sieved 2.40 4.74
Average laterite ore N/A 2.425 4.75
62
3.3.2 Hydraulic Conductivity
According to ASTM-D5084−10 [20], the presence of air in the pores of the materials
decreases hydraulic conductivity. Therefore, for accurate hydraulic conductivity results all
samples were consistently saturated prior to testing. As the flow rate for the column with
no media was only, at the most, double the value of columns containing media, the
reported results for media can be deemed conservative. The hydraulic conductivity results
are presented in Table 3.4 for zeolites and Table 3.5 for laterite, including flow rates and the
results calculated with different Δh reference points (the top of the media bed or the
tailwater level). Overall, the hydraulic conductivity results calculated with Δh at the
tailwater level were consistently lower than those using Δh at the media top, and the
former values show less variance than the latter. In fact, for most results with the tailwater
as a reference point, the differences between samples was only discernible on the fourth
decimal place; therefore the results were presented with four decimal places for both
methods to enable a direct comparison.
The results found with the media top as a reference point for Δh are comparable to those
found by Kandra et al. [39], where media in a particle size of 2 mm gave a hydraulic
conductivity of 0.0278 m/s (reported as 102 m/h), on a constant head test. It was important
to note that Wilson et al. [24] compared the values of hydraulic conductivity for the same
media when tested by falling and constant head and found very similar results (5.67x10-5
and 5.55 x 10-5 m/s, respectively). The tailwater reference point is seen more frequently in
literature, however results which are comparable to those obtained in this study using the
tailwater as reference were not found. However, Briaud et al. [29] described the hydraulic
conductivity of gravels to be in the order of 10-2 m/s, for laboratory tests, which is in fact
supportive of all results found in the present study. These authors remarked that the
hydraulic conductivity of gravels may actually be in the range of 10-2 and 10-4 m/s.
consequently, the results described were in agreement with those reported in other studies
for similar media grain sizes (which supported the methodology employed in this study).
Nonetheless, the experiments and results presented here are derived from a combination of
ASTM methods guidelines and peer reviewed academic literature [20, 21, 24, 31]; and as
such, the results also show some points that have not been extensively reported.
63
The results are discussed in terms of the parameters which were found to have an influence
upon them: (1) particle size, (2) density and (3) bed height. The media’s particle size and
density have shown the same effects on hydraulic conductivity results, regardless of the
reference point (for Δh) used for calculation of k. Hence, for the sake of simplicity, when
these parameters are discussed below, the first value presented is the one calculated with
the media as reference, followed by the value relating to the tailwater measurement in
square brackets, e.g. 0.0268 m/s [0.0092 m/s]. The last parameter discussed is the media
bed height (or filter depth). Varying the bed height has proven to have an effect on
hydraulic conductivity results and to be much more expressive for one Δh approach than the
other.
64
Table 3.4: Summary of hydraulic conductivity results for natural zeolite
Zeolite
Size tested (mm)
Original size (mm)
Test day Bed height
(cm) Bed height fluctuation
Flow rate (L/h)
k (m/s) for each Δh Pre-treatment
Media bed Tailwater
Control – no media Test 1
N/A
204.50 N/A N/A
N/A
N/A Test 2 N/A 205.44 N/A
Averages: 204.97 N/A N/A
0.5 - 1.2 0.5 - 1.2 Day 1 62.3 70.04 0.0307 0.0296 0.0289
0.0068
None Day 2 62.1 ± 0.04 68.27 0.0067
Day 3 61.9 67.37 0.0066
Averages: 68.56 0.0297 0.0067
1.0 - 2.0 1.0 - 2.0 Day 1 64.3 ± 3.13 105.79 0.0617 0.0436
0.0107
Sieved Day 2 61.8 101.30 0.0097
Averages: 103.55 0.0526 0.0102
1.2 - 2.2 1.2 - 2.2 Day 1 61.5 111.57 0.0462 0.0397 0.0382
0.01707
None Day 2 60.8 ± 0.36 101.47 0.0096
Day 3 60.3 100.31 0.0095
Averages: 104.45 0.043 0.0100
1.0 - 2.8 1.0 - 3.0 Day 1 63.8 77.44 0.0385* 0.0488 0.0478
0.0077*
Sieved Day 2 63.8 ± 0.00 97.83 0.0098
Day 3 63.8 96.42 0.0096
Averages: 90.56 0.047 0.0091
1.2 - 3.0 1.2 - 3.0 Day 1 52.6 112.34 0.0269 0.0268 0.0266
0.0091
None Day 2 52.2 ± 0.05 113.82 0.0092
Day 3 52.2 113.10 0.0092
Averages: 113.09 0.0268 0.0092
65
Table 3.5: Summary of hydraulic conductivity results for laterite ore
Laterite ore
ID Size tested
(mm) Original size
(mm) Test day
Bed height (cm)
Bed height Fluctuation
Flow rate (L/h)
k (m/s) for each Δh Pre-treatment
Media bed Tailwater
Control – no media Test 1 N/A N/A
204.50 N/A N/A
N/A
N/A Test 2 205.44 N/A
Averages: 204.97 N/A
A 1.0 - 2.8 1.0 - 3.2 Day 1 55.30
0.0
119.94 0.0338 0.0337 0.0333
0.0103
Sieved Day 2 55.30 118.74 0.0102
Day 3 55.30 117.89 0.0101
Averages: 118.86 0.0336 0.0102
B 1.0 - 3.2 1.0 - 3.2 Day 1 53.80
0.0
124.28 0.0323 0.0325 0.0329
0.0105
None Day 2 53.80 124.76 0.0105
Day 3 53.80 126.35 0.0107
Averages: 125.13 0.0326 0.0106
C 1.0 - 3.2 1.0 - 3.2 Day 1 54.80
0.0
120.45 0.0355 0.0361 0.0358
0.0105
None Day 2 54.80 121.81 0.0107
Day 3 54.80 121.58 0.0106
Averages: 121.28 0.0358 0.0106
D 1.0 - 3.2 1.0 - 3.2 Day 1 58.80
0.0
110.65 0.0342 0.0347 0.0344
0.0099
None Day 2 58.80 112.82 0.0101
Day 3 58.80 112.07 0.0100
Averages: 111.85 0.0344 0.0100
E 1.0 - 3.2 1.0 - 3.2 Day 1 67.30
0.0
100.67 0.0584 0.0579 0.0575
0.0103
None Day 2 67.30 99.65 0.0102
Day 3 67.30 99.58 0.0101
Averages: 99.97 0.0579 0.0102
F 1.0 - 3.2 < 1.0 - 3.2 Day 1 51.90
0.0
128.38 0.0302 0.0304 0.0305
0.0105
<1.0 out Day 2 51.90 129.93 0.0105
Day 3 51.90 129.34 0.0106
Averages: 129.22 0.0303 0.0105
66
G 1.0 - 3.2 < 1.0 - 3.2 Day 1 57.60
0.0
121.43 0.0349 0.0352 0.0335
0.0107
<1.0 out Day 2 57.60 123.69 0.0108
Day 3 57.60 120.93 0.0103
Averages: 122.02 0.0335 0.0106
H 1.0 - 3.2 < 1.0 - 3.2 Day 1 67.30
0.0
100.19 0.0574 0.0570 0.0581
0.0101
<1.0 out Day 2 67.30 99.53 0.0100
Day 3 67.30 101.25 0.0102
Averages: 100.32 0.0575 0.0101
67
3.3.2.1 Media density
The laterite ore was a much denser material than zeolite (2.425 compared to 1.08 mg/L),
which resulted in significantly less air bubbles being trapped between particles in the
column (visual inspection). Further evidence for this latter supposition was based on the
fact that no difference in bed height was observed after overnight settling and saturation. It
should also be noted that the laterite ore had a much smaller surface area than zeolite (5.5
compared to 12.8 m²/g) and thus less air bubbles would be trapped in the pores of the
laterite ore. The zeolite samples did show some variation in bed height after saturation due
to their higher porosity and lower bulk density (section 3.2.3). Visual inspection of the
column showed air bubbles trapped between zeolite particles (Supplementary Information).
It is proposed that the bulk density and surface area of a material did not have a direct
influence on hydraulic conductivity; however, the packing of denser materials in the column
and a reduction in porosity has been observed to reduce fluctuations in bed height and the
amount of air bubbles trapped in the column. This latter behaviour was clearly observed for
the laterite ore which showed no variation in bed height over a 3 day period [Table 3.5],
compared to the zeolite samples [Table 3.4]. A comparison of the hydraulic conductivity
values for laterite (1.0 - 3.2 mm) and zeolite (1.2 - 3.0 mm) with similar particle sizes and
bed heights showed minimal differences.
The shape, smoothness and packing of particles in the filter media have been discussed as
factors of influence on flow paths through the media and consequently on hydraulic
conductivity; as well as other parameters such as clogging and turbidity removal [5, 40, 41].
Meanwhile, porosity has been pointed out as a smaller contributor to the variability of
hydraulic conductivity by Deb and Shukla [18]. In fact, the influence of porosity in
permeability testing has been discussed and related to saturation levels by Briaud [29],
where the presence of air in the media was described as being as much of a barrier for the
water as solid particles would be, hence deeply affecting hydraulic conductivity values.
Therefore, the results observed in the present study were in agreement with other studies
and the saturation of the media proved to be satisfactory and essential, given the similar
results for the two similar experiments using zeolite and laterite.
68
Additionally, two test results in Table 4 make the point of variability in results due to
saturation, settling and disturbances on the media bed affecting the less dense material but
not the denser media. For the size range 1.0 – 2.0 mm, the first test day was also the day
the column was packed, i.e. no settling or saturation time allowed. The significant variance
in relation to the same media tested on the next day demonstrated the slow diffusion of air
bubbles through the media during extended saturation periods. Similarly, for particle size
1.0 – 2.8 mm, although the bed height reported was constant, the Day 1 result for flow rate
clearly indicated an issue: the bed lifted due to a pump set at too high a speed to fill the
column. By the end of Day 1, the bed was nearly 1.0 cm higher in relation to the start. This
datum clearly expressed the sensitivity of very porous media to disturbances on the bed,
particularly when involving an upward flow mode. The laterite’s high density ensured high
levels of stability in this sense.
3.3.2.2 Media particle size
Consistent with other studies [5, 35, 42, 43], larger grain sizes had higher hydraulic
conductivities. This outcome was due to void spaces between particles being larger,
providing more area/path for water to flow. Smaller particles are known to settle better
and pack more closely together, which decreases the ease with which water can flow
through the media [17, 24, 35]. It was noted that the study by Kandra et al. [43] evaluated
zeolite hydraulic conductivity using three different sample sizes (0.5, 2.0 and 5.0 mm) in
different bed arrangements, i.e. single-size, layered with different sizes and mixed-sizes
beds. These authors found that a mixed-sizes bed performed in the same manner as a
triple-layered filter (layers vertically arranged), showing less clogging (and consequent
increased life span) and better removal of sediment in relation to the single-size bed. The
media with varied size (the three sizes mixed together) compared to the size-range strategy
used in the present study, and it was recommended by Kandra et al. [43] over the layered
system for being less costly and more practical.
Due to the similar size of all particles in the size fraction, a relatively loose packing
arrangement in the column occurred, thus facilitating a greater amount of water to pass
through. The large size distribution of the “1.2 – 3.0 mm” allowed for the smaller particles
69
to pack tightly between the larger size particles. This tight packing arrangement reduced
the distance between particles and therefore results in a lower hydraulic conductivity
(0.0268 m/s [0.0092 m/s]). A similar tight packing arrangement occurred for a very narrow
size distribution (0.5 – 1.2 mm), which resulted in a smaller hydraulic conductivity (0.0297
m/s [0.0067 m/s]). Lee et al. [35] also noted the influence of varied particle sizes within the
filter bed on the measurements of hydraulic conductivity, whereby the media with more
diverse sizes showed lower and more variable k, compared to more homogeneous
distributions. A good illustration of this effect is found in the study by Park et al. [42], where
the media was a mixture of zeolite and sand, using three different zeolite particle sizes. The
largest zeolite size mixed with sand had low hydraulic conductivity, as opposed to the most
likely result of having high k for larger particles. This happened because with the larger
zeolite particles, larger void spaces were present between the particles and these spaces
filled with sand, which resulted in a significantly lower hydraulic conductivity for the
mixture.
3.3.2.3 Bed height
The issue of varying hydraulic conductivity values and testing methods have previously been
noted; including a technical note from ASTM’s journal which investigated parameters that
could influence hydraulic conductivity test results [26]. The assessment in this case was by
application of a ruggedness test, which is a method for testing the robustness of another
test, in this case hydraulic conductivity testing. In other words, the idea was to test the
effect of different parameters on a given result, when following the same test method. The
results of a ruggedness test can determine certain criteria for a given test by establishing
tolerances for specific parameters. Not all parameters considered of influence have the
same relevance, and many can be adapted if need be to suit various situations. In the case
of Peirce et al. [44], the hydraulic conductivity was tested using a falling head permeameter
with clay media and an acetone solution for permeant liquid. Their results suggested that a
testing method applied to the same specimen in different laboratories could produce very
different results, if parameters such as media water content and back-pressure varied from
one experiment to the next [26]. Notably, the bed height was kept constant for all tests.
This latter condition may be because the standard tests were carried out in a permeameter,
70
which is a closed cylinder containing the media with a filter depth that remained constant.
As a consequence of the permeameter’s design, the effect of bed height was rarely
observed in hydraulic conductivity testing.
The current standard ASTM-D5084−10 [20] states the minimum specimen height and
diameter as 25 mm, however a maximum limit is not stated. In fact, none of the standard
test methods specifies a bed height protocol, and thus a range of bed heights were used in
the present study to evaluate its effect. The effect of bed height focused on the laterite ore
as it was denser and less porous than the zeolite sample, which made for a more reliable
bed height as variations due to settling, air trapping or disturbances when refilling the
column were substantially eliminated. It was observed that samples consisting of the same
media exhibited significantly different hydraulic conductivity results when different bed
heights were employed. When calculated using the media top as a reference point for Δh,
the hydraulic conductivity consistently increased with higher bed heights. In contrast, when
calculated with the tailwater as a reference point, hydraulic conductivity values remained
the same with discreet variations within the 5 % accuracy range.
Table 3.6 shows the percentage variation on bed height, flow rate and hydraulic
conductivity for pairs of samples with identical media and pre-treatments (as per Table 3.5).
Between samples D and E for instance, an increase of 14.56% in bed height (from 58.80 to
67.30 cm) resulted in a 10.62% reduction on average flow rate and a substantial 68.31 %
increase in hydraulic conductivity.
Table 3.6: Variation in flow rate and hydraulic conductivity when varying bed height for pairs
of equal samples
Sample
pair
Size Pre
treatment
Bed
height
Flow rate k
(Δh media)
K
(Δh tailwater)
D and E 1.0 – 3.2 None + 14.56 % - 10.62 % + 68.31% + 2 %
F and G 1.0 – 3.2 < 1.0 out + 10.98 % - 5.57 % + 10.56% + 0.95 %
G and H 1.0 – 3.2 < 1.0 out + 16.84 % - 17.78 % + 71.64 % - 4.71%
F and H 1.0 – 3.2 < 1.0 out + 29.67 % - 22.36 % + 89.77 % - 3.81 %
71
Although there is a lack of literature on the effect of bed height in hydraulic conductivity,
there have been a number of studies on a filter’s performance for pollutant removal with
bed height [2, 3, 5]. These studies have shown that the higher the bed height the better the
contaminant removal [2, 3, 5]. This effect is largely due to the fact that having a higher bed
provides more extensive surface area for the media to capture the soluble contaminants,
while particulates will generally be retained in the top portion of the filter [3]. Similarly,
Kandra et al. [43] investigated the initial bed height (or filter depth) in relation to the
clogging of the system when removing sediment from stormwater. Hydraulic conductivity
(or infiltration rate) was monitored during column tests with zeolite media permeated by
synthetic stormwater. As the sediments were trapped in the media, hydraulic conductivity
decreased until eventually the flow through was too low and the filter needed maintenance,
i.e. until the end of the system’s lifespan. At first, the zeolite columns were saturated and
flushed with clean water to remove free dust from the media bed; clean water was also
used to determine the initial infiltration rate prior to the test with stormwater [43]. These
latter procedures directly relate to the manner in which the present study was conducted.
Interestingly, Kandra et al. [39] also found that deeper filters had higher initial hydraulic
conductivity than their shallow counterparts, whereby the zeolite media with a 2 mm size
gave a hydraulic conductivity of 0.03 m/s in a 50 cm deep filter; 0.03 m/s in a 30 cm deep
filter; and 0.02 m/s in a 10 cm deep filter. In the case of the present study, the difference in
hydraulic conductivity values were observed with much smaller changes in bed height; it is
proposed that this was due to the varied particle sizes in the media (1.0 to 3.2 mm)
compared to the Kandra et al. [5] study where all particles were 2 mm in size.
In the aforementioned study, the hydraulic conductivities were determined by the constant
head method ASTM D2434, which is currently inactive. Albeit the length of the specimen (L)
is part of the equation, the tailwater Δh method failure in responding to the bed height
changes, when the flow rate was clearly affected, implied an insensitiveness to design
changes. The majority of studies that have investigated filter media depth were focussed
upon pollutant removal performance, whereby the flow contained particulates and solids
which impacted hydraulic conductivity values and eventually clogged the media. However,
this study has shown that the bed height had an influence on the hydraulic conductivity in
the absence of particulates in solution. In a more recent work, Deb and Shukla [18] pointed
72
out that the hydraulic conductivity of saturated soils is subject to a high statistical variability.
The differences in results were related to measurement methods, sample holder and factors
like the number of people involved, field and lab measurements and even the decisions and
skill level of the investigators [18].
The study of bed height influence in the hydraulic conductivity of media is clearly of interest,
especially for those designing filters. The possibility of designing the equipment for the
desired hydraulic conductivity based on the height of the bed is important.
3.4 CONCLUSIONS
The main finding of this study was the gap in standard technical test methods for
determining hydraulic conductivity, particularly for coarse media. By adapting Standard
Test Methods (ASTM) to an open filter design, the determination of hydraulic conductivity
was possible, albeit relying on a combination of methods and literature sources. It was
found that the bed height exerted significant influence on the results, with a 68.31%
increase in hydraulic conductivity and 10.62% reduction on average flow rate observed after
a 14.56% increase in bed height for an untreated laterite ore sample (samples D and E).
Analogous behaviour was observed for other samples, and further investigation is required
in order to determine the underlying theory of the bed height effect.
Hydraulic conductivity testing by means of columns is frequently used and this study
provided an insight as to the variability in testing methods. In developing new methods, it
was found that different reference points were used to calculate hydraulic conductivity
under the falling head method and that application of different reference points had a
significant impact on the end results. Overall, the hydraulic conductivity was consistently
lower when calculated using Δh at the tailwater level compared to Δh at the media top
level; e.g. 0.0092 and 0.0268 m/s respectively (zeolite 1.2-3.0 mm). The figures, schemes
and definitions in ASTM methodology are singularly for permeameters, but the methods
may benefit from a statement which disambiguates the interpretation of the directions,
such as an expression of “total head = pressure head + elevation head”. This simple line
may avoid erroneous results and in fact expand the application of the standard to special
73
cases, such as the case of the columns used in this study. Specifically for filter cartridges, a
pre-design is recommended so as to allow the bed height to be adjusted accordingly (or
proportionally).
The media particle size and saturation level were demonstrated to have strong influence on
the results, corroborating the idea that hydraulic conductivity is highly sensible to a range of
parameters, which may interfere with one another as well. The bulk density difference
between zeolite and laterite (1.08 and 2.425 mg/L respectively) did not reflect in hydraulic
conductivity values, but did support the recommendation for thorough saturation of the
media prior to testing. A further study of each of the outlined parameters individually is
suggested for future work and the results may pave the way for a standard test method
applicable for open filters and column studies.
3.5. ACKNOWLEDGEMENTS
The financial support of the Energy and Process Engineering Discipline of the Science and
Engineering Faculty, Queensland University of Technology is gratefully acknowledged.
Zeolite Australia Pty Ltd supplied natural zeolite which originated from the Werris Creek
mine in NSW and magnetite from Chillagoe.
74
3.6 REFERENCES
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in: O.o.R.a. Development (Ed.), Environmental Protection Agency
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[2] B.E. Hatt, T.D. Fletcher, A. Deletic, Treatment performance of gravel filter media:
Implications for design and application of stormwater infiltration systems, Water Research,
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[3] K. Bratières, C. Schang, A. Deletić, D.T. McCarthy, Performance of enviss™ stormwater
filters: Results of a laboratory trial, Water Science and Technology, 66 (2012) 719-727.
[4] P. Misaelides, Application of natural zeolites in environmental remediation: A short
review, Microporous and Mesoporous Materials, 144 (2011) 15-18.
[5] H.S. Kandra, D. McCarthy, T.D. Fletcher, A. Deletic, Assessment of clogging phenomena in
granular filter media used for stormwater treatment, Journal of Hydrology, 512 (2014) 518-
527.
[6] K.R. Reddy, T. Xie, S. Dastgheibi, Removal of heavy metals from urban stormwater runoff
using different filter materials, Journal of Environmental Chemical Engineering, 2 (2014)
282-292.
[7] Y.L. Li, D.T. McCarthy, A. Deletic, Stable copper-zeolite filter media for bacteria removal
in stormwater, Journal of Hazardous Materials, 273 (2014) 222-230.
[8] K.R. Reddy, T. Xie, S. Dastgheibi, Adsorption of mixtures of nutrients and heavy metals in
simulated urban stormwater by different filter materials, Journal of Environmental Science
and Health - Part A Toxic/Hazardous Substances and Environmental Engineering, 49 (2014)
524-529.
[9] K.R. Reddy, T. Xie, S. Dastgheibi, PAHs removal from urban storm water runoff by
different filter materials, Journal of Hazardous, Toxic, and Radioactive Waste, 18 (2014).
[10] L. Zhang, W. Wu, J. Liu, Q. Zhou, J. Luo, J. Zhang, X. Wang, Removal of phosphate from
water using raw and activated laterite: Batch and column studies, Desalination and Water
Treatment, 52 (2014) 778-783.
[11] S.K. Maji, A. Pal, T. Pal, Arsenic removal from aqueous solutions by adsorption on
laterite soil, J. Environ. Sci. Health Part A Toxic Hazard. Subst. Environ. Eng., 42 (2007) 453-
462.
75
[12] L. Craig, L.L. Stillings, D.L. Decker, J.M. Thomas, Comparing activated alumina with
indigenous laterite and bauxite as potential sorbents for removing fluoride from drinking
water in Ghana, Applied Geochemistry, 56 (2015) 50-66.
[13] A. Maiti, B.K. Thakur, J.K. Basu, S. De, Comparison of treated laterite as arsenic
adsorbent from different locations and performance of best filter under field conditions,
Journal of Hazardous Materials, 262 (2013) 1176-1186.
[14] A. Porteous, Dictionary of environmental science and technology, in, Wiley, New York;
Chichester, England, 2008.
[15] J. Song, X. Chen, C. Cheng, D. Wang, S. Lackey, Z. Xu, Feasibility of grain-size analysis
methods for determination of vertical hydraulic conductivity of streambeds, Journal of
Hydrology, 375 (2009) 428-437.
[16] V. Bagarello, M. Iovino, D. Elrick, A Simplified Falling-Head Technique for Rapid
Determination of Field-Saturated Hydraulic Conductivity, Soil Science Society of America
Journal, 68 (2004) 66-73.
[17] A.H. Ören, T. Özdamar, Hydraulic conductivity of compacted zeolites, Waste
Management and Research, 31 (2013) 634-640.
[18] S.K. Deb, M.K. Shukla, Variability of hydraulic conductivity due to multiple factors,
American Journal of Environmental Sciences, 8 (2012) 489-502.
[19] Ecosol, Ecosol(TM) Technical Specifications for Cartridge Filter, in: E.W.F. Systems (Ed.)
www.ecosol.com.au, Australia, 2014.
[20] ASTM-D5084−10, Standard Test Methods for Measurement of Hydraulic Conductivity of
Saturated Porous Materials Using a Flexible Wall Permeameter, in, ASTM International,
2010.
[21] ASTM-D5856-95, Standard Test Method for Measurement of Hydraulic Conductivity of
Porous Material using a Rigid-Wall Compaction-Mold Permeameter, in, ASTM International,
2007.
[22] ASTM-C702/C702M−11, Standard Practice for Reducing Samples of Aggregate to
Testing Size, in, ASTM International, 2011.
[23] ASTM-E2396−11, Standard Test Method for Saturated Water Permeability of Granular
Drainage Media [Falling-Head Method] for Vegetative (Green) Roof Systems, in, ASTM
International, 2011.
76
[24] M.A. Wilson, W.D. Hoff, R.J.E. Brown, M.A. Carter, A falling head permeameter for the
measurement of the hydraulic conductivity of granular solids, Review of Scientific
Instruments, 71 (2000) 3942-3946.
[25] W.R. Wise, R.D. Myers, Modified falling head permeameter analyses of soils from two
south Florida wetlands, J. Am. Water Resour. Assoc., 38 (2002) 111-117.
[26] J.J. Peirce, G. Sallfors, E. Peterson, Parameter Sensitivity of Hydraulic Conductivity
Testing Procedure, Geotechnical Testing Journal, 10 (1987) 223-228.
[27] P.C. Carman, Fluid flow through granular beds, Chemical Engineering Research and
Design, 75, Supplement (1997) S32-S48.
[28] D.O.A. Johnson, F J;Lowery, Birl, Automation of a Falling Head Permeameter for Rapid
Determination of Hydraulic Conductivity of Multiple Samples
Soil Science Society of America Journal, 69 (2005) 828-833.
[29] J.-L. Briaud, Flow of Fluid and Gas Through Soils, in: Geotechnical Engineering, John
Wiley & Sons, Inc., 2013, pp. 370-400.
[30] ASTM-International, ASTM International, in, 2015.
[31] D.O. Johnson, F.J. Arriaga, B. Lowery, Automation of a falling head permeameter for
rapid determination of hydraulic conductivity of multiple samples, Soil Science Society of
America Journal, 69 (2005) 828-833.
[32] ASTM-D7100−11, Standard Test Method for Hydraulic Conductivity Compatibility
Testing of Soils with Aqueous Solutions, in, ASTM International, 2011.
[33] ASTM-D2434-06, Standard Test Method for Permeability of Granular Soils (Constant
Head) (Withdrawn 2015, no replacement), in, ASTM International, 2006.
[34] S. Le Coustumer, T.D. Fletcher, A. Deletic, S. Barraud, P. Poelsma, The influence of
design parameters on clogging of stormwater biofilters: A large-scale column study, Water
Research, 46 (2012) 6743-6752.
[35] S.H. Lee, H.Y. Jo, S.T. Yun, Y.J. Lee, Evaluation of factors affecting performance of a
zeolitic rock barrier to remove zinc from water, Journal of Hazardous Materials, 175 (2010)
224-234.
[36] ASTM-C136/C136M-14, Standard Test Method for Sieve Analysis of Fine and Coarse
Aggregates, in, ASTM International, 2014.
[37] ASTM-D5057–10, Standard Test Method for Screening Apparent Specific Gravity and
Bulk Density of Waste, in, ASTM International, 2010.
77
[38] ASTM-D5357-03, Standard Test Method for Determination of Relative Crystallinity of
Zeolite Sodium A by X-ray Diffraction, in, ASTM International, 2013.
[39] H.S. Kandra, A. Deletic, D. McCarthy, Impact of filter design variables on clogging in
stormwater filters, in: WSUD 2012 - 7th International Conference on Water Sensitive Urban
Design: Building the Water Sensitive Community, Final Program and Abstract Book, 2012.
[40] G.S. Beavers, E.M. Sparrow, D.E. Rodenz, Influence of bed size on the flow
characteristics and porosity of randomly packed beds of spheres, J Appl Mech Trans ASME,
40 Ser E (1973) 655-660.
[41] S. Suthaker, D.W. Smith, S.J. Stanley, Evaluation of filter media for upgrading existing
filter performance, Environmental Technology, 16 (1995) 625-643.
[42] J.B. Park, S.H. Lee, J.W. Lee, C.Y. Lee, Lab scale experiments for permeable reactive
barriers against contaminated groundwater with ammonium and heavy metals using
clinoptilolite (01-29B), Journal of Hazardous Materials, 95 (2002) 65-79.
[43] H.S. Kandra, A. Deletic, D. McCarthy, Assessment of Impact of Filter Design Variables on
Clogging in Stormwater Filters, Water Resources Management, 28 (2014) 1873-1885.
[44] J.J. Peirce, G. Sallfors, E. Peterson, Parameter sensitivity of hydraulic conductivity
testing procedure, Geotechnical Testing Journal, 10 (1987) 223-228.
78
SUPPLEMENTARY INFORMATION
Table S1: X-ray fluorescence of laterite ore
Element % Element %
Fe2O3 93.0 Na2O 0.11
SiO2 1.80 TiO2 0.07
MnO 0.57 P2O5 0.06
Al2O3 0.46 SO3 0.03
CaO 0.80 K2O 0.002
MgO 0.02 Loss on ignition (LOI)* 2.17
Total 99.15
* Loss on ignition refers to the amount of mass loss when the sample is heated to 950°C to
remove water and carbonates.
Figure S1: Air bubbles in zeolite packed column after overnight settling and saturation
79
Chapter 4: Comparison of natural and modified zeolites with SAC resins for ammonium removal from landfill leachates
Comparison of natural and modified zeolites with SAC resins for ammonium removal from
landfill leachates
Marita G. Bertholini, Sara J. Couperthwaite and Graeme J. Millar*
School of Chemistry, Physics & Mechanical Engineering, Science and Engineering
Faculty, Queensland University of Technology (QUT), Gardens Point Campus,
Brisbane, Queensland 4001, Australia.
Landfill leachate is often viewed as a problem in terms of safe disposal to the environment.
However, leachate also has potential for nutrient recovery as it contains significant
concentrations of ammoniacal nitrogen. Ion exchange using either zeolites or resins can
potentially sustainably recover ammonia species from leachate, however, insufficient
information is currently available regarding the selectivity and applicability of these
materials. This study focussed on low ammonium concentration solutions of simulated and
actual landfill leachate, and their equilibrium and column performance. Sodium forms of
zeolite media proved to be more effective for ammonium uptake (12.24 g NH4/kg zeolite)
than natural (9.18 g NH4/kg zeolite) and acid forms (6.12 g NH4/kg zeolite). Acid treatment
appeared to have dealuminated the zeolite framework. Multicomponent equilibria with
zeolites revealed complex relationships between sorbing and desorbing species and resins
did not exhibit significant uptake of ammonium ions in the presence of relatively large
concentrations of competing ions. Hence, post reverse osmosis treated leachate was
considered the best option for application of resin technology. Column trials showed that
natural zeolite performance was limited by very slow diffusion of ammonium ions in the
80
microporous channels. In contrast, strong acid cation resin was characterized by
considerably faster diffusion processes and could remove ammonium ions from solution to
low levels. However, the resin was not selective towards ammonium species and also
removed the other cations present in the post reverse osmosis treated leachate.
KEYWORDS: ammonium; natural zeolite; isotherm; resin; landfill leachate
*Corresponding author:
Professor Graeme J. Millar
Science and Engineering Faculty, Queensland University of Technology, P Block, 7th
Floor, Room 706, Gardens Point Campus, Brisbane, Queensland 4000, Australia
ph (+61) 7 3138 2377 : email [email protected]
81
4.1 INTRODUCTION
Landfilling of waste remains a common practice, but this is not without environmental
hazards despite careful practices to ensure responsible operation [1, 2]. For example, upon
events such as: rainfall; waste decomposition; surface water run-off; and, groundwater
ingression; liquid flows through the waste and becomes contaminated [3]. Landfill
leachates vary from one facility to the next, due to waste characteristics which are
dependent on population, landfill temperature, age, and the local climate [3, 4].
As such, a range of leachate treatment strategies have been proposed or implemented [3].
Conventional biological methods such as constructed wetlands, lagoons, activated sludge,
sequencing batch reactors, fluidized beds and trickling filters have all been contemplated for
landfill leachate treatment [1]. For more mature landfills, leachate may be better treated by
membrane bioreactors [5], albeit it is evident that biological processes have not only
advantages but also several drawbacks to their use. Sustainability is of importance in
relation to nutrient recovery from leachates, which provides a challenge to biological
processes which do not recover nutrients such as ammonia. Therefore, alternate
technologies based upon physical methods are of interest. Kumar and Pal [6] recently
assessed the possibility of recovering nitrogen and phosphorous nutrients from waste
resources such as landfill leachate in the form of struvite (MgNH4PO4·6H2O). Struvite was
reported to be suitable for agricultural applications and potentially economically attractive.
Activated carbon adsorbents have also received attention as they can demonstrate
effectiveness for control of dissolved organic carbon, ammonia and phosphate [7]. As
outlined by Kurniawan et al. [8] coagulant addition to landfill leachate can also promote the
removal of compounds responsible for chemical oxygen demand (COD).
Cotman and Gotvajn [9] concluded that a combination of biological and physical methods
was required for successful remediation of landfill leachates and that ion exchange may be a
key technology. Ion exchange using common media such as zeolites or synthetic resins
appears eminently suitable to landfill leachate treatment as ammonium species can be
recovered in the regeneration step. There have been many studies of ammonium ion
exchange with natural zeolite samples from simple solutions of ammonium salts [10]. In
terms of multi-component solutions, representative of landfill leachate, the number of
82
studies is considerably more limited. For example, Ye et al. [11] performed batch and
column studies of a leachate obtained from Wuhan, China with Chinese natural zeolite.
These authors found that the optimal solution pH was 7.1 and that diffusion constraints
limited the rate of ammonium exchange with the zeolite surface sites. Karadag et al. [12]
showed that the uptake of ammonium species on the zeolite was accompanied primarily by
the release of calcium and potassium ions from the surface exchange sites.
Synthetic resins have also been proposed to exhibit potential for ammonium exchange from
aqueous solutions. For example, Sica et al. [13] reported a kinetic evaluation of ammonium
exchange with Purolite C105H resin from ammonium chloride solutions and noted that high
removal efficiencies could be achieved under a range of pH conditions. Li et al. [14] tested a
variety of strong acid cation resins for the uptake of ammonium ions from solutions, which
simulated wastewater from a fertilizer plant. It was found that equilibrium occurred within
50 minutes and that isotherms could be modelled by the Freundlich equation. The identity
of the resin was discovered to be important with respect to attainment of low content of
ammonium species in the treated water. In contrast, Luo et al. [15] determined that the
Langmuir model best fitted ammonium exchange isotherms generated for resins treating
vanadium smelting wastewater. The application of strong acid cation resin in municipal
wastewater treatment plants for ammoniacal nitrogen removal has also been advocated by
Malovanyy et al. [16]. When used in conjunction with the Anammox process, almost
complete removal of ammonium species was observed. Copper loaded resin has also been
studied in an effort to create a resin system which is more selective for ammonia uptake
from wastewater solutions [17]. Typically, a chelating resin is complexed with copper
species, loaded with ammonia/ammonium species and then regenerated with an acidic
solution. However, this system has not been applied commercially and is still under
development. Mumford et al. [18] compared the performance of a synthetic resin with a
zeolite material to exchange either copper or ammonium species from solution. Zeolite was
observed to prefer ammonium ions whereas the chelating resin loaded copper ions in
preference to ammonium species. Wirthensohn et al. [19] also compared the ability of
resins with natural zeolite to remove ammonium ions from anaerobic digester effluent.
Notably, the solution examined was that produced after the effluent had been passed
through a reverse osmosis unit. The solution comprised of 85 mg/L potassium; 67 mg/L
83
sodium; 0.74 mg/L magnesium; and, 1.79 mg/L calcium ions in addition to 467 mg/L
ammonium ions. Resins were claimed to be acceptable in this instance in meeting low
emission targets for ammonium species (< 2 mg/L).
Bashir et al. [20] studied the fundamental behaviour of a cationic resin (INDION 225 Na) for
ammonia removal from landfill leachate and concluded that the equilibrium data was best
simulated using a Langmuir fit, and that the pseudo second order kinetic expression
optimally represented the exchange process. In practice, Bashir et al. [21] indicated that a
combination of cation and anion resins could be used to reduce the concentration of
ammonia, chemical oxygen demand (COD) and colour from landfill leachate. In terms of
solutions comprising of relatively low concentrations of ammonium species, Yoon et al. [22]
studied cationic resins in both the sodium and “proton” forms for drinking water treatment.
The H+-resin was suggested to be more effective at ammonium uptake and resin
performance was dependent upon solution temperature and ammonium concentration.
From the above discussion it was apparent that both resins and zeolites may have
application for landfill leachate treatment. Of particular interest were landfill leachates
which contain relatively low concentrations of ammonium (<100 mg/L) and especially those
which have been desalinated using reverse osmosis prior to the ion exchange stage [23, 24].
Questions relating to the possible advantage of zeolite selectivity to ammonium ions
compared to resins, and the increased diffusion limitation associated with zeolites [11] were
of importance. Inherently, this meant that the impact of competing ions in solution upon
ammonium loading had to be examined. As previous studies regarding landfill leachate
treatment using natural zeolites have been used as received from the supplier [12], it was
also pertinent to examine the use of zeolites wherein the exchange sites had been modified.
In addition, it was evident from inspection of previous literature that a detailed evaluation
of the nature and extent of desorbing species during the ammonium ion uptake process had
not been reported. Consequently, this study evaluated natural zeolites and strong acid
cation resin for ammonium removal from a variety of simulated and actual landfill leachate.
The equilibrium exchange and column behaviour were assessed.
84
4.2 MATERIALS AND METHODS
4.2.1 Materials
4.2.1.1 Zeolite
All zeolite samples were from the Werris Creek mine in New South Wales (NSW), as supplied
by Zeolite Australia. Both natural zeolite as mined from the ground and modified forms
were used. Natural zeolite was sieved by the supplier and the particle size was in the range
0.5 to 2.0 mm. Zeolite Australia also supplied modified natural zeolite (termed “sodium
zeolite or Na+ zeolite) which was produced by washing natural zeolite with a NaCl solution.
The particle size of this sample ranged from 1.2 to 2.2 mm. Finally, Zeolite Australia also
supplied an acid modified zeolite (termed “H+ zeolite”) that was made by contacting natural
zeolite with a dilute hydrochloric acid solution. The particle size of this sample was 1.0 to
2.0 mm.
4.2.1.2 Resin
Synthetic resin of the strong acid cation (SAC) type from Lanxess (Lanxess S108 SAC Resin)
was used in these studies. Two exchange forms of the resin were employed, namely the as
received acid form (“H+ resin”), and the sodium version (“Na+ resin”) which was obtained by
flushing the resin with a concentrated NaCl solution located in a uPVC column until the
effluent pH and conductivity indicated the exchange was complete.
4.2.1.3 Test Solutions
Three test solutions were used: (1) an NH4Cl solution with NH4+ concentration of 250 mg/L
that was prepared using analytical grade NH4Cl (Rowe Scientific) dissolved in ultrapure
water (milli-Q); (2) landfill leachate provided by an operator in Victoria, Australia. The last
stage of treatment at this landfill was reverse osmosis, hence samples were collected from
the pre-RO stream and the post-RO stream; (3) a multi-cation solution with NH4+
concentration of 15 mg/L was prepared by dissolving analytical salts (chloride forms) in
deionised (DI) water. The composition of these aforementioned solutions is shown in Table
4.1.
85
Table 4.1: Concentration of cations in the solutions tested
Cation
(mg/L)
Ammonium chloride
Pre-RO Landfill
Leachate
Simulated Post-RO
Leachate
Ammonium 240 54.3 13.63
Magnesium - 102.7 4.65
Calcium - 53.8 1.40
Potassium - 215 16.65
Sodium - 536.8 42.35
4.2.2 Equilibrium Exchange Isotherms
Equilibrium experiments were conducted using not only ammonium chloride solution but
also pre-RO leachate. Each equilibrium experiment consisted of a series of twelve bottles
containing the sorbent (zeolite or resin) and 100 ml of the test solution. The first bottle was
the control and had no sorbent added to the solution, while the remaining bottles had the
sorbent added in mass increments [25, 26]. The masses of each material were measured
using an analytical balance with 0.0001 g precision. The bottles were agitated on a rotary
mixer at ambient temperature (ca. 294 K) for a period of 72 hours for zeolites and 2 hours
for resins, which was sufficient to ensure equilibrium had been attained. The pH and
conductivity of the solution/adsorbent mix in each bottle was measured before and after
equilibrium using a TPS® smartCHEM-LAB™ multi-parameter analyser. At the end of the
equilibration period the samples were syringe-filtered (0.20 µm) in preparation for solution
characterisation. The concentration of metal cations after equilibrium was determined by
ICP-OES and the ammonium content was determined by Kjeldahl distillation. Cation
concentrations after equilibration (Ce (mg/L)) were then recorded and the equilibrium
sorbate concentration in the zeolite or resin material (qe (mg/g) or g/kg) calculated using
Equation 4.1;
Equation 4.1:
Where, qe is the equilibrium loading of the cation on the sorbent (mg/g); V is the solution
volume (L); m is the mass of material (g); and Co is the initial concentration of the cation of
interest in solution (mg/L).
86
Equilibrium data was primarily fitted using the Competitive Langmuir model which observes
the stoichiometric restrictions imposed by an ion exchange process [27] [Equation 4.2].
Equation 4.2
Where R = either resin or zeolite exchange site. By the law of mass action and the mass
balance qt = qeNH4 + qeH and Co = CeH + CeNH4, we can derive Equation 4.3.
Equation 4.3
Non-linear least squares fitting procedures were conducted in accord with previous work
which has shown the validity of this approach [26, 28]. The Solver add-on in Microsoft Excel
was used to process the equilibrium exchange data.
The Aranovich-Donohue (AD) equation is suitable for modelling isotherms with the ‘L3’
profile according to the classification reported by Hinz [29]. Aranovich and Donohue [30]
developed a general sorption equilibrium expression which allowed for not only monolayer
uptake of sorbate but also multi-layer formation. In theory, any isotherm equation can be
incorporated into the Aranovich and Donohue model and in this study we have used the
Langmuir version [Equation 4.4].
Equation 4.4:
4.2.3 Column Trials
Two u-PVC columns of 50 mm diameter and 0.5 m height, were loaded with media and
arranged in parallel. The stock solutions of interest were transferred from a holding vessel
into both columns simultaneously. Two separate peristaltic Masterflex® pumps were used
to maintain the flow rate through each column at ten bed volumes per hour (10 BV/h). The
bed heights in the two columns were approximately the same: 0.330 m for zeolite and 0.317
m for resin. Samples were collected at regular intervals for analysis.
4.2.4 Analysis
Solutions were analysed using a VISTA-MPX CCD simultaneous ICP-OES instrument using an
integration time of 0.15 seconds with 3 replications, with the following wavelengths (nm): Al
87
(396.152), Ca (422.673), Mg (285.213) and sodium (589.592). A certified standard from
Australian Chemical Reagents (ACR) containing 1000 mg/L of aluminium, calcium,
magnesium, and sodium was diluted to form a multi-level calibration curve using a Hamilton
auto-diluter. Ammonium content of the solutions treated was determined using a VELP
Scientifica UDK 149 Automatic Distillation Unit followed by titration. Boric acid indicator
was used with sulphuric acid to titrate the distilled solutions. The ammonium concentration
on each sample was determined by Equation 4.5:
Equation 4.5:
Where: Ctitrant = Concentration of acid titrant in (mol/L); Vtitrant = Volume of acid titrant (ml);
Ntitrant = Normality of acid titrant; Vsample = Volume of sample distilled (ml); Mr = Molar mass
of NH4+ (g/mol); Acid titrant: sulphuric acid (H2SO4).
88
4.3 RESULTS AND DISCUSSION
4.3.1 Ammonium exchange equilibria from NH4Cl solutions – Natural Zeolite
4.3.1.1 “As Received” Natural Zeolite
Approximately 234 mg/L ammonium ions from ammonium chloride solution were
equilibrated with natural zeolite as supplied by the producer. Not only was the ammonium
uptake calculated but the corresponding release of ions present on the natural zeolite was
recorded [Figure 4.1]. In particular, the test solution showed an increase in the presence of
Ca2+, Mg2+, Na+ and K+. This latter observation was consistent with previously disclosed
analysis of the natural zeolite material from Zeolite Australia [10]. The sorption profiles
were all plotted on the same x-axis scale, namely the equilibrium concentration of
ammonium ions in solution as this facilitated interpretation of the exchange process. From
the shape of the profiles it was apparent that there was only marginal affinity of the natural
zeolite for ammonium ions as the best fit of the equilibrium data was almost linear. Lins et
al. [31] reported that ammonium ion exchange isotherms generated from contacting landfill
leachate with Argentinian natural zeolite were convex in character. However, as revealed
by Millar et al. [10] the shape of the ammonium equilibrium profile was dependent upon
the composition of the zeolite exchange sites, with the presence of lower quantities of
sorbed calcium ions leading to convex isotherms and higher calcium loadings promoting
linear isotherms. In addition the bottle-point method employed to create equilibrium
exchange isotherms has also been demonstrated to be important in relation to the shape of
the isotherm profile [25, 26]. It was noted that Lins et al. [31] used a constant mass of
zeolite in their equilibrium experiments and this approach has been shown to be
problematic in terms of generating isotherm profiles which may not be representative of
the exchange process occurring.
The maximum loading of ammonium ions was predicted to be 0.51 eq NH4/kg zeolite (9.18
g/kg). This latter value was in accord with previous studies such as that by Tosun [32] which
concluded that Turkish natural zeolite loaded up to ca. 14 g NH4/kg. It was of interest to
examine the balance between equivalents of ammonium loaded on the natural zeolite
sample in relation to the amount of cations released during the exchange process [Table
4.2]. It was noted in every instance that the quantity of cations desorbed from the zeolite
89
surface was slightly in excess of the amount expected based purely upon an ion exchange
process. One possible explanation of this latter observation may have been related to the
fact that under the long agitation times required for equilibrium to occur with zeolite
materials, the zeolite samples were noted to structurally degrade. Thus, elements from the
zeolite material may have partially dissolved in the solution and thus been detected during
the analysis procedures. Ye et al. [11] also described zeolite sample degradation occurring
during batch tests of ammonium uptake from landfill leachate, hence this fact should be
taken into account when interpreting experimental data.
Table 4.2: Mass balance between sorbing and desorbing species when 234 mg/L ammonium
ions from ammonium chloride solution contacted with natural zeolite (eq)
Zeolite Mass Added (g)
0 0.1250 0.2504 0.3753 0.5000 0.7500
Desorbed (eq) 0.00 0.89 1.51 1.81 2.31 3.14
Adsorbed (eq) 0.00 0.73 1.26 1.55 1.99 2.75
Zeolite Mass Added (g)
1.0005 1.2503 2.5002 3.7503 6.2504 9.3753
Desorbed (eq) 3.92 4.62 6.95 8.46 10.34 11.53
Adsorbed (eq) 3.55 4.27 6.29 7.58 9.37 10.51
In terms of the ease of desorption of the ions displaced by ammonium species, sodium ions
(K = 4.4 L/mg) appeared to be released easier than potassium (K = 0.62 L/mg) which was in
turn more facile to desorption than magnesium (K = 0.25 L/mg) or calcium ions (K = 0.24
L/mg) [Figure 4.1]. Hankins et al. [33] reported a detailed study of the exchange behaviour
of ammonium ions in the presence of competing cations with natural zeolite. These authors
proposed that the preference of natural zeolites for sorbing ions from solution was K+ >
NH4+ > Na+ > Ca2+ > Fe3+ > Al3+ > Mg2+. This current study is in agreement with Hankins et al.
[33] in that potassium ions are more preferred by the zeolite surface than sodium ions.
Weatherley and Miladinovic [34] in contrast, found that the order of preference for cations
by natural zeolite was Ca2+ > K+ > Mg2+. Milan et al. [35] determined from column studies of
ammonium ion exchange with homoionic natural zeolites that the selectivity series was
Mg2+ > K+ > Ca2+. Alternatively, Lei et al. [36] examined binary mixtures of ammonium and
90
selected competing cations with natural zeolite and deduced that sodium ions were
responsible for the greatest inhibition of ammonium uptake, followed by potassium,
calcium and magnesium ions, respectively. Apparently, there is not a universally accepted
selectivity order for natural zeolites in relation to ammonium uptake in the presence of
competing cations.
Figure 4.1: Equilibrium exchange of 234 mg/L ammonium ions from ammonium chloride
solution with natural zeolite
The reasons for this latter deduction are not well understood at present, but Milan et al.
[35] discussed differences in mobilities of cations on zeolite exchange sites due to variations
in ion coordination to the zeolite surface and mobilities of ions in solutions. Wang et al. [37]
also noted that modification of natural zeolite structure using sodium hydroxide under
91
hydrothermal conditions resulted in changes of the order of selectivity for the zeolite with
sodium, potassium, calcium and magnesium ions. This latter observation was correlated
with differences in the pore structure and dimensions.
The solution pH was also monitored during the ammonium exchange process in order to
evaluate the sorption behaviour in more detail [Figure 4.1]. It was noted that the
equilibrium solutions were invariably slightly higher in pH value compared to the initial
solution/zeolite mixtures, albeit the pH remained below 7 and typical pH elevation was in
the range 0.5 to 0.8 pH units. This latter observation was consistent with a reduction in the
presence of the acidic ammonium chloride species in solution and an increase of neutral
sodium chloride and potassium chloride species, in addition to less acidic species such as
calcium and magnesium chloride.
4.3.1.2 Sodium Natural Zeolite
Pre-treatment of natural zeolites with sodium-containing brine has often been described in
relation to providing a more uniform exchange material [33] or in relation to regeneration
strategies [38]. Sodium exchanged zeolites are also claimed to be more effective at
removing ammonium ions from solution [39]. Consequently, a natural zeolite was used
which had been subjected to washing with concentrated sodium chloride solutions by the
supplier, Zeolite Australia. Figure 4.2 shows that the sodium exchange procedure has some
impact upon the nature of the ammonium ion exchange with natural zeolite. The shape of
the ammonium isotherm profile was now slightly convex which was in harmony with the
previous literature which suggested that the greater presence of sodium ions on the zeolite
sites promoted ammonium ion uptake on the zeolite surface sites [10]. Consistent with this
latter statement was the improved ammonium uptake of 0.68 eq/kg (12.24 g NH4/kg
zeolite) and observation of an increased quantity of sodium ions released from the sodium
modified zeolite sample [c.f. Figure 4.1 & 4.2]. The maximum desorbed quantity of sodium
ions was 0.30 eq/kg zeolite for the modified zeolite compared to 0.15 eq/kg for the “as
received” zeolite. The quantity of magnesium ions desorbed from the sodium modified
zeolite was substantially reduced by the pre-treatment method employed, with maximum
desorbed quantity of 0.02 eq/kg zeolite compared to 0.10 eq/kg for the “as received”
92
material. Interestingly, the desorbed quantities of calcium and potassium ions were similar
in both the cases of “as received” and sodium modified zeolite [c.f. Figure 4.1 & 4.2],
however, there was a difference in the shape of the isotherm profiles. Most notably,
desorption of calcium ions from the sodium modified zeolite surface appeared to be more
facile as the shape of the desorption curve was less unfavourable. Similarly to the situation
for the as received natural zeolite, the solution pH was observed to be slightly raised
following equilibration of the ammonium ions with the sodium modified sample [Figure
4.2].
Figure 4.2: Equilibrium exchange of 245 mg/L ammonium ions from ammonium chloride solution with sodium modified natural zeolite
93
Overall, it could be ascertained that simple washing of natural zeolite with concentrated
sodium chloride solutions may not be adequate to convert all the exchange sites to the
sodium form. Leyva-Ramos et al. [40] found that the exchange of sodium ions from sodium
chloride solution was relatively slow when attempting to modify natural zeolites, with
typically days of contact time required. Instead, acid washing of the natural zeolite may be
required to effectively remove the more strongly held cations on the zeolite exchange sites,
then sodium exchange from sodium chloride or sodium hydroxide solutions may be more
effective. This latter conclusion is supported by work published by Wang et al. [41] who
treated natural zeolite with dilute acid solutions and found that the material performance
for demineralizing coal seam water samples was significantly improved. Millar et al. [10]
similarly reported that ammonium ion exchange with natural zeolite initially washed with
dilute acid and subsequently sodium loaded by using a sodium hydroxide solution exhibited
enhanced uptake of ammonium species.
4.3.1.3 Acid Pre-Treated Natural Zeolite
In line with our previous observations regarding the relative ineffectiveness of sodium
chloride treatment of natural zeolite samples it was pertinent to examine the ammonium
exchange behaviour of acid modified natural zeolite samples. Figure 4.3 shows the
exchange behaviour of natural zeolite which has been washed with a dilute hydrochloric
acid solution. It was apparent that the overall exchange capacity of the natural zeolite had
been reduced as the maximum loading value for ammonium (qmax) was reduced to 0.34
eq/kg zeolite. This latter observation was in contrast to the study of Sarioglu [42] who
noted a 22 % increase in the capacity of a Turkish natural zeolite for ammonium ions,
relative to an unmodified material. Reasons for the outlined discrepancy may be explained
as follows. It was evident that the original capacity of the Turkish natural zeolite prior to
acid treatment was relatively low, 0.7 to 1.08 g NH4/kg zeolite.
94
Figure 4.3: Equilibrium exchange of 244 mg/L ammonium ions from ammonium chloride solution with acid modified natural zeolite
Examination of the physical composition of the Turkish zeolite indicated that it was not
markedly different from the Australian zeolite in that the main extra-framework cations
present were calcium, sodium, potassium, and magnesium. Consequently, one must
consider the possibility that the channels within the zeolite sample may have been blocked
by the presence of non-zeolitic materials. This latter situation was proposed by Sarioglu [42]
as a reason for the increase in zeolite capacity after acid treatment as it was envisaged that
amorphous material obstructing entrance of ions to the internal pore structure may have
been dissolved. Wang et al. [41] also postulated that unblocking of zeolite pores with dilute
acids was at least partially responsible for improvement in zeolite exchange properties.
95
These authors also discussed the possibility of pore enlargement in the zeolites, due to
dealumination by the presence of acid, as being another factor which could promote ion
uptake. As for the decrease in ammonium ion uptake after acid treatment observed in this
study, excess dealumination of the zeolite structure due to exposure to too concentrated
acid may explain the experimental data. As noted by Wang et al. [41] higher acid
concentrations could destroy pore structures in natural zeolites, and thus reduced ion
exchange capacity could be expected. Leyva-Ramos et al. [40] discovered that
dealumination of natural zeolites was accelerated below pH 2.5 and by a pH of 0.76, 55 % of
the aluminium had been removed from the zeolite material. Notably, the solution pH was
decreased after equilibration due to release of protons from the zeolite exchange sites
[Figure 4.3].
Table 4.3: Mass balance between sorbing and desorbing species when 244 mg/L Ammonium
ions from ammonium chloride solution contacted with acid modified natural zeolite (eq)
Zeolite Mass Added (g)
0 0.1276 0.2520 0.3759 0.5093 0.7518
Desorbed (eq) 0.00 0.55 1.00 1.45 1.81 2.53
Adsorbed (eq) 0.00 0.31 0.84 1.39 1.79 2.23
Zeolite Mass Added (g)
1.0025 1.2578 2.5164 3.7507 6.2517 9.3786
Desorbed (eq) 3.14 3.61 5.68 7.09 9.04 10.63
Adsorbed (eq) 2.84 3.40 5.54 6.88 8.80 10.28
4.3.2 Ammonium exchange equilibria from NH4Cl solutions – Resin
4.3.2.1 H+ - Resin
Exchange of ammonium ions with Lanxess S108H SAC resin resulted in the isotherm shown
in Figure 4.4. The Competitive Langmuir fit of the data suggested that the maximum loading
of ammonium ions was 2.24 eq/kg resin (40.3 g/kg). The stated capacity of the resin by the
manufacturer was 1.8 mol/L for H+ resin with a packing density of 0.79 g/L. Hence, it was
calculated that the maximum loading of ammonium ions of 31.8 g/L in this study was in
agreement with the value of 32.4 g/L outlined by Lanxess. The convex nature of the
96
isotherm indicated that the exchange of ammonium ions with SAC resin was favourable.
Due to the ejection of protons from the resin surface by ammonium ions, the solution pH
was observed to reduce in accord with the ion exchange process [Figure 4.4].
Figure 4.4: Exchange of 250 mg/L ammonium ions from ammonium chloride solution with Lanxess S108H resin
4.3.2.2 Na+ - Resin
The impact of changing the cation on the resin surface from H+ to Na+ was examined in
relation to ammonium ion uptake. Figure 4.5 shows the sorption of ammonium ions and
concomitant desorption of sodium ions from the resin surface when a 252 mg/L solution of
ammonium ions from ammonium chloride solution was equilibrated with Na+-SAC resin.
The maximum loading of ammonium ions was estimated to be 2.65 eq/kg resin (47.7 g/kg or
37.7 g/L). As the resin contracted by ca. 10 % volume when transforming from the proton
to sodium exchanged form, the stated capacity as indicated by Lanxess increased from 1.8
to 2.0 eq/L. Consequently, the predicted loading capacity of the sodium exchanged SAC
resin was 36 g NH4/L resin, which was comparable to the measured value of 37.7 g NH4/L
determined in this study. The desorption isotherm for sodium ions released from the resin
surface due to uptake of ammonium ions was very similar in profile to the ammonium
isotherm [Figure 4.5], which indicated that the selectivity of the resin for sodium ions was
comparable to that for ammonium ions. The slightly convex nature of the ammonium
isotherm suggested that it was slightly preferred relative to the more linear sodium
isotherm. The quoted selectivity by the resin manufacturer relative to protons on an 8 %
divinylbenzene crosslinked sulphonated polystyrene resin is 1.56 and 2.01 for sodium and
97
ammonium ions, respectively, which confirmed our latter deduction. Likewise, Yoon et al.
[22] also mentioned that the selectivity order for strong acid cation resin was NH4+ > Na+ >
H+. Correlation of the amount of ammonium ions sorbed and the quantity of sodium ions
desorbed from the resin surface [Figure 4.5] showed a linear relationship, as expected if a
stoichiometric ion exchange process occurred. The only caveat was the fact that the
gradient of the trend line was not 1 but ca. 1.1, which suggested that a small excess of
sodium ions had been ejected from the resin. One explanation was that there was an error
in the measurement of sodium ions, however, this does not seem feasible as considerable
care was taken to ensure accuracy of the ICP-OES analysis. Alternatively, the possibility of
the presence of some non-structural sodium ions in the original resin sample should be
considered. During the conversion of the proton resin material to the sodium exchanged
form, it is possible that despite extensive washing some sodium salt remained within the
resin pores. Sica et al. [13] analysed sodium exchanged SAC resin both prior to and after
ammonium exchange. As expected, the amount of sodium present in the resin sample
decreased and the quantity of ammonium loaded increased. However, there was not a
stoichiometric relationship between the two exchanging ions. Sica et al. [13] noted the
outlined discrepancy and suggested that adsorption may also have accompanied the ion
exchange process. Helfferich [43] emphasised that ion exchange processes were unlikely to
occur without the co-presence of adsorption phenomena. The pH of the equilibrated
solutions increased due, at least in part, to replacement of acidic ammonium chloride
solution with neutral sodium chloride species [Figure 4.5], and perhaps due to expulsion of
some sodium hydroxide species if some residual moieties remained after the
aforementioned conversion process.
98
Figure 4.5: Exchange of 252 mg/L ammonium ions from ammonium chloride solution with sodium exchanged Lanxess S108 resin
Yoon et al. [22] reported that H+ SAC resin had a slightly higher efficiency for ammonium ion
removal from solution than Na+ exchanged SAC resin. In our study this behaviour was not
apparent as the ammonium capacities calculated were simply in agreement with volume
changes of the resin due to the change in identity of the major counter ion present on the
resin exchange sites. One key difference may have related to the fact that Yoon et al. [22]
compared the performance of resins from different suppliers, whereas we used the same
type of resin in this study. The variation in resin performance noted by Yoon et al. [22] may
have been at least in part due to the properties of the different SAC resins used.
4.3.3 Landfill Leachate Equilibria - Pre-RO (field sample)
4.3.3.1 “As Received” Natural Zeolite
The sorption and desorption profiles resultant from equilibrium exchange of landfill
leachate with “as received” natural zeolite [Figure 4.6] were considerably more complex
when compared with the situation wherein only aqueous solutions of ammonium chloride
were studied [Figure 4.1].
99
Figure 4.6: Equilibrium exchange of cations from landfill leachate with “as received” natural
zeolite
The maximum loading of ammonium ions on the zeolite was only 0.07 eq/kg zeolite (1.26
g/kg zeolite) which represented a dramatic reduction compared to the value of 9.18 g
NH4/kg zeolite deduced from the ammonium chloride equilibria. Reasons for the
diminishment in ammonium loading include the fact that the concentration of ammonium
ions present was significantly lower (54 mg/L compared to 234 mg/L) [10] and the presence
of more preferred competing cations in solution. Calcium desorption from the zeolite
surface was suppressed by use of the landfill leachate with maximum desorbed amount only
ca. 0.2 instead of ca. 0.49 eq/kg for ammonium chloride solution. Interestingly, the sodium
100
ion desorption isotherm was now of the unfavourable concave type (K = 0.04 L/mg) instead
of the favourable convex profile noted for the situation for ammonium chloride exchange
[Figure 4.1]. In agreement with this study, Karadag et al. [12] found that calcium and
sodium ions were the principal ions released from the surface of a natural zeolite when
loading ammonium ions from leachate solutions.
The isotherm profile for magnesium ion uptake on the zeolite sample was complex as
evidenced by the fact that the Competitive Langmuir fit did not simulate the experimental
data to a high degree. Visual inspection of the data in Figure 4.6 suggested that the
magnesium may have initially sorbed on the zeolite surface in a favourable manner (convex
profile) followed by a period where the magnesium uptake was less favourable (concave
profile). Potassium ions were also noted to load on the zeolite surface under the test
conditions, and similar to the case with magnesium ions the sorption profile was not
simulated by the Competitive Langmuir fit. The data of Guo et al. [44] was in harmony with
this study as they noted that with mixtures of ammonium and potassium ions in solution,
both species loaded on natural zeolite with potassium ions more preferred. Karadag et al.
[12] noted in column studies of natural zeolite for landfill leachate treatment that potassium
ions which were initially loaded onto the zeolite during the early stages of the exchange
process were actually desorbed once the breakthrough point for ammonium ions was
attained. In terms of equilibrium isotherm profiles, one would expect these changes in
sorption behaviour to be observed at high values of Ce. For the potassium isotherm in
Figure 4.6 it was noted that potassium ions were loaded onto the zeolite exchange sites at
low values of Ce and that the potassium ions then desorbed as evidenced by the rapid
reduction in overall potassium loading on the zeolite. Concomitant to desorption of
potassium ions from the zeolite surface at high Ce, was an accelerated release of sodium
ions from the zeolite surface [Figure 4.6]. The underlying reason for the complex sorption
behaviour observed in Figure 4.6 is not clear. Inspection of Table 4.4 showed that the
equivalence balance was reasonably good at high zeolite doses (corresponding to low values
of Ce) and significantly divergent at low zeolite dose rates. In particular, the amount of
species desorbing from the zeolite surface was substantially higher than the quantity of
species loading on the exchange sites.
101
Table 4.4: Mass balance between sorbing and desorbing species when landfill leachate
equilibrated with as received natural zeolite (eq)
Zeolite Mass Added (g)
0 0.1008 0.2176 0.3136 0.6036 0.9013
Desorbed (eq) 0.00 0.65 1.55 2.41 1.31 2.11
Adsorbed (eq) 0.00 0.28 0.18 0.46 1.36 1.88
Zeolite Mass Added (g)
1.5015 2.5162 5.0101 10.0122 15.0000 20.0107
Desorbed (eq) 4.35 4.83 6.95 8.14 9.86 10.52
Adsorbed (eq) 2.72 4.04 6.26 8.40 9.24 10.11
The discrepancy in the equivalence balance indicates that a simple, stoichiometric ion
exchange process cannot explain the recorded data. Instead we must consider other
possibilities such as degradation of the zeolite structure during the experiment or the
presence of species which are not exchanged with zeolite surface sites. The issue of sample
degradation interfering with the sorption data does not appear to explain the disparity in
the equivalence balance. For example, when high masses of zeolite were present in the
solution, the equivalence balance was good yet the extent of attrition was significant (due
to the zeolite particles contacting each other during the agitation/equilibration period).
One must then consider if the uptake of the ammonium ions displaced non-structural
species encapsulated in the zeolite pore system. This latter suggestion appears plausible in
light of the known ability of sorbent materials to exhibit super equivalent ion exchange
(SEIX) [45, 46]. Notably, the lowest zeolite masses used corresponded with the point where
the zeolite exchange sites were saturated.
4.3.3.2 Sodium Exchanged Natural Zeolite
A similar set of experiments to those revealed in Section 4.3.3.1 were conducted for the
zeolite sample which had been modified by contact with sodium chloride solutions. As
shown in Figure 4.7, there was minimal difference in the extent of ammonium uptake on the
zeolite which was in contrast to the data for the solutions which only comprised of
ammonium ions [c.f. Figure 4.1 & 4.2]. Surprisingly, the release of calcium ions into solution
102
was actually enhanced (0.39 eq Ca/kg zeolite) relative to the case when an “as received”
zeolite sample was tested (0.20 eq Ca/kg zeolite) [Figure 4.6].
Figure 4.7: Equilibrium exchange of cations from landfill leachate with “sodium modified” natural zeolite
Correspondingly, the release of sodium ions from the zeolite sample was suppressed. The
loading of magnesium ions on the zeolite was slightly enhanced by the sodium chloride pre-
treatment of the sample, and the overall isotherm profile was similar to the case when “as
received” natural zeolite was used [Figure 4.6]. The potassium ion loading on the zeolite
was remarkably similar to the “as received” natural zeolite situation and supported the
concept that the distinct lowering in potassium loading at high values of Ce, which was
103
followed by a final increase in potassium ion uptake, was a real event and not random
scatter in the data.
Table 4.5 illustrates the balance between species sorbed on the zeolite surface and species
released into solution once equilibrium was achieved. As in the case with the “as received”
natural zeolite the quantity of ions desorbed from the zeolite material was consistently
higher than the amount of ions sorbed on the zeolite. Indeed, the discrepancy observed
was greater after the sodium chloride washing of the zeolite sample. It is tentatively
suggested that the washing process perhaps disturbed some of the mineral phases present
(noting that the sample was not comprised of only clinoptilolite but also contained various
materials based upon silicon, iron, titania and calcium [10]).
Table 4.5: Mass balance between sorbing and desorbing species when landfill leachate
equilibrated with sodium modified natural zeolite (eq)
Zeolite Mass Added (g)
0 0.1000 0.2084 0.3002 0.6004 0.9006
Desorbed (eq) 0.00 1.56 0.97 2.19 3.64 3.73
Adsorbed (eq) 0.00 0.27 0.82 1.28 1.95 3.38
Zeolite Mass Added (g)
1.5002 2.5021 5.0065 10.0056 15.0028 20.0079
Desorbed (eq) 6.02 7.93 10.15 13.33 15.53 16.42
Adsorbed (eq) 4.75 6.20 8.44 10.30 11.26 11.93
Overall, the ammonium exchange data with natural zeolite samples was more complex than
may have previously been ascertained from previous studies. The behaviour of the ions
loading on the zeolite and desorbing from the zeolite exhibited changing modes of
interaction with the zeolite as apparent from the inability to fit the data with fundamental
sorption models such as the Langmuir expression.
104
4.3.3.3 Sodium Exchanged SAC Resin
Figure 4.8 shows the equilibrium exchange data when pre-RO treated landfill leachate was
contacted with a sodium exchanged strong acid cation resin. The desorption behaviour of
sodium ions was illustrated using the equilibrium concentration of calcium ions as the x-axis
value. It was apparent that the sorption of ammonium and potassium ions was
unfavourable (concave isotherm profiles) relative to the case for calcium and magnesium
ions (convex isotherm profiles). Malovanyy et al. [47] studied the exchange of ammonium
ions from municipal water using a strong acid cation resin and suggested the selectivity
series was in the order Ca2+ > Mg2+ > K+ ≈ NH4+ > Na+ > H+. This latter trend was generally in
agreement with the current results. However, greater detail was evident in Figure 4.8 which
highlighted more complex behaviour when multicomponent solutions interacted with SAC
resin. The sodium desorption before could not be modelled using a simple Langmuir type
isotherm, instead the Aranovich-Donohue expression was employed to simulate the
experimental data. Both calcium and magnesium ions were observed to be displaced from
the resin surface at high values of Ce, whereas ammonium and potassium ions
concomitantly were loaded onto the resin exchange sites. Interestingly, a large amount of
sodium ions was desorbed from the resin surface at high Ce values, and clearly the number
of equivalents released was in excess of the stoichiometric number expected if this
behaviour was simply due to ion exchange. This latter observation was consistent with the
idea that non-framework sodium ions were present within the resin pore structure, i.e.
dissolved sodium chloride species. The possibility that dissolved salts may be present in
resins has been detailed by Millar et al. [28] in studies of water softening using weak acid
cation resins. The sodium chloride species would have been introduced during the
preparation stage where the resin was exchanged with concentrated sodium chloride
solutions in order to create the sodium exchanged form. It was noted that extensive
flushing of the resin was performed after the outlined exchange process. Hence, the sodium
chloride species appear to be retained in the resin pores due to forces such as electrostatic
attraction.
105
Figure 4.8: Equilibrium exchange of cations from landfill leachate with sodium exchanged strong acid cation resin
4.3.4 Column Trials
Column trials were restricted to studies of sodium modified zeolites and resins as in
commercial operations regeneration of exchange media is commonly conducted using
concentrated solutions of sodium ions, such as sodium chloride brines [47]. Due to the
limited uptake of ammonia observed in Figure 4.8 regarding treatment of pre-RO treated
leachate, column studies were conducted using the post-RO leachate as it was envisaged
that competition with competing ions would have been reduced in this instance.
106
4.3.4.1 Sodium Modified Natural Zeolite
A simulated post-RO treated landfill leachate solution [Table 4.1] was passed through a
column of sodium modified natural zeolite (bed volume 650 cm3) at a flow rate of 6.55 L/h
(10.1 BV/h). At the end of the experiment, a total of 400 bed volumes had passed through
the column. The general shape of the ammonium ion breakthrough curve was comparable
to that presented by Cyrus and Reddy [48] in that a sharp breakthrough of ammonium ions
was not observed. Interruption tests [43] were conducted at various times to determine the
influence of intraparticle diffusion upon the exchange process. For this latter test, typically
the flow was stopped for a period of at least 10 hours and then restarted. Figure 4.9 shows
that with every interruption of the flow of RO treated landfill leachate the concentration of
cations at the outlet of the zeolite column reduced significantly. As such, it was inferred
that the ion exchange process was significantly limited by intraparticle diffusion. This
conclusion was supported by previous work by Malekian et al. [49] and Lei et al. [36] whose
kinetics studies indicated that ammonium exchange with natural zeolite was diffusion
limited. Sprynskyy et al. [50] also presented data where they had stopped the flow in
columns of mordenite and observed a significant decrease in the ammonium content of the
effluent. As only ammonium ion concentration was impacted and not that of the other
cations detected in solution, this latter effect was correlated with intraparticle diffusion.
The column study indicated that ammonium, potassium, and magnesium were all loaded on
the zeolite material, whereas sodium and calcium ions were released from the zeolite;
which was consistent with the equilibrium studies shown in Figure 4.7. The sodium ion
concentration in the effluent appeared to correlate with the ammonium uptake behaviour
i.e. the less sodium ions released the lower the quantity of ammonium ions loaded onto the
zeolite. The concentration of calcium ions in the effluent was overall not significantly
reduced after 400 BV treated [Figure 4.9]. The removal of potassium ions remained
substantial (< 5 mg/L) at the end of the column test which reflected the higher selectivity of
the zeolite for this ion. In contrast, magnesium ion concentrations were substantially
increased in the effluent as the exchange process proceeded [Figure 4.9] which indicated
that the affinity of the zeolite for ammonium ions and magnesium ions was similar.
107
Figure 4.9: Column trials of post-RO treated landfill leachate with sodium modified natural zeolite; flow rate 10.1 BV/h
4.3.4.2 Sodium Exchanged Strong Acid Cation Resin
Simulated post-RO treated landfill leachate was in this instance flowed through a 5 cm
diameter u-PVC column packed with sodium exchanged strong acid cation resin. The
average flow rate was 6.25 L/h which equated to 10.1 BV/h. In contrast to the situation
where natural zeolite performance in column trials was evaluated, there was no noticeable
deviation in the concentration of ions in the effluent after an interruption test was
performed [Figure 4.10]. This latter result does not preclude that intraparticle diffusion
controlled the kinetics of ammonium ion exchange in SAC resin, but it does imply that the
overall kinetics for the resin process was substantially faster than that for zeolite samples.
108
For example, Sica et al. [31] proposed that both film and intraparticle diffusion were
important in relation to exchange of ammonium ions with sodium exchanged SAC resin.
The only ion to increase in concentration during the column test was sodium, which was
released from the resin upon replacement with more favourable cations. Ammonium,
potassium, calcium, and magnesium ions were all reduced in concentration in the effluent
to < 0.1 mg/L during the entire test period where 400 BV of water were treated. In contrast,
column data presented by Yoon et al. [22] for ammonium ion exchange with SAC resin did
not show essentially complete removal of ammonium ions from solution at any stage of the
test. However, as noted by these latter authors the bed depth they employed was very
short (< 2 cm) which may have been less than the critical bed depth required under the
applied experimental conditions to prevent immediate breakthrough of ammonium ions.
Malovanyy et al. [47] employed significantly larger bed volumes of cation resins in column
tests devised to understand the ammonium ion removal behaviour from wastewater
solutions. With the aforementioned relatively deep beds (ca. 30 to 40 cm) comparable to
this study, in excess of 100 BV of water was treated with almost complete removal of
ammonium ions from solution.
Integration of the breakthrough curves using the Trapezoid rule revealed that the quantities
of the various cations sorbed on the resin when the test was stopped were as follows:
ammonium 189.4 meq; potassium 105.9 meq; calcium 16.8 meq and magnesium 95.5 meq.
Similarly, the quantity of sodium ions expelled from the resin surface was 399.5 meq.
Examination of the total amounts of cations sorbed on and released from the resin revealed
an acceptable balance in the number of equivalents (399.5 meq released; 407.6 meq
sorbed) confirming stoichiometric ion exchange. As none of the cations had achieved a
breakthrough point (normally 10 % of the incoming feed value [48]) it was not possible to
make any comment regarding the relative selectivity of the ions with the strong acid cation
resin from the present study. However, previous work [47] suggested the appropriate
selectivity series for SAC resin was Ca2+ > Mg2+ > K+ ≈ NH4+ > Na+ > H+ and this selectivity
series was supported by the exchange data shown in section 4.3.3.3.
109
Figure 4.10: Column trials of post-RO treated landfill leachate with sodium exchanged SAC
resin; flow rate 10.1 BV/h
It was apparent that for the landfill leachate after reverse osmosis treatment that synthetic
strong acid cation resin was highly efficient at demineralizing this solution to produce water
of very low total dissolved solids content. In contrast, the natural zeolite sample was not
able to meet low levels (e.g. < 1 mg/l) of ammonium concentration in the effluent for a
significant number of bed volumes. Malovanyy et al. [47] also concluded from column trials
of ammonium exchange from municipal water with either cation resins or natural zeolites
that synthetic resins were superior in performance. The one caveat was that resins did not
exhibit selectivity preference to ammonium ions and that if this latter criteria is required to
be satisfied then zeolites may be advantageous.
110
The performance of the natural zeolite could be promoted by improved operational
strategies. For example, consider a situation where two columns of zeolite are available;
one of which is shut down and the other is in operation. At an appointed time such as when
ammonium concentration in the outlet reached a prescribed value, the column in operation
would be shut down and the flow diverted to the other zeolite column. This latter
procedure would allow the diffusion of ions to occur into the zeolite pore structure of the
now shutdown column and thus promote overall sorption capacity. Again, once the
ammonium concentration in the effluent attained a set value, the column now in operation
would be shut down once more. Finally, the difference in cost between resin and natural
zeolite should be considered when deciding which exchange media to use. For example,
reasonable prices for resin and zeolite are A$5/L (A$6.33/kg) and A$0.5/kg, respectively.
Clearly, the total cost of a zeolite charge is substantially cheaper than the comparable resin
quantity. Other factors should also be analysed such as capital cost differences due to
different media capacities and regeneration efficiencies. In summary, the ultimate choice of
material to use needs to be based upon economic as well as technical specifications.
4.4 CONCLUSIONS
This research has elucidated the differences in performance between natural zeolites and
synthetic resins for the treatment of landfill leachate. It was our hypothesis that, in certain
circumstances, synthetic resins may have application for removal of ammonium species
from leachate solutions despite their inherently low selectivity to ammonium species. In
addition, we were of the opinion that exchange of ammonium ions with natural zeolites
may be more complicated than previously disclosed. Ammonium ion uptake (from simple
solutions of ammonium chloride) on zeolites depended upon the identity of the exchanging
ions on the zeolite surface sites, with sodium chloride pre-treatment enhancing ammonium
ion loading. Acid treatment of the zeolite in an effort to increase zeolite capacity for
ammonium ions actually reduced the ability to capture ammonium species, presumably due
to destruction of the zeolite framework. SAC resins did load ammonium ions but displayed
minimal affinity for these latter species. It was discovered that multi-component exchange
of ions from landfill leachate solutions with natural zeolite exhibited complex behaviour.
111
The quantity of species desorbed from the zeolite material did not correlate to the amount
of species loaded on the zeolite, and this may have been due to the species not attached to
the zeolite framework.
Column trials indicated that strong acid cation resin could effectively purify landfill leachate
post reverse osmosis treatment. Despite the limited selectivity of resins for ammonium
species in the presence of competing cations such as potassium, calcium, and magnesium;
the resin was able to perform adequately for at least 400 BV with simulated post-RO treated
landfill leachate. The inherently higher cation exchange capacity and reduced limitations
imposed by intraparticle diffusion associated with resins appeared to be the main factors of
importance. In contrast, zeolites were characterized by limitations associated with diffusion
of the ammonium ions into the zeolite micropores.
Future research to further develop the application of zeolites or resins for landfill leachate
remediation may focus on the following aspects. More complete conversion of natural
zeolites to the sodium exchanged form would allow interpretation of exchange data which
was comparable to commercial operations. In such cases, the zeolite exchange sites would
be expected to be predominantly loaded with sodium ions due to the regeneration process
which would normally employ sodium chloride or sodium hydroxide solutions. Optimization
of acid pre-treatment of zeolite materials to enhance ammonium uptake is warranted.
From this study it appears that there is a threshold exposure to acid (in terms of acid
concentration and contact time) which if surpassed results in excessive dealumination of the
zeolite. Exploration of various landfill leachate compositions to determine the differences in
resin and zeolite performance would shed light upon which solutions could viably be
treated. Finally, examination of several loading and regeneration cycles for both zeolites
and resins would be interesting to determine not only the process robustness but also the
efficiency of various regeneration strategies.
112
4.5. ACKNOWLEDGMENTS
The Science and Engineering Faculty, Queensland University of Technology is gratefully
acknowledged for provision of some of the equipment used in this study. We thank Zeolite
Australia for award of a student scholarship.
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117
Chapter 5: Core-Shell Structures based upon Natural and Synthetic Zeolites
Core-Shell Structures based upon Natural and Synthetic Zeolites
Marita Guarino Bertholini1, Sara J. Couperthwaite*, Graeme J. Millar, Kenneth Nuttall, Ian
Mackinnon
Institute for Future Environments and 1School of Chemistry, Physics and Mechanical
Engineering, Science and Engineering Faculty, Queensland University of Technology,
Brisbane, Queensland 4000, Australia.
We report for the first time core-shell zeolite materials which have a natural zeolite core
and a shell comprising of synthetic zeolites with an affinity for ammonium ions. Both non-
hydrothermal and hydrothermal methods were employed to make the modified zeolite at
bench and floor scale. A layer of zeolite N and zeolite W was observed as the shell
component around the natural zeolite core. The cation exchange capacity for ammonium
ions was enhanced compared to natural zeolite alone (>200 meq/100 g). Synthesis at bench
scale was relatively robust with numerous samples made in a repeatable manner. However,
it was noted that the agitation mode and intensity applied was an important factor,
especially when scaling up the zeolite synthesis to floor scale. Problems with mixing the
zeolite and process solution were noted as seen by the formation of dense clusters of
material at the end of the reaction. It was recommended that an intermittent agitation
strategy be adopted as this aided mixing of the reaction components but minimised attrition
of the zeolite material. Ammonium exchange tests confirmed the enhanced capacity of the
zeolite shell for ammonium ions but revealed issues with the resultant solution pH which
was highly alkaline. The alkaline conditions adjusted the ammonium/ammonia equilibrium
to predominantly ammonia gas which was not exchanged by the zeolite material. Possible
improvements could relate to adjustment of the Si/Al ratio of the zeolite shell.
118
KEYWORDS: core shell; ammonia; zeolite N; natural zeolite; hydrothermal synthesis
119
5.1 INTRODUCTION
Australia is one of the driest continents on Earth and as such there exist many challenges to
secure water supply. In addition, Australia is blessed with abundant natural resources and it
is imperative that these are not only used wisely but that maximum economic return is
achieved. Of particular concern is removal of ammonium species in water and wastewater,
which despite intensive research remains a substantial problem [1-3]. Ammonium
contamination primarily arises from industries which involve: animal farming (feedlots;
poultry farms; piggeries; aquaculture; abattoirs), urban wastewater (sewage treatment
plants; landfill), and chemical production. Recently, particular emphasis has been placed
upon the impact of ammonium run-off into the Great Barrier Reef catchment [4-6]. In
addition, commercial ammonia production using the Haber process is highly energy
intensive, expensive and a major consumer of natural resources [7]. To solve the
aforementioned issues, ideally, what is needed is to develop economical, effective, and
sustainable technologies which can not only remove ammoniacal nitrogen from polluted
water sources but also recover ammonium species as a usable fertilizer product.
Ion exchange materials satisfy the condition of enabling recovery of ammonium species
from wastewater streams. In contrast, biological methods, are normally destructive and
have limitations in terms of resistance to system shocks, lengthy treatment periods and the
requirement for larger construction footprints than ion exchange [8-10]. Natural zeolites
have received considerable interest for remediation of ammoniacal nitrogen contamination.
For example, Millar et al. [11] examined the ammonium uptake performance of two
different Australian natural zeolites and found that the loading depended upon the identity
of the exchangeable cations on the zeolite surface. Pilot plant studies by Cooney et al. [12]
demonstrated that ammonium ions could be removed from sewage effluent at relatively
low flow rates (1-2 bed volumes per hour) and that the active sites could be regenerated by
use of an alkaline solution of 0.6 M sodium chloride. However, natural zeolites exhibit
disadvantages such as slow diffusion of ions through the microporous structure [13] and
limited selectivity for ammonium ions in the presence of common competing ions in
wastewater such as calcium, magnesium, sodium, and potassium [14]. The synthetic zeolite
termed “Zeolite N” was discovered to exhibit excellent capacity for ammonium ions
recovered from wastewater and a high selectivity for ammonium ions in the presence of
120
calcium and magnesium ions [15, 16]. The synthesis of zeolite N from solutions of
potassium hydroxide and potassium chloride with aluminosilicate materials such as kaolin
has been described in depth [17, 18]. However, synthetic zeolites are invariably
substantially more costly than natural zeolites which are simply dug out of the ground,
crushed, and sieved before use. Alternate exchange media such as synthetic resins are
characterized by more favourable ion diffusion kinetics but they are inherently more
expensive and not selective to ammonium ions [19 - 21].
Consequently, modification of natural zeolites appears to be one research approach which
could improve zeolite performance. For example, Wang and Lin [22] reacted sodium
hydroxide with clinoptilolite and increased the ammonium exchange capacity from 97
meq/100 g to 275 to 355 meq/100 g. The enhanced capacity correlated with the synthesis
of zeolites Na-Y and Na-P. Another approach could be to create a new material with natural
zeolite as a core and an ammonium selective zeolite as the shell [Figure 5.1]. Core shell
structures have been shown to demonstrate improved properties in many catalyst
applications [23]. There are several studies of core-shell zeolite structures involving for
example coating a polymeric shell around the zeolite core to promote carbon dioxide
sorption in the presence of water [24] and zeolites with magnetic cores [25] as recyclable
biocatalysts. However, the production of a core-shell hybrid natural-synthetic zeolite has
not been previously disclosed, despite analogous core-shell resins exhibiting improved
operational capacity and reduced regenerant consumption [26].
Figure 5.1: Concept for Core-Shell Structure for Modified Natural Zeolites
121
The core-shell zeolite concept has the potential to create shorter diffusion paths for the
exchanging ammonium species which should enhance zeolite performance. In addition, the
use of natural zeolite as a silicon/aluminium source has benefits such as allowing synthesis
of Si:Al ratios higher than 1, which should reduce the alkalinity of the zeolite material. The
natural zeolite core would also provide excellent structural stability as it is a renowned
robust material.
Consequently, the challenge was how to optimally synthesise novel, core shell zeolites
materials and to determine if they do indeed offer advantages for ammonium ion removal
from solution. This study therefore examined bench and small pilot plant synthesis of core
shell zeolites to provide an insight to the impact of manufacturing conditions upon zeolite
composition and what issues needed to be addressed if the process was scaled up. In
addition, equilibrium tests were conducted for an ammonium chloride solution and a landfill
leachate sample.
122
5.2 Materials and Methods
5.2.1 Natural Zeolite
The natural zeolite used in this investigation was supplied by Zeolite Australia and originated
from the Werris Creek mine in New South Wales. The natural zeolite was crushed and pre-
sieved at the mine to size ranges of 0.5 to 1.0 mm and 1.0 to 2.0 mm. For this study a size
range of 0.5 to 2.0 mm was used, which was obtained by mixing equal parts (in mass) of
each of the supplied sizes. After mixing, the zeolite was washed with de-ionised (DI) water
to remove dust and very fine particles, then oven-dried overnight at 105°C. Note that in
some tests fines were still present i.e. natural zeolite was used without the washing and
drying stages being conducted.
5.2.2 Core-Shell Zeolite Synthesis
Zeolite N synthesis typically requires the use of solutions comprising of specified quantities
of potassium hydroxide and potassium chloride [17, 18]. Analytical grade reagents (Rowe
Scientific) were dissolved in deionised water in the order potassium hydroxide followed by
potassium chloride. The exothermic dissolution of potassium hydroxide promoted the
subsequent dissolution of potassium chloride. Then natural zeolite was added to give a
mass ratio of components of 1 NZ:1 KOH:1 KCl:5 H2O. The reactor was then sealed and the
heating and agitation initiated.
The agitation conditions (agitation rate; the agitation mode and the agitator type) of the
mixture in the reactor vessel were assumed to potentially influence the zeolite synthesis
procedure. Both static and continuously agitated methods were examined, including semi-
static modes in which the mixture was stirred intermittently. The reactors employed for
pilot plant studies (Parr) and bench evaluations (Berghof) were equipped with different
types of stirrer paddles. The former was an “x-type” while the latter was a dual, vertical
impeller [Figure 5.2]. A modification was made to the Parr agitator in order to potentially
enhance the extent of agitation in the reactor vessel [Figure 5.2].
123
(a) Bench top reactor
(b) Pilot Plant Reactor
(c) Modified Pilot Plant
Reactor
Figure 5.2: Illustration of agitator systems employed in bench and pilot plant reactors
Reaction temperatures to employ were based upon previous work by Mackinnon et al. [17,
18]. Non-hydrothermal reactions at 95 oC were previously conducted when using clay as an
aluminosilicate source [17], however, the reaction was considerably accelerated when
hydrothermal synthesis conditions were employed [18]. An upper limit of 200 oC was found,
whereupon the zeolite N was not stable and converted to kaliophilite (KAlSiO4). Low
temperature (95 oC) zeolite synthesis experiments were carried out in the open vessel
described in Figure 5.3 under continuous stirring. In this case, quantities of material used
were: natural zeolite 200 g (0.5 to 2.0 mm); potassium hydroxide 200 g; potassium chloride
200 g; water 1000 g.
Figure 5.3: low temperature synthesis reactor scheme
124
The stainless steel beaker used for this experiment had a total volume of 1200 mL and was
equipped with a perforated paddle for agitation of the solution. The solution temperature
was raised by means of a heating plate. Aluminium foil was placed over the reactor vessel
in order to minimise evaporation losses, albeit it was required to add approximately 100 mL
deionized water per day.
At the end of the selected reaction time, the reactor was switched off and left to cool at
room temperature overnight or until the internal temperature was below 80°C. The
product was unloaded from the vessel and washed with deionised water until the pH was
below 10. The washed product was then vacuum-filtered using a Büchner funnel with
additional washing, and oven-dried at approximately 105°C overnight.
In order to establish a case where all other experiments could be compared, multiple zeolite
batches were prepared at a temperature of 175˚C and a continuous stirring speed of 50 rpm
in the bench scale reactor unit. The reaction time was 400 minutes (6.7 h), and it was noted
that generally it took 40 minutes to reach the required process temperature of 175˚C. The
masses of components added were: natural zeolite 30 g (0.5 to 2.0 mm); potassium
hydroxide 30 g; potassium chloride 30 g; water 150 g. The base case process was repeated
in excess of 30 times and produced consistent results. Generally, each batch in this reactor
produced about 20 g of core shell zeolite. The base-case core shell zeolite (coded H1)
consisted of a homogeneous mixture of the products of 25 batches.
5.2.3 Zeolite Scale up
The behaviour of the zeolite manufacturing process at a larger scale was tested using a Parr
reactor as described in Table 5.1. The equipment was a floor stand unit with a pivotal
vessel, which could be used either in a vertical or a horizontal position. Table 5.1 compares
the bench and floor scale reactor details. Typically, quantities of material used were:
natural zeolite 240 g (0.5 to 2.0 mm); potassium hydroxide 240 g; potassium chloride 240 g;
water 1200 g.
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Table 5.1: Comparison between reactors used for zeolite synthesis
Bench reactor - Berghof® Floor stand reactor - Parr
Instrument Company
Capacity (ml) 200 1800
Vessel lining PTFE Stainless steel
Agitator type Central paddle x-shaped + PTFE baffles
Agitator material PTFE Stainless steel
Heating mode Heating block jacket
Water volume (mL) 150 1200
Table 5.2 summarizes the experiments, sample codes, and the conditions for each test.
Table 5.2: Scale up and agitation experiments
Sample Code Reactor Agitation mode Comments Agitator
H1 Bench (Berghof) Continuous Base case Central paddle
SH1 Bench (Berghof) Static - Central paddle
SH2 Bench (Berghof) Semi-static - Central paddle
IH1 Bench (Berghof) Intermittent 30 s/h - 50 rpm Central paddle
IH2 Bench (Berghof) Intermittent 30 s/2h - 50 rpm Central paddle
SP1 Medium (Parr) Static Horizontal none
SP2 Medium (Parr) Semi-static - X-shape
IP1 Medium (Parr) Intermittent 30 s/h - 10 rpm X extended
CP1 Medium (Parr) Continuous 5 rpm X extended
5.2.4 Process Optimization
Commercially, zeolites would be made using tap water and not ultra-pure water due to cost
considerations. Hence, in some tests deionised water was replaced with tap water.
Another parameter evaluated was the impact of pre-washing the natural zeolite sample. By
avoiding the pre-washing step, costs could again be reduced. However, fines present in the
natural zeolite as supplied would remain and thus may influence zeolite synthesis. The
natural zeolite initial ratio was increased to test the effect of treating more natural zeolite
with the same amount of liquor. The composition of the final liquor was analysed to allow
126
determination whether the effluent could be recycled; as a major cost in the process was
the potassium hydroxide and potassium chloride chemicals which were present in excess of
stoichiometric amounts. The effect of extended reaction time upon the zeolite formed was
also studied with the floor scale unit. A summary of the tests concerning zeolite synthesis
optimization is presented in Table 5.3.
Table 5.3: Details of experiments involving process optimization of zeolite synthesis
Process Optimization
Sample Code Key Parameter Change Agitation Reactor
SH3 Tap water Semi-static Bench
SHR1 Ratio: 1.3NZ:1KOH:1KCl:5H2O Semi-static Bench
SHU1 No pre-wash – fines included Semi-static Bench
CPt1 Reaction time: 13 h Continuous Standalone
5.2.5 Core Shell Zeolite Performance
Equilibrium isotherms were generated in order to assess the performance of the
synthesised core shell zeolite for ammonium removal from test solutions. Each isotherm
series comprised of twelve plastic bottles filled with increasing masses of zeolite, starting
from 0.0 g (control) and with 100 mL of the test solution. The test solutions were: (1)
ammonium chloride solution containing 250 mg/L of NH4+; and (2) landfill leachate
containing a mixture of competing cations and approximately 50 mg/L of NH4+ [Table 5.4].
Table 5.4: Ammonium chloride and landfill leachate test solutions compositions
Cation (mg/L) Ammonium Magnesium Calcium Sodium Potassium
Ammonium chloride 248.25 - - - -
Landfill leachate 54.3 102.7 53.8 536.8 215
The bottles were arranged in an incubator (Innova 42R) at 30°C with constant horizontal
agitation at 200 rpm. The pH and conductivity of each bottle with mixture were measured
before and after the equilibration time; the mixtures were allowed to equilibrate for 72
hours. After checking pH and solution conductivity, the solutions were syringe-filtered and
tested for their ammonium content by Kjeldahl distillation (VELP Scientifica UDK 149)
127
followed by titration with sulphuric acid and boric acid indicator. Finally, the concentrations
of the other cations were determined by ICP-OES. The mechanical stability of the core shell
zeolite was visually assessed at the end of the equilibration period.
5.2.6 Analysis
5.2.6.1 Inductively Coupled Plasma – Optical Emission Spectroscopy (ICP-OES)
The analysis of solutions used in this study was done by ICP-OES with an integration time of
0.15 seconds and 3 repetitions. The instrument was a VISTA-MPX CCD and the wavelengths
applied were Al (396.152), Ca (422.673), Mg (285.213) and sodium (589.592). The standard
for this analysis was certified by Australian Chemical Reagents (ACR) and it contained 1000
mg/L of aluminium, calcium, magnesium, and sodium. A Hamilton Auto-Diluter was used to
dilute the standard into a multi-level calibration curve.
5.2.6.2 X-ray Diffraction
A PANalytical X’pert PRO MPD X-ray diffractometer was used to collect x-ray diffraction
(XRD) patterns of the powdered samples. The powdered samples were micronized with a
corundum standard and prepared as discs onto an aluminium holder for analysis. The
following instrumental conditions were applied: Cu Kα radiation, 40 kV, 40 mA, between 3.5
and 75 degrees two theta. The patterns obtained were analysed using X’pert HighScore Plus
software, matching the obtained patterns with ICSD reference patterns.
5.2.6.3 Cation Exchange Capacity
Ammonium Cation Exchange Capacity (CEC) of zeolites was determined by systematically
saturating a sample of zeolite with 1 M NH4Cl and extracting NH4+ with 1 M KCl. The
ammonium concentration in the extracts was determined by the Kjeldhal method.
5.2.6.4 Particle Size Distribution
The particle size distribution tests were undertaken as follows. A mechanical sieve shaker
was used to separate the different particle sizes; the amount of particles in each size range
was determined by weight using an analytical balance with 0.001 g precision. The sieve
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shaker meshes ranged from 2.0 mm to 0.5 mm and were stacked in order of size, decreasing
from top to bottom. All sieves were pre-weighed before loading. After each loading the
sieve was shaken for 20 minutes; followed by lightly tapping the whole stack to ensure all
particles went through the meshes. The sieves loaded with the different sample sizes were
weighed again and the sample mass determined by subtracting the empty sieve mass.
5.2.6.5 Optical Microscopy
Optical images were collected using a Zeiss Axio Imager M2m microscope. Samples were
prepared by mounting in resin, cutting, and polishing to give cross sections of the hybrid
materials or imaged on a microscope slide without any modification.
5.2.6.6 Laser Ablation Inductively Coupled Mass Spectroscopy (LA-ICP-MS)
Cross sections of the hybrid materials were obtained via mounting the synthesised zeolites
in resin, cutting, and polishing the surface of the resin. Representative sections of the resins
were analysed using an Agilent 8800 Laser Ablation Inductively Coupled Plasma Mass
Spectrometer at rate of 3 µM sec-1 with a spot size of 10 µM, from the internal core through
the outer shell.
5.2.6.7 Scanning Electron Microscopy (SEM)
The microstructure of the samples are to be observed by scanning electron microscopy
(SEM / EDS) using a Zeiss Sigma VP Field Emission Scanning Electron Microscope.
5.2.6.8 Surface Area Analysis
The specific surface area and pore volume of the materials were measured at 77 K by a
TriStar 3020 instrument using a BET algorithm for data reduction and standard procedures
for adsorption and desorption of nitrogen. Degas conditions included: 10.0 K/min
temperature ramp with a target temperature of 303 K; 10,000 min hold time.
129
5.3 Results and Discussion
5.3.1 Core-shell Zeolite Production
Figure 5.4 shows the zeolite materials before and after modification. The natural zeolite
was clearly non-homogenous as evidenced by the variety of individual crystals which were
incorporated in each grain. This observation was consistent with published literature
regarding Australian natural zeolite which suggested the particles were actually a mixture of
clinoptilolite, quartz and amorphous material [11, 27]. Following modification, the material
appeared lighter in appearance and more uniform in colour.
Figure 5.4: Optical images of 0.5-2.0 mm particle size natural zeolite (left) and zeolite N core
shell material (right)
Inspection of cross-section images of not only the unmodified zeolite particles but also the
material formed after 6 h reaction are shown in Figure 5.5. It was noted that the
unmodified zeolite images revealed a random assortment of individual grains distributed
throughout the larger granules. Whereas, in contrast the material where it was attempted
to grow zeolite N displayed a distinct region around the large granules which was
presumably the zeolite N shell.
In order to explore the core-shell structure in more depth, SEM micrographs and EDS
analysis of the fresh and modified natural zeolite samples were conducted [Figure 5.6]. The
surface of the modified sample was decorated with trapezoidal prisms which were more
apparent in the higher resolution image presented. Mackinnon et al. [17] similarly observed
prismatic shapes when they made zeolite N from kaolin clay.
130
(a) Fresh Natural Zeolite
(b) Modified Natural Zeolite
Figure 5.5: Unmodified and zeolite N core shell material mounted in resin following 6 h
reaction
(a) Fresh Natural Zeolite (b) Modified Natural Zeolite
(c) High resolution image of Modified Natural Zeolite
Figure 5.6: SEM images of as received natural zeolite and modified natural zeolite
131
Energy Dispersive Spectroscopy (EDS) analysis also revealed distinct differences in the
surface composition of the fresh natural zeolite and modified zeolite samples, as
summarized in Table 5.5. Key observations for the modified zeolite included: the removal of
calcium and magnesium ions from the zeolite material; substantial incorporation of
potassium; reduction in Si/Al ratio; presence of chloride ions. The formula for zeolite N is
K12Al10Si10O40Cl2⋅8H2O when made from kaolin [17]. As such, the analysis of the modified
zeolite surface was generally consistent with zeolite N which had a slightly higher Si/Al ratio
than that recorded when the material was made from kaolin. A Si/Al ratio of 4.5:1 for the
fresh natural zeolite was in the expected range for clinoptilolite materials [28]. The
reduction in Si/Al ratio at the surface of the modified zeolite may indicate dissolution of
silica containing species such as quartz during the synthesis procedure. The K/Cl ratio of
5.9:1 was consistent with the octahedral geometry for zeolite N proposed by Christensen
and Fjellvag [29].
Table 5.5: Elemental composition of natural zeolite and modified zeolite from EDS
measurements
Apparent Atomic % Concentration
Element Fresh Natural Zeolite Modified Zeolite
O 52.61 27.17
Na 0.27 0.37
Mg 0.93 0
Al 9.04 14.93
Si 40.76 21.84
K 1.18 24.44
Ca 3.27 0
Fe 0.56 0.59
Cl 0 3.80
132
Figure 5.7: [001] View of Clinoptilolite Structure (|Na3K
3 (H
2O)
24| [Al
6Si
30O
72])
Figure 5.8: [001] view of Zeolite N Structure (|K10
(H2O)
8Cl
2| [Al
12Si
12O
40])
133
It was also found that EDS examination of “lighter” deposits observed in images of both
materials studied suggested a significant quantity of calcium phosphate was present. The
differences in the framework structures of Clinoptilolite and Zeolite N are shown in Figures
5.7 and 5.8.
The N2 BET surface area for the modified zeolite sample was 14.06 m2g-1 and the cation
exchange capacity was 188 meq/100 g. An interesting aspect of zeolite N is that the largest
pores are of dimensions which do not allow entry of nitrogen molecules. Consequently,
surface area measurements only relate to the external surface area of the material. The
CEC of the modified zeolite was significantly higher than the natural zeolite as received (65
meq/100 g) but much less than the CEC values of ca. 500 meq/100 g recorded for pure
zeolite N samples made from clay [18].
5.3.1.1 Impact of Synthesis Temperature
Initially, non-hydrothermal methods were employed to make the core shell zeolite material.
In this case an open reactor was used which operated at a temperature of 95°C. The
reaction was conducted for a period up to 9 days. Visually, it was noted that a significant
build-up of well adhered material clusters occurred on the corners of the reactor vessel,
which were removed daily. These clusters appeared to become harder as a function of
reaction time and may not have been so substantial with a more appropriate agitator design
that encompassed the entire width of the reactor vessel. Figure 5.9 revealed that the XRD
patterns of the materials made did not exhibit major changes during the 9 day reaction
period. Indeed, Table 5.6 showed that XRD quantitative analysis indicated that only a small
fraction of the clinoptilolite converted to zeolite N. Correspondingly, the cation exchange
capacity (CEC) of the samples collected over the duration of the experiment was relatively
small (< 50 meq/100 g) until zeolite N was found to be present in the material.
The core shell zeolite synthesis at temperatures below 100°C was unsuccessful. The
experiment demonstrated the need for excessively long periods of time to create Zeolite N.
In contrast, the synthesis of zeolite N from clay starting materials has been observed to
proceed considerably faster [17]. When using kaolin clay, reaction periods as low as 6 hours
134
were required to create a well-developed zeolite N structure. Clays readily dissolve in
concentrated caustic solutions whereas the more robust, tetrahedrally coordinated zeolite
materials do not. The base case sample (H1) which was conducted at 175 oC [Figure 5.9]
confirmed that hydrothermal conditions were required if core-shell structures were to be
made.
Figure 5.9: XRD patterns for materials made at non-hydrothermal conditions
Table 5.6: Open reactor zeolite synthesis results
Sample ID Reaction time
(days)
Quantitative XRD CEC (meq/100g)
H1 17.7% N (+W) 188.38
C1 1 Zeolite N - not present.
Present: Corundum, Quartz,
Albite, Sanidine, Clinoptilolite,
Amorphous.
30.97
C2 2 37.41
C3 3 46.06
C4 4 48.67
C5 5 47.26
C6 6 8% zeolite N 50.48
C7 7 11.04% zeolite N 125.43
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5.3.1.2 Agitation
The influence of the agitation mode employed was investigated relative to the base case
(Sample H1). In terms of the recorded CEC values [Table 5.7] the agitation mode did not
substantially influence the exchange capacity of the zeolite. There were some variations in
the amount of zeolite W and zeolite N detected.
Table 5.7: Summary of the results from zeolite modification using different agitation
approaches
Sample
Code
Agitation
mode
Agitator
type
CEC
(meq/100g)
Zeolite N
(%)
Zeolite
W
(%)
0.5-
2.0mm
(%)
H1 Continuous - Base case Central
paddle
188.38 17.70 12.12 93.17
Bench scale (Berghof)
SH1 Static Central
paddle
214.73 33.50 0 -
SH2 Semi-static Central
paddle
206.84 18.59 7.77 97.56
IH1 Intermittent 30 s/h - 50
rpm
Central
paddle
200.85 21.57 8.98 91.24
IH2 Intermittent 30 s/2h -
50 rpm
Central
paddle
211.10 16.63 4.90 -
Standalone scale (floor stand Parr)
SP1 Static – horizontal None 180.18 - -
SP2 Semi-static X-shape 89.62 - 93.83
IP1 Intermittent 30 s/h - 10
rpm
X extended 154.46 - 95.0
CP1 Continuous - 5 rpm X extended 162.35 - 74.69
At the bench scale, all agitation modes tested produced satisfactory yield results, with all
batches having over 90% of particles within the 0.5 to 2.0 mm range. However, at the larger
scale the different agitation modes were found to influence the product quality. A
repetition of the base case, with continuous agitation in the floor stand reactor, albeit with
136
a much lower agitation rate (5 rpm), produced significant quantities of undersize material as
demonstrated by the fact that only 74.69 % of the particles were within the desired range of
0.5 - 2.0 mm. In an effort to reduce attrition of the zeolite sample, intermittent agitation
was performed and this did indeed raise the yield of zeolite in the appropriate size range (95
% yield). Nevertheless, the CEC of the modified zeolite was found to have decreased to only
162 meq/100 g compared to 188 meq/100 g for the base case sample made at small scale.
The static test with the Parr reactor in a horizontal position was problematic. The
hypothesis behind this test was that an increased contact surface would be available
between the zeolite and liquor. However, in the absence of motion, the zeolite grains
clustered together forming one single block [Figure 5.10]. The cohesive zeolite cluster
prevented the contact of the liquor with the central particles, thus leading to a lack of
conversion of the natural zeolite material.
Figure 5.10: Zeolite block formed during use of the floor scale reactor in a horizontal
configuration
The semi-static cycle (SP2) also had problems with clustering, but the end product was not a
single block, instead a number of separate clusters were present. Since agitation was
applied at the end of the cycle it was most likely the case that if a single block was formed, it
was broken at this stage into smaller clusters. In summary, intermittent agitation appears
to be the best approach when making modified zeolite at larger scale. Further research
should be addressed to optimising the agitation period and intensity.
137
Insight as to the chemistry which occurred during the various agitation tests was gained by
examination of the XRD traces for the materials made. Figure 5.11 and Table 5.8 show that
at bench scale static conditions actually resulted in the largest amount of zeolite N and
zeolite W. This observation correlated with the fact that the recorded CEC value was largest
for this sample, suggesting that these zeolites had improved capacity for ammonium ions.
Figure 5.11: XRD for bench scale agitation series - H1 is the base case
Generally, for the bench scale synthesis it could be said that static, intermittent and
continuous agitation are all satisfactory in terms of cation exchange capacity and yield, with
all CEC values above 200 meq/100g and particle size yields above 90 %. For this case, an
intermittent agitation would be the best approach as it ensures homogeneity in the product
while maintaining resource efficiency with minimal agitation. It is proposed that the size of
the vessel in relation to the volume of zeolite is the main reason for the good results under
static or semi-static conditions for the bench scale.
138
Table 5.8: Analysis of products formed for bench scale agitation series
Phase Name Wt% in sample
SH1 H1 IH1 SH2 IH2
Sanidine Na0.07 14.5 6.189 12.646 4.532 0.879
Zeolite N 33.5 17.699 21.57 18.59 16.632
Clinoptilolite 0.5 - - - -
Microcline maximum 1.6 - 4.106 3.017 11.348
Quartz 0.6 1.648 1.181 2.033 1.076
Muscovite 2M1 5.7 6.453 - - -
Albite high K0.16* 7.4 2.123 4.096 - -
Zeolite W - 12.159 8.984 7.768 4.903
Anorthite - 2.074 8.129 9.152 9.749
Andesine An50 C-1structure - 4.179 - 3.202 -
Ca montmorillonite - 1.752 3.436 2.64 -
Non-diffracting/unidentified 36.2
Figure 5.12: XRD for stand unit (Parr) - H1 is the base case
Figure 5.12 also revealed differences in the quality of the material made, with overall less of
the desired zeolite N and zeolite W formed.
139
The distribution of the zeolite in the vessel is such that the depth of the media is small
enough for the potassium solution to penetrate through virtually every void, effectively
saturating all of the natural zeolite. In contrast, with the standalone reactor the depth of the
zeolite bed is proposed to prevent the diffusion of the solution during the reaction time.
Upon loading the Parr reactor, the solution was loaded first and the zeolite added in a
progressive manner so as to ensure that all the particles were in contact with the solution
before they settled. However, it was apparent that once the zeolite settled, the particles
were too closely packed to allow the liquor to diffuse.
Moving on to a mid-point between static and agitated processes, the intermittent condition
proved a better solution. With agitation applied for 30 seconds at hourly intervals, the
product still presented some clusters; similar to the static, but these were much smaller and
dispersed. This observation indicated that an incremental increase in agitation (either
longer agitation periods or shorter intervals between them) would eventually avoid the
formation of clusters. The particle sizing of the product under intermittent conditions was
the best possible, with over 95% of the grains within the 0.5 -2.0 mm size range.
5.3.1.3 Process Optimization
Table 5.9 revealed some interesting insights in regard to the influence of process
parameters for zeolite synthesis. The experiment with tap water (SH3) was found to
increase the cation exchange capacity of the modified zeolite material to 219 meq/100 g.
This latter result suggested that the use of deionized water was not a requirement for
zeolite synthesis. Given the large quantity of potassium salts added to the water used in the
zeolite synthesis step it is perhaps not surprising that small amounts of dissolved species did
not detrimentally impact the process. Whether the presence of ions such as sodium,
calcium, and magnesium which are found in tap water promote the synthesis of modified
zeolite is not proven as yet. It is known from previous studies that zeolite N can be made if
small quantities of sodium ions are present in the starting mixture, but that when large
quantities are present the formation of zeolite N is hindered [17, 18]. The use of a larger
ratio of natural zeolite to the potassium salts appeared to slightly diminish the CEC value
(170 meq/100 g). In harmony, the corresponding XRD pattern indicated a lower percentage
140
of zeolite N in the product [Figure 5.13]. Pre-washing the natural zeolite had minimal
impact upon the modified zeolite quality.
Table 5.9: Process Optimization Results
ID Change Agitation CEC (meq/100g)
H1 Base case Continuous 188.38
SH3 Tap water Semi-static 219.15
SHR1 Ratio: 1.3NZ:1KOH:1KCl:5H2O Semi-static 170.56
SHU1 No pre-wash – fines included Semi-static 204.95
Figure 5.13: XRD patterns for samples prepared during process optimization tests
In addition to the starting conditions, the process efficiency can be increased by reducing
waste and disposal costs. In particular for this synthesis, the mother liquor left after the
zeolite synthesis process was extremely alkaline and contained a significant concentration of
metal cations; therefore, it requires treatment prior to disposal. Elemental analysis of this
liquor and of the initial solution were compared in order to identify the effect of the process
on the solution composition as well as to study avenues for dealing with this component.
141
The heat produced in the synthesis reaction can also have value as it can be reused in the
system. Figure 5.14 is an example of such use within the process flow: the hot mother
liquor is recirculated to a heat exchanger and reused to aid in pre-heating the next batch.
Figure 5.14: Process flow scheme - illustrative of liquor reuse option
ICP results [Figure 5.15] compare the metal cations concentration on the initial solution
(MZ0), which is simply a potassium solution, with that of the final solution, i.e. the post-
reaction liquor. The potassium left in the solution indicated that the concentration of this
cation in the beginning might be excessive. Approximately half of the initial concentration
of that cation was found in the end liquor, suggesting that use of different solution
conditions for the zeolite manufacture may be more economic. For the resource efficiency
in this aspect, the reuse of this liquor for the next batch would mean simply completing the
required potassium load, rather than starting from zero.
142
However, aluminium, silicon and sodium are also present in the final liquor in elevated
concentrations. This observation may indicate dissolution of some of the materials present
in the zeolite sample. Therefore, the recycling or disposal of the final liquor needs
consideration of the effects of these outlined contaminants.
The analysis of the end liquors in relation to the starting solution also revealed a
consistently lower load of cations in the liquor left after a semi-static process in relation to
the agitated batches. The lower cation load was suggestive that the majority of the cations
in the liquor were derived from the zeolite breakage and eventual dissolution, in particular
with the case of silicon.
Figure 5.15: Analysis of solution compositions for zeolite synthesis tests
143
5.3.2 Modified Zeolite Performance
5.3.2.1 Ammonium Chloride solution – 250 ppm NH4+
The pH of the solution-zeolite mixture was relatively high even before equilibration, as can
be seen in Figure 5.16. Zeolites are best thought of as salts of weak acids, and as such are
susceptible to hydrolysis in water. Materials of low Si/Al ratio are especially prone to
hydrolysis in solution according to the following reaction:
Experience with zeolites such as sodium exchanged zeolite A has shown that excessive
washing with water can result in up to 15 % of the sodium ions being replaced by H3O+ ions.
Breck et al. [30] reported that the hydrolytic equilibrium could be reversed by addition of
small amounts of sodium chloride and consequently a pH of 6.0 to 6.5 was recorded
(compared to the original value of pH 10.0 to 10.5). Hence, the modified zeolite which had
a Si/Al ratio of less than 2 was expected to exhibit a high solution pH in solution. The issue
with high solution pH was that the dominant form of ammoniacal nitrogen would be
ammonia rather than ammonium. Ion exchange ability was thus predicted to be limited as a
gas cannot be removed by cation exchange. Figure 5.16 supported this latter conclusion as
ammoniacal nitrogen concentration was only reduced by a maximum of 38.82 %, using a
modified zeolite with a CEC of 154.46 meq/100g (sample IP1).
Figure 5.16: pH and ammonium concentrations for ammonium exchange with modified zeolite (sample IP1)
144
In addition to issues with high solution pH, the modified zeolite presented very low
mechanical resistance. In this test, by the end of the 72 hours of agitation the solution had
substantial suspended solids and required extensive filtering prior to analysis. The same
experiment was carried out for natural zeolite and modified zeolites and the breakage with
the former sample was significantly less than that observed with the modified zeolite
sample.
5.3.2.2 Pre-RO landfill leachate – 50 ppm NH4+
A modified zeolite sample of the same batch (IP1) was equilibrated with a pre-RO treated
landfill leachate solution with the composition shown in Table 5.4. The Langmuir Vageler fit
of the data shown in Figure 5.17 (a) suggested that the maximum uptake of ammoniacal
nitrogen on the modified zeolite sample was 0.41 mol/kg. Notably, this was substantially
greater than the corresponding value of 0.07 mol/kg we previously observed for natural
zeolite. Hence, the presence of the synthetic zeolite N shell appeared to have promoted the
uptake of ammonium ions from solution.
However, examination of the isotherm profile in Figure 5.17 (b) revealed interesting
behaviour which was not recorded when an unmodified natural zeolite was used. It was
apparent that as the zeolite mass used in the equilibrium tests was increased Ce initially
decreased to ca. 21.5 mg/L and then increased to 37.6 mg/L [Figure 5.17 (b)]. This latter
behaviour can be rationalised if one considers the changing pH of the exchanging solution
[Figure 5.17 (c)]. It was evident that as the mass of zeolite added increased, the solution pH
significantly increased to highly alkaline values. At elevated pH, the equilibrium between
ammonium and ammonia species was such that ammonia became the dominant entity in
solution. As such ammonium ions available for ion exchange were greatly reduced in
concentration.
145
(a)
(b)
(c)
Figure 5.17: Ammonium removal from pre-RO leachate solution by core-shell zeolite
(sample IP1)
5.4 CONCLUSIONS
It was deduced from this study that the production of a modified zeolite was possible but
that careful consideration of process parameters such as reactor heating type, vessel
construction and stirrer design was required. Static and semi-static conditions gave good
results at bench scale (214.73 and 206.84 meq/100g) but that was not repeated on the
larger scale, where the formation of clusters prevented the diffusion of the liquor through
the zeolite particles resulting on a CEC as low as 89.62 meq/100g. Thus the intermittent
agitation mode was deemed the best option for this latter situation, giving satisfactory CEC
and size results (154.46meq/100g, with 95% of particles within the size range). On this
note, the agitator model and design influenced the product’s quality and quantity and the
optimum type has not been defined as yet. However, agitator requirements such as
ensuring that the side and bottom surfaces of the vessel are swept by the agitator to
146
prevent cluster formation, may be recommended (e.g. the modified stirrer on the Parr
reactor).
It was similarly evident that the temperature was important, with non-hydrothermal tests
(<100˚C) probably not feasible when using a natural zeolite as a precursor material. There
was also scope to minimize costs of core shell zeolite synthesis by reducing the excessive
amount of potassium species in the starting mixture.
The modified zeolite performance proved to be limited by the material’s high pH (over 10)
and its weak mechanical resistance. Although the ammonium exchange capacity was shown
to be much higher than that of natural zeolite (188.38 meq/100g compared to 65 meq/100g
for natural zeolite), the application of this innovative product may be limited unless
improvements are made in regards to use of a higher Si/Al ratio (to lessen pH of the
solution) and incorporation of a binder to inhibit sample degradation.
5.5 ACKNOWLEDGEMENTS
Zeolite Australia is thanked for the funding and provision of zeolite materials for this
research. We are grateful to Mr. Kenneth Nuttall for conducting the optical and electron
microscopy measurements. The financial and infra-structure support of the Energy and
Process Engineering Discipline of the Science and Engineering Faculty and the Institute of
Future Environments, Queensland University of Technology is gratefully acknowledged.
5.6 REFERENCES
[1] A. Mojiri, H.A. Aziz, N.Q. Zaman, S.Q. Aziz, A review on anaerobic digestion, bio-reactor
and nitrogen removal from wastewater and landfill leachate by bio-reactor, Advances in
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Chapter 6: Conclusions and Future Work
Water resources are ever scarcer with increase in global population, manufactured goods,
and food and agriculture. Also increasing with the expressively increase in urban population
is the wastewater produced, e.g. municipal and industrial wastewaters, landfill leachate,
contaminated stormwater runoff and so on.
Zeolites are remarkable materials and their properties are used in several different
industries, whereby the most prominent field is the treatment of wastewaters. The ion
exchange capacity of zeolites is the underlying reason for its suitability for these processes,
as it allows the zeolite to selectively remove contaminants from effluent streams. They are
cheap, robust and naturally available world-wide. Moreover, zeolites can be regenerated
after being exhausted, the contaminants can be harvested for reuse, and the treatment can
continue with the same media.
The application of zeolites in stormwater runoff filters was explored with emphasis on the
hydraulic properties of zeolites regarding filter design. A filter design with layers of different
materials was the base point for the determination of hydraulic conductivity in zeolite and
laterite media.
Natural zeolite and its sodium and acid modified forms were used for landfill leachate
treatment to meet stringent ammonium discharge limits. The performance of the zeolites
was compared with that of commercial SAC resins, also in sodium and acid forms. Isotherms
and column studies were performed using real and synthetic landfill leachates, before and
after reverse osmosis treatment.
Finally, the production of a new zeolite consisting of a zeolite N shell with a natural zeolite
core was tested under the process perspective with the goal of determining the scalability
of the process to eventual commercial quantities. The operational performance of the core
shell zeolite was tested by isotherm equilibria using real landfill leachate and a synthetic
single-cation solution. The constant agitation employed on the isotherm test also provided
input on the mechanical strength of the material.
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6.1 Conclusions
Natural zeolites have been a prolific field of study since their discovery and it is apparent
that at times, assumptions were made or new discoveries took the interest away from one
topic or another, leaving zeolite literature permeated with gaps. Often new studies present
deeper understanding of these minerals and identify new gaps to be filled.
It was evident in this program of research that methods for the determination of hydraulic
conductivity do not necessarily encompass all conditions and applications. In particular the
case of open columns, as opposed to sealed permeameters, was found to be overlooked in
standard test methods. However, the use of this configuration is typical in filters and in
column studies. The variability found around testing methods and systems was substantial,
implying that the accuracy of results may be tainted by the choices of the investigator.
Application of different reference points for the Δh parameter, i.e. the tailwater and the
media top, produced significantly distinct results. Notably, the Δh at the tailwater method
was insensitive to variations in bed height. Meanwhile, the Δh at the media top method
revealed marked changes, such as an increase of 68.31% in hydraulic conductivity and
reduction of 10.62% in flow rate, in response to a raise of 14.56% in bed height.
Zeolites are commonly used in filters and the hydraulic conductivity is a guiding parameter
in the design of these devices. The gaps and conflicting literature on the topic were perhaps
the most important outcome of this study.
The treatment of landfill leachate was found to be effective in a column filter arrangement
using sodium exchanged resin, although the resin has not shown selectivity for ammonium
in a mixture of competing cations. During the entire treatment of 400 BV, the Na+ resin
removed NH4+, Ca2+, Mg2+ and K+ alike to < 0.1mg/L, while desorbing Na+ ions. The Na+
zeolite on the other hand, showed high selectivity; loading NH4+, K+ and Mg2+, and desorbing
Ca2+ and Na+; but lesser efficiency, which is associated to intraparticle diffusion limits on the
zeolite and not on the resin. However, the zeolite performed well and is a valid option
considering cost effectiveness. It is also worth noting that the modifications of natural
zeolite were found to be flawed. In the case of the sodium exchange, not all of the zeolite
sites were completely loaded with sodium, as would be case for a zeolite that has been
regenerated sufficient times in loading and regeneration cycles. The acid zeolite exchange is
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slightly more complex, whereby the excessive exposure to acid may remove aluminium from
the zeolite framework. The superior efficiency of the resin, although not selective, makes
this material a direct competitor for zeolites.
The use of synthetic zeolites is another case in which efficiency may be superior to natural
zeolites but production costs still favour the natural option. Thus the creation of a core shell
zeolite with the production cost closer to the natural material and the efficiency similar to
synthetic materials is an interesting proposition.
The production of core shell zeolite has been proven to be scalable with due adjustment of
equipment and reaction conditions, giving best results for a batch under intermittent
agitation with a modified stirrer on a scale eight times larger than the initial bench reactor.
The efficiency of the process has also shown good response to the preliminary assessment
with production yielding up to 95% of particles within the size range and a CEC up to
154.46meq/100g; these results are the baseline for optimisation of the process. The core
shell zeolite’s operational performance was found to be limited by two main factors:
chemically, the high pH of the material (consistently over 8) prevents the high removal
efficiency, and mechanically, the lower density causes the zeolite to break easily under
agitation or pressure. Nonetheless, the concept is valid and prosperous and the applications
of the new material may amplify the range of natural zeolite uses.
Overall, this research program has advanced the knowledge of natural zeolites regarding
their possible applications and has shown that innovations are promising in the area of
zeolite modifications.
6.3 Future research and recommendations
With the conclusions that hydraulic conductivity values are influenced by parameters not
previously discussed, it is clear that a further study of each of these parameters individually
is needed. The bed height effect in particular has been studied in terms of its influence in
contaminant removal and clogging of media, but not on hydraulic conductivity as such.
Another beneficial study for the understanding of hydraulic conductivity is the comparison
of open column and permeameter systems, as well as constant and falling head systems for
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the same media. In particular for zeolites, the assessment of hydraulic conductivity for a
range of particle sizes larger than 0.5 mm and smaller than 10 mm is required.
In light of the hydraulic conductivity variations found when using different methods to
determine its value, it is highly recommended that the values of flow rate are seen with
more weight than those of hydraulic conductivity. It is extremely important to keep in mind
when designing filters that the relationship between lab and field (or large scale) hydraulic
conductivity results has not been established.
The ion exchange treatment of landfill leachate raised questions about the zeolite
modifications. Thus, a study where the zeolite modifications are better controlled is a clear
path in the future. In addition the study presented went as far as to breakthrough and the
effects of regeneration cycles were not investigated, thus a longer study including various
loading and regeneration cycles is required.
Another important investigation is the use of real leachate on a column filter, preferably
testing leachate from different landfills.
The future of the core shell zeolite appears to be promising and not too simple. In terms of
the manufacturing process, future research includes process efficiency in terms of resource
consumption and by-product recycling, and in terms of ratio of reactants. The core shell
material itself requires further development of its performance as an ion exchanger and
exploration of other applications. Additionally, the mechanical performance of the core shell
zeolite will be a defining parameter for most of its applications therefore a full assessment
regarding the operational performance of this new material is required. Future research can
also include applications for the unavoidable fines generated from agitation, such as land
remediation.