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Defluoridation of drinking water using adsorption processes
Accepted Manuscript Title: Defluoridation of drinking water using adsorption processes Authors: Paripurnanda Loganathan, Saravanamuthu Vigneswaran, Jaya Kandasamy, Ravi Naidu PII: DOI: Reference:
S0304-3894(12)01211-3 doi:10.1016/j.jhazmat.2012.12.043 HAZMAT 14806
To appear in:
Journal of Hazardous Materials
Received date: Revised date: Accepted date:
2-8-2012 18-12-2012 26-12-2012
Please cite this article as: P. Loganathan, S. Vigneswaran, J. Kandasamy, R. Naidu, Defluoridation of drinking water using adsorption processes, Journal of Hazardous Materials (2010), doi:10.1016/j.jhazmat.2012.12.043 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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Research Highlights
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! ! Comprehensive and critical literature review on various adsorbents used for defluoridation ! ! pH, temperature, kinetics and co-existing anions effects on F adsorption
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! ! Choice of adsorbents for various circumstances
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! ! Adsorption thermodynamics and mechanisms
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! ! Future research on efficient, low cost adsorbents which are easily regenerated
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Defluoridation of drinking water using adsorption processes
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Paripurnanda Loganathan1, Saravanamuthu Vigneswaran1*, Jaya Kandasamy1, Ravi Naidu2 1
Faculty of Engineering and Information Technology, University of Technology, Sydney, NSW, 2007, Australia
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CRC CARE, University of South Australia, Adelaide, SA 5095, Australia
*Corresponding author: ph: 612 9514 2641, E-mail: [email protected]
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ABSTRACT
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Excessive intake of fluoride (F), mainly through drinking water, is a serious health hazard
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affecting humans worldwide. There are several methods used for the defluoridation of
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drinking water, of which adsorption processes are generally considered attractive because of
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their effectiveness, convenience, ease of operation, simplicity of design, and for economic
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and environmental reasons. In this paper, we present a comprehensive and a critical literature
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review on various adsorbents used for defluoridation, their relative effectiveness, mechanisms
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and thermodynamics of adsorption, and suggestions are made on choice of adsorbents for
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various circumstances. Effects of pH, temperature, kinetics and co-existing anions on F
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adsorption are also reviewed. Because the adsorption is very weak in extremely low or high
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pHs, depending on the adsorbent, acids or alkalis are used to desorb F and regenerate the
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adsorbents. However, adsorption capacity generally decreases with repeated use of the
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regenerated adsorbent. Future research needs to explore highly efficient, low cost adsorbents
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that can be easily regenerated for reuse over several cycles of operations without significant
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loss of adsorptive capacity and which have good hydraulic conductivity to prevent filter
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clogging during the fixed-bed treatment process.
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1. Introduction
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2. Adsorption mechanisms
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3. Factors influencing adsorption
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3.1. pH
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3.2. Co-existing anions
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3.3. Temperature
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3.4. Adsorption kinetics
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Keywords: adsorption, defluoridation, fluoride, fluoride toxicity, water treatment
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4. Adsorbents 4.1. Metal oxides and hydroxides
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4.2. Layered double hydroxides
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4.3. Ion exchange resins and fibres
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4.4. Zeolites
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4.5. Carbon materials
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4.6. Natural materials
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4.7. Industrial by-products
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5. Adsorption thermodynamics
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6. Fluoride desorption and adsorbent regeneration
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7. Conclusions
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8. Acknowledgements
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9. References
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1. Introduction
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Fluoride (F) has beneficial effects on teeth at low concentrations in drinking water (0.4-
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1.0 mg/L), especially for young children in that it promotes calcification of dental enamel and
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protects teeth against tooth decay. Excessive levels of F on the other hand can cause a
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number of problems ranging from mild dental fluorosis to crippling skeletal fluorosis as the
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level and period of exposure to F increases [1]. The World Health Organisation [2] has
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recommended a guideline value of 1.5 mg/L as the concentration above which dental
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fluorosis is likely. However, it is also important to consider climatic conditions and quantity
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of water intake and other factors such as F intake from certain diets when establishing F
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limits. Water consumption in hot humid regions is generally higher than in temperate regions
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and therefore the F concentration limit for likely fluorosis should be lower. For example, the
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US Public Health Service [3] has recommended that the upper limit for F concentration in
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drinking water should be decreased from 1.7 to 0.8 mg/L with increases of the average
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maximum daily air temperature.
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High F intake has been suspected being involved in a range of adverse health problems in
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addition to fluorosis, including cancer, impaired kidney function, digestive and nervous
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disorders, reduced immunity, Alzheimer’s disease, nausea, adverse pregnancy outcomes,
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respiratory problems, lesions of the endocrine glands, thyroid, liver and other organs [1,4-9].
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However, there appears to be no convincing evidence for F being directly involved in causing
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these conditions [1].
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Elevated concentration of F in drinking water is due to its natural occurrence or industrial
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activities. Many rocks and minerals in the earth’s crust contain F [10,11] which can be
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leached out by natural weathering and rainwater, causing F contamination of surface and
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ground waters. Besides this natural source, F also enters the water bodies from waste waters
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produced from industries such as aluminium, steel, glass, semiconductors, electronic, tooth
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paste, fertiliser and insecticide manufacturing plants [9,12-15]. 4
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Fluorosis due to excessive concentration of F has been reported in at least 28 countries
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from South Asia, Africa, the Middle East, North, Central and South America, and Europe [1].
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In India, it was estimated that 56.2 million people were affected by fluorosis and this problem
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was prevalent in 17-19 out of the 32 States [16,17]. The major source of F in a majority of
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countries is rocks and minerals such as fluorospar, cyyolite, and fluorapatite containing F
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[17]. For example, Choi and Chen [18] reported extremely high F concentration (> 1000
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mg/L) in surface water in areas with F-rich volcanic rocks.
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One method of reducing excessive concentrations of F in water is to blend water having a
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high F concentration with water that has a low F concentration from an alternative source. If
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such a source is not readily available, defluoridation is the only means remaining to prevent
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fluorosis [1]. For contaminants other than F, water treatment methods are used to remove
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contaminants to below the maximum level permissible but defluoridation is special in that the
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treated water should have an optimum F concentration to derive the beneficial effects of F.
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The main methods of defluoridation are precipitation/coagulation, adsorption, ion
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exchange, reverse osmosis, and electrodialysis. Of these, precipitation/coagulation and
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adsorption are convenient methods and are widely used, especially in developing countries’
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rural areas. The scale and treatment site differ between industrialised countries and
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developing countries. In industrialised countries the treatment of water is generally performed
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at water treatment plants close to the water source but in developing countries it is carried out
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at a village community level or at a household level [1] using simple inexpensive locally
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available adsorptive media [9,19-21]. Industrialised countries generally use more efficient but
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more costly adsorption media including synthetic ion exchange resins as well as advanced
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techniques such as reverse osmosis and electrodialysis.
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The advantages and shortcomings of the various methods of defluoridation are presented
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in Fig. 1. Of these various methods that of adsorption is generally considered attractive 5
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because of its effectiveness, convenience, ease of operation, simplicity of design, and
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economic and environmental considerations, provided low-cost adsorbents which can
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effectively remove F around the neutral pH of drinking water are used. The
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precipitation/coagulation method where lime and Al salts are used to remove F as a CaF2
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precipitate followed by removal of left over F in solution by co-precipitation with and
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adsorption on to the precipitated Al(OH)3 is further developed into a technique called
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Nalgonda process in India [17,24,25]. This process is extensively used in India and Africa
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[1]. However, the main drawback of this technique is the low effectiveness of F removal and
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production of toxic AlF complexes in solution [8].
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Previous reviews on defluoridation of water presented timely information on several
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adsorbents used for defluoridation, but they generally did not focus on the chemical
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mechanisms and solution factors influencing the adsorption processes, and regeneration of
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the adsorbents [1,17,26,27]. The objective of this paper is to compile and present current
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information on the potential of the various adsorbents used for defluoridation of drinking
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water, their relative effectiveness, mechanisms and thermodynamics of adsorption, factors
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influencing adsorption and methods of adsorbent regeneration for reuse. Based on the review
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adsorbents are selected for specific circumstances.
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2. Adsorption mechanisms
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The capacity, energy and kinetics of adsorption of F by adsorbents are controlled by the
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mechanism of adsorption. Understanding these mechanisms can provide useful information
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on the optimisation of the adsorption process in water treatment plants and the subsequent
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desorption/adsorbent regeneration process for reuse. There are mainly five mechanisms of F
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adsorption, namely: (1) van der Waals forces (outer-sphere surface complexation, (2) ion 6
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exchange (outer-sphere surface complexation), (3) hydrogen bonding (H-bonding) (inner-
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sphere surface complexation), (4) ligand exchange (inner-sphere surface complexation), and
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(5) chemical modification of the adsorbent surface. The first two mechanisms are governed
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by weak physical adsorption and are non-specific to F, whereas the third and fourth
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mechanisms are governed by strong chemical adsorption specific to F. The fifth mechanism
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is governed by both specific and non-specific adsorption. In the presence of the other anions
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in water, F cannot be easily removed by adsorbents using the first two mechanisms. Fluoride
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adsorption resulting from the third and fourth mechanisms selectively removes F from water
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in the presence of most types of anions; only anions (e.g. phosphate) which also adsorb
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specifically on the adsorbent compete with F for adsorption sites.
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Van der Waals forces are weak short range forces acting between two atoms. The larger
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the adsorbate size the greater the force of attraction. Therefore adsorbates with high
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molecular weights such as dissolved organic matter are adsorbed on adsorbents having high
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surface area through van der Waals forces. This is the reason for the weak adsorption of F
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and strong adsorption of dissolved organic matter on activated carbon [28]. Fluoride was
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considered to be adsorbed on manganese oxide-coated alumina by van der Waals forces at
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high pH [29].
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Ion exchange is a stoichiometric process where any counter ion leaving the ion exchanger
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surface is replaced by an equivalent number of moles of another counter ion to maintain
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electro-neutrality of the ion exchanger. The ions are adsorbed physically by fully retaining
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their inner hydration shell and the adsorption is due to electrostatic or Coulombic attraction.
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The ion exchange process is rapid and reversible. Fluoride removal by ion exchange resins
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[30] and ion exchange fibres [31] is mainly governed by the ion exchange mechanism as
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illustrated in Fig. 2a. Ion exchange tends to prefer counter ions of higher valency, higher
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concentration and ions of smaller hydrated equivalent volume [32]. Therefore F removal 7
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using ion exchange resins is difficult because the order of selectivity for anions by anion
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exchange resins is as follows: citrate > SO42- , oxalate > I- > NO3- > CrO42- > Br- > SCN- > Cl-
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> acetate > F- [32]. H-bonding is a strong dipole-dipole attractive force between bonding of the strong
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electropositive H atom in a molecule in an adsorbent or adsorbate and a strong
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electronegative atom such as oxygen or fluorine in another molecule [33]. The energy of
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adsorption in H-bonding is stronger than in van der Waals forces and ion exchange but
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weaker than in the ligand exchange process discussed in the next paragraph. H-bonding
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occurs in the adsorption of F on ion exchange resins [30,34] and coal-based adsorbents [35]
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as illustrated in Fig. 2b.
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In a ligand exchange mechanism, the adsorbing anion such as F- forms a strong covalent
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chemical bond with a metallic cation at the adsorbent surface resulting in the release of other
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potential determining ions such as OH- ions previously bonded to the metallic cation (Fig.
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2c). Thus, F is said to form an inner sphere complex or is specifically adsorbed on the
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adsorbent surface. The adsorption of F on several multivalent metal oxides near neutral pH
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was reported to have increased the pH of solutions as a result of release of OH- ions from the
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adsorbents by ligand exchange of OH groups on the adsorbent surface with F in solution
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[36,37]. Adsorbents with a ligand exchange mechanism have the particular advantage of
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combining high adsorption capacity with high selectivity for anions. These adsorbents can
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remove large proportions of anions having higher selectivity for adsorption from very dilute
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solutions of the anions even in the presence of higher concentrations of competing anions of
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lower selectivity. Adsorption of F by ligand exchange is illustrated in Fig. 2c [23,37,38].
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The adsorption capacity of F on adsorbents can be increased by chemical modification of
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adsorbent surfaces (Fig. 2d). This is particularly of advantage in the case of adsorbents
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possessing negative surface charges which tend to repel the similarly charged F- ions. In such 8
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adsorbents, positively charged multivalent cations such as Al3+, La4+, Zr4+, Fe3+, and Ce3+ are
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impregnated onto the adsorbent to create positive charges on the adsorbent surface for
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attracting F- by coulombic forces as well as producing adsorption sites capable of chemical
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interaction with F [13,23,39-44] (Fig. 2d). These metallic cations act as a bridge in adsorbing
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F onto the adsorbent.
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Adsorption of F on many microporous adsorbents is recognised as a 2-step process; an
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initial rapid adsorption (mostly within an hour) at the outer surface of the adsorbent that
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reaches a pseudo-equilibrium at the solid-solution interface followed by a much slower
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process (hours to days) where the F moves by intra-particle diffusion into the interior pores
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and channels of the adsorbent [45]. Intra-particle diffusion rate is directly related to the
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square root of time of adsorption. Therefore if a straight line relationship is obtained between
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the rate of adsorption and square root of time with the line passing through the origin, it can
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be inferred that the diffusion process controls the adsorption, especially at longer times. Such
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a relationship was obtained in many studies for the adsorption of F on different adsorbents
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(granular ferric hydroxide [7], manganese oxide-coated alumina [29], fly ash [46], activated
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alumina [47], and alum sludge [48]).
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It could be concluded that ligand exchange is the predominant mechanism of F adsorption
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for inorganic adsorbents having high adsorption capacities. For organic adsorbents having
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high adsorption capacities, H-bonding seems to be the predominant mechanism.
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3. Factors influencing adsorption
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3.1. pH
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As for other contaminants, the removal of F from water by adsorbents is influenced
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by several factors including pH, co-existing ions, temperature, adsorption kinetics, and 9
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adsorbent particle size and activation. Of these, pH is generally considered to be the most
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important factor [16]. Fluoride adsorption is low at both very low and very high pH. The pH
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at which the maximum amount of F is removed depends on the adsorbent characteristics but
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generally it is between 4 and 8 (Table 1). One of the important properties of the adsorbent
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influencing the extent of F adsorption is the pH at the point of zero charge (PZC). For
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example, Fe and Al oxides having a PZC at around 7-8 remove the maximum amount of F at
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pH 6-8 [7,47,49] and activated carbon with a PZC of 3.9-4.7 was reported to have removed
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the maximum amount of F at pH 3-4 [42]. At pH values above the PZC the surface of
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adsorbents is negatively charged, therefore the negatively charged F ions are not attracted to
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the adsorbent surface. At pH values lower than the PZC the surface is positively charged,
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therefore F ions are adsorbed. In some situations, at low pH values, F exist as positively
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charged AlF complexes and this can reduce F adsorption [11,49,75].
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The main reason for a reduction in F adsorption below pH 4 is that F forms weakly
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ionised HF at these pH values [13,47,48,50]. At a pH above 7-8 the removal of F decreases
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not only because the adsorbent surface becomes negatively charged but also because the
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concentrations of hydroxyl, bicarbonate and silicates ions increase so that they compete with
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F for adsorption. In addition to increased number of positive surface charges at low pH values
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increasing F adsorption, surface protonation at a pH less than the pH of PZC provides an
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increased number of H atoms at the adsorbent surface leading to an increased number of H-
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bonding between the H atoms and F in solution resulting in increased F adsorption [76].
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Adsorption by a ligand exchange mechanism is also favoured at low pH because of a stronger
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attractive force between F and the adsorbent surface and the presence of more hydroxylated
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sites for exchange with F than at a high pH [4,29,61].
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It is apparent from the literature review that F adsorption is generally lowest at
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extremely low and high pH values. The adsorption is highest around the neutral pH that is 10
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commonly found in natural water. Therefore prior pH adjustment is not normally required for
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effective removal of F in treatment plants.
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3.2. Co-existing anions In natural water, several anions including PO43-, Cl-, SO42-, Br-, NO3- are
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simultaneously present with F at different concentrations which can compete with F for
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removal by adsorbents. The extent of the competition depends on the relative concentrations
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of the ions and their affinity for the adsorbent. Meenakshi and Viswanathan [30] reported that
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increased concentrations of Cl-, SO42-, Br-, NO3- , and HCO3- decreased F adsorption by an
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anion exchange resin which adsorbs anions by an ion exchange mechanism, whereas these
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anions had no effect on F adsorption by a chelating resin which adsorbed F selectively by a
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H-bonding mechanism. The preferential order of adsorption of anions by the chelating resin
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was reported to be F- > Cl- > NO3- > SO42-.
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Solangi et al. [34] studied the adsorption of F by a thio-urea incorporated amberlite
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resin in the presence of PO43-, Cl-, SO42-, Br-, NO3-, NO2-, HCO3-, and CO32- at five times the
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molar concentration of F and observed that Br-, NO2-, and PO43- had little interference with F
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adsorption; the other ions had no interference. This was explained as due to the strong H-
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bonding of F with the amide groups in the resin.
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Multivalent metal oxides are known to adsorb F selectively by the ligand exchange
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specific adsorption mechanism. An alum sludge containing a high percentage of Al, Ti and
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Fe oxides adsorbed F selectively in the presence of SO42- and NO3- [48]. At 50 mg/L, SO42-
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and NO3- reduced F adsorption from 85% to 40% and 62%, respectively from a solution
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containing 20 mg F/L. In contrast to these anions, PO43- and selenate concentrations at 20
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mg/L reduced F adsorption to 25% of F in solution. Based on these results, Sujana et al. [48]
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proposed that the decreased order of competition of anions for F adsorption to be PO43- ≥ 11
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selenate > SO42- > NO3-. Das et al. [55] also found that PO43- interfered more than SO42- in F
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removal by a calcined Zn/Al layered double hydroxide (LDH) consisting of Zn/Al oxide. Kumar et al. [7] investigated the adsorption of F by a granular ferric hydroxide in the
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presence of competing anions such as Cl-, SO42-, BrO3-, NO3-, CO32- and PO43-, each having
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concentrations of 20 to 100 mg/L with an initial F concentration of 20 mg/L. There was no
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significant influence of competing anions on F removal when the adsorbent dose was 10 g/L,
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which was attributed to the availability of plenty of sorption sites. However, when the
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adsorbent concentration was reduced to 5g/L, PO43-, CO32-, and SO42- , each at concentrations
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of 100 mg/L, this reduced the F adsorption capacity by 35, 25, and 20% of F in solution,
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respectively. The other anions did not significantly reduce the F adsorption capacity.
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Similarly, Raichur and Basu [50] found that F adsorption by a mixture of naturally occurring
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rare earth oxides (oxides of La, Ce, Pr, Nd, Sm and Y) was not significantly affected by the
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presence of SO42- or NO3- in water at a concentration equal to that of F (100 mg/L).
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It can be concluded that the non-specifically adsorbing anions (e.g. nitrate, chloride)
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do not compete with F for adsorption on adsorbents that adsorb F using specific adsorption.
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Only anions that are adsorbed by specific adsorption (e.g. phosphate, selenate, arsenate)
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compete with F for adsorption. When F is adsorbed by non-specific adsorption the non-
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specifically adsorbing anions can compete with F.
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3.3. Temperature
Temperature had no consistent effect on F adsorption. Adsorption on many adsorbents
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increased with temperature showing an endothermic nature of adsorption (granular ferric
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hydroxide [7], fly ash [46], calcined Mg/Al/CO3 LDH [54], LDH/chitosan [63], spent catalyst
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[75]). In contrast with many others it decreased with temperature showing an exothermic
283
nature of adsorption (chelating resin [30], alum sludge [48], calcined Zn/Al LDH [55], 12
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modified activated carbon [59], LDH/chitin [65], geo-materials [68]). Temperature has also
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been reported to have no significant effect on adsorption by some adsorbents (trivalent
286
cations/zeolite [40]). The reasons for the differences in the effect of temperature were not clearly stated in
288
the studies. It may depend on the temperature range studied, the nature of the adsorbent, and
289
the conditions used in the studies. For example, at extremely low temperatures (5, 10oC) the
290
rate of adsorption was reported to be low because the rate of movement of F to the adsorption
291
sites is low. Lai and Liu [75] reported that F adsorption on spent catalyst increased
292
significantly when the temperature increased from 5oC to 25oC but very little further increase
293
in adsorption was observed when the temperature rose to 50oC. Sujana et al. [48] reported
294
that the exothermic nature of adsorption of F on alum sludge existed because the rising
295
temperature increased the tendency for F to escape from the adsorbent. Another reason given
296
was an increase in thermal energy of adsorbed F at higher temperatures, causing increased
297
desorption.
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3.4. Adsorption kinetics
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Fluoride adsorption studies have shown that the rate of removal of F by adsorbents is
301
high in the initial 5-120 min where generally more than 90% of F is adsorbed, but thereafter
302
the rate significantly levels off and eventually approaches zero denoting the attainment of
303
equilibrium. This is because initially the adsorption sites are vacant and the F concentration
304
gradient between solution and adsorbent surface is high. Subsequently the rate decreases
305
because of the decrease in vacant sites. A fast rate of adsorption helps the adsorbent to treat
306
large quantities of water [54]. A slow rate causes operational, control, and maintenance
307
problems in the adsorption process of the filter bed [31].
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The rate of F adsorption increases with an increase in concentration of adsorbent [54]
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and a decrease in initial F concentration [29,47,54]. It also depends on the structural
310
properties of the adsorbents and the interaction between F and the sites of adsorption.
311
Adsorption kinetics has been described by pseudo-first order and pseudo-second order kinetic
312
models and the diffusion model and rate constants have been determined (Table 1). These
313
models have also provided information on adsorption mechanisms.
4. Adsorbents
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The effectiveness of F adsorption from drinking water by various adsorbents, the
318
methods used for the assessment, and the kinetics and equilibrium models that best explained
319
the adsorption process are presented in Table 1. Caution needs to be exercised in comparing
320
the adsorption capacities of adsorbents because of the inconsistencies in data presentation
321
including differences in methodology and parameters used in the studies (pH, temperature, F
322
concentration range, competing ions, etc.). An ideal adsorbent that can be used to remove F
323
must have the following characteristics: low-cost, a high F adsorption capacity, rapid
324
adsorption of F, easily regenerated after its removal capacity is exhausted and good physical
325
characteristics (rapid water flow without filter clogging).
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4.1. Metal oxides and hydroxides Oxides and hydroxides, also called hydrous oxides or oxyhydroxides, of trivalent and
329
tetravalent metals such as Fe, Al, La, Mn, and Zr are used to remove both anionic and
330
cationic contaminants from water and wastewaters because of their strong ability to adsorb
331
these ions [37,77]. The predominant mechanism of adsorption of F on oxides and hydroxides
332
is ligand exchange by the formation of inner sphere complexes (specific adsorption) as 14
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discussed previously. Most metal oxides and hydroxides have their PZC above the natural
334
water pH of 7 (granulated ferric hydroxide pH 7.5-8.0 [7], activated alumina pH 8.25 [36], γ-
335
alumina pH 8 [49]). Therefore, at the neutral pH of natural water, these adsorbents have
336
positive surface charge which is favourable for the adsorption of the negatively charged F. Of the oxides and hydroxides of metals, Al oxide, especially the activated form
338
(activated alumina) has been the most commonly used adsorbent for the removal of F
339
[9,29,36,37,42]. Activated alumina is produced by thermal degradation of aluminium
340
hydroxide to obtain materials with high specific surface area and a distribution of micro- and
341
macro-pores [78]. The specific surface area (m2/g) of activated alumina has been reported to
342
be 160 [49], 297 [29], and > 200 [26]. The adsorption capacity of activated alumina varies
343
with the structure of the alumina. For example, γ-Al2O3 has a much higher adsorption
344
capacity than α-Al2O3 [42] and therefore this form of activated alumina is commonly used for
345
defluoridation of water. The maximum F adsorption capacity of activated γ-Al2O3 (mg/g) has
346
been reported to be 1.1 [36], 12.0 [41], 2.41 [47], and 16.3 [49].
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Aluminium oxides have poor adsorption capacity for F in acidic conditions because of
348
their tendency to dissolve and form positively charged AlF complexes (AlF2+, AlF2+) which
349
are repelled by the positively charged surfaces of Al oxides at these conditions (PZC of Al
350
oxides > pH 7) [49]. In alkaline conditions, Al oxides have negatively charged surfaces
351
which repel the negatively charged F ions present at these pHs. Also, as stated previously, the
352
increase in the concentrations of OH- in alkaline condition competes with F- for adsorption.
353
Therefore the optimum pH for F adsorption is considered to be near neutral pH [47,49].
Ac ce
pt
347
354
Recent studies have shown that surface coatings of adsorbents with other materials
355
have enhanced the adsorption of F [29]. Because manganese oxides have a high specific
356
surface area, a micro-porous structure [79] and a high adsorption capacity towards anions
15
Page 15 of 62
357
[80], activated alumina was coated with these oxides and F adsorption capacities have been
358
studied [29,36]. Maliyekkal et al. [36] compared the F adsorption capacity and rate of adsorption of
360
activated alumina (AA) and a granular manganese oxide coated AA (MOCAA) at pH 7 and
361
found that in the batch study, most of the adsorption was complete in 3 h in the case of
362
MOCAA compared to 10 h for AA. The pseudo-first order and pseudo-second order rate
363
constants were also higher for MOCAA. The Langmuir adsorption capacity of MOCAA was
364
2.85 mg/g compared to 1.08 mg/g for AA. Maximum F adsorption was found to occur at a
365
wider pH range of 4-7 for MOCAA compared to that for AA (4-6). The column study also
366
showed that MOCAA had higher F adsorption capacity. The superiority of MOCAA over AA
367
in the adsorption of F was reported to be not due to the surface area difference because the
368
specific surface area of AA was 204 m2/g and that of MOCAA was 170 m2/g. Maliyekkal et
369
al. [36] suggested that the reason for this could be the increased zeta potential (surface
370
charge) of the MOCAA, although no supporting data were presented.
ed
M
an
us
cr
ip t
359
Teng et al. [29] used a redox process to coat AA with an amorphous MnO2. The
372
granular MOCAA produced had a higher surface area (316 m2/g) compared to that for AA
373
(297 m2/g) and a significantly rougher surface with plenty of pores. Maximum F adsorption
374
was obtained at the pH range of 4-6. One of the two mechanisms of F adsorption at a pH less
375
than 6 was considered to be chemical adsorption by a ligand exchange of surface OH groups
376
in MOCAA by F in solution resulting in an increase of pH due to the release of OH-. The
377
other mechanism was intra-particle diffusion of F. The Langmuir adsorption maximum at pH
378
5.2 was 7.09 mg/g which was much higher than the value of 1.08 mg/g reported for AA by
379
Maliyekkal et al. [36] and values reported for many other granulated adsorbents. For initial F
380
concentrations less than 21 mg/L, most of the adsorption was completed within 30 min
Ac ce
pt
371
16
Page 16 of 62
381
compared to 10 h for AA. Faster adsorption by MOCAA was considered to be due to the
382
larger surface area and porous surface of this adsorbent. Iron oxides are also known to have large capacity to remove anions from water by
384
mechanisms similar to those operating in Al oxide adsorbents [81]. Kumar et al. [7] studied
385
the adsorptive removal of F by a highly porous and poorly crystalline granulated ferric
386
hydroxide (GFH) (β-FeOOH) with a specific surface area of 250-300 m2/g, a PZC of pH 8,
387
and a granular size of 0.32-2.0 mm. Nearly 95% adsorption of F from solution was achieved
388
within the first 10 min of agitation of 10-20 mg F/L with 10 g GFH/L at pH 6-7. The
389
maximum F adsorption was observed at the pH range of 3-8. A sharp decrease of F
390
adsorption was observed above the PZC pH of 8 as the GFH surface became more negatively
391
charged causing electrostatic repulsion of the negatively charged F- ions in addition to
392
increasing concentration of OH- ions which competed with F- for adsorption.
M
an
us
cr
ip t
383
Hydroxyapatite, the most abundant of phosphate minerals, has been used for
394
defluoridation of water. Fan et al. [52] compared the F adsorption capacities of several
395
natural minerals and found that the adsorption capacities at pH 6 followed the order:
396
hydroxyapatite > fluorspar > quartz activated using ferric ions > calcite > quartz. The highest
397
F adsorption capacity of hydroxyapatite was explained as owing to F exchanging with a OH
398
group at the surface and inside the apatite mineral. The F adsorption on the other minerals
399
was deemed to be a surface adsorption process.
pt
Ac ce
400
ed
393
It is evident from the literature review that Fe and Al oxides and hydroxides are the
401
commonly used adsorbents for defluoridation. They have a moderate level of F adsorption
402
capacity (1-16 mg/g) and if locally available at low cost, can potentially be employed in rural
403
areas, especially in developing countries.
404 405
4.2. Layered double hydroxides 17
Page 17 of 62
The majority of clay minerals such as kaolinite, mica, montmorillonite, vermiculite
407
and zeolite, carry predominantly negative charges and therefore adsorb very small amounts of
408
anions. Layered double hydroxide (LDH) or hydrotalcite (HTlc) is another type of clay
409
mineral, but has positive charges and therefore adsorb significant quantities of anions and
410
oxyanions (e.g. fluoride, arsenite, arsenate, chromate, phosphate, selenite, selenate, nitrate,
411
etc) and monoatomic anions (e.g. fluoride, chloride) from aqueous solutions [82].
412
Structurally, LDHs are composed of positively charged brucite-like sheets compensated by a
413
large number of exchangeable charge-balancing anions in the hydrated interlayer regions
414
[55,56,82]. Charges can also be produced by the ionisation of the OH groups at the surface of
415
the LDH particles. The PZC of LDHs is in the region of pH 9-12 [82] and therefore at the
416
neutral pH of natural water, LDHs carry positive charges and act as anion exchangers.
an
us
cr
ip t
406
Calcined LDHs have higher F adsorption capacities than the uncalcined LDHs [53].
418
The optimum temperature of calcination is generally considered to be 450-500oC [53-56]. Lv
419
et al. [54] reported that the F adsorption capacity of a Mg/Al LDH increased from 65 to 70
420
and then to 80 mg/g when the calcination temperature was increased from 200oC to 400oC
421
and then to 500oC, respectively, but decreased to 62 and 50 mg/g when the temperature was
422
increased to 600oC and 800oC, respectively. Wang et al. [53] suggested that the increase in F
423
adsorption capacity of LDH as a result of calcination was due to the higher specific surface
424
area, porosity, and surface reactivity of the Mg/Al oxide produced by calcination. Another
425
reason reported was the incorporation of F into the structure of Mg/Al oxide resulting in the
426
formation of the original structure of the LDH. The decrease in F adsorption capacity at
427
calcining temperatures higher than 500oC was considered to be due to the transformation of
428
the Mg/Al oxides into a spinel structure that does not exhibit the property of LDH structural
429
reconstruction [54].
Ac ce
pt
ed
M
417
18
Page 18 of 62
The adsorptive property of LDH depends on the metallic constitution of the LDH
431
structure. Lv et al. [54] showed that the F adsorption capacity of calcined Mg/Al LDH was
432
higher than that of calcined Ni/Al LDH and calcined Zn/Al LDH because of the higher
433
atomic weights of Ni and Zn compared to Mg. Among the calcined Mg/Al LDHs, the one
434
having Mg/Al molar ratio of 2 was found to have the highest F adsorption capacity. The
435
maximum F adsorption capacity of 213 mg/g was obtained at pH 6. This adsorption capacity
436
is the highest of all adsorbents listed in Table 1. However, Das et al. [55] reported a lower
437
Langmuir adsorption capacity of 17 mg/g at pH 6 for a calcined Zn/Al LDH. The F
438
adsorption data of Wang et al. [53] on a calcined Mg/Al LDH did not fit the Langmuir
439
adsorption model but the amount of F adsorbed continued to increase with equilibrium F
440
concentration with an adsorption capacity of 35 mg/g at the highest tested F equilibrium
441
concentration of 70 mg/L. The rate of F adsorption on calcined LDH is variable [53,55,56].
M
an
us
cr
ip t
430
The literature review reveals that LDHs can have very high F adsorption capacity (17-
443
213 mg/g) if calcined to 500oC. The adsorption capacity varies depending on the type and
444
proportion of the metals in the LDH structure. Because LDHs have high adsorption capacity,
445
they are useful adsorbents for defluoridation of waters with high F concentrations.
447 448
pt
Ac ce
446
ed
442
4.3. Ion exchange resins and fibres Ion exchange resins and fibres are an important class of adsorbents used to remove
449
anionic and cationic pollutants from water and wastewater. Their framework or matrix
450
consists of irregular, macromolecular, three dimensional network of hydrocarbon chain [32].
451
The cation exchange resins and fibres have negatively charged functional groups whereas the
452
anion exchange resins and fibres have positively charged functional groups such as -NH3+,
453
NH2+, ≡N+, ≡S+. Therefore the cation exchangers adsorb cations and the anion exchangers
454
adsorb anions such as F-. The cation exchangers can also be made to adsorb anions if they are 19
Page 19 of 62
455
impregnated with positively charged metallic cations that have strong affinity for anions
456
[39,57]. Ku et al. [57] used an Al incorporated cation exchange resin (Amberlite IR-120) to
458
remove F from water and found that the Langmuir maximum F adsorption capacity at pH 4
459
was 4.6 mg/g. Within the pH range of 4-9 tested, F adsorption emerged as the greatest at pH
460
5-7. Metals other than Al have also been incorporated into cation exchangers to enhance the F
461
adsorption capacity. For example, Luo and Inoue [39] compared the F adsorption capacities
462
of the cation exchange resin, Amberlite 200 CT modified by the incorporation of a number of
463
trivalent metal ions, La, Ce, Y, Fe, and Al. The F adsorption capacity of the metals
464
incorporated resins in the pH range of 4-7 was in the order: La ≤ Ce > Y > Fe ~ Al. More
465
than 80% of F was adsorbed by 20 g La/L resin from a 15 mL solution containing 15 mg F/L.
466
The resin without any metal incorporation adsorbed only about 5% of F from the solution.
M
an
us
cr
ip t
457
The majority of anion exchange materials are not effective in adsorbing F from
468
natural water containing other anions because F- affinity to anion exchange materials is the
469
least of all anions (citrate > SO42- > oxalate > I- > NO3- > CrO42- > Br- > SCN- > Cl- > formate
470
> acetate > F- [32]). Consistent with this order of affinity, a strong base anion exchanger was
471
shown to effectively adsorb NO3-, Br-, and SO42-, but not Cl- and F- [83].
pt
Ac ce
472
ed
467
The F adsorption capacity of anion exchangers can be enhanced by modifying the
473
adsorbing sites on the anion exchangers. Generally, anion exchange resins impregnated with
474
chelating agents that can form H-bonding with F are used. Meenakshi and Viswanathan [30]
475
compared the F adsorption capacity of a chelating resin having sulfonic acid functional group
476
and an anion exchange resin and showed that 1 g of the chelating resin adsorbed 95% of F
477
from a 50 mL solution containing 3 mg F/L in 40 min compared to 65% adsorption by the
478
anion exchange resin.
20
Page 20 of 62
Solangi et al. [34] reported that 100 mg of the ion exchange resin, Amberlite XAD-4
480
modified by incorporation with thio-urea binding sites removed 90% of the F from a 10 mL
481
solution containing 16 mg F/L compared to 30% by the unmodified resin at pH 7. The higher
482
F adsorption capacity of the modified resin was explained as being due to H-bonding between
483
the amide groups in thio-urea and F. The Langmuir maximum F adsorption capacity of the
484
modified resin at pH 7 was much higher than the F adsorptive capacities of many other
485
adsorbents in the literature (Table 1). Interference of other co-existing anions present at the
486
concentration ratio of 1:5 (F:other ions) was insignificant. Therefore the authors concluded
487
that the modified resin can be effectively used for defluoridation of water.
us
cr
ip t
479
Fibrous adsorbents, due to their physico-chemical structure, generally have rapid rates
489
of adsorption of ions. If their adsorptive capacity and affinity can also be enhanced, they can
490
be a useful class of adsorbents for removal of ions from water. Ruixia et al. [31] introduced
491
functional groups into a polyacrylonitrile ion exchange fibre and studied its F adsorption
492
behaviour. They found that the F adsorption reached equilibrium in 5 min with 40% of F
493
adsorption from a 500 mL solution containing 5, 34, and 50 mg/L of F, As(V), and P,
494
respectively when 1 g fibre modified with functional groups was added. The adsorption by
495
the unmodified fibre was less than 5% of solution F. The rate of adsorption was considered to
496
be faster than that observed in many other adsorbents. Maximum adsorption was obtained at
497
pH 3. The mechanism of adsorption was considered to be ion exchange.
M
ed
pt
Ac ce
498
an
488
It could be concluded from this review that ion exchange resins and fibres have poor
499
F adsorption capacities but the adsorption capacities and selectivity for F adsorption in the
500
presence of other ions can be significantly increased (up to 61 mg/g) by surface modification
501
of the adsorbents by loading with organic functional groups and metals. These adsorbents are
502
relatively expensive and therefore they are useful only for water treatment in industrial
503
countries. 21
Page 21 of 62
504 505
4.4. Zeolites Because zeolites have negative surface charges at all pH values, they have high
507
adsorption capacity for cations, but have low adsorption capacity for anions. Nonetheless
508
their adsorption capacity for anions can be increased by modifying the zeolite surface with
509
cationic surfactants or multivalent metallic cations [85]. For F adsorption, only metallic
510
cations incorporated with zeolites have been used. There appear to be no studies conducted
511
on using surfactant- or organic compounds-modified zeolites on defluoridation.
us
cr
ip t
506
Onyango et al. [40] modified the surface of a synthetic zeolite by exchanging Na+ in
513
the zeolite with Al3+ and La3+ to create active sites for F adsorption. The introduction of Al
514
and La opened up the pores in zeolite leading to increased porosity. The Langmuir adsorption
515
capacity is slightly larger for La-zeolite than for Al-zeolite (Table 1). However, within the F
516
concentration range studied (10-80 mg/L), Al-zeolite had twice as much adsorption capacity
517
as the La-zeolite. The adsorption capacities obtained for these zeolites were reported to be
518
much higher than many other adsorbents including the commonly used activated alumina
519
(Table 1). This suggested that the mechanism of adsorption of F onto Al-zeolite was mostly
520
by a chemical adsorption process (ligand exchange) and adsorption onto La-zeolite was
521
mostly by a physical adsorption process (coulombic attraction).
M
ed
pt
Ac ce
522
an
512
The Na-zeolite had no PZC because it was negatively charged at all pHs between 3
523
and 13 and therefore the adsorption of F- is poor due to charge repulsion. In contrast, the Al-
524
zeolite and the La-zeolite had PZC at pHs of 8.15 and 4-5.25, respectively indicating that
525
below these pHs these zeolites were positively charged. In accordance with the surface
526
charge on La-zeolite the F adsorption was the highest at pH 5; and decreased above and
527
below this pH. In the case of the Al-zeolite, the increase of pH increased F adsorption with a
528
plateau forming above pH 5. The F adsorption on Al-zeolite at pHs above 5 did not decrease 22
Page 22 of 62
529
as observed for La-zeolite up to the highest pH 9 tested because F was reported to be
530
adsorbed by chemical adsorption. The authors stated that, over all, Al-zeolite was superior to
531
La-zeolite for defluoridation. In a subsequent study by Onyango et al. [41] on F adsorption by four types of Al pre-
533
treated low-silica synthetic zeolite it was observed that F adsorption on all four types
534
increased from pH 2 to 4 and remained constant up to pH 8, before decreasing from pH 8 to
535
11. The pH effect on F adsorption was explained by the following reactions.
cr
ip t
532
an
us
536
537
M
538
At low pH, part of the F is removed as weakly ionised HF and therefore F adsorption is
540
reduced (first equation). Another reason could be that some of the F is complexed to Al
541
solubilised at low pHs to form AlF+ complex which has low tendency to adsorb on the
542
positively charged adsorbent at the low pHs. At high pH, the negatively charged adsorbent
543
(last equation) repels the negatively charged F- as well as increased competition of OH- for
544
adsorption.
pt
Ac ce
545
ed
539
Samatya et al. [23] using a natural zeolite from Turkey pre-treated with La, Al, and Zr
546
to study the removal of F from tap water spiked with NaF. They showed that the F adsorption
547
capacities at the equilibrium F concentration range of 0-12 mg/L were largest for Zr-zeolite
548
and smallest for La-zeolite. The values for the La and Al zeolites were lower than the values
549
reported by Onyango et al. [40], probably because the metal loadings were much lower and
550
the zeolite used was natural compared to the synthetic zeolite of Onyango et al. [40]. 23
Page 23 of 62
551
However, all three metal zeolites removed 95% of F from an aqueous solution containing 2.5
552
mg F/L at an adsorbent dose of 6 g/L. In contrast to the above studies which showed high F removal capacities of
554
multivalent metal incorporated zeolites, Diaz-Nava et al. [85] found that La and Eu treated
555
natural zeolite from Mexico had only slightly higher F removal capacities than when this
556
zeolite was treated with Na and Ca. The weight percentages of the metals in the zeolite were
557
1.7, 3.20, 0.32, and 1.49 for Na, Ca, La, and Eu, respectively.
cr
ip t
553
The review revealed that zeolites on their own had very low adsorption capacities for
559
F but when they were loaded with multivalent metallic cations they produced moderate to
560
very high F adsorptive capacities (2-45 mg/g). The adsorptive capacity depends on the type
561
and amount of metal loading. Zeolites can also provide good physical properties for practical
562
use.
M
an
us
558
564
4.5. Carbon materials
ed
563
Activated carbon (AC) is an important carbon material commonly used as an
566
adsorbent for the removal of a wide range of aquatic pollutants due to its exceptionally high
567
surface area (500-1500 m2/g), highly developed internal microporosity, presence of a range of
568
functional groups, low cost and ready availability [86,87]. However, it displays poor
569
adsorption capacity towards anionic pollutants because of its low PZC (pH 1.6-3.5, [88]).
Ac ce
570
pt
565
The amount of anions adsorbed onto AC depends on the pore size distribution
571
because the adsorption occurs mainly in the pores. Abe et al. [60] reported that F adsorption
572
capacity increased with the specific surface area in 11 out of the 12 carbonaceous materials
573
studied by them. The adsorption of F on bone char did not fit this pattern because the
574
mechanism of adsorption was different from the rest. In bone char, F is adsorbed chemically
575
by ligand exchange with the OH group in the hydroxyapatite compound [1]. The 24
Page 24 of 62
ineffectiveness of carbon in adsorbing F was also shown by Srimurali et al. [70] who found
577
that lignite coal and char fines, a by-product obtained during the making of coke, adsorbed
578
only 7.9 and 19% of F, respectively from a 50 mL solution containing 5.0 mg F/L and when
579
100 mg of adsorbents were added and mixed for 5 h. In comparison, bentanite clay removed a
580
much larger percentage of 33% of solution F.
ip t
576
A new type of carbon material called aligned carbon nanotubes (ACNT) made up of
582
needle-like cylindrical tubules of concentric graphitic carbon capped by fullerene-like
583
hemispheres was developed as a promising adsorbent material for the adsorption of
584
contaminants in water [12]. Li et al. [12] found that ACNT had higher F adsorption capacity
585
than AC (Table 1), despite the ACNT having a much lower specific area and pore volume
586
than the AC. Li et al. [12] stated that though the adsorption capacity of the ACNT was high,
587
the cost too was high which may limit its full commercial utilisation.
M
an
us
cr
581
The surface of the carbon particles can be modified to improve the F adsorptive
589
properties. Such modifications have been brought about by creating new functional groups
590
which have strong affinity towards F. Daifullah et al. [59] modified the structure of an AC
591
product by steam pyrolysis of rice straw and oxidised the product using HNO3, H2O2, and
592
KMnO4. Of these treatments, the material obtained by KMnO4 oxidation produced the
593
highest F adsorption. The adsorption mechanism was considered to be ligand exchange of
594
OH groups on the carbon surface with F. The presence of MnO2 on the carbon surface caused
595
by reduction of the KMnO4 may have also participated in the removal of F. Thermodynamic
596
studies showed that the adsorption was chemical in nature. The Langmuir adsorption
597
maximum was 15.9 mg F/g at the natural pH of water. This value was considered to be higher
598
than the values reported in the literature for many other adsorbents.
Ac ce
pt
ed
588
599
Ramos et al. [42] studied the F adsorption behaviour of Al-impregnated AC (AlAC)
600
produced from coconut shells followed by calcination to 300oC. The AlAC had a F 25
Page 25 of 62
adsorption capacity more than four times higher than the plain carbon at an equilibrium
602
concentration of 2-8 mg F/L. The Langmuir adsorption capacity of AlAC (1.07 mg/g) at
603
natural water pH, though higher than the value obtained for plain carbon (0.49 mg F/g), was
604
lower than the values obtained for some other adsorbents especially activated alumina. The F
605
adsorption reached its maximum at pH 3 because the PZC of the AlAC was 4.1-4.8 (above
606
this pH, AlAC had negative surface charges which repel F-). However, there was appreciable
607
adsorption of F at pH 6-7 (0.6 mg F/g at the equilibrium F concentration of 6 mg F/L) due to
608
chemical adsorption of F onto Al which does not involve electrostatic attractive forces.
us
cr
ip t
601
In a later study, Li et al. [61] reported that the adsorption capacity of aligned carbon
610
nanotubes (ACNT) at pH 6 increased from 2.3 mg F/g to 9.6 mg F/g when Al was loaded
611
onto ACNT and calcined to 450oC at an optimum Al2O3 rate of 30% by weight. The
612
adsorption capacity fell to less than 6 mg F/g when the loading rate was 20 and 40%. The
613
maximum adsorption capacity was obtained at pH 6.0-9.0. The mechanism of F adsorption at
614
pH less than the PZC of 7.5 was considered to be by electrostatic attraction of negatively
615
charged F- ions onto the positively charged adsorbent as well as by ligand exchange of the
616
OH group by F. At the PZC and slightly above this pH the adsorption was reported to be
617
mainly by ligand exchange. At pHs much higher than the PZC (pH > 9), OH- ions in solution
618
competed with F- ions for adsorption and therefore F adsorption decreased.
M
ed
pt
Ac ce
619
an
609
It can be concluded from the review that AC is a poor adsorbent for F. However, its
620
adsorption capacity can be increased to moderate levels (up to 16 mg/g) by chemical
621
modification of the carbon surface. Specialised carbon materials such as ACNT have higher
622
adsorption capacity than AC. Their adsorption capacity can also be increased by surface
623
modification but this will incur additional cost.
624 625
4.6. Natural materials 26
Page 26 of 62
626
Several natural inorganic materials (soils, clays, minerals, and building materials)
627
have been used in the defluoridation of water. The adsorption capacities of these materials
628
have been discussed in previous reviews (17, 26, 27) and also presented in Table 1. Bioadsorbents contain a variety of functional groups such as carboxyl, imidazole,
630
sulphydryl, amino, thioether, phenol, carbonyl, amide, and hydroxyl moieties capable of
631
adsorbing the pollutants, especially the metal cations [78]. Considering this feature of
632
bioadsorbents, metals having strong affinity towards F were loaded onto selected
633
bioadsorbents (Zr4+ on collagen fibres [13], La3+ on geletin [44]) for enhancing the adsorption
634
of F from drinking water.
us
cr
ip t
629
Chitin and chitosan have been shown to constitute another effective group of
636
bioadsorbents for removing a range of aquatic pollutants due to their low cost and high
637
contents of amino and hydroxyl functional groups [89]. Chitin is a polysaccharide found in a
638
wide range of organisms but most commonly extracted from shellfish processing waste,
639
while chitosan is a copolymer of glucosamines derived from chitin by deacetylation in hot
640
alkali [67]. Sahli et al. [22] reported that the adsorption of F on chitosan from a synthetic
641
water was rapid and reached a maximum at 5 min. Increase of pH from 2 to 6 increased
642
adsorption followed by a decrease up to pH 10. The F adsorption from a brackish water
643
containing 3.25 mg F/L and several other anions at much higher concentrations (Cl- 1083
644
mg/L, SO42- 215 mg/L, HCO3- 171 mg/L) showed that the selectivity of chitosan towards
645
various anions was as follows: F- > HCO3- >NO3- > Cl- > SO42-. In a batch adsorption study of
646
15 min duration the F concentration fell to 1.82 mg/L, but with a successive batch adsorption
647
the concentration reduced to 0.9 mg/L which was lower than the WHO standard of 1.5 mg
648
F/L.
Ac ce
pt
ed
M
an
635
649
Ma et al. [66] prepared a magnetic-chitosan adsorbent by co-precipitation of chitosan
650
with Fe2+ and Fe3+ salts and used it to remove F from a synthetic F solution in a batch study. 27
Page 27 of 62
The adsorbent was shown to have a higher F adsorption capacity (Langmuir adsorption
652
capacity 22.49 mg F/g) than a commercial activated alumina (Table 1). Adsorption capacity
653
was highest in the pH range 5-9. The FTIR data of the adsorbent showed that the main groups
654
involved in the adsorption process were amine and iron oxides in the magnetic-chitosan. The
655
main advantage of having magnetic properties in the adsorbent is to easily separate the
656
adsorbent from the treated water after its use using an external magnetic field so that the
657
adsorbent can be reused.
cr
ip t
651
A research group in India has reported that the F adsorption capacity of nano-
659
hydroxyapatite (nHAp) can be increased by incorporating chitin or chitosan onto nHAp [63-
660
65,90]. In a batch experiment conducted by Sundaram et al. [65], it was observed that mixing
661
0.25 g each of chitosan, nHAp, and nHAp/chitosan composite with 50 mL solution of NaF
662
containing 10 mg F/L, removed 0.05, 1.30, and 1.56 mg F/g adsorbent, respectively after 30
663
min of shaking the suspensions. Defluoridation of water from a fluoride-endemic village
664
using nHAp-chitosan composite was found to be higher compared to using nHAp. Similar
665
results were obtained from a subsequent study where a composite of nHAp and chitin was
666
used [64]. However, the adsorption capacity of this composite was higher than that of the
667
composite made with chitosan (Table 1).
an
M
ed
pt
Ac ce
668
us
658
The review revealed that although the majority of natural materials have low
669
effectiveness in removing F (< 1 mg/g), they are inexpensive and consequently can be used in
670
less developed countries, especially in rural areas. Furthermore, their F removal capacities
671
can be increased through the process of surface modification, such as loading with
672
multivalent metallic cations, as reported previously for other adsorbents.
673 674
4.7. Industrial by-products
28
Page 28 of 62
675
Several types of industrial by-products have been used for the adsorptive removal of
676
pollutants including fluoride from water [27,28,89,91,92]. Most of the adsorbents fall into the
677
categories of by-products from the mining industry (e.g. red mud), steel industry (e.g. slag
678
materials), and power plant industry (e.g. fly ash). Cengeloğlu et al. [4] investigated the adsorption of F from aqueous solutions using a
680
red mud containing 18.7% Al2O3, 39.7% Fe2O3, and 14.5% SiO2 with (ARM) and without
681
(RM) activation by 20% HCl treatment in a batch study. The amount of F removed from a
682
solution (pH 5.5) containing 3.8-28.5 mg F/L by ARM at a dose of 0.2 g/50 mL was at least
683
three times higher than that by RM at the same dose. The main mechanism of F removal was
684
considered to be that of specific adsorption caused by ligand exchange of F and OH groups
685
on the metal oxide surfaces in red mud.
an
us
cr
ip t
679
Red mud particles are too fine for use in filter beds because of their poor hydraulic
687
conductivity. This problem can be overcome either by mixing red mud with coarse-size
688
particles such as sand [93] or by granulation [94] before use. Tor et al. [72] used the method
689
of Zhu et al. [94] for the granulation of red mud by mixing it with fly ash, sodium carbonate,
690
quicklime and sodium silicate. The maximum F adsorption (2.2 mg/g) from a NaF solution
691
containing 15 mg F/L and granulated red mud (GRM) of 5 g/L was obtained at a pH of 4.7 in
692
the pH test range of 2.5-7.3. The adsorption capacity determined from a column study was
693
2.05 mg F/g for a flow rate of 2 mL/min compared to 0.644 mg F/g obtained in the batch
694
study at the same GRM dosage (5g/L) and initial F concentration (5 mg/L).
ed
pt
Ac ce
695
M
686
Fly ash is a major by-product that is produced from the combustion of coal in power
696
stations. It consists of fine and powdery materials (1.0-100 µm) made up of a mixture of
697
amorphous and crystalline alumina-silicates and several compounds of Si, Al, Fe, Ca, and Mg
698
[95] and therefore a good candidate material for F adsorption from water. Chaturvedi et al.
699
[46] reported that the adsorption of F from water by a fly ash containing 56.0% SiO2, 25.9% 29
Page 29 of 62
Al2O3, 2.22% CaO, and 1.26% Fe2O3 had a maximum Langmuir adsorption capacity of 20
701
mg/g at pH 6.5. The F adsorption increased from 79 to 94% when the pH of the F solution
702
(10 mg/L) increased from 2.0 to 6.5 and then decreased with further increase in pH up to 9.5.
703
However, Nemade et al. [62] reported that F adsorption by a fly ash decreased continuously
704
from pH 2 to 12. The difference in the pH effects between the two studies could be due to the
705
difference in chemical characteristics of the fly ashes and experimental conditions used.
ip t
700
In a column study, Piekos and Paslawska [74] found that an alkaline fly ash (pH ≥ 10,
707
9.1% CaO) effectively removed F from water containing 1-100 mg F/L when the solution
708
was passed through a column (400 mm length) packed with 450 g fly ash at a flow rate of ≤ 2
709
mL/h. Complete adsorption of F was obtained after 120 h. The F adsorption mechanism was
710
suggested to be chemical binding of F onto Ca(OH)2 and physical adsorption onto the
711
residual carbon particles in the fly ash.
M
an
us
cr
706
Steel industry by-products such as blast furnace slag, electric arc furnace slag, basic
713
oxygen furnace slag and converter slag have been shown to adsorb many pollutants,
714
especially phosphate, from water due to the presence of high contents of Ca, Fe, Al, Mg, and
715
Mn oxides [28]. However, limited studies have been conducted in their use in defluoridation
716
of water [16,27]. Islam and Patel [71] showed that a basic oxygen furnace slag (BOFS)
717
containing 46.5% CaO, 16.7% Fe oxides, and 13.8% SiO2 had a good capacity to adsorb F.
718
Thermal activation of the BOFS by heating to 1000oC for 24 h increased the porosity and
719
surface area leading to increased F adsorption. The percentage adsorption of F (initial F
720
concentration 10 mg/L, 0.5 g/100 mL adsorption dosage) was higher for the thermally
721
activated BOFS (TABOFS) (93%) than for the BOFS (70%). Unlike the case with most
722
adsorbents, increase of pH from 2 to 10 increased F adsorption on TABOFS. The increased
723
adsorption at high pH values was considered to be due to F adsorption onto Ca(OH)2 in the
Ac ce
pt
ed
712
30
Page 30 of 62
724
slag. Presence of other anions reduced F adsorption in the order, PO43- > HCO3- > CO32- >
725
SO42- > Cl- > NO3-. Large quantities of alum sludge, a waste product, are generated during the
727
manufacture of alum from bauxite [48] and kaolin [96] by the sulphuric acid process in many
728
countries. As this sludge mainly consists of oxides of Al, it has a high adsorption capacity to
729
remove many aquatic pollutants. Sujana et al. [48] used a sludge (47.2% Al2O3 and 20.7%
730
TiO2) produced from an alum manufacturing plant using bauxite raw material for the removal
731
of F from water. The sludge after washing with water and calcining to 400oC for 3 h adsorbed
732
more than twice the amount of F that was adsorbed by the uncalcined sludge from a solution
733
containing 10 mg F/L of pH 6 (adsorbent dosage 20-300 mg/g). The higher adsorption
734
capacity of the calcined sludge was considered to be due the increased surface area produced
735
by calcination. The F was considered to be specifically adsorbed according to a two-step
736
ligand exchange mechanism. Within a pH range of 3 to 9, the maximum F adsorption was
737
observed at pH 6; and at both above and below pH 6 the adsorption decreased, as also
738
reported for activated alumina by others. The presence of other anions in solution reduced F
739
adsorption according to the order, PO43- > silicate > SO42- > NO3-.
pt
ed
M
an
us
cr
ip t
726
In contrast to F adsorption on sludge produced during the manufacture of alum from
741
bauxite discussed in the previous paragraph, Nigussie et al. [96] reported that F adsorption by
742
a sludge produced during the manufacture of alum from kaolin raw material containing
743
primarily SiO2 did not increase when the sludge was heated to 300oC. At temperatures above
744
300oC up to 700oC, the adsorption decreased. Also, F adsorption was nearly constant from
745
pH 3 to 8 and significantly decreased above pH 10. The difference in the adsorption
746
behaviour of the two types of sludge may be due to the differences in the physico-chemical
747
characteristics of the sludge types. However, the F adsorption capacities of the sludge in the
748
two studies were not too different. The presence of HCO3- at concentrations higher than those
Ac ce
740
31
Page 31 of 62
749
of F, decreased F adsorption efficiency, while other anions (PO43-, SO42- , Cl- , NO3-) had no
750
significant effect [96]. Lai and Liu [75] used a by-product of the petrochemical industry called spent catalyst
752
consisting mainly of porous silica and alumina (specific surface area 130 m2/g, PZC 5.2) to
753
remove F from aqueous solutions. In a batch adsorption experiment on pH effect, they
754
observed that F adsorption decreased with increase in pH from 2 to 9.5, with a plateau in the
755
pH range of 4 to 7. The maximum adsorption capacity at pH 4 was found to be 28 mg F/g at
756
25 and 50oC. This value was considered to be comparable to that of activated alumina.
757
Because the activation energy calculated using the Arrhenius equation was very low (3.2
758
J/mol) the mechanism of adsorption was considered to be non-specific adsorption involving
759
coulombic forces.
an
us
cr
ip t
751
Many of the industrial by-products are wastes that need to be disposed of. Their
761
beneficial use can save disposal cost, prevent environmental pollution arising from the
762
disposal sites, and release disposal lands for alternative uses. In this respect, though these
763
materials have low F adsorption capacities, they are cost-effective and therefore can
764
potentially be used in developing countries where cost of operation is a major factor in the
765
choice of adsorbents. Furthermore, many of these materials have been shown to be good
766
adsorbents for other aquatic pollutants and therefore it is possible to simultaneously remove
767
many pollutants using these materials.
769
ed
pt
Ac ce
768
M
760
5. Adsorption thermodynamics
770
The strength and spontaneous nature of adsorption and information on whether the
771
adsorption process is exothermic or endothermic are provided by determining thermodynamic
772
parameters such as changes in Gibbs free energy (ΔG0), enthalpy (ΔH0), and entropy (ΔS0).
32
Page 32 of 62
773
These thermodynamic parameters are calculated from data on adsorption at different
774
temperatures using standard thermodynamics equations (see references in Table 2). A negative value for ΔG0 indicates spontaneous and thermodynamically favourable
776
adsorption, while a negative ΔH0 value indicates an exothermic adsorption process. An
777
exothermic reaction means that the amount of adsorption decreases with increasing
778
temperature. An endothermic reaction (positive ΔH0 value) has been explained as due to
779
enlargement of pore sizes and/or activation of the adsorbent surface [97,98]. Very low ΔH0
780
values are generally associated with physical adsorption and very high values with chemical
781
adsorption [99]. However, no definite value for distinguishing the two forms of adsorption
782
exists. ΔG0 has also been used to distinguish between the two forms of adsorption. For
783
example, Meenakshi and Viswanathan [30] reported that ΔG0 values of up to -20 kJ/mol were
784
indicative of physical adsorption, while values less than -40 kJ/mol involved chemical
785
adsorption. Positive ΔS0 values indicate good affinity of adsorbed species towards the
786
adsorbent and increased randomness at the solid-solution interface associated with structural
787
changes at the adsorption sites during the adsorption process [99,100].
ed
M
an
us
cr
ip t
775
All studies, except the one on F adsorption at the highest temperature of 55oC on
789
activated carbon modified by H2O2/KMnO4 oxidation [59] presented in Table 2 had negative
790
ΔG0 values. The negative values indicate good affinity of F for the adsorbents and the
791
adsorption process was spontaneous. The small positive value obtained for the activated
792
carbon was attributed to the disruption of the MnO2 formed at the carbon surfaces and
793
increase F solubility at high temperature [59]. The ΔS0 values were mostly positive, again
794
indicating a strong affinity of F towards the adsorbents. The ΔH0 values, however, were
795
positive and negative indicating the endothermic and exothermic nature of adsorption,
796
respectively.
Ac ce
pt
788
797 33
Page 33 of 62
798
6. Fluoride desorption and adsorbent regeneration A suitable adsorbent for F removal should not only have high F adsorption capacity
800
and cost-effectiveness but also be amenable to easy desorption of the adsorbed F and capable
801
of efficient regeneration for multiple reuse of the adsorbent. Also, the desorbing agent should
802
not produce any damage to the adsorbent that causes a reduction in its adsorption capacity.
803
Only adsorbents that can be reused have practical value in real systems for economic and
804
environmental reasons. However, inexpensive adsorbents, such as industrial by-products and
805
natural materials, may be used only once and they need not be regenerated for reuse because
806
usually the costs of desorbing chemicals and regeneration process are higher than that of
807
these adsorbents.
an
us
cr
ip t
799
Desorption of F is carried out by leaching of adsorbed F by acids, bases and salts
809
(Table 3). The selection of a desorbing agent largely depends on the influence of pH on F
810
adsorption and the strength of adsorption. Adsorption on most of the adsorbents decreases at
811
high pHs, and therefore bases having high pHs are commonly used for desorption of F from
812
such adsorbents (Table 3). As explained previously, the desorption of F at high pHs is due to
813
the increased repulsive forces between the negatively charged adsorbent surface and
814
negatively charged F- ions in solution as well as the competition between increased
815
concentration of OH- at high pH values and F- for adsorption. In materials where F adsorption
816
increases with pH, acids have been found to be more effective as desorbing agents. For
817
example, in a hydrous manganese oxide coated alumina [29] and a thio-urea modified
818
amberlite resin [34], F adsorption decreased at lower pH values. Consistent with this
819
adsorption pattern, the percentage of F desorption was found to be higher at low pH values.
820
Adsorption of F by most adsorbents is strong and not easily reversible, partly due to the
821
chemical nature of the adsorption process. Therefore, stronger acids and bases, as well as a
Ac ce
pt
ed
M
808
34
Page 34 of 62
822
longer time of their interaction with the adsorbent are required for efficient F desorption
823
[36,41,50,55]. Water and neutral salts are generally found to be less effective [31,41].
824 825
7. Conclusions Adsorption and precipitation/coagulation methods are widely used for the
827
defluoridation of water. The precipitation/coagulation method has the drawback of not being
828
efficient in waters having low F concentration, and the process produces toxic AlF complexes
829
in the treated water as well as large volumes of waste. The adsorption processes are generally
830
considered attractive because of their effectiveness, convenience, ease of operation,
831
simplicity of design, and for economic and environmental reasons.
an
us
cr
ip t
826
Multivalent metal oxides and hydroxides and layered double hydroxides generally
833
have high F adsorption capacities. Ion exchange resins and fibres, zeolites, and carbon
834
materials have low adsorption capacities on their own, but when their surfaces are modified
835
by incorporating organic functional groups or multivalent metallic cations, the adsorption
836
capacity increased. Some of the adsorbents (e.g. layered double hydroxides, alumina) had to
837
be activated by calcining at high temperatures to increase the adsorption capacity. The above
838
materials have potential for use in industrial countries and in areas where the F concentration
839
in water is very high. Natural and industrial by-products have low adsorption capacities but
840
because they are inexpensive, they have potential for use in rural areas, especially in
841
developing countries.
ed
pt
Ac ce
842
M
832
The pH of water is a dominant factor influencing F adsorption. Generally, F
843
adsorption increases from acidic to near neutral pH and then decreases with increase in pH.
844
Consistent with this pH effect, acids and alkali have been successfully used to desorb F and
845
consequently regenerate the adsorbent for reuse. However, multiple regeneration and reuse
35
Page 35 of 62
846
reduce the adsorptive capacity. Another important factor influencing F adsorption is the type
847
and concentration of other ions present in water. Future research needs to explore highly efficient, low cost adsorbents that can be
849
easily regenerated for reuse over several operational cycles without significant loss of
850
adsorptive capacity and have good hydraulic conductivity to prevent filters clogging during a
851
fixed-bed treatment process. Surface modifications of the adsorbents can be explored to
852
increase the capacity and strength of adsorption without significantly increasing the cost. The
853
majority of studies reported have been conducted in batch trials on synthetic waters. These
854
trials need to be extended using continuous mode column trials which have more relevance to
855
real operating systems on natural waters containing other ions as well.
an
us
cr
ip t
848
856
Acknowledgements
M
857
This study was funded by cooperative Research Centre for Contamination Assessment
859
and Remediation of the Environment (CRC CARE) (Project Number 4.1.12.11/12). We thank
860
Phil Thomas for editing this manuscript.
861
863 864 865
References
Ac ce
862
pt
ed
858
[1] J. Fawell, K. Bailey, E. Chilton, E. Dahi, L. Fewtrell, Y. Magara, Fluoride in Drinking Water, World Health Organization, IWA Publishing, UK, 2006.
866
[2] World Health Organisation, Guidelines for Drinking-water Quality, Volume 2. Health
867
Criteria and Other Supporting Information, 2nd Edition, World Health Organisation,
868
Geneva, 1996.
36
Page 36 of 62
869
[3] US Public Health Service, US Public Health Service Drinking Water Standards, US
870
Government Printing Office, Department of Health Education and Welfare, Washington
871
DC, 1962.
876 877 878
ip t
875
[5] N.J. Chinoy, Effects of fluoride on physiology of animals and human beings, Indian J. Environ. Toxicol. 1 (1991), 17-32.
cr
874
mud, Sep. Purif. Technol. 28 (2002) 81-86.
[6] P.T.C. Harrison, Fluoride in water: a UK perspective, J. Fluor. Chem. 126 (2005) 1448-
us
873
[4] Y. Cengeloğlu, E. Kir, M. Ersöz, Removal of fluoride from aqueous solution by using red
1456.
[7] E. Kumar, A. Bhatnagar, J. Minkyu, W. Jung, S. Lee, S. Kim, G. Lee, H. Song, J. Choi, J.
an
872
Yang, B. Jeon, Defluoridation from aqueous solutions by granular ferric hydroxide
880
(GFH), Water Res. 43 (2009), 490-499.
884 885 886
ed
883
Mater. B137 (2006) 456-463.
[9] S. Vigneswaran, C. Visvanathan, Water Treatment Processes. Simple Options, CRC Press, Inc., Florida, 1995.
pt
882
[8] Meenakshi, R.C. Maheshwari, Fluoride in drinking water and its removal, J. Hazard.
[10] J.J. Murray, Appropriate Use of Fluorides for Human Health, World Health
Ac ce
881
M
879
Organisation, Geneva. 1986.
887
[11] P. Loganathan, M.J. Hedley, N.D. Grace, J. Lee, S.J. Cronin, N.S. Bolan, J.M. Zanders,
888
Fertiliser contaminants in New Zealand grazed pasture with special reference to
889
cadmium and fluorine - a review, Aust. J. Soil Res. 41 (2003) 501-532.
890
[12] Y.H. Li, S. Wang, X. Zhang, J. Wei, C. Xu, Z. Luan, D. Wu, Adsorption of fluoride
891
from water by aligned carbon nanotubes, Materials Res. Bull. 38 (2003) 469-476.
892
[13] X. Liao, B. Shi, Adsorption of fluoride on zirconium(IV)-impregnated collagen fiber.
893
Environ. Sci. Technol. 39 (2005) 4628-4632. 37
Page 37 of 62
894
[14] S.P. Mishra, M. Das, U.N. Dash, Review on adverse effects of water contaminants like
895
arsenic, fluoride and phosphate and their remediation, J. Scientific Industrial Res. 69
896
(2010) 249-253. [15] E.G. Paulson, Reducing fluoride in industrial wastewater, Chem. Eng. 84 (1977) 89-94.
898
[16] S.S. Babu, S. Kumar, P.D. Nemade, T. Roychowdhury, Impacts of fluoride in drinking
ip t
897
water and role of pH in its removal efficiency: a critical review, Crit. Rev. Environ. Sci.
900
Technol. (2012) (in press).
904 905
us
903
defluoridation of water, Chem. Rev. (2012) 2454-2466.
[18] W.W. Choi, K.Y. Chen, The removal of fluoride from waters by adsorption, J. Amer.
an
902
[17] S. Jagtap, M.K. Yenkie, N. Labhsetwar, S. Rayalu, Fluoride in drinking water and
Water Works Assoc. 71 (1979) 562-570.
[19] K.B.P.N. Jinadasa, S.W.R. Weerasooriya, C.B. Dissanayake, A rapid method for the
M
901
cr
899
defluoridation of fluoride-rich drinking waters at village level, Inter. J. Environ. Studies.
907
31 (1988) 305-312.
ed
906
[20] J.P. Padmasiri, C.B. Dissanayake, A simple defluoridator for removing excess fluorides
909
from fluoride-rich drinking water, Intern. J. Environ. Health Res. 5 (1995) 153-160.
910
[21] C. Zevenbergen, L.P. van Reeuwijk, G. Frapporti, R.J. Louws, R.D. Schuiling, A simple
Ac ce
pt
908
911
method for defluoridation of drinking water at village level by adsorption on Ando soil
912
in Kenya, Sci. Total Environ. 188 (1996) 225-232.
913
[22] M.A.M. Sahli, S. Annouar, M. Tahaikt, M. Mountadar, A. Soufiane, A. Elmidaoui,
914
Fluoride removal for underground brackish water by adsorption on the natural chitosan
915
and by electrodialysis, Desalination 21 (2007) 37-45.
916
[23] S. Samatya, U. Yüksel, M. Yüksel, N. Kabay, Removal of fluoride from water by metal
917
ions (Al3+, La3+ and ZrO2+) loaded natural zeolite, Sep. Sci. Technol. 42 (2007) 2033-
918
2047. 38
Page 38 of 62
919
[24] K.R. Bulusu, B.B. Sundaresan, B.N. Pathak, W.G. Nawlakhe, D.N. Kulkarni, V.P.
920
Thergaonkar, Fluorides in water, defluoridation methods and their limitations, J. Inst.
921
Engineers (India) Environ. Engineer. Div. 60 (1979) 1-25.
926 927 928
ip t
925
[26] M. Mohapatra, S. Anand, B.K. Mishra, D.E. Giles, P. Singh, Review of fluoride removal from drinking water, J. Environ. Manage. 91 (2009) 67-77.
cr
924
alum by Nalgonda technique, Indian. J. Environ. Health 17 (1975) 26-65.
[27] A. Bhatnagar, E. Kumar, M. Sillanpää, Fluoride removal from water by adsorption-A
us
923
[25] W.G. Nawlakhe, D.N. Kulkarni, B.N. Pathak, K.R. Bulusu, Defluoridation of water with
review, Chem. Eng. J. 171 (2011) 811-840.
[28] V.K. Gupta, J.M. Carrott, M.M.L.R. Carrott, Suhas, Low-cost adsorbents: Growing
an
922
approach to wastewater treatment – a review. Crit. Rev. Environ. Sci. Technol. 39
930
(2009) 783-842.
M
929
[29] S. Teng, S. Wang, W. Gong, X. Liu, B. Gao, Removal of fluoride by hydrous manganese
932
oxide-coated alumina: Performance and mechanism, J. Hazard. Mater. 168 (2009), 1004-
933
1011.
936 937
pt
935
[30] S. Meenakshi, N. Viswanathan, Identification of selective ion-exchange resin for fluoride sorption, J. Colloid Interface Sci. 308 (2007) 438-450.
Ac ce
934
ed
931
[31] L. Ruixia, G. Jinlong, T. Hongxiao, Adsorption of fluoride, phosphate, and arsenate ions on a new type of ion exchange fiber, J. Colloid Interface Sci. 248 (2002) 268-274.
938
[32] F. Helferich, Ion Exchange, Dover Publication Inc., USA, 1995.
939
[33] E.R. Weiner, Applications of Environmental Aquatic Chemistry. CRC Press, Taylor and
940 941 942
Francis Group, Florida, 2008. [34] I.B. Solangi, S. Memon, M.I. Bhanger, An excellent fluoride sorption behaviour of modified amberlite resin, J. Hazard. Mater. 176 (2010) 186-192.
39
Page 39 of 62
943
[35] A. Sivasamy, K.P. Singh, D. Mohan, M. Maruthamuthu, Studies on defluoridation of
944
water by coal-based sorbents, J. Chem. Technol. Biotechnol. 76 (2001) 717-722.
946 947
[36] S.M. Maliyekkal, A.K. Sharma, L. Philip, Manganese-oxide-coated alumina: A promising sorbent for defluoridation of water, Water Res. 40 (2006) 3497-3506. [37] S. Tokunaga, S.A. Haron, S.A. Wasay, K.F. Wong, K. Laosangthum, A. Uchiumi,
ip t
945
Removal of fluoride ions from aqueous solutions by multivalent metal compounds, Inter.
949
J. Environ. Studies 48 (1995) 17-28.
953 954
us
952
of drinking water, Appl. Geochem. 16 (2001) 531-539.
[39] F. Luo, K. Inoue, The removal of fluoride ion by using metal (III)-loaded Amberlite
an
951
[38] Y. Wang, E.J. Reardon, Activation and regeneration of a soil sorbent for defluoridation
resins, Solvent Extract. Ion Exch. 22 (2004) 305-322.
[40] M.S. Onyango, Y. Kojima, O. Aoyi, E.C. Bernardo, H. Matsuda, Adsorption equilibrium
M
950
cr
948
modelling and solution chemistry dependence of fluoride removal from water by
956
trivalent-cation-exchanged zeolite F-9, J. Coloid Interface Sci. 279 (2004) 341-350.
957
ed
955
[41] M.S. Onyango, Y. Kojima, A. Kumar, D. Kuchar, M. Kubota, H. Matsuda, Uptake of fluoride by Al3+ pretreated low-silica synthetic zeolites: adsorption equilibrium and rate
959
studies, Sep. Sci. Technol. 41 (2006) 683-704.
961
Ac ce
960
pt
958
[42] R.L. Ramos, J. Ovalle-Turrubiartes, M.A.Sanchez-Castillo, Adsorption of fluoride from aqueous solution on aluminium-impregnated carbon, Carbon 37 (1999) 609-617.
962
[43] S.A. Wasay, M.J. Haron, S. Tokunaga, Adsorption of fluoride, phosphate, and arsenate
963
ions on lanthanum-impregnated silica gel, Water Environ. Res. 68 (1996) 295-300.
964 965 966 967
[44] Y. Zhou, C. Yu, Y. Shan, Adsorption of fluoride from aqueous solution on La3+impregnated cross-linked gelatin, Sep. Purif. Technol. 36 (2004) 89-94. [45] P. Trivedi, L. Axe, Long-term fate of metal contaminants in soils and sediments: role of intraparticle diffusion in hydrous metal oxides, in: R. Hamon, M. McLaughlin, E. Lombi 40
Page 40 of 62
968
(Eds.), Natural Attenuation of Trace Element Availability in Soils, CRC Taylor and
969
Francis Group, New York, 2006, pp. 57-71.
975 976 977 978 979
ip t
974
fluoride on activated alumina, Sep. Purif. Technol. 42 (2005) 265-271.
[48] M.G. Sujana, R.S. Thakur, S.B. Rao, Removal of fluoride from aqueous solution by
cr
973
[47] S. Ghorai, K.K. Pant, Equilibrium, kinetics and breakthrough studies for adsorption of
using alum sludge, J. Colloid Interface Sci. 206 (1998) 94-101.
us
972
adsorption on fly ash, Water Air Soil Pollut. 49 (1990) 51-61.
[49] Y. Ku, H. Chiou, The adsorption of fluoride ion from aqueous solution by activated alumina, Water Air Soil Pollut. 133 (2002) 349-360.
an
971
[46] A.K. Chaturvedi, K.P. Yadava, K.C. Pathak, V.N. Singh, Defluoridation of water by
[50] A.M. Raichur, M.J. Basu, Adsorption of fluoride onto mixed rare earth oxides, Sep. Purif. Tech. 24 (2001) 121-127.
M
970
[51] L.E.L. Hammari, A. Laghzizil, P. Barboux, K. Lahlil, A. Saoiabi, Retention of fluoride
981
ions from aqueous solution using porous hydroxyapatite. Structure and conductive
982
properties, J. Hazard. Mater. B114 (2004) 41-44.
pt
984
[52] X. Fan, D.J. Parker, M.D. Smith, Adsorption kinetics of fluoride on low cost materials, Water Res. 37 (2003) 4929-4937.
Ac ce
983
ed
980
985
[53] H. Wang, J. Chen, Y. Cai, J. Ji, L. Liu, H.H. Teng, Defluoridation of drinking water by
986
Mg/Al hydrotalcite-like compounds and their calcined products, Appl. Clay Sci. 35
987
(2007) 59-66.
988
[54] L. Lv, J. He, M. Wei, D.G. Evans, X. Duan, Factors influencing the removal of fluoride
989
from aqueous solution by calcined Mg-Al-CO3 layered double hydroxides, J. Hazard.
990
Mater. B133 (2006) 119-128.
41
Page 41 of 62
991
[55] D.P. Das, J. Das, K. Parida, Physicochemical characterization and adsorption behaviour
992
of calcined Zn/Al hydrotalcite-like compound (HTlc) towards removal of fluoride from
993
aqueous solution, J. Colloid Interface Sci. 261 (2003) 213-220. [56] C. Diaz-Nava, M. Solache-Rios, M.T. Olguin, Sorption of fluoride ions from aqueous
995
solutions and well drinking water by thermally treated hydrotalcite, Sep. Sci. Technol.
996
38 (2003) 131-147.
cation synthetic resin, Sep. Sci. Technol. 37 (2002) 89-103.
cr
998
[57] Y. Ku, H. Chiou, W. Wang, The removal of fluoride ion from aqueous solution by a
us
997
ip t
994
[58] K.M. Popat, P.S. Anand, B.D. Dasare, Selective removal of fluoride ions from water by
1000
the aluminium form of the aminomethylphosphonic acid-type ion exchanger, React.
1001
Polym. 23 (1994) 23-32.
[59] A.A.M. Daifullah, S.M. Yakout, S.A. Elreefy, Adsorption of fluoride in aqueous
M
1002
an
999
solutions using KMnO4-modified activated carbon derived from steam pyrolysis of rice
1004
straw, J. Hazard. Mater. 147 (2007) 633-643.
1007
fluoride ions onto carbonaceous materials, J. Colloid Interface Sci. 275 (2004) 35-39.
pt
1006
[60] I. Abe, S. Iwasaki, T. Tokimoto, N. Kawasaki, T. Nakamura, S. Tanada, Adsorption of
[61] Y.H. Li, S. Wang, X. Zhang, J. Wei, C. Xu, Z. Luan, D. Wu, B. Wei, Removal of
Ac ce
1005
ed
1003
1008
fluoride from water by carbon nanotube supported alumina, Environ. Technol. 24 (2003)
1009
391-398.
1010 1011 1012 1013
[62] P.D. Nemade, A.V. Rao, B.J. Alappat, Removal of fluorides from water using low cost adsorbents, Water Sci. Tech. Water Supply 2 (2002) 311-317. [63] N. Viswanathan, S. Meenakshi, Selective fluoride adsorption by a hydrotalcite/chitosan composite, Appl. Clay Sci. 48 (2010) 607-611.
1014
[64] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Fluoride sorption by nano-
1015
hydroxyapatite/chitin composite, J. Hazard Mater. 172 (2009) 147-151. 42
Page 42 of 62
1016
[65] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Uptake of fluoride by nano-
1017
hydroxyapatite/chitosan, a bioinorganic composite, Bioresource Technol. 99 (2008)
1018
8226-8230.
1019
[66] W. Ma, F. Ya, M. Han, R. Wang, Characteristics of equilibrium, kinetics studies for adsorption of fluoride on magnetic-chitosan particle, J. Hazard. Mater. 143 (2007) 296-
1021
302.
[67] S.P. Kamble, S. Jagtap, N.K. Labhsetwar, D. Thakare, S. Godfrey, S. Devotta, S.S.
cr
1022
ip t
1020
Rayalu, Defluoridation of drinking water using chitin, chitonsan and lanthanum-
1024
modified chitosan, Chem. Eng. J. 129 (2007) 173-180.
us
1023
[68] M.G. Sujana, H.K. Pradhan, S. Anand, Studies on sorption of some geomaterials for
1026
fluoride removal from aqueous solutions, J. Hazard. Mater. 161 (2009) 120-125.
1027
[69] M. Agarwal, K. Rai, R. Shrivastav, S. Dass, Defluoridation of water using amended
1031 1032 1033
M
ed
1030
[70] M. Srimurali, A. Pragathi, J. Karthikeyan, A study on removal of fluorides from drinking water by adsorption onto low-cost materials, Environ. Pollut. 99 (1998) 285-289. [71] M. Islam, R. Patel, Thermal activation of basic oxygen furnace slag and evaluation of its
pt
1029
clay, J. Cleaner Production 11 (2003) 439-444.
fluoride removal efficiency, Chem. Eng. J. 169 (2011) 68-77.
Ac ce
1028
an
1025
[72] A. Tor, N. Danaoglu, G. Arslan, Y. Cengeloglu, Removal of fluoride from water by
1034
using granular red mud: Batch and column studies, J. Hazard. Mater. 164 (2009) 271-
1035
278.
1036 1037 1038 1039
[73] V.K. Gupta, I. Ali, V.K. Saini, Defluoridation of wastewaters using waste carbon slurry, Water Res. 41 (2007) 3307-3316. [74] R. Piekos, S. Paslawska, Fluoride uptake characteristics of fly ash, Fluoride 32 (1999) 14-19.
43
Page 43 of 62
1040 1041 1042
[75] Y.D. Lai, J.C. Liu, Fluoride removal from water with spent catalyst, Sep. Sci. Technol. 31 (1996) 2791-2803. [76] Y. Tang, X. Guanc, J. Wang, N. Gaob, M.R. Mcphaild, C.C. Chusuei, Fluoride adsorption onto granular ferric hydroxide: effects of ionic strength, pH, surface loading,
1044
and major co-existing anions, J. Hazard. Mater. 171 (2009) 774-779.
ip t
1043
[77] Y. Zhou, R.J. Haynes, Sorption of heavy metals by inorganic and organic components of
1046
solid wastes: Significance to use of wastes as low-cost adsorbents and immobilising
1047
agents, Crit. Rev. Environ. Sci. Technol. 40 (2011) 909-977.
us
1048
cr
1045
[78] N. Chubar, Physico-chemical treatment of micropollutants: Adsorption and ion exchange, in: J. Virkutyte, R.S. Varma, V. Jegatheesan, (Eds.), Treatment of
1050
Micropollutants in Water and Wastewater, IWA publishing, London, 2010, pp. 165-203. [79] W.H. Zou, R.P. Han, Z.Z. Chen, J. Shi, H.M. Liu, Characterization and properties of
M
1051
an
1049
manganese oxide coated zeolite as adsorbent for removal of copper (II) and lead (II) ions
1053
from solution, J. Chem. Eng. Data 51 (2006) 534-541.
1056 1057 1058 1059 1060 1061 1062 1063
the accumulation of arsenic in lake sediments, Water Res. 9 (1985) 1029-1032.
pt
1055
[80] T. Takamatsu, M. Kawashima, M. Koyama, The role of Mn2+ rich manganese oxide in
[81] E.N. Peleka, E.A. Deliyanni, Adsorptive removal of phosphates from aqueous solutions,
Ac ce
1054
ed
1052
Desalination 245 (2009) 357-371. [82] K. Goh, T. Lim, Z. Dong, Application of layered double hydroxides for removal of oxyanions: A review, Water Res. 42 (2008) 1343-1368. [83] K. Vaaramma, J. Lehto, Removal of metals and anions from drinking water by ion exchange, Desalination 155 (2003) 157-170. [84] S. Wang, Y. Peng, Natural zeolites as effective adsorbents in water and wastewater treatment, Chem. Eng. J. 156 (2010) 11-24.
44
Page 44 of 62
1064 1065 1066 1067
[85] C. Diaz-Nava, M.T. Olguin, M. Solache-Rios, Water defluoridation by Mexican heulandite-clinoptilolite, Sep. Sci. Technol. 37 (2002) 3109-3128. [86] P. Chingombe, B. Saha, R.J. Wakeman, Surface modification and characterisation of a coal-based activated carbon, Carbon 43 (2005) 3132-3143. [87] C.Y. Yin, M.K. Aroua, W.M.A.W. Daud, Review of modifications of activated carbon
1069
for enhancing contaminant uptakes from aqueous solutions, Sep. Purif. Technol. 52
1070
(2007) 403-415.
1074 1075
cr
us
1073
Purif. Technol. 70 (2010) 329-337.
[89] A. Bhatnagar, M. Sillanpää, A review of emerging adsorbents for nitrate removal from water. Chem. Eng. J. 168 (2011) 493-504.
an
1072
[88] R. Mahmudov, C.P. Huang, Perchlorate removal by activated carbon adsorption. Sep.
[90] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Defluoridation chemistry of synthetic
M
1071
ip t
1068
hydroxyapatite at nano scale: equilibrium and kinetic studies, J. Hazard Mater. 155
1077
(2008) 206-215.
ed
1076
[91] A. Bhatnagar, V.J.R. Vilar, C.M.S. Botelho, R.A.R. Boaventura, A review of the use of
1079
red mud as adsorbent for the removal of toxic pollutants from water and wastewater,
1080
Environ. Technol. 32 (2011) 231-249.
1082 1083 1084
Ac ce
1081
pt
1078
[92] D. Mohan, C.U. Pittman Jr, Arsenic removal from water/wastewater using adsorbents-A critical review. J. Hazard. Mater. 142 (2007) 1-53. [93] H. Genc-Fuhrman, H. Bregnhoj, D. McConchie, Arsenate removal from water using sand-red mud columns, Water Res. 39 (2005) 2944-2954.
1085
[94] C. Zhu, Z. Luan, Y. Wang, X. Shan, Removal of cadmium from aqueous solutions by
1086
adsorption on granular red mud (GRM), Sep. Purif. Technol. 57 (2007) 161-169.
45
Page 45 of 62
1087
[95] I.A.M. Yunusa, P. Loganathan, S.P. Nissanka, V. Manoharan, M.D. Burchett, C.G.
1088
Skilbert, D. Eamus, Application of coal fly ash in agriculture: a strategic perspective,
1089
Crit. Rev. Environ. Sci. Technol.42 (2012) 559-600. [96] W. Nigussie, F. Zewge, B.S. Chandravanshi, Removal of excess fluoride from water
1091
using waste residue from alum manufacturing process, J. Hazard. Mater. 147 (2007)
1092
954-963.
ip t
1090
[97] R.H. Masue, T.A. Loeppert, T.A. Kramer, Arsenate arsenite adsorption and desorption
1094
behaviour on coprecipitated aluminium:iron hydroxides, Environ. Sci. Technol. 41
1095
(2007) 837-842.
us
cr
1093
[98] L. Yan, Y. Xu, H. Yu, X. Xin, Q. Wei, B. Du, Adsorption of phosphate from aqueous
1097
solution by hydroxyl-aluminum, hydroxyl-iron and hydroxyl-iron-aluminium pillared
1098
bentonites, J. Hazard. Mater. 179 (2010) 244-250.
1102
M
ed
1101
1989.
[100] N.Y. Mezenner, A. Bensmaili, Kinetics and thermodynamic study of phosphate adsorption on iron hydroxide-eggshell waste, Chem. Eng. J. 147 (2009) 87-96.
pt
1100
[99] S.D. Faust, O.M. Aly, Adsorption Processes for Water Treatment, Butterworths, Boston,
Ac ce
1099
an
1096
46
Page 46 of 62
ip t cr
an
Initial (I), pH; Adsorption (Ads) capacity (mg/g) and Temperat- Equilibrium other results. Maximum (max) o (E) ure ( C) concentration (mg/L); Adsorbent concentration (g/L); column ht, height; d, diameter; FL, flow rate Metal oxides and hydroxides 3-12; I, 1-100; Max ads pH 3-8, then decreased with 10, 25 2 increased pH to 12; Langmuir ads max (pH 6-7) 10oC, 3.68; 25oC, 5.97
Best kinetic model to fit data;
Best equilibrium model to fit data
Reference
Pseudo-first order, Bangham model Pseudo-first, pseudosecond, diffusion
Langmuir
[7]
Langmuir
[29]
Pseudo-first, pseudosecond, diffusion
Langmuir
[36]
Equilibrium (equil)
M
Adsorption method: batch (B), column (C); Water type: synthetic (S), wastewater (W)
B; S
Manganese oxide coated alumina
B, C; S
Manganese oxide coated alumina
B, C; S
B, 3-12; 25 C, 5.2; 25
Ac c
Granular ferric hydroxide
ep te
d
Adsorbent
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents
B, 3-12; 30 C, 7; 30
B, I, 6.242.1; C, I 5; 140; 50cm ht, 2.4 cm d, FL 2.39 m3/m2h B, I, 2.5-30 C, 3.56; 0.5x0.028m d, 0.3mht, FL 2.19 m3/m2h
B, Max ads pH 4-6, decreased with increased pH from 6 to 12. Langmuir ads max 7.1 at pH 5.2 C, breakthrough point 669 bed volume B, max ads pH 4-7, lowest at pH 12; Langmuir ads max 2.85 at pH 7 C, Bed saturation F concentration 1.25 g/L 47
Page 47 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
C, 3.56; 0.5x0.028m d, 0.3mht, FL 2.19 m3/m2h B, I, 2.5-14; 4-40 C, 5; 200 550 mm ht, 51mm d, FL 20,30mL/min -; 12.5
Pseudo-first, pseudosecond, diffusion
Langmuir
[36]
Surface adsorption, diffusion
Langmuir and Freundlich
[47]
Max ads pH 5-7 Langmuir ads max 16.3 at pH 6
Pseudosecond order
[49]
First 5 min most F ads; 40 min equil -
Langmuir and Freundlich Langmuir -
[51]
Langmuir
[52]
us
B, max ads pH 4-7, lowest at pH 12; Langmuir ads max 1.08 at pH 7
B, I, 2.5-30
C, Bed saturation F concentration 0.47 g/L max ads pH 7 Langmuir ads max 2.41 at pH 7
B; S
Rare earth oxides
B; S
3-11; 29
E 1-30; 1-8
Max ads pH 6-6.5 Langmuir ads max 196 at pH 6.5
Hydroxyapatite (HA)
B; S
I, 19-19000; 10
Hydroxyapatite (HA)
B; S
-; Room temp 6; Room temp
At initial F concentration of ≤190 mg/L, 60-80% F removed by porous HA and 30-35% by crystalline HA >90% F adsorption; Langmuir ads max 4.54 compared to fluorspar 1.79, activated quartz 1.16, calcite 0.39, quartz 0.19
M
Activated alumina
B, 4-10; Room temp C, 7; room temp 4-11; 30
96% removal
d
ep te
B, C; S
B, 3-12; 30 C, 7; 30
Ac c
Activated alumina
B, C; S
an
Activated alumina
I;2.5 x 10-5-6 x 10-2; 17
Pseudosecond order
[50]
48
Page 48 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
7; 25
B; S
5-10; 30
Zn/Al LDH calcined (450oC) Na/Mg/Al LDH calcined (500oC)
B; S
4-10; 30
B; S, well water
S: 5,7,9; Room temp W:8.4,8.5 Room temp
us
B; S
an
Mg/Al LDH calcined and uncalcined Mg/Al LDH calcined (200-800oC)
B, I, 10.4-104 Ads max pH 4.7-7.3, Langmuir ads max 0.1 3.8 at pH 7 C, I, 72; 4; 9 cm ht, 1 cm Column ads capacity 3.4 d, FL 0.5 mL/min Layered double hydroxides (LDH) E, 0-70; Calcination increased F ads (130oC 2.5 optimum). Ads capacity 35 at equilibrium F concentration 70 mg/L E, 0-250; Max ads pH 6.0; Langmuir ads max 213 1 at pH 6. Ads highest for calcination temp 500oC. Ads Mg/Al LDH>Ni/Al, Zn/Al LDHs, Mg/Al molar ratio 2:1 best E, 2-60; Max ads pH 6; Langmuir ads max 17 at 1 pH 6. Max ads 13.43 for adsorbent dose 0.2g/L, F concentration 10 mg/L, pH 6 S: I, 5; Distribution coefficient between solid 10 and solution, S: pH 5, 3501; pH 7, W: I, 5.9, 6.9; 1047; pH 9, 1155 10 W: 414-437
M
B 4-9; 20 C 7.5; 20
Ac c
Pseudo-firast order
Langmuir
[43]
adsorption reached max at 15 min -
-
[53]
Freundlich
[54]
ads reached equil in 4 h
Langmuir
[55]
80-97% F removal in 1h
-
[56]
d
B; W C; W
ep te
La impregnated silica gel
49
Page 49 of 62
ip t cr
Chelating resin (CR), anion exchange resin (AER) Metals loaded Amberlite resin
B; S, field water
3-11; 30
B, C; Field water
B: 1-8; 30 C: 6.0; 30
Al-Amberlite resin
B, C; S
B: E, 2-13; 10 C: I, 16; 2; 0.8 mm d, 50 mm ht, FL,1 mL/min I, 2-10; 20
ep te
B: E, 0-60; 1.6 C: I, 7.9; 2; 8 mm d, FL 6 mL/h -; B: 4-9.1; 30 C: 5.5, 6.7 C: I, 40; 30 20-50; 2cm d, 16cm ht, FL 280700 mL/h
Ac c
an
B: 1-10; 25 C: 7; 25
M
B, C; S
Ion exchange resins and fibres B: ads max at pH 7 for TUA and pH 910 for amberlite (A). At pH 7, ads capacity of TUA three times that of A. Langmuir max for TUA at pH 7, 61. C: column ads capacity 50 pH had no significant effect on ads; Langmuir ads max at pH 7: CR 1.3, AER 1.5. At low F concentration CR removed 30% more F than AER. Field water: CR had higher F removal B: Langmuir ads max at pH 7, resin with La, Ce, Al, 25; Fe, 49; Y, 19. pH for ads max: Ce 4-7, Fe 3, Al 5-9, L 3-7 C: column ads capacity for La resin at pH 6, 20. Bed saturation 300 bed vol. B: max ads at pH 4-7, low ads pH 9, Langmuir ads max 4.6 at pH 4 C: column ads capacity at pH 5.5 and 6.7 were 1.13 and 0.72 (FL 460 mL/h)
d
Thio-urea modified Amberlite (TUA)
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
Ads reached equil in 30 min
Langmuir
[34]
Pseudosecond order
Langmuir
[30]
Ads equil time: La 2 h; Ce 1 h; Fe, Al, Y > 5 h
Langmuir
[39]
Pseudosecond order
Langmuir
[57]
50
Page 50 of 62
ip t cr
B: 2-6; 25 C: -; 25
B: E, 0-60; 4 C: I, 10; 1 1.5cm d, 5cm ht, FL 2.5-3.5 mL/min E, 0-100; 5
B, Tap water
2-9; 30
Al, La, Zr loaded natural zeolites (Z) Al loaded four synthetic zeolites Al, La loaded synthetic zeolites
B; S, tap water (T)
-; 30
B; S
2-11; 25
S: I, 1-20; 2 T: I, 2.9; 6 I, 5-80; 0.05-0.2
B; S, field water (F)
S: 3.5-9; 20-40 F: 7.4; 30
S: I, 10-80; 2 F: I, 3.3, 4; 1-4
ads max at pH 2.6-6.9. Ads capacity 30 for 5 g adsorbent at E = 100 mg/L Zeolites S: Langmuir ads max : Zr (Z) 3.4-4.1, La(Z) 2.4-2.6, Al(Z) 2.0-2.4 T: F removal(%): Zr(Z) 91.1, La(Z) 90.4, Al(Z) 89.7 Ads max pH 4-8; F ads at equilibrium concentration 40 mg F/L and pH 4-6 was 6-16 S: ads max at pH 6-9; Langmuir ads max 20oC, Al(Z) 34, La(Z) 45.At 40 mg/L equilibrium conc., ads capacity 16 vs 8 for Al(Z) vs La(Z). (F): F conc reduction more by Al(Z) than La(Z)
ep te
d
Al-chelating porous anion exchanger
Ac c
B: ads max at pH 3.0 and decreased with increased pH to 6; Freundlich adsorption constant PO43- > AsO43- > FC: column ads capacity for PO43, AsO43, F- 156, 96, 45, respectively
an
B, C; S
M
Ion exchange fibre
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d) Pseudosecond order
Freundlich
[31]
1 h equil time (95% adsorption)
-
[58]
-
Langmuir
[23]
Elovich
RedlichPeterson
[41]
-
RedlichPeterson
[40]
Equil reached in 5 min
51
Page 51 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
4.6-9.2; 25
E, 0-20
B; S
3-11; 25
E, 0-16
B; S
3-11; 25
-
B; S
2-12; -
I, 4-30; -
B; S
us
B; S
ep te
3-7; 25
Carbon materials Ads max at pH 2, then decreased with pH. Langmuir ads max at 25oC, 15.9. Ads decreased with temperature increase Ads capacity of bone char increased with pH. Freundlich adsorption constant highest for bone char and lowest for carbon black Ads max at pH 7; ads capacity at equilibrium concentration 10 mg F/L for activated carbon, γ Al2O3, and ACNT: 0.32, 3.7, and 4.1, respectively Ads max at pH 6-9; ads max of 9.6 at pH 6 and calcined temp 450oC for 0.5 mg/L adsorbent and initial concentration 6 mg F/L Ads decreased from pH 2 to pH 12; Fish bone charcoal had the highest F ads
an
E, 5-20; -
M
2-10; 25, 45, 55
d
B; S
Ac c
KMnO4 modified activated carbon Bone char, activated carbon, carbon black Aligned carbon nanotubes (ACNT) Alumina loaded CNT (calcined, 250-1050oC) Wood, animal , fish bone activated charcoal Al impregnated activated carbon
I, 0.5-15 E, 0-6.5; -
Ads decreased from pH 3 to 7. Calcining at 300oC gave the highest ads among 300-1000oC. Langmuir ads max 1.07, plain carbon 0.49
Pseudosecond order
LangmuirFreundlich
[59]
-
Freundlich
[60]
-
Freundlich
[12]
Equil reached at 20 h
Freundlich
[61]
Most of F removed at 2 h
Langmuir, Freundlich
[62]
-
Langmuir, Freundlich
[42]
52
Page 52 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
Chitin/nanohydroxyapatite composite Chitosan/nano hydroxyapatite composite
B; S
3-11; 30
I, 6-12; 2
B; S
3-11; 30
I, 9-15; 2
Magnetic (Fe)-chitosan
B; S
S: 5-9; 30
I, 5-140; 1
La loaded chitosan
B; S, field water (F)
S: 5-9; 30 F: 7; 30
S: E, 0-15; F: I, 10.2; -
Natural chitosan Zr loaded collogen
B; S B; S
2-10; 20 3.5-11; 30
-
Natural materials Ads max at pH 3, decreased to pH 11. At pH 7, 10 mg F/L, 2 g adsorbent/L, composite ads 1.3; LDH alone 1.0, chitosan alone 0.05; Langmuir ads max 1.9 for composite Ads max at pH 3, decreased with increased pH to pH 11; Langmuir ads max 8.4 at pH 7 Ads max at pH 3, decreased with increased pH to pH 11; At pH 7, 10 mg F/L, adsorbent 5 mg/L, composite ads capacity 1.56, hydroxyapatite 1.30, chitosan 0.05; Langmuir ads max 2.04 pH no significant effect. Ads higher than activated alumina. Langmuir twosite max ads 24
us
I, 9-15; 2
an
3-11; 30
M
B; S
Ac c
ep te
d
Chitosan/ LDH composite
I, 19-95; 1
S: Ads max pH 6.7; max ads 5.5 at pH 6.7 and equilibrium conc.15 mg F/L F: ads capacity 1 compared to 2 for distilled water at equilibrium conc. 8 mg F/L Max ads at pH 6; Langmuir ads max 1.39 at pH 6 Max ads pH 5-8, drastic decrease from pH 9 onwards. Langmuir ads max 2.18 at pH 5-8
PseudoLangmuir second order, diffusion
[63]
PseudoLangmuir second order, Freundlich diffusion PseudoLangmuir second order, diffusion
[64]
Langmuir Equil at 90 min; Pseudo- (one and second order two sites)’ Bradley Pseudo-first order, diffusion
[66]
Equil at 5 min Equil at 500 min
[22]
Langmuir, Freundlich Langmuir
[65]
[67]
[13]
53
Page 53 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
B; S, field water
Al, Fe, Ca loaded silty clay (C)
B; S
S: E, 5-40; 2 F: I, 10.25; 14 I, 10; 100
Soil with high Fe oxide (calcined 500oC)
B, C; S
S: 2.5-10; 32 F: 7.75; 32 Al, Ca-C, C alone: 7-8; Fe-C 2.7-3.0; 30 B: 7; 25 C: 7; 25
Volcanic ash soil
B; S
2.5-9.5; 32
I, 5; 1-8
Increased pH, decreased ads, Be: 46 to 29%, Ka: 38 to 5% for 2g/L adsorbent
ep te
B: max ads 1 at equilibrium concentration 60 mg/L
-; 18
B:I, 4.8-95; 33 C: I, 4; 47; 50 mL column, FL 0.2 mL/min I, 0-50; 40
Ac c
Bentonite B; (Be), S kaolinite (Ka)
pH 2-5, > 90% ads, then decreased with increased pH to pH 12 (17%); Langmuir ads max 21.3 at pH 5-7 S: ads max at pH 5-6 and 3-5; Langmuir ads max 12-15 at pH 5 F: 3 repeated stages of ads reduced F concentration to <0.5 mg/L % F removal: C alone 25, Al-C 78-94’ Ca-C 32-52, Fe-C 94-98
us
E: 0-50; 4
an
2-12; 29
M
B; S
d
La loaded cross-linked gelatin Laterites with Ni
C: F concentration reached 1 mg/L after 120 pore volumes. Column ads capacity 0.15 At equilibrium concentration of 19 mg/L, ads capacity was 2.9 Langmuir ads max 5.5
Equil at 40 min; pseudofirst order Pseudo-first order, diffusion
Langmuir
[44]
Langmuir
[68]
-
-
[69]
-
Freundlich
[38]
Equil in ~2h Langmuir at low conc, >24 h at high conc -
[21]
[70]
54
Page 54 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
us
B; S C; S
B:2.5-7.3; 25 C: 4.7; 25
B; S
1-10; -
B: I, 5-150; 2.5 C: I, 5;10, 0.635 cm2 area,15cm ht, FL 2 mL/min I, 100-1000; 1-8.4
B; S C; W
B:2- 11; 25 C: -; 25
Fly ash (class F, 9.1% CaO)
C; S
Fly ash (2.22 % CaO) Alum sludge (calcined and uncalcined)
B; S B; S
10.1; 20
2-9.5; 30 3.5-8.8; 32
Industrial by-products Ads increased from 21 to 93% as pH increased from 2 to 7, then nearly constant; Langmuir ads max at 25oC, 4.6 and 45oC, 8.1 B: ads max at pH 4.7; Langmuir ads max 8.92 at pH 4.7 C: column total ads capacity 2.05 (0.64 by batch trial for initial F, 5 mg/L
an
I, 1-50; 5
M
2-10; 25
Ads max at pH 5.5; ads capacity A 4.8, UA 1.0 at equilibrium concentration 20 mg/L; Langmuir ads max A 6.3, UA 3.1at pH 5.5 B: ads max at pH 7.58; Langmuir ads max 4.3 at pH 7.5
d
ep te
Activated red mud (A), unactivated (UA) Waste carbon slurry 450oC (activated)
B; S
Ac c
Basic oxygen furnace slag (BOFS) heated1000oC Granular red mud
I, 1-11;1 C: I, 11; 0.5; 0.9 cm2 area, 3.1 cm ht, FL 1.5 mL/min I, 0-100; 450; 40 cm ht, 4.5 cm d, FL 2 mL/h E, 0-3; 20 E, 0-15; 0.5-16
C: breakthrough column ads capacity 4.16 F concentration in effluent reached 0 mg/L after 120-168 h Ads max at pH 6.5; Langmuir ads max 20 at pH 6.5 Ads max at pH 6; calcined higher ads capacity than uncalcined; Langmuir ads max 5.39 at pH 6
Equil at 35 min; pseudofirst order
Langmuir
[71]
Equil at 6 h; pseudosecond order
Freundlich, RedlichPeterson
[72]
Equil at 2 h
Langmuir
[4]
Equil at 1 h; Pseudo-first order
RedlichPeterson
[73]
-
-
[74]
pseudo-first order pseudo-first order; diffusion
Langmuir
[46]
-
[48]
55
Page 55 of 62
ip t cr
E, 0-40; 10
Ads max (1.6 for initial F 20 mg/L, pseudo-first adsorbent 10 g/L) at pH 2, then continue order; 70% to decrease. Max ads 28 ads in 10min
-
[75]
d
M
an
us
2-9.5; 25
ep te
B; S
Ac c
Spent catalyst
56
Page 56 of 62
ΔH0 (kJ/mol)
ΔS0 (J/mol/0C)
Reference
22.7
8.4
[71]
8.23
0.1
[63]
Granular ferric hydroxide Geomaterials
-39.2 25 -5.8 to -15.9 32 to 62
7.34
15.8
[7]
-0.1 to -30
-3 to -10
Nanohydroxyapatite/ chitin composite Nanohydroxyapatite/ chitosan composite
-7.1 30
-7.0 40
-6.8 50
11.8
-3.8 30
-3.4 40
-3.4 50
8.6
Modified activated carbon Waste carbon slurry
-3.6 -0.9 25 45 -25.4 -26.5 25 35
Chelating ion exchange resin
-5.7 30
Calcined Zn/Al LDH
-25.0 -25.6 30 40
Fly ash
-1.3 30
-1.6 40
-26.3 50
-2.0 50
us
[64] [65]
-48.6
-15.1
[59]
7.3
11
[73]
-2.4
4.2
[30]
-5.7
6.4
[55]
6.5-6.9
17.1
[46]
an
-5.6 50
1.5
[68]
1.6
M
0.9 55 -27.6 45
ed
-5.5 40
Ac ce
Basic oxygen furnace slag LDH/chitosan composite
pt
Adsorbent
ip t
ΔG0 (kJ/mol) ----------------------Temperature (oC) -0.38 -0.72 -1.94 25 35 45 -6.81 -6.74 -6.72 30 40 50
cr
Table 2. Thermodynamic parameters for fluoride adsorption by different adsorbents
57
Page 57 of 62
Table 3. Fluoride desorption and adsorbent regeneration Desorption/regeneration reagent (BV, bed volume; d and ht, column internal diameter and height; FL, flow rate
Results
Reference
Amberlite modified with thio-urea (ATU) (B) and (C)
(B) 100 g F-loaded ATU shaken 40 min with 5 mL HNO3, H2SO4, HCl, NaOH (0.01, 0.001 M), 25oC. (C) 2 g ATU, 0.8 mm d, 50 mm ht, FL 1 mL/min, 25oC
(B) maximum desorption [34] (99.9%) with 0.01 M HCl. Next best 0.01 M H2SO4 (77.6%). (C) maximum recovery 92.3% (0.01 M HCl), 35 mL volume used
Manganese oxide coated alumina (MOCA) (B)
0.25 g MOCA adsorbed with F, 50 85% desorption at pH 12. Little mL water with pH adjusted to 3.0- F desorbed at pH < 10 13.5, shaken 24 h, 25oC
Granular red mud (GRM) (C)
10 g GRM, 0.635 cm2 cross sectional area, 10 cm ht, 10 mL 0.2 M NaOH, FL 1 ml/min, 25oC
MOCA and activated alumina (AA) (C)
0.028 m d, 0.1, 0.2 and 0.3 m ht, 2.5% NaOH. 3 cycles of regeneration/adsorption
Al treated zeolite (B)
0.1 g F loaded adsorbent, 50 mL acetic acid, water, NaHCO3, NaOH shaken for 1 d. 4 desorption steps. 25oC
AA (C)
50.8 mm d, 550 mm ht, repeated cycles of regeneration/reactivation
ip t
Adsorbent (batch (B), column (C) method used)
us
cr
[29]
[72]
94.6%F desorbed for MOCA, 91%F for AA in Ist cycle. 2nd cycle no reduction in adsorption, 3rd cycle 5% reduction
[36]
NaHCO3 most effective (63%desorption), Acetic acid least effective (5%). 4 zeolites with NaHCO3 desorption, 1st step 61-67%, total 75-95%
[41]
Loss of 2% F uptake capacity after 5 cycles of regeneration
[47]
Ac ce
pt
ed
M
an
Adsorption (mg/g) decreased from 2.05 to 0.82 in 4 regeneration/adsorption cycles. %desorption from 87 to 46
Zr loaded collagen fibre (B)
0.1 g, 100 mL water with pH adjusted with NaOH, HNO3, 0.5 h shaking, 30oC
pH < 9 very little desorption, pH [13] 11.5 97% desorption
La loaded gelatin (B)
1st wash 1 M NaOH, then water wash or acid wash with 1:1 HNO3 to neutral pH
After adsorption/regeneration process for 3 times, maximum adsorption capacity reduced from 98.5 to 82.3%
[44]
Calcined Zn/Al LDH (B)
NaOH (0.001-0.04 M), 1 g adsorbent treated with maximum F/L of NaOH for 6 h, 30oC.
F desorption increased from 5.12 to 8.55 mg/g (100%). Desorption increased with NaOH concentration
[55]
Ion exchange fibre (B), (C)
(B) 0.1 g adsorbent F loaded, mixed with 25 mL 0.1 M NaCl, NaOH, HCl for 2 h, 25oC. (C) 1.5 cm d, 5
(B) Desorption 20% for HCl, 60% for NaCl, 80% for NaOH. (C) 100% desorbed with 5
[31]
58
Page 58 of 62
mLNaOH
Rare earth oxides (B)
0.2 g adsorbent, 100 mL water, pH adjusted, shaken for 30 min. 29oC
pH<6, F desorbed to ~ 0. pH ~ 12 >95%. Regeneration decreased adsorption 98 to 91%
[50]
Activated alum sludge (B)
F loaded sludge 4g/L, water pH adjusted, 4 h shaking, 30oC
pH 2-7 desorption ~ 0. > pH 7 desorption increased, highest at pH 12
[48]
La loaded silica gel (C)
Adsorbent loaded with F eluted with dilute NaOH (pH 8.5). 1 cm d, 9 cm h, FL 0.5 mL/min, 20oC.
Column was regenerated
[43])
Ac ce
pt
ed
M
an
us
cr
ip t
cm ht, eluted with 0.5 M NaOH, FL 1 mL/min, 25oC
59
Page 59 of 62
List of figures
ip t
Fig. 1. Common technologies for defluoridation of drinking water [8,22,23]
Ac ce
pt
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M
an
us
cr
Fig. 2. Mechanisms of F adsorption (, Adsorbent; Me, multivalent metallic cation)
60
Page 60 of 62
Adsorption/Ion Exchange
! ! Most widely used ! ! Medium cost ! ! Low effectiveness; cannot remove F below 5 mg/L because of high solubility product of CaF2; need secondary treatment ! ! Large amounts of chemicals required ! ! Precise control of chemicals additions (frequent testing of feed and treated water) ! ! Costs of chemicals, chemical storage and feeding system ! ! Large volumes of waste sludge; disposal problem ! ! Acid neutralisation of treated water required ! ! Toxic chemicals left in treated water (AlF complexes, SO4)
! ! Most widely used ! ! Can be costly (especially ionexchange resins), but can use low-cost adsorbent (including certain waste materials) ! ! Effective even at low F concentration ! ! Simplicity and flexibility of design ! ! Ease of operation ! ! No waste production ! ! Low selectivity against some anions for adsorption/all anions for ion exchange competing ions ! ! Frequent adsorbent regeneration or replacement required ! ! Granular adsorbent better for good hydraulic flow ! ! Effective, mostly at pH < 7 for adsorption
cr
us
an M
ed
Reverse osmosis
ip t
Precipitation/coagulation
Ac ce
pt
! ! Excellent removal ! ! Very high capital cost. Very high operational (energy) cost ! ! No chemicals required ! ! No waste production ! ! No ion selectively; beneficial nutrients and other contaminants removed together with F ! ! Some membranes pH sensitive ! ! F concentrated residue disposal problem ! ! Water wasted ! ! Clogging, scaling and fouling problems
Electrodialysis
! ! Excellent removal ! ! High capital cost. High operational (energy) cost ! ! No chemicals required ! ! No waste production ! ! No ion selectively; beneficial nutrients and other contaminants removed together with F ! ! Skilled labour required ! ! Polarization problem
Fig. 1.
61
Page 61 of 62
(b)
Hydrogen Bonding
ip t
Ion Exchange
an
us
cr
(a)
Ligand Exchange
(d)
Adsorbent surface chemical modification
Ac ce
pt
ed
M
(c)
Fig 2.
62
Page 62 of 62
S0304-3894(12)01211-3 doi:10.1016/j.jhazmat.2012.12.043 HAZMAT 14806
To appear in:
Journal of Hazardous Materials
Received date: Revised date: Accepted date:
2-8-2012 18-12-2012 26-12-2012
Please cite this article as: P. Loganathan, S. Vigneswaran, J. Kandasamy, R. Naidu, Defluoridation of drinking water using adsorption processes, Journal of Hazardous Materials (2010), doi:10.1016/j.jhazmat.2012.12.043 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
1
Research Highlights
2 3 4
! ! Comprehensive and critical literature review on various adsorbents used for defluoridation ! ! pH, temperature, kinetics and co-existing anions effects on F adsorption
6
! ! Choice of adsorbents for various circumstances
7
! ! Adsorption thermodynamics and mechanisms
8
! ! Future research on efficient, low cost adsorbents which are easily regenerated
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5
9
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1
Page 1 of 62
10
Defluoridation of drinking water using adsorption processes
11
13
Paripurnanda Loganathan1, Saravanamuthu Vigneswaran1*, Jaya Kandasamy1, Ravi Naidu2 1
Faculty of Engineering and Information Technology, University of Technology, Sydney, NSW, 2007, Australia
14
16
2
CRC CARE, University of South Australia, Adelaide, SA 5095, Australia
*Corresponding author: ph: 612 9514 2641, E-mail: [email protected]
cr
15
ip t
12
us
17
ABSTRACT
19
Excessive intake of fluoride (F), mainly through drinking water, is a serious health hazard
20
affecting humans worldwide. There are several methods used for the defluoridation of
21
drinking water, of which adsorption processes are generally considered attractive because of
22
their effectiveness, convenience, ease of operation, simplicity of design, and for economic
23
and environmental reasons. In this paper, we present a comprehensive and a critical literature
24
review on various adsorbents used for defluoridation, their relative effectiveness, mechanisms
25
and thermodynamics of adsorption, and suggestions are made on choice of adsorbents for
26
various circumstances. Effects of pH, temperature, kinetics and co-existing anions on F
27
adsorption are also reviewed. Because the adsorption is very weak in extremely low or high
28
pHs, depending on the adsorbent, acids or alkalis are used to desorb F and regenerate the
29
adsorbents. However, adsorption capacity generally decreases with repeated use of the
30
regenerated adsorbent. Future research needs to explore highly efficient, low cost adsorbents
31
that can be easily regenerated for reuse over several cycles of operations without significant
32
loss of adsorptive capacity and which have good hydraulic conductivity to prevent filter
33
clogging during the fixed-bed treatment process.
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2
Page 2 of 62
1. Introduction
36
2. Adsorption mechanisms
37
3. Factors influencing adsorption
38
3.1. pH
39
3.2. Co-existing anions
40
3.3. Temperature
41
3.4. Adsorption kinetics
42
cr
35
ip t
Keywords: adsorption, defluoridation, fluoride, fluoride toxicity, water treatment
us
34
4. Adsorbents 4.1. Metal oxides and hydroxides
44
4.2. Layered double hydroxides
45
4.3. Ion exchange resins and fibres
46
4.4. Zeolites
47
4.5. Carbon materials
48
4.6. Natural materials
49
4.7. Industrial by-products
pt
ed
M
an
43
5. Adsorption thermodynamics
51
6. Fluoride desorption and adsorbent regeneration
52
7. Conclusions
53
8. Acknowledgements
54
9. References
Ac ce
50
55 56 57
1. Introduction
58 3
Page 3 of 62
Fluoride (F) has beneficial effects on teeth at low concentrations in drinking water (0.4-
60
1.0 mg/L), especially for young children in that it promotes calcification of dental enamel and
61
protects teeth against tooth decay. Excessive levels of F on the other hand can cause a
62
number of problems ranging from mild dental fluorosis to crippling skeletal fluorosis as the
63
level and period of exposure to F increases [1]. The World Health Organisation [2] has
64
recommended a guideline value of 1.5 mg/L as the concentration above which dental
65
fluorosis is likely. However, it is also important to consider climatic conditions and quantity
66
of water intake and other factors such as F intake from certain diets when establishing F
67
limits. Water consumption in hot humid regions is generally higher than in temperate regions
68
and therefore the F concentration limit for likely fluorosis should be lower. For example, the
69
US Public Health Service [3] has recommended that the upper limit for F concentration in
70
drinking water should be decreased from 1.7 to 0.8 mg/L with increases of the average
71
maximum daily air temperature.
M
an
us
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ip t
59
High F intake has been suspected being involved in a range of adverse health problems in
73
addition to fluorosis, including cancer, impaired kidney function, digestive and nervous
74
disorders, reduced immunity, Alzheimer’s disease, nausea, adverse pregnancy outcomes,
75
respiratory problems, lesions of the endocrine glands, thyroid, liver and other organs [1,4-9].
76
However, there appears to be no convincing evidence for F being directly involved in causing
77
these conditions [1].
Ac ce
pt
ed
72
78
Elevated concentration of F in drinking water is due to its natural occurrence or industrial
79
activities. Many rocks and minerals in the earth’s crust contain F [10,11] which can be
80
leached out by natural weathering and rainwater, causing F contamination of surface and
81
ground waters. Besides this natural source, F also enters the water bodies from waste waters
82
produced from industries such as aluminium, steel, glass, semiconductors, electronic, tooth
83
paste, fertiliser and insecticide manufacturing plants [9,12-15]. 4
Page 4 of 62
Fluorosis due to excessive concentration of F has been reported in at least 28 countries
85
from South Asia, Africa, the Middle East, North, Central and South America, and Europe [1].
86
In India, it was estimated that 56.2 million people were affected by fluorosis and this problem
87
was prevalent in 17-19 out of the 32 States [16,17]. The major source of F in a majority of
88
countries is rocks and minerals such as fluorospar, cyyolite, and fluorapatite containing F
89
[17]. For example, Choi and Chen [18] reported extremely high F concentration (> 1000
90
mg/L) in surface water in areas with F-rich volcanic rocks.
cr
ip t
84
One method of reducing excessive concentrations of F in water is to blend water having a
92
high F concentration with water that has a low F concentration from an alternative source. If
93
such a source is not readily available, defluoridation is the only means remaining to prevent
94
fluorosis [1]. For contaminants other than F, water treatment methods are used to remove
95
contaminants to below the maximum level permissible but defluoridation is special in that the
96
treated water should have an optimum F concentration to derive the beneficial effects of F.
M
an
us
91
The main methods of defluoridation are precipitation/coagulation, adsorption, ion
98
exchange, reverse osmosis, and electrodialysis. Of these, precipitation/coagulation and
99
adsorption are convenient methods and are widely used, especially in developing countries’
100
rural areas. The scale and treatment site differ between industrialised countries and
101
developing countries. In industrialised countries the treatment of water is generally performed
102
at water treatment plants close to the water source but in developing countries it is carried out
103
at a village community level or at a household level [1] using simple inexpensive locally
104
available adsorptive media [9,19-21]. Industrialised countries generally use more efficient but
105
more costly adsorption media including synthetic ion exchange resins as well as advanced
106
techniques such as reverse osmosis and electrodialysis.
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97
107
The advantages and shortcomings of the various methods of defluoridation are presented
108
in Fig. 1. Of these various methods that of adsorption is generally considered attractive 5
Page 5 of 62
because of its effectiveness, convenience, ease of operation, simplicity of design, and
110
economic and environmental considerations, provided low-cost adsorbents which can
111
effectively remove F around the neutral pH of drinking water are used. The
112
precipitation/coagulation method where lime and Al salts are used to remove F as a CaF2
113
precipitate followed by removal of left over F in solution by co-precipitation with and
114
adsorption on to the precipitated Al(OH)3 is further developed into a technique called
115
Nalgonda process in India [17,24,25]. This process is extensively used in India and Africa
116
[1]. However, the main drawback of this technique is the low effectiveness of F removal and
117
production of toxic AlF complexes in solution [8].
us
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109
Previous reviews on defluoridation of water presented timely information on several
119
adsorbents used for defluoridation, but they generally did not focus on the chemical
120
mechanisms and solution factors influencing the adsorption processes, and regeneration of
121
the adsorbents [1,17,26,27]. The objective of this paper is to compile and present current
122
information on the potential of the various adsorbents used for defluoridation of drinking
123
water, their relative effectiveness, mechanisms and thermodynamics of adsorption, factors
124
influencing adsorption and methods of adsorbent regeneration for reuse. Based on the review
125
adsorbents are selected for specific circumstances.
127 128
M
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126
an
118
2. Adsorption mechanisms
129
The capacity, energy and kinetics of adsorption of F by adsorbents are controlled by the
130
mechanism of adsorption. Understanding these mechanisms can provide useful information
131
on the optimisation of the adsorption process in water treatment plants and the subsequent
132
desorption/adsorbent regeneration process for reuse. There are mainly five mechanisms of F
133
adsorption, namely: (1) van der Waals forces (outer-sphere surface complexation, (2) ion 6
Page 6 of 62
exchange (outer-sphere surface complexation), (3) hydrogen bonding (H-bonding) (inner-
135
sphere surface complexation), (4) ligand exchange (inner-sphere surface complexation), and
136
(5) chemical modification of the adsorbent surface. The first two mechanisms are governed
137
by weak physical adsorption and are non-specific to F, whereas the third and fourth
138
mechanisms are governed by strong chemical adsorption specific to F. The fifth mechanism
139
is governed by both specific and non-specific adsorption. In the presence of the other anions
140
in water, F cannot be easily removed by adsorbents using the first two mechanisms. Fluoride
141
adsorption resulting from the third and fourth mechanisms selectively removes F from water
142
in the presence of most types of anions; only anions (e.g. phosphate) which also adsorb
143
specifically on the adsorbent compete with F for adsorption sites.
an
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134
Van der Waals forces are weak short range forces acting between two atoms. The larger
145
the adsorbate size the greater the force of attraction. Therefore adsorbates with high
146
molecular weights such as dissolved organic matter are adsorbed on adsorbents having high
147
surface area through van der Waals forces. This is the reason for the weak adsorption of F
148
and strong adsorption of dissolved organic matter on activated carbon [28]. Fluoride was
149
considered to be adsorbed on manganese oxide-coated alumina by van der Waals forces at
150
high pH [29].
Ac ce
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144
151
Ion exchange is a stoichiometric process where any counter ion leaving the ion exchanger
152
surface is replaced by an equivalent number of moles of another counter ion to maintain
153
electro-neutrality of the ion exchanger. The ions are adsorbed physically by fully retaining
154
their inner hydration shell and the adsorption is due to electrostatic or Coulombic attraction.
155
The ion exchange process is rapid and reversible. Fluoride removal by ion exchange resins
156
[30] and ion exchange fibres [31] is mainly governed by the ion exchange mechanism as
157
illustrated in Fig. 2a. Ion exchange tends to prefer counter ions of higher valency, higher
158
concentration and ions of smaller hydrated equivalent volume [32]. Therefore F removal 7
Page 7 of 62
159
using ion exchange resins is difficult because the order of selectivity for anions by anion
160
exchange resins is as follows: citrate > SO42- , oxalate > I- > NO3- > CrO42- > Br- > SCN- > Cl-
161
> acetate > F- [32]. H-bonding is a strong dipole-dipole attractive force between bonding of the strong
163
electropositive H atom in a molecule in an adsorbent or adsorbate and a strong
164
electronegative atom such as oxygen or fluorine in another molecule [33]. The energy of
165
adsorption in H-bonding is stronger than in van der Waals forces and ion exchange but
166
weaker than in the ligand exchange process discussed in the next paragraph. H-bonding
167
occurs in the adsorption of F on ion exchange resins [30,34] and coal-based adsorbents [35]
168
as illustrated in Fig. 2b.
an
us
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162
In a ligand exchange mechanism, the adsorbing anion such as F- forms a strong covalent
170
chemical bond with a metallic cation at the adsorbent surface resulting in the release of other
171
potential determining ions such as OH- ions previously bonded to the metallic cation (Fig.
172
2c). Thus, F is said to form an inner sphere complex or is specifically adsorbed on the
173
adsorbent surface. The adsorption of F on several multivalent metal oxides near neutral pH
174
was reported to have increased the pH of solutions as a result of release of OH- ions from the
175
adsorbents by ligand exchange of OH groups on the adsorbent surface with F in solution
176
[36,37]. Adsorbents with a ligand exchange mechanism have the particular advantage of
177
combining high adsorption capacity with high selectivity for anions. These adsorbents can
178
remove large proportions of anions having higher selectivity for adsorption from very dilute
179
solutions of the anions even in the presence of higher concentrations of competing anions of
180
lower selectivity. Adsorption of F by ligand exchange is illustrated in Fig. 2c [23,37,38].
Ac ce
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169
181
The adsorption capacity of F on adsorbents can be increased by chemical modification of
182
adsorbent surfaces (Fig. 2d). This is particularly of advantage in the case of adsorbents
183
possessing negative surface charges which tend to repel the similarly charged F- ions. In such 8
Page 8 of 62
adsorbents, positively charged multivalent cations such as Al3+, La4+, Zr4+, Fe3+, and Ce3+ are
185
impregnated onto the adsorbent to create positive charges on the adsorbent surface for
186
attracting F- by coulombic forces as well as producing adsorption sites capable of chemical
187
interaction with F [13,23,39-44] (Fig. 2d). These metallic cations act as a bridge in adsorbing
188
F onto the adsorbent.
ip t
184
Adsorption of F on many microporous adsorbents is recognised as a 2-step process; an
190
initial rapid adsorption (mostly within an hour) at the outer surface of the adsorbent that
191
reaches a pseudo-equilibrium at the solid-solution interface followed by a much slower
192
process (hours to days) where the F moves by intra-particle diffusion into the interior pores
193
and channels of the adsorbent [45]. Intra-particle diffusion rate is directly related to the
194
square root of time of adsorption. Therefore if a straight line relationship is obtained between
195
the rate of adsorption and square root of time with the line passing through the origin, it can
196
be inferred that the diffusion process controls the adsorption, especially at longer times. Such
197
a relationship was obtained in many studies for the adsorption of F on different adsorbents
198
(granular ferric hydroxide [7], manganese oxide-coated alumina [29], fly ash [46], activated
199
alumina [47], and alum sludge [48]).
pt
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189
It could be concluded that ligand exchange is the predominant mechanism of F adsorption
201
for inorganic adsorbents having high adsorption capacities. For organic adsorbents having
202
high adsorption capacities, H-bonding seems to be the predominant mechanism.
203 204
Ac ce
200
3. Factors influencing adsorption
205 206
3.1. pH
207
As for other contaminants, the removal of F from water by adsorbents is influenced
208
by several factors including pH, co-existing ions, temperature, adsorption kinetics, and 9
Page 9 of 62
adsorbent particle size and activation. Of these, pH is generally considered to be the most
210
important factor [16]. Fluoride adsorption is low at both very low and very high pH. The pH
211
at which the maximum amount of F is removed depends on the adsorbent characteristics but
212
generally it is between 4 and 8 (Table 1). One of the important properties of the adsorbent
213
influencing the extent of F adsorption is the pH at the point of zero charge (PZC). For
214
example, Fe and Al oxides having a PZC at around 7-8 remove the maximum amount of F at
215
pH 6-8 [7,47,49] and activated carbon with a PZC of 3.9-4.7 was reported to have removed
216
the maximum amount of F at pH 3-4 [42]. At pH values above the PZC the surface of
217
adsorbents is negatively charged, therefore the negatively charged F ions are not attracted to
218
the adsorbent surface. At pH values lower than the PZC the surface is positively charged,
219
therefore F ions are adsorbed. In some situations, at low pH values, F exist as positively
220
charged AlF complexes and this can reduce F adsorption [11,49,75].
M
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209
The main reason for a reduction in F adsorption below pH 4 is that F forms weakly
222
ionised HF at these pH values [13,47,48,50]. At a pH above 7-8 the removal of F decreases
223
not only because the adsorbent surface becomes negatively charged but also because the
224
concentrations of hydroxyl, bicarbonate and silicates ions increase so that they compete with
225
F for adsorption. In addition to increased number of positive surface charges at low pH values
226
increasing F adsorption, surface protonation at a pH less than the pH of PZC provides an
227
increased number of H atoms at the adsorbent surface leading to an increased number of H-
228
bonding between the H atoms and F in solution resulting in increased F adsorption [76].
229
Adsorption by a ligand exchange mechanism is also favoured at low pH because of a stronger
230
attractive force between F and the adsorbent surface and the presence of more hydroxylated
231
sites for exchange with F than at a high pH [4,29,61].
Ac ce
pt
ed
221
232
It is apparent from the literature review that F adsorption is generally lowest at
233
extremely low and high pH values. The adsorption is highest around the neutral pH that is 10
Page 10 of 62
234
commonly found in natural water. Therefore prior pH adjustment is not normally required for
235
effective removal of F in treatment plants.
236 237
3.2. Co-existing anions In natural water, several anions including PO43-, Cl-, SO42-, Br-, NO3- are
239
simultaneously present with F at different concentrations which can compete with F for
240
removal by adsorbents. The extent of the competition depends on the relative concentrations
241
of the ions and their affinity for the adsorbent. Meenakshi and Viswanathan [30] reported that
242
increased concentrations of Cl-, SO42-, Br-, NO3- , and HCO3- decreased F adsorption by an
243
anion exchange resin which adsorbs anions by an ion exchange mechanism, whereas these
244
anions had no effect on F adsorption by a chelating resin which adsorbed F selectively by a
245
H-bonding mechanism. The preferential order of adsorption of anions by the chelating resin
246
was reported to be F- > Cl- > NO3- > SO42-.
M
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238
Solangi et al. [34] studied the adsorption of F by a thio-urea incorporated amberlite
248
resin in the presence of PO43-, Cl-, SO42-, Br-, NO3-, NO2-, HCO3-, and CO32- at five times the
249
molar concentration of F and observed that Br-, NO2-, and PO43- had little interference with F
250
adsorption; the other ions had no interference. This was explained as due to the strong H-
251
bonding of F with the amide groups in the resin.
pt
Ac ce
252
ed
247
Multivalent metal oxides are known to adsorb F selectively by the ligand exchange
253
specific adsorption mechanism. An alum sludge containing a high percentage of Al, Ti and
254
Fe oxides adsorbed F selectively in the presence of SO42- and NO3- [48]. At 50 mg/L, SO42-
255
and NO3- reduced F adsorption from 85% to 40% and 62%, respectively from a solution
256
containing 20 mg F/L. In contrast to these anions, PO43- and selenate concentrations at 20
257
mg/L reduced F adsorption to 25% of F in solution. Based on these results, Sujana et al. [48]
258
proposed that the decreased order of competition of anions for F adsorption to be PO43- ≥ 11
Page 11 of 62
259
selenate > SO42- > NO3-. Das et al. [55] also found that PO43- interfered more than SO42- in F
260
removal by a calcined Zn/Al layered double hydroxide (LDH) consisting of Zn/Al oxide. Kumar et al. [7] investigated the adsorption of F by a granular ferric hydroxide in the
262
presence of competing anions such as Cl-, SO42-, BrO3-, NO3-, CO32- and PO43-, each having
263
concentrations of 20 to 100 mg/L with an initial F concentration of 20 mg/L. There was no
264
significant influence of competing anions on F removal when the adsorbent dose was 10 g/L,
265
which was attributed to the availability of plenty of sorption sites. However, when the
266
adsorbent concentration was reduced to 5g/L, PO43-, CO32-, and SO42- , each at concentrations
267
of 100 mg/L, this reduced the F adsorption capacity by 35, 25, and 20% of F in solution,
268
respectively. The other anions did not significantly reduce the F adsorption capacity.
269
Similarly, Raichur and Basu [50] found that F adsorption by a mixture of naturally occurring
270
rare earth oxides (oxides of La, Ce, Pr, Nd, Sm and Y) was not significantly affected by the
271
presence of SO42- or NO3- in water at a concentration equal to that of F (100 mg/L).
M
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261
It can be concluded that the non-specifically adsorbing anions (e.g. nitrate, chloride)
273
do not compete with F for adsorption on adsorbents that adsorb F using specific adsorption.
274
Only anions that are adsorbed by specific adsorption (e.g. phosphate, selenate, arsenate)
275
compete with F for adsorption. When F is adsorbed by non-specific adsorption the non-
276
specifically adsorbing anions can compete with F.
278 279
pt
Ac ce
277
ed
272
3.3. Temperature
Temperature had no consistent effect on F adsorption. Adsorption on many adsorbents
280
increased with temperature showing an endothermic nature of adsorption (granular ferric
281
hydroxide [7], fly ash [46], calcined Mg/Al/CO3 LDH [54], LDH/chitosan [63], spent catalyst
282
[75]). In contrast with many others it decreased with temperature showing an exothermic
283
nature of adsorption (chelating resin [30], alum sludge [48], calcined Zn/Al LDH [55], 12
Page 12 of 62
284
modified activated carbon [59], LDH/chitin [65], geo-materials [68]). Temperature has also
285
been reported to have no significant effect on adsorption by some adsorbents (trivalent
286
cations/zeolite [40]). The reasons for the differences in the effect of temperature were not clearly stated in
288
the studies. It may depend on the temperature range studied, the nature of the adsorbent, and
289
the conditions used in the studies. For example, at extremely low temperatures (5, 10oC) the
290
rate of adsorption was reported to be low because the rate of movement of F to the adsorption
291
sites is low. Lai and Liu [75] reported that F adsorption on spent catalyst increased
292
significantly when the temperature increased from 5oC to 25oC but very little further increase
293
in adsorption was observed when the temperature rose to 50oC. Sujana et al. [48] reported
294
that the exothermic nature of adsorption of F on alum sludge existed because the rising
295
temperature increased the tendency for F to escape from the adsorbent. Another reason given
296
was an increase in thermal energy of adsorbed F at higher temperatures, causing increased
297
desorption.
298
3.4. Adsorption kinetics
pt
299
ed
M
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287
Fluoride adsorption studies have shown that the rate of removal of F by adsorbents is
301
high in the initial 5-120 min where generally more than 90% of F is adsorbed, but thereafter
302
the rate significantly levels off and eventually approaches zero denoting the attainment of
303
equilibrium. This is because initially the adsorption sites are vacant and the F concentration
304
gradient between solution and adsorbent surface is high. Subsequently the rate decreases
305
because of the decrease in vacant sites. A fast rate of adsorption helps the adsorbent to treat
306
large quantities of water [54]. A slow rate causes operational, control, and maintenance
307
problems in the adsorption process of the filter bed [31].
Ac ce
300
13
Page 13 of 62
The rate of F adsorption increases with an increase in concentration of adsorbent [54]
309
and a decrease in initial F concentration [29,47,54]. It also depends on the structural
310
properties of the adsorbents and the interaction between F and the sites of adsorption.
311
Adsorption kinetics has been described by pseudo-first order and pseudo-second order kinetic
312
models and the diffusion model and rate constants have been determined (Table 1). These
313
models have also provided information on adsorption mechanisms.
4. Adsorbents
us
315
cr
314
ip t
308
316
The effectiveness of F adsorption from drinking water by various adsorbents, the
318
methods used for the assessment, and the kinetics and equilibrium models that best explained
319
the adsorption process are presented in Table 1. Caution needs to be exercised in comparing
320
the adsorption capacities of adsorbents because of the inconsistencies in data presentation
321
including differences in methodology and parameters used in the studies (pH, temperature, F
322
concentration range, competing ions, etc.). An ideal adsorbent that can be used to remove F
323
must have the following characteristics: low-cost, a high F adsorption capacity, rapid
324
adsorption of F, easily regenerated after its removal capacity is exhausted and good physical
325
characteristics (rapid water flow without filter clogging).
327 328
M
ed
pt
Ac ce
326
an
317
4.1. Metal oxides and hydroxides Oxides and hydroxides, also called hydrous oxides or oxyhydroxides, of trivalent and
329
tetravalent metals such as Fe, Al, La, Mn, and Zr are used to remove both anionic and
330
cationic contaminants from water and wastewaters because of their strong ability to adsorb
331
these ions [37,77]. The predominant mechanism of adsorption of F on oxides and hydroxides
332
is ligand exchange by the formation of inner sphere complexes (specific adsorption) as 14
Page 14 of 62
333
discussed previously. Most metal oxides and hydroxides have their PZC above the natural
334
water pH of 7 (granulated ferric hydroxide pH 7.5-8.0 [7], activated alumina pH 8.25 [36], γ-
335
alumina pH 8 [49]). Therefore, at the neutral pH of natural water, these adsorbents have
336
positive surface charge which is favourable for the adsorption of the negatively charged F. Of the oxides and hydroxides of metals, Al oxide, especially the activated form
338
(activated alumina) has been the most commonly used adsorbent for the removal of F
339
[9,29,36,37,42]. Activated alumina is produced by thermal degradation of aluminium
340
hydroxide to obtain materials with high specific surface area and a distribution of micro- and
341
macro-pores [78]. The specific surface area (m2/g) of activated alumina has been reported to
342
be 160 [49], 297 [29], and > 200 [26]. The adsorption capacity of activated alumina varies
343
with the structure of the alumina. For example, γ-Al2O3 has a much higher adsorption
344
capacity than α-Al2O3 [42] and therefore this form of activated alumina is commonly used for
345
defluoridation of water. The maximum F adsorption capacity of activated γ-Al2O3 (mg/g) has
346
been reported to be 1.1 [36], 12.0 [41], 2.41 [47], and 16.3 [49].
ed
M
an
us
cr
ip t
337
Aluminium oxides have poor adsorption capacity for F in acidic conditions because of
348
their tendency to dissolve and form positively charged AlF complexes (AlF2+, AlF2+) which
349
are repelled by the positively charged surfaces of Al oxides at these conditions (PZC of Al
350
oxides > pH 7) [49]. In alkaline conditions, Al oxides have negatively charged surfaces
351
which repel the negatively charged F ions present at these pHs. Also, as stated previously, the
352
increase in the concentrations of OH- in alkaline condition competes with F- for adsorption.
353
Therefore the optimum pH for F adsorption is considered to be near neutral pH [47,49].
Ac ce
pt
347
354
Recent studies have shown that surface coatings of adsorbents with other materials
355
have enhanced the adsorption of F [29]. Because manganese oxides have a high specific
356
surface area, a micro-porous structure [79] and a high adsorption capacity towards anions
15
Page 15 of 62
357
[80], activated alumina was coated with these oxides and F adsorption capacities have been
358
studied [29,36]. Maliyekkal et al. [36] compared the F adsorption capacity and rate of adsorption of
360
activated alumina (AA) and a granular manganese oxide coated AA (MOCAA) at pH 7 and
361
found that in the batch study, most of the adsorption was complete in 3 h in the case of
362
MOCAA compared to 10 h for AA. The pseudo-first order and pseudo-second order rate
363
constants were also higher for MOCAA. The Langmuir adsorption capacity of MOCAA was
364
2.85 mg/g compared to 1.08 mg/g for AA. Maximum F adsorption was found to occur at a
365
wider pH range of 4-7 for MOCAA compared to that for AA (4-6). The column study also
366
showed that MOCAA had higher F adsorption capacity. The superiority of MOCAA over AA
367
in the adsorption of F was reported to be not due to the surface area difference because the
368
specific surface area of AA was 204 m2/g and that of MOCAA was 170 m2/g. Maliyekkal et
369
al. [36] suggested that the reason for this could be the increased zeta potential (surface
370
charge) of the MOCAA, although no supporting data were presented.
ed
M
an
us
cr
ip t
359
Teng et al. [29] used a redox process to coat AA with an amorphous MnO2. The
372
granular MOCAA produced had a higher surface area (316 m2/g) compared to that for AA
373
(297 m2/g) and a significantly rougher surface with plenty of pores. Maximum F adsorption
374
was obtained at the pH range of 4-6. One of the two mechanisms of F adsorption at a pH less
375
than 6 was considered to be chemical adsorption by a ligand exchange of surface OH groups
376
in MOCAA by F in solution resulting in an increase of pH due to the release of OH-. The
377
other mechanism was intra-particle diffusion of F. The Langmuir adsorption maximum at pH
378
5.2 was 7.09 mg/g which was much higher than the value of 1.08 mg/g reported for AA by
379
Maliyekkal et al. [36] and values reported for many other granulated adsorbents. For initial F
380
concentrations less than 21 mg/L, most of the adsorption was completed within 30 min
Ac ce
pt
371
16
Page 16 of 62
381
compared to 10 h for AA. Faster adsorption by MOCAA was considered to be due to the
382
larger surface area and porous surface of this adsorbent. Iron oxides are also known to have large capacity to remove anions from water by
384
mechanisms similar to those operating in Al oxide adsorbents [81]. Kumar et al. [7] studied
385
the adsorptive removal of F by a highly porous and poorly crystalline granulated ferric
386
hydroxide (GFH) (β-FeOOH) with a specific surface area of 250-300 m2/g, a PZC of pH 8,
387
and a granular size of 0.32-2.0 mm. Nearly 95% adsorption of F from solution was achieved
388
within the first 10 min of agitation of 10-20 mg F/L with 10 g GFH/L at pH 6-7. The
389
maximum F adsorption was observed at the pH range of 3-8. A sharp decrease of F
390
adsorption was observed above the PZC pH of 8 as the GFH surface became more negatively
391
charged causing electrostatic repulsion of the negatively charged F- ions in addition to
392
increasing concentration of OH- ions which competed with F- for adsorption.
M
an
us
cr
ip t
383
Hydroxyapatite, the most abundant of phosphate minerals, has been used for
394
defluoridation of water. Fan et al. [52] compared the F adsorption capacities of several
395
natural minerals and found that the adsorption capacities at pH 6 followed the order:
396
hydroxyapatite > fluorspar > quartz activated using ferric ions > calcite > quartz. The highest
397
F adsorption capacity of hydroxyapatite was explained as owing to F exchanging with a OH
398
group at the surface and inside the apatite mineral. The F adsorption on the other minerals
399
was deemed to be a surface adsorption process.
pt
Ac ce
400
ed
393
It is evident from the literature review that Fe and Al oxides and hydroxides are the
401
commonly used adsorbents for defluoridation. They have a moderate level of F adsorption
402
capacity (1-16 mg/g) and if locally available at low cost, can potentially be employed in rural
403
areas, especially in developing countries.
404 405
4.2. Layered double hydroxides 17
Page 17 of 62
The majority of clay minerals such as kaolinite, mica, montmorillonite, vermiculite
407
and zeolite, carry predominantly negative charges and therefore adsorb very small amounts of
408
anions. Layered double hydroxide (LDH) or hydrotalcite (HTlc) is another type of clay
409
mineral, but has positive charges and therefore adsorb significant quantities of anions and
410
oxyanions (e.g. fluoride, arsenite, arsenate, chromate, phosphate, selenite, selenate, nitrate,
411
etc) and monoatomic anions (e.g. fluoride, chloride) from aqueous solutions [82].
412
Structurally, LDHs are composed of positively charged brucite-like sheets compensated by a
413
large number of exchangeable charge-balancing anions in the hydrated interlayer regions
414
[55,56,82]. Charges can also be produced by the ionisation of the OH groups at the surface of
415
the LDH particles. The PZC of LDHs is in the region of pH 9-12 [82] and therefore at the
416
neutral pH of natural water, LDHs carry positive charges and act as anion exchangers.
an
us
cr
ip t
406
Calcined LDHs have higher F adsorption capacities than the uncalcined LDHs [53].
418
The optimum temperature of calcination is generally considered to be 450-500oC [53-56]. Lv
419
et al. [54] reported that the F adsorption capacity of a Mg/Al LDH increased from 65 to 70
420
and then to 80 mg/g when the calcination temperature was increased from 200oC to 400oC
421
and then to 500oC, respectively, but decreased to 62 and 50 mg/g when the temperature was
422
increased to 600oC and 800oC, respectively. Wang et al. [53] suggested that the increase in F
423
adsorption capacity of LDH as a result of calcination was due to the higher specific surface
424
area, porosity, and surface reactivity of the Mg/Al oxide produced by calcination. Another
425
reason reported was the incorporation of F into the structure of Mg/Al oxide resulting in the
426
formation of the original structure of the LDH. The decrease in F adsorption capacity at
427
calcining temperatures higher than 500oC was considered to be due to the transformation of
428
the Mg/Al oxides into a spinel structure that does not exhibit the property of LDH structural
429
reconstruction [54].
Ac ce
pt
ed
M
417
18
Page 18 of 62
The adsorptive property of LDH depends on the metallic constitution of the LDH
431
structure. Lv et al. [54] showed that the F adsorption capacity of calcined Mg/Al LDH was
432
higher than that of calcined Ni/Al LDH and calcined Zn/Al LDH because of the higher
433
atomic weights of Ni and Zn compared to Mg. Among the calcined Mg/Al LDHs, the one
434
having Mg/Al molar ratio of 2 was found to have the highest F adsorption capacity. The
435
maximum F adsorption capacity of 213 mg/g was obtained at pH 6. This adsorption capacity
436
is the highest of all adsorbents listed in Table 1. However, Das et al. [55] reported a lower
437
Langmuir adsorption capacity of 17 mg/g at pH 6 for a calcined Zn/Al LDH. The F
438
adsorption data of Wang et al. [53] on a calcined Mg/Al LDH did not fit the Langmuir
439
adsorption model but the amount of F adsorbed continued to increase with equilibrium F
440
concentration with an adsorption capacity of 35 mg/g at the highest tested F equilibrium
441
concentration of 70 mg/L. The rate of F adsorption on calcined LDH is variable [53,55,56].
M
an
us
cr
ip t
430
The literature review reveals that LDHs can have very high F adsorption capacity (17-
443
213 mg/g) if calcined to 500oC. The adsorption capacity varies depending on the type and
444
proportion of the metals in the LDH structure. Because LDHs have high adsorption capacity,
445
they are useful adsorbents for defluoridation of waters with high F concentrations.
447 448
pt
Ac ce
446
ed
442
4.3. Ion exchange resins and fibres Ion exchange resins and fibres are an important class of adsorbents used to remove
449
anionic and cationic pollutants from water and wastewater. Their framework or matrix
450
consists of irregular, macromolecular, three dimensional network of hydrocarbon chain [32].
451
The cation exchange resins and fibres have negatively charged functional groups whereas the
452
anion exchange resins and fibres have positively charged functional groups such as -NH3+,
453
NH2+, ≡N+, ≡S+. Therefore the cation exchangers adsorb cations and the anion exchangers
454
adsorb anions such as F-. The cation exchangers can also be made to adsorb anions if they are 19
Page 19 of 62
455
impregnated with positively charged metallic cations that have strong affinity for anions
456
[39,57]. Ku et al. [57] used an Al incorporated cation exchange resin (Amberlite IR-120) to
458
remove F from water and found that the Langmuir maximum F adsorption capacity at pH 4
459
was 4.6 mg/g. Within the pH range of 4-9 tested, F adsorption emerged as the greatest at pH
460
5-7. Metals other than Al have also been incorporated into cation exchangers to enhance the F
461
adsorption capacity. For example, Luo and Inoue [39] compared the F adsorption capacities
462
of the cation exchange resin, Amberlite 200 CT modified by the incorporation of a number of
463
trivalent metal ions, La, Ce, Y, Fe, and Al. The F adsorption capacity of the metals
464
incorporated resins in the pH range of 4-7 was in the order: La ≤ Ce > Y > Fe ~ Al. More
465
than 80% of F was adsorbed by 20 g La/L resin from a 15 mL solution containing 15 mg F/L.
466
The resin without any metal incorporation adsorbed only about 5% of F from the solution.
M
an
us
cr
ip t
457
The majority of anion exchange materials are not effective in adsorbing F from
468
natural water containing other anions because F- affinity to anion exchange materials is the
469
least of all anions (citrate > SO42- > oxalate > I- > NO3- > CrO42- > Br- > SCN- > Cl- > formate
470
> acetate > F- [32]). Consistent with this order of affinity, a strong base anion exchanger was
471
shown to effectively adsorb NO3-, Br-, and SO42-, but not Cl- and F- [83].
pt
Ac ce
472
ed
467
The F adsorption capacity of anion exchangers can be enhanced by modifying the
473
adsorbing sites on the anion exchangers. Generally, anion exchange resins impregnated with
474
chelating agents that can form H-bonding with F are used. Meenakshi and Viswanathan [30]
475
compared the F adsorption capacity of a chelating resin having sulfonic acid functional group
476
and an anion exchange resin and showed that 1 g of the chelating resin adsorbed 95% of F
477
from a 50 mL solution containing 3 mg F/L in 40 min compared to 65% adsorption by the
478
anion exchange resin.
20
Page 20 of 62
Solangi et al. [34] reported that 100 mg of the ion exchange resin, Amberlite XAD-4
480
modified by incorporation with thio-urea binding sites removed 90% of the F from a 10 mL
481
solution containing 16 mg F/L compared to 30% by the unmodified resin at pH 7. The higher
482
F adsorption capacity of the modified resin was explained as being due to H-bonding between
483
the amide groups in thio-urea and F. The Langmuir maximum F adsorption capacity of the
484
modified resin at pH 7 was much higher than the F adsorptive capacities of many other
485
adsorbents in the literature (Table 1). Interference of other co-existing anions present at the
486
concentration ratio of 1:5 (F:other ions) was insignificant. Therefore the authors concluded
487
that the modified resin can be effectively used for defluoridation of water.
us
cr
ip t
479
Fibrous adsorbents, due to their physico-chemical structure, generally have rapid rates
489
of adsorption of ions. If their adsorptive capacity and affinity can also be enhanced, they can
490
be a useful class of adsorbents for removal of ions from water. Ruixia et al. [31] introduced
491
functional groups into a polyacrylonitrile ion exchange fibre and studied its F adsorption
492
behaviour. They found that the F adsorption reached equilibrium in 5 min with 40% of F
493
adsorption from a 500 mL solution containing 5, 34, and 50 mg/L of F, As(V), and P,
494
respectively when 1 g fibre modified with functional groups was added. The adsorption by
495
the unmodified fibre was less than 5% of solution F. The rate of adsorption was considered to
496
be faster than that observed in many other adsorbents. Maximum adsorption was obtained at
497
pH 3. The mechanism of adsorption was considered to be ion exchange.
M
ed
pt
Ac ce
498
an
488
It could be concluded from this review that ion exchange resins and fibres have poor
499
F adsorption capacities but the adsorption capacities and selectivity for F adsorption in the
500
presence of other ions can be significantly increased (up to 61 mg/g) by surface modification
501
of the adsorbents by loading with organic functional groups and metals. These adsorbents are
502
relatively expensive and therefore they are useful only for water treatment in industrial
503
countries. 21
Page 21 of 62
504 505
4.4. Zeolites Because zeolites have negative surface charges at all pH values, they have high
507
adsorption capacity for cations, but have low adsorption capacity for anions. Nonetheless
508
their adsorption capacity for anions can be increased by modifying the zeolite surface with
509
cationic surfactants or multivalent metallic cations [85]. For F adsorption, only metallic
510
cations incorporated with zeolites have been used. There appear to be no studies conducted
511
on using surfactant- or organic compounds-modified zeolites on defluoridation.
us
cr
ip t
506
Onyango et al. [40] modified the surface of a synthetic zeolite by exchanging Na+ in
513
the zeolite with Al3+ and La3+ to create active sites for F adsorption. The introduction of Al
514
and La opened up the pores in zeolite leading to increased porosity. The Langmuir adsorption
515
capacity is slightly larger for La-zeolite than for Al-zeolite (Table 1). However, within the F
516
concentration range studied (10-80 mg/L), Al-zeolite had twice as much adsorption capacity
517
as the La-zeolite. The adsorption capacities obtained for these zeolites were reported to be
518
much higher than many other adsorbents including the commonly used activated alumina
519
(Table 1). This suggested that the mechanism of adsorption of F onto Al-zeolite was mostly
520
by a chemical adsorption process (ligand exchange) and adsorption onto La-zeolite was
521
mostly by a physical adsorption process (coulombic attraction).
M
ed
pt
Ac ce
522
an
512
The Na-zeolite had no PZC because it was negatively charged at all pHs between 3
523
and 13 and therefore the adsorption of F- is poor due to charge repulsion. In contrast, the Al-
524
zeolite and the La-zeolite had PZC at pHs of 8.15 and 4-5.25, respectively indicating that
525
below these pHs these zeolites were positively charged. In accordance with the surface
526
charge on La-zeolite the F adsorption was the highest at pH 5; and decreased above and
527
below this pH. In the case of the Al-zeolite, the increase of pH increased F adsorption with a
528
plateau forming above pH 5. The F adsorption on Al-zeolite at pHs above 5 did not decrease 22
Page 22 of 62
529
as observed for La-zeolite up to the highest pH 9 tested because F was reported to be
530
adsorbed by chemical adsorption. The authors stated that, over all, Al-zeolite was superior to
531
La-zeolite for defluoridation. In a subsequent study by Onyango et al. [41] on F adsorption by four types of Al pre-
533
treated low-silica synthetic zeolite it was observed that F adsorption on all four types
534
increased from pH 2 to 4 and remained constant up to pH 8, before decreasing from pH 8 to
535
11. The pH effect on F adsorption was explained by the following reactions.
cr
ip t
532
an
us
536
537
M
538
At low pH, part of the F is removed as weakly ionised HF and therefore F adsorption is
540
reduced (first equation). Another reason could be that some of the F is complexed to Al
541
solubilised at low pHs to form AlF+ complex which has low tendency to adsorb on the
542
positively charged adsorbent at the low pHs. At high pH, the negatively charged adsorbent
543
(last equation) repels the negatively charged F- as well as increased competition of OH- for
544
adsorption.
pt
Ac ce
545
ed
539
Samatya et al. [23] using a natural zeolite from Turkey pre-treated with La, Al, and Zr
546
to study the removal of F from tap water spiked with NaF. They showed that the F adsorption
547
capacities at the equilibrium F concentration range of 0-12 mg/L were largest for Zr-zeolite
548
and smallest for La-zeolite. The values for the La and Al zeolites were lower than the values
549
reported by Onyango et al. [40], probably because the metal loadings were much lower and
550
the zeolite used was natural compared to the synthetic zeolite of Onyango et al. [40]. 23
Page 23 of 62
551
However, all three metal zeolites removed 95% of F from an aqueous solution containing 2.5
552
mg F/L at an adsorbent dose of 6 g/L. In contrast to the above studies which showed high F removal capacities of
554
multivalent metal incorporated zeolites, Diaz-Nava et al. [85] found that La and Eu treated
555
natural zeolite from Mexico had only slightly higher F removal capacities than when this
556
zeolite was treated with Na and Ca. The weight percentages of the metals in the zeolite were
557
1.7, 3.20, 0.32, and 1.49 for Na, Ca, La, and Eu, respectively.
cr
ip t
553
The review revealed that zeolites on their own had very low adsorption capacities for
559
F but when they were loaded with multivalent metallic cations they produced moderate to
560
very high F adsorptive capacities (2-45 mg/g). The adsorptive capacity depends on the type
561
and amount of metal loading. Zeolites can also provide good physical properties for practical
562
use.
M
an
us
558
564
4.5. Carbon materials
ed
563
Activated carbon (AC) is an important carbon material commonly used as an
566
adsorbent for the removal of a wide range of aquatic pollutants due to its exceptionally high
567
surface area (500-1500 m2/g), highly developed internal microporosity, presence of a range of
568
functional groups, low cost and ready availability [86,87]. However, it displays poor
569
adsorption capacity towards anionic pollutants because of its low PZC (pH 1.6-3.5, [88]).
Ac ce
570
pt
565
The amount of anions adsorbed onto AC depends on the pore size distribution
571
because the adsorption occurs mainly in the pores. Abe et al. [60] reported that F adsorption
572
capacity increased with the specific surface area in 11 out of the 12 carbonaceous materials
573
studied by them. The adsorption of F on bone char did not fit this pattern because the
574
mechanism of adsorption was different from the rest. In bone char, F is adsorbed chemically
575
by ligand exchange with the OH group in the hydroxyapatite compound [1]. The 24
Page 24 of 62
ineffectiveness of carbon in adsorbing F was also shown by Srimurali et al. [70] who found
577
that lignite coal and char fines, a by-product obtained during the making of coke, adsorbed
578
only 7.9 and 19% of F, respectively from a 50 mL solution containing 5.0 mg F/L and when
579
100 mg of adsorbents were added and mixed for 5 h. In comparison, bentanite clay removed a
580
much larger percentage of 33% of solution F.
ip t
576
A new type of carbon material called aligned carbon nanotubes (ACNT) made up of
582
needle-like cylindrical tubules of concentric graphitic carbon capped by fullerene-like
583
hemispheres was developed as a promising adsorbent material for the adsorption of
584
contaminants in water [12]. Li et al. [12] found that ACNT had higher F adsorption capacity
585
than AC (Table 1), despite the ACNT having a much lower specific area and pore volume
586
than the AC. Li et al. [12] stated that though the adsorption capacity of the ACNT was high,
587
the cost too was high which may limit its full commercial utilisation.
M
an
us
cr
581
The surface of the carbon particles can be modified to improve the F adsorptive
589
properties. Such modifications have been brought about by creating new functional groups
590
which have strong affinity towards F. Daifullah et al. [59] modified the structure of an AC
591
product by steam pyrolysis of rice straw and oxidised the product using HNO3, H2O2, and
592
KMnO4. Of these treatments, the material obtained by KMnO4 oxidation produced the
593
highest F adsorption. The adsorption mechanism was considered to be ligand exchange of
594
OH groups on the carbon surface with F. The presence of MnO2 on the carbon surface caused
595
by reduction of the KMnO4 may have also participated in the removal of F. Thermodynamic
596
studies showed that the adsorption was chemical in nature. The Langmuir adsorption
597
maximum was 15.9 mg F/g at the natural pH of water. This value was considered to be higher
598
than the values reported in the literature for many other adsorbents.
Ac ce
pt
ed
588
599
Ramos et al. [42] studied the F adsorption behaviour of Al-impregnated AC (AlAC)
600
produced from coconut shells followed by calcination to 300oC. The AlAC had a F 25
Page 25 of 62
adsorption capacity more than four times higher than the plain carbon at an equilibrium
602
concentration of 2-8 mg F/L. The Langmuir adsorption capacity of AlAC (1.07 mg/g) at
603
natural water pH, though higher than the value obtained for plain carbon (0.49 mg F/g), was
604
lower than the values obtained for some other adsorbents especially activated alumina. The F
605
adsorption reached its maximum at pH 3 because the PZC of the AlAC was 4.1-4.8 (above
606
this pH, AlAC had negative surface charges which repel F-). However, there was appreciable
607
adsorption of F at pH 6-7 (0.6 mg F/g at the equilibrium F concentration of 6 mg F/L) due to
608
chemical adsorption of F onto Al which does not involve electrostatic attractive forces.
us
cr
ip t
601
In a later study, Li et al. [61] reported that the adsorption capacity of aligned carbon
610
nanotubes (ACNT) at pH 6 increased from 2.3 mg F/g to 9.6 mg F/g when Al was loaded
611
onto ACNT and calcined to 450oC at an optimum Al2O3 rate of 30% by weight. The
612
adsorption capacity fell to less than 6 mg F/g when the loading rate was 20 and 40%. The
613
maximum adsorption capacity was obtained at pH 6.0-9.0. The mechanism of F adsorption at
614
pH less than the PZC of 7.5 was considered to be by electrostatic attraction of negatively
615
charged F- ions onto the positively charged adsorbent as well as by ligand exchange of the
616
OH group by F. At the PZC and slightly above this pH the adsorption was reported to be
617
mainly by ligand exchange. At pHs much higher than the PZC (pH > 9), OH- ions in solution
618
competed with F- ions for adsorption and therefore F adsorption decreased.
M
ed
pt
Ac ce
619
an
609
It can be concluded from the review that AC is a poor adsorbent for F. However, its
620
adsorption capacity can be increased to moderate levels (up to 16 mg/g) by chemical
621
modification of the carbon surface. Specialised carbon materials such as ACNT have higher
622
adsorption capacity than AC. Their adsorption capacity can also be increased by surface
623
modification but this will incur additional cost.
624 625
4.6. Natural materials 26
Page 26 of 62
626
Several natural inorganic materials (soils, clays, minerals, and building materials)
627
have been used in the defluoridation of water. The adsorption capacities of these materials
628
have been discussed in previous reviews (17, 26, 27) and also presented in Table 1. Bioadsorbents contain a variety of functional groups such as carboxyl, imidazole,
630
sulphydryl, amino, thioether, phenol, carbonyl, amide, and hydroxyl moieties capable of
631
adsorbing the pollutants, especially the metal cations [78]. Considering this feature of
632
bioadsorbents, metals having strong affinity towards F were loaded onto selected
633
bioadsorbents (Zr4+ on collagen fibres [13], La3+ on geletin [44]) for enhancing the adsorption
634
of F from drinking water.
us
cr
ip t
629
Chitin and chitosan have been shown to constitute another effective group of
636
bioadsorbents for removing a range of aquatic pollutants due to their low cost and high
637
contents of amino and hydroxyl functional groups [89]. Chitin is a polysaccharide found in a
638
wide range of organisms but most commonly extracted from shellfish processing waste,
639
while chitosan is a copolymer of glucosamines derived from chitin by deacetylation in hot
640
alkali [67]. Sahli et al. [22] reported that the adsorption of F on chitosan from a synthetic
641
water was rapid and reached a maximum at 5 min. Increase of pH from 2 to 6 increased
642
adsorption followed by a decrease up to pH 10. The F adsorption from a brackish water
643
containing 3.25 mg F/L and several other anions at much higher concentrations (Cl- 1083
644
mg/L, SO42- 215 mg/L, HCO3- 171 mg/L) showed that the selectivity of chitosan towards
645
various anions was as follows: F- > HCO3- >NO3- > Cl- > SO42-. In a batch adsorption study of
646
15 min duration the F concentration fell to 1.82 mg/L, but with a successive batch adsorption
647
the concentration reduced to 0.9 mg/L which was lower than the WHO standard of 1.5 mg
648
F/L.
Ac ce
pt
ed
M
an
635
649
Ma et al. [66] prepared a magnetic-chitosan adsorbent by co-precipitation of chitosan
650
with Fe2+ and Fe3+ salts and used it to remove F from a synthetic F solution in a batch study. 27
Page 27 of 62
The adsorbent was shown to have a higher F adsorption capacity (Langmuir adsorption
652
capacity 22.49 mg F/g) than a commercial activated alumina (Table 1). Adsorption capacity
653
was highest in the pH range 5-9. The FTIR data of the adsorbent showed that the main groups
654
involved in the adsorption process were amine and iron oxides in the magnetic-chitosan. The
655
main advantage of having magnetic properties in the adsorbent is to easily separate the
656
adsorbent from the treated water after its use using an external magnetic field so that the
657
adsorbent can be reused.
cr
ip t
651
A research group in India has reported that the F adsorption capacity of nano-
659
hydroxyapatite (nHAp) can be increased by incorporating chitin or chitosan onto nHAp [63-
660
65,90]. In a batch experiment conducted by Sundaram et al. [65], it was observed that mixing
661
0.25 g each of chitosan, nHAp, and nHAp/chitosan composite with 50 mL solution of NaF
662
containing 10 mg F/L, removed 0.05, 1.30, and 1.56 mg F/g adsorbent, respectively after 30
663
min of shaking the suspensions. Defluoridation of water from a fluoride-endemic village
664
using nHAp-chitosan composite was found to be higher compared to using nHAp. Similar
665
results were obtained from a subsequent study where a composite of nHAp and chitin was
666
used [64]. However, the adsorption capacity of this composite was higher than that of the
667
composite made with chitosan (Table 1).
an
M
ed
pt
Ac ce
668
us
658
The review revealed that although the majority of natural materials have low
669
effectiveness in removing F (< 1 mg/g), they are inexpensive and consequently can be used in
670
less developed countries, especially in rural areas. Furthermore, their F removal capacities
671
can be increased through the process of surface modification, such as loading with
672
multivalent metallic cations, as reported previously for other adsorbents.
673 674
4.7. Industrial by-products
28
Page 28 of 62
675
Several types of industrial by-products have been used for the adsorptive removal of
676
pollutants including fluoride from water [27,28,89,91,92]. Most of the adsorbents fall into the
677
categories of by-products from the mining industry (e.g. red mud), steel industry (e.g. slag
678
materials), and power plant industry (e.g. fly ash). Cengeloğlu et al. [4] investigated the adsorption of F from aqueous solutions using a
680
red mud containing 18.7% Al2O3, 39.7% Fe2O3, and 14.5% SiO2 with (ARM) and without
681
(RM) activation by 20% HCl treatment in a batch study. The amount of F removed from a
682
solution (pH 5.5) containing 3.8-28.5 mg F/L by ARM at a dose of 0.2 g/50 mL was at least
683
three times higher than that by RM at the same dose. The main mechanism of F removal was
684
considered to be that of specific adsorption caused by ligand exchange of F and OH groups
685
on the metal oxide surfaces in red mud.
an
us
cr
ip t
679
Red mud particles are too fine for use in filter beds because of their poor hydraulic
687
conductivity. This problem can be overcome either by mixing red mud with coarse-size
688
particles such as sand [93] or by granulation [94] before use. Tor et al. [72] used the method
689
of Zhu et al. [94] for the granulation of red mud by mixing it with fly ash, sodium carbonate,
690
quicklime and sodium silicate. The maximum F adsorption (2.2 mg/g) from a NaF solution
691
containing 15 mg F/L and granulated red mud (GRM) of 5 g/L was obtained at a pH of 4.7 in
692
the pH test range of 2.5-7.3. The adsorption capacity determined from a column study was
693
2.05 mg F/g for a flow rate of 2 mL/min compared to 0.644 mg F/g obtained in the batch
694
study at the same GRM dosage (5g/L) and initial F concentration (5 mg/L).
ed
pt
Ac ce
695
M
686
Fly ash is a major by-product that is produced from the combustion of coal in power
696
stations. It consists of fine and powdery materials (1.0-100 µm) made up of a mixture of
697
amorphous and crystalline alumina-silicates and several compounds of Si, Al, Fe, Ca, and Mg
698
[95] and therefore a good candidate material for F adsorption from water. Chaturvedi et al.
699
[46] reported that the adsorption of F from water by a fly ash containing 56.0% SiO2, 25.9% 29
Page 29 of 62
Al2O3, 2.22% CaO, and 1.26% Fe2O3 had a maximum Langmuir adsorption capacity of 20
701
mg/g at pH 6.5. The F adsorption increased from 79 to 94% when the pH of the F solution
702
(10 mg/L) increased from 2.0 to 6.5 and then decreased with further increase in pH up to 9.5.
703
However, Nemade et al. [62] reported that F adsorption by a fly ash decreased continuously
704
from pH 2 to 12. The difference in the pH effects between the two studies could be due to the
705
difference in chemical characteristics of the fly ashes and experimental conditions used.
ip t
700
In a column study, Piekos and Paslawska [74] found that an alkaline fly ash (pH ≥ 10,
707
9.1% CaO) effectively removed F from water containing 1-100 mg F/L when the solution
708
was passed through a column (400 mm length) packed with 450 g fly ash at a flow rate of ≤ 2
709
mL/h. Complete adsorption of F was obtained after 120 h. The F adsorption mechanism was
710
suggested to be chemical binding of F onto Ca(OH)2 and physical adsorption onto the
711
residual carbon particles in the fly ash.
M
an
us
cr
706
Steel industry by-products such as blast furnace slag, electric arc furnace slag, basic
713
oxygen furnace slag and converter slag have been shown to adsorb many pollutants,
714
especially phosphate, from water due to the presence of high contents of Ca, Fe, Al, Mg, and
715
Mn oxides [28]. However, limited studies have been conducted in their use in defluoridation
716
of water [16,27]. Islam and Patel [71] showed that a basic oxygen furnace slag (BOFS)
717
containing 46.5% CaO, 16.7% Fe oxides, and 13.8% SiO2 had a good capacity to adsorb F.
718
Thermal activation of the BOFS by heating to 1000oC for 24 h increased the porosity and
719
surface area leading to increased F adsorption. The percentage adsorption of F (initial F
720
concentration 10 mg/L, 0.5 g/100 mL adsorption dosage) was higher for the thermally
721
activated BOFS (TABOFS) (93%) than for the BOFS (70%). Unlike the case with most
722
adsorbents, increase of pH from 2 to 10 increased F adsorption on TABOFS. The increased
723
adsorption at high pH values was considered to be due to F adsorption onto Ca(OH)2 in the
Ac ce
pt
ed
712
30
Page 30 of 62
724
slag. Presence of other anions reduced F adsorption in the order, PO43- > HCO3- > CO32- >
725
SO42- > Cl- > NO3-. Large quantities of alum sludge, a waste product, are generated during the
727
manufacture of alum from bauxite [48] and kaolin [96] by the sulphuric acid process in many
728
countries. As this sludge mainly consists of oxides of Al, it has a high adsorption capacity to
729
remove many aquatic pollutants. Sujana et al. [48] used a sludge (47.2% Al2O3 and 20.7%
730
TiO2) produced from an alum manufacturing plant using bauxite raw material for the removal
731
of F from water. The sludge after washing with water and calcining to 400oC for 3 h adsorbed
732
more than twice the amount of F that was adsorbed by the uncalcined sludge from a solution
733
containing 10 mg F/L of pH 6 (adsorbent dosage 20-300 mg/g). The higher adsorption
734
capacity of the calcined sludge was considered to be due the increased surface area produced
735
by calcination. The F was considered to be specifically adsorbed according to a two-step
736
ligand exchange mechanism. Within a pH range of 3 to 9, the maximum F adsorption was
737
observed at pH 6; and at both above and below pH 6 the adsorption decreased, as also
738
reported for activated alumina by others. The presence of other anions in solution reduced F
739
adsorption according to the order, PO43- > silicate > SO42- > NO3-.
pt
ed
M
an
us
cr
ip t
726
In contrast to F adsorption on sludge produced during the manufacture of alum from
741
bauxite discussed in the previous paragraph, Nigussie et al. [96] reported that F adsorption by
742
a sludge produced during the manufacture of alum from kaolin raw material containing
743
primarily SiO2 did not increase when the sludge was heated to 300oC. At temperatures above
744
300oC up to 700oC, the adsorption decreased. Also, F adsorption was nearly constant from
745
pH 3 to 8 and significantly decreased above pH 10. The difference in the adsorption
746
behaviour of the two types of sludge may be due to the differences in the physico-chemical
747
characteristics of the sludge types. However, the F adsorption capacities of the sludge in the
748
two studies were not too different. The presence of HCO3- at concentrations higher than those
Ac ce
740
31
Page 31 of 62
749
of F, decreased F adsorption efficiency, while other anions (PO43-, SO42- , Cl- , NO3-) had no
750
significant effect [96]. Lai and Liu [75] used a by-product of the petrochemical industry called spent catalyst
752
consisting mainly of porous silica and alumina (specific surface area 130 m2/g, PZC 5.2) to
753
remove F from aqueous solutions. In a batch adsorption experiment on pH effect, they
754
observed that F adsorption decreased with increase in pH from 2 to 9.5, with a plateau in the
755
pH range of 4 to 7. The maximum adsorption capacity at pH 4 was found to be 28 mg F/g at
756
25 and 50oC. This value was considered to be comparable to that of activated alumina.
757
Because the activation energy calculated using the Arrhenius equation was very low (3.2
758
J/mol) the mechanism of adsorption was considered to be non-specific adsorption involving
759
coulombic forces.
an
us
cr
ip t
751
Many of the industrial by-products are wastes that need to be disposed of. Their
761
beneficial use can save disposal cost, prevent environmental pollution arising from the
762
disposal sites, and release disposal lands for alternative uses. In this respect, though these
763
materials have low F adsorption capacities, they are cost-effective and therefore can
764
potentially be used in developing countries where cost of operation is a major factor in the
765
choice of adsorbents. Furthermore, many of these materials have been shown to be good
766
adsorbents for other aquatic pollutants and therefore it is possible to simultaneously remove
767
many pollutants using these materials.
769
ed
pt
Ac ce
768
M
760
5. Adsorption thermodynamics
770
The strength and spontaneous nature of adsorption and information on whether the
771
adsorption process is exothermic or endothermic are provided by determining thermodynamic
772
parameters such as changes in Gibbs free energy (ΔG0), enthalpy (ΔH0), and entropy (ΔS0).
32
Page 32 of 62
773
These thermodynamic parameters are calculated from data on adsorption at different
774
temperatures using standard thermodynamics equations (see references in Table 2). A negative value for ΔG0 indicates spontaneous and thermodynamically favourable
776
adsorption, while a negative ΔH0 value indicates an exothermic adsorption process. An
777
exothermic reaction means that the amount of adsorption decreases with increasing
778
temperature. An endothermic reaction (positive ΔH0 value) has been explained as due to
779
enlargement of pore sizes and/or activation of the adsorbent surface [97,98]. Very low ΔH0
780
values are generally associated with physical adsorption and very high values with chemical
781
adsorption [99]. However, no definite value for distinguishing the two forms of adsorption
782
exists. ΔG0 has also been used to distinguish between the two forms of adsorption. For
783
example, Meenakshi and Viswanathan [30] reported that ΔG0 values of up to -20 kJ/mol were
784
indicative of physical adsorption, while values less than -40 kJ/mol involved chemical
785
adsorption. Positive ΔS0 values indicate good affinity of adsorbed species towards the
786
adsorbent and increased randomness at the solid-solution interface associated with structural
787
changes at the adsorption sites during the adsorption process [99,100].
ed
M
an
us
cr
ip t
775
All studies, except the one on F adsorption at the highest temperature of 55oC on
789
activated carbon modified by H2O2/KMnO4 oxidation [59] presented in Table 2 had negative
790
ΔG0 values. The negative values indicate good affinity of F for the adsorbents and the
791
adsorption process was spontaneous. The small positive value obtained for the activated
792
carbon was attributed to the disruption of the MnO2 formed at the carbon surfaces and
793
increase F solubility at high temperature [59]. The ΔS0 values were mostly positive, again
794
indicating a strong affinity of F towards the adsorbents. The ΔH0 values, however, were
795
positive and negative indicating the endothermic and exothermic nature of adsorption,
796
respectively.
Ac ce
pt
788
797 33
Page 33 of 62
798
6. Fluoride desorption and adsorbent regeneration A suitable adsorbent for F removal should not only have high F adsorption capacity
800
and cost-effectiveness but also be amenable to easy desorption of the adsorbed F and capable
801
of efficient regeneration for multiple reuse of the adsorbent. Also, the desorbing agent should
802
not produce any damage to the adsorbent that causes a reduction in its adsorption capacity.
803
Only adsorbents that can be reused have practical value in real systems for economic and
804
environmental reasons. However, inexpensive adsorbents, such as industrial by-products and
805
natural materials, may be used only once and they need not be regenerated for reuse because
806
usually the costs of desorbing chemicals and regeneration process are higher than that of
807
these adsorbents.
an
us
cr
ip t
799
Desorption of F is carried out by leaching of adsorbed F by acids, bases and salts
809
(Table 3). The selection of a desorbing agent largely depends on the influence of pH on F
810
adsorption and the strength of adsorption. Adsorption on most of the adsorbents decreases at
811
high pHs, and therefore bases having high pHs are commonly used for desorption of F from
812
such adsorbents (Table 3). As explained previously, the desorption of F at high pHs is due to
813
the increased repulsive forces between the negatively charged adsorbent surface and
814
negatively charged F- ions in solution as well as the competition between increased
815
concentration of OH- at high pH values and F- for adsorption. In materials where F adsorption
816
increases with pH, acids have been found to be more effective as desorbing agents. For
817
example, in a hydrous manganese oxide coated alumina [29] and a thio-urea modified
818
amberlite resin [34], F adsorption decreased at lower pH values. Consistent with this
819
adsorption pattern, the percentage of F desorption was found to be higher at low pH values.
820
Adsorption of F by most adsorbents is strong and not easily reversible, partly due to the
821
chemical nature of the adsorption process. Therefore, stronger acids and bases, as well as a
Ac ce
pt
ed
M
808
34
Page 34 of 62
822
longer time of their interaction with the adsorbent are required for efficient F desorption
823
[36,41,50,55]. Water and neutral salts are generally found to be less effective [31,41].
824 825
7. Conclusions Adsorption and precipitation/coagulation methods are widely used for the
827
defluoridation of water. The precipitation/coagulation method has the drawback of not being
828
efficient in waters having low F concentration, and the process produces toxic AlF complexes
829
in the treated water as well as large volumes of waste. The adsorption processes are generally
830
considered attractive because of their effectiveness, convenience, ease of operation,
831
simplicity of design, and for economic and environmental reasons.
an
us
cr
ip t
826
Multivalent metal oxides and hydroxides and layered double hydroxides generally
833
have high F adsorption capacities. Ion exchange resins and fibres, zeolites, and carbon
834
materials have low adsorption capacities on their own, but when their surfaces are modified
835
by incorporating organic functional groups or multivalent metallic cations, the adsorption
836
capacity increased. Some of the adsorbents (e.g. layered double hydroxides, alumina) had to
837
be activated by calcining at high temperatures to increase the adsorption capacity. The above
838
materials have potential for use in industrial countries and in areas where the F concentration
839
in water is very high. Natural and industrial by-products have low adsorption capacities but
840
because they are inexpensive, they have potential for use in rural areas, especially in
841
developing countries.
ed
pt
Ac ce
842
M
832
The pH of water is a dominant factor influencing F adsorption. Generally, F
843
adsorption increases from acidic to near neutral pH and then decreases with increase in pH.
844
Consistent with this pH effect, acids and alkali have been successfully used to desorb F and
845
consequently regenerate the adsorbent for reuse. However, multiple regeneration and reuse
35
Page 35 of 62
846
reduce the adsorptive capacity. Another important factor influencing F adsorption is the type
847
and concentration of other ions present in water. Future research needs to explore highly efficient, low cost adsorbents that can be
849
easily regenerated for reuse over several operational cycles without significant loss of
850
adsorptive capacity and have good hydraulic conductivity to prevent filters clogging during a
851
fixed-bed treatment process. Surface modifications of the adsorbents can be explored to
852
increase the capacity and strength of adsorption without significantly increasing the cost. The
853
majority of studies reported have been conducted in batch trials on synthetic waters. These
854
trials need to be extended using continuous mode column trials which have more relevance to
855
real operating systems on natural waters containing other ions as well.
an
us
cr
ip t
848
856
Acknowledgements
M
857
This study was funded by cooperative Research Centre for Contamination Assessment
859
and Remediation of the Environment (CRC CARE) (Project Number 4.1.12.11/12). We thank
860
Phil Thomas for editing this manuscript.
861
863 864 865
References
Ac ce
862
pt
ed
858
[1] J. Fawell, K. Bailey, E. Chilton, E. Dahi, L. Fewtrell, Y. Magara, Fluoride in Drinking Water, World Health Organization, IWA Publishing, UK, 2006.
866
[2] World Health Organisation, Guidelines for Drinking-water Quality, Volume 2. Health
867
Criteria and Other Supporting Information, 2nd Edition, World Health Organisation,
868
Geneva, 1996.
36
Page 36 of 62
869
[3] US Public Health Service, US Public Health Service Drinking Water Standards, US
870
Government Printing Office, Department of Health Education and Welfare, Washington
871
DC, 1962.
876 877 878
ip t
875
[5] N.J. Chinoy, Effects of fluoride on physiology of animals and human beings, Indian J. Environ. Toxicol. 1 (1991), 17-32.
cr
874
mud, Sep. Purif. Technol. 28 (2002) 81-86.
[6] P.T.C. Harrison, Fluoride in water: a UK perspective, J. Fluor. Chem. 126 (2005) 1448-
us
873
[4] Y. Cengeloğlu, E. Kir, M. Ersöz, Removal of fluoride from aqueous solution by using red
1456.
[7] E. Kumar, A. Bhatnagar, J. Minkyu, W. Jung, S. Lee, S. Kim, G. Lee, H. Song, J. Choi, J.
an
872
Yang, B. Jeon, Defluoridation from aqueous solutions by granular ferric hydroxide
880
(GFH), Water Res. 43 (2009), 490-499.
884 885 886
ed
883
Mater. B137 (2006) 456-463.
[9] S. Vigneswaran, C. Visvanathan, Water Treatment Processes. Simple Options, CRC Press, Inc., Florida, 1995.
pt
882
[8] Meenakshi, R.C. Maheshwari, Fluoride in drinking water and its removal, J. Hazard.
[10] J.J. Murray, Appropriate Use of Fluorides for Human Health, World Health
Ac ce
881
M
879
Organisation, Geneva. 1986.
887
[11] P. Loganathan, M.J. Hedley, N.D. Grace, J. Lee, S.J. Cronin, N.S. Bolan, J.M. Zanders,
888
Fertiliser contaminants in New Zealand grazed pasture with special reference to
889
cadmium and fluorine - a review, Aust. J. Soil Res. 41 (2003) 501-532.
890
[12] Y.H. Li, S. Wang, X. Zhang, J. Wei, C. Xu, Z. Luan, D. Wu, Adsorption of fluoride
891
from water by aligned carbon nanotubes, Materials Res. Bull. 38 (2003) 469-476.
892
[13] X. Liao, B. Shi, Adsorption of fluoride on zirconium(IV)-impregnated collagen fiber.
893
Environ. Sci. Technol. 39 (2005) 4628-4632. 37
Page 37 of 62
894
[14] S.P. Mishra, M. Das, U.N. Dash, Review on adverse effects of water contaminants like
895
arsenic, fluoride and phosphate and their remediation, J. Scientific Industrial Res. 69
896
(2010) 249-253. [15] E.G. Paulson, Reducing fluoride in industrial wastewater, Chem. Eng. 84 (1977) 89-94.
898
[16] S.S. Babu, S. Kumar, P.D. Nemade, T. Roychowdhury, Impacts of fluoride in drinking
ip t
897
water and role of pH in its removal efficiency: a critical review, Crit. Rev. Environ. Sci.
900
Technol. (2012) (in press).
904 905
us
903
defluoridation of water, Chem. Rev. (2012) 2454-2466.
[18] W.W. Choi, K.Y. Chen, The removal of fluoride from waters by adsorption, J. Amer.
an
902
[17] S. Jagtap, M.K. Yenkie, N. Labhsetwar, S. Rayalu, Fluoride in drinking water and
Water Works Assoc. 71 (1979) 562-570.
[19] K.B.P.N. Jinadasa, S.W.R. Weerasooriya, C.B. Dissanayake, A rapid method for the
M
901
cr
899
defluoridation of fluoride-rich drinking waters at village level, Inter. J. Environ. Studies.
907
31 (1988) 305-312.
ed
906
[20] J.P. Padmasiri, C.B. Dissanayake, A simple defluoridator for removing excess fluorides
909
from fluoride-rich drinking water, Intern. J. Environ. Health Res. 5 (1995) 153-160.
910
[21] C. Zevenbergen, L.P. van Reeuwijk, G. Frapporti, R.J. Louws, R.D. Schuiling, A simple
Ac ce
pt
908
911
method for defluoridation of drinking water at village level by adsorption on Ando soil
912
in Kenya, Sci. Total Environ. 188 (1996) 225-232.
913
[22] M.A.M. Sahli, S. Annouar, M. Tahaikt, M. Mountadar, A. Soufiane, A. Elmidaoui,
914
Fluoride removal for underground brackish water by adsorption on the natural chitosan
915
and by electrodialysis, Desalination 21 (2007) 37-45.
916
[23] S. Samatya, U. Yüksel, M. Yüksel, N. Kabay, Removal of fluoride from water by metal
917
ions (Al3+, La3+ and ZrO2+) loaded natural zeolite, Sep. Sci. Technol. 42 (2007) 2033-
918
2047. 38
Page 38 of 62
919
[24] K.R. Bulusu, B.B. Sundaresan, B.N. Pathak, W.G. Nawlakhe, D.N. Kulkarni, V.P.
920
Thergaonkar, Fluorides in water, defluoridation methods and their limitations, J. Inst.
921
Engineers (India) Environ. Engineer. Div. 60 (1979) 1-25.
926 927 928
ip t
925
[26] M. Mohapatra, S. Anand, B.K. Mishra, D.E. Giles, P. Singh, Review of fluoride removal from drinking water, J. Environ. Manage. 91 (2009) 67-77.
cr
924
alum by Nalgonda technique, Indian. J. Environ. Health 17 (1975) 26-65.
[27] A. Bhatnagar, E. Kumar, M. Sillanpää, Fluoride removal from water by adsorption-A
us
923
[25] W.G. Nawlakhe, D.N. Kulkarni, B.N. Pathak, K.R. Bulusu, Defluoridation of water with
review, Chem. Eng. J. 171 (2011) 811-840.
[28] V.K. Gupta, J.M. Carrott, M.M.L.R. Carrott, Suhas, Low-cost adsorbents: Growing
an
922
approach to wastewater treatment – a review. Crit. Rev. Environ. Sci. Technol. 39
930
(2009) 783-842.
M
929
[29] S. Teng, S. Wang, W. Gong, X. Liu, B. Gao, Removal of fluoride by hydrous manganese
932
oxide-coated alumina: Performance and mechanism, J. Hazard. Mater. 168 (2009), 1004-
933
1011.
936 937
pt
935
[30] S. Meenakshi, N. Viswanathan, Identification of selective ion-exchange resin for fluoride sorption, J. Colloid Interface Sci. 308 (2007) 438-450.
Ac ce
934
ed
931
[31] L. Ruixia, G. Jinlong, T. Hongxiao, Adsorption of fluoride, phosphate, and arsenate ions on a new type of ion exchange fiber, J. Colloid Interface Sci. 248 (2002) 268-274.
938
[32] F. Helferich, Ion Exchange, Dover Publication Inc., USA, 1995.
939
[33] E.R. Weiner, Applications of Environmental Aquatic Chemistry. CRC Press, Taylor and
940 941 942
Francis Group, Florida, 2008. [34] I.B. Solangi, S. Memon, M.I. Bhanger, An excellent fluoride sorption behaviour of modified amberlite resin, J. Hazard. Mater. 176 (2010) 186-192.
39
Page 39 of 62
943
[35] A. Sivasamy, K.P. Singh, D. Mohan, M. Maruthamuthu, Studies on defluoridation of
944
water by coal-based sorbents, J. Chem. Technol. Biotechnol. 76 (2001) 717-722.
946 947
[36] S.M. Maliyekkal, A.K. Sharma, L. Philip, Manganese-oxide-coated alumina: A promising sorbent for defluoridation of water, Water Res. 40 (2006) 3497-3506. [37] S. Tokunaga, S.A. Haron, S.A. Wasay, K.F. Wong, K. Laosangthum, A. Uchiumi,
ip t
945
Removal of fluoride ions from aqueous solutions by multivalent metal compounds, Inter.
949
J. Environ. Studies 48 (1995) 17-28.
953 954
us
952
of drinking water, Appl. Geochem. 16 (2001) 531-539.
[39] F. Luo, K. Inoue, The removal of fluoride ion by using metal (III)-loaded Amberlite
an
951
[38] Y. Wang, E.J. Reardon, Activation and regeneration of a soil sorbent for defluoridation
resins, Solvent Extract. Ion Exch. 22 (2004) 305-322.
[40] M.S. Onyango, Y. Kojima, O. Aoyi, E.C. Bernardo, H. Matsuda, Adsorption equilibrium
M
950
cr
948
modelling and solution chemistry dependence of fluoride removal from water by
956
trivalent-cation-exchanged zeolite F-9, J. Coloid Interface Sci. 279 (2004) 341-350.
957
ed
955
[41] M.S. Onyango, Y. Kojima, A. Kumar, D. Kuchar, M. Kubota, H. Matsuda, Uptake of fluoride by Al3+ pretreated low-silica synthetic zeolites: adsorption equilibrium and rate
959
studies, Sep. Sci. Technol. 41 (2006) 683-704.
961
Ac ce
960
pt
958
[42] R.L. Ramos, J. Ovalle-Turrubiartes, M.A.Sanchez-Castillo, Adsorption of fluoride from aqueous solution on aluminium-impregnated carbon, Carbon 37 (1999) 609-617.
962
[43] S.A. Wasay, M.J. Haron, S. Tokunaga, Adsorption of fluoride, phosphate, and arsenate
963
ions on lanthanum-impregnated silica gel, Water Environ. Res. 68 (1996) 295-300.
964 965 966 967
[44] Y. Zhou, C. Yu, Y. Shan, Adsorption of fluoride from aqueous solution on La3+impregnated cross-linked gelatin, Sep. Purif. Technol. 36 (2004) 89-94. [45] P. Trivedi, L. Axe, Long-term fate of metal contaminants in soils and sediments: role of intraparticle diffusion in hydrous metal oxides, in: R. Hamon, M. McLaughlin, E. Lombi 40
Page 40 of 62
968
(Eds.), Natural Attenuation of Trace Element Availability in Soils, CRC Taylor and
969
Francis Group, New York, 2006, pp. 57-71.
975 976 977 978 979
ip t
974
fluoride on activated alumina, Sep. Purif. Technol. 42 (2005) 265-271.
[48] M.G. Sujana, R.S. Thakur, S.B. Rao, Removal of fluoride from aqueous solution by
cr
973
[47] S. Ghorai, K.K. Pant, Equilibrium, kinetics and breakthrough studies for adsorption of
using alum sludge, J. Colloid Interface Sci. 206 (1998) 94-101.
us
972
adsorption on fly ash, Water Air Soil Pollut. 49 (1990) 51-61.
[49] Y. Ku, H. Chiou, The adsorption of fluoride ion from aqueous solution by activated alumina, Water Air Soil Pollut. 133 (2002) 349-360.
an
971
[46] A.K. Chaturvedi, K.P. Yadava, K.C. Pathak, V.N. Singh, Defluoridation of water by
[50] A.M. Raichur, M.J. Basu, Adsorption of fluoride onto mixed rare earth oxides, Sep. Purif. Tech. 24 (2001) 121-127.
M
970
[51] L.E.L. Hammari, A. Laghzizil, P. Barboux, K. Lahlil, A. Saoiabi, Retention of fluoride
981
ions from aqueous solution using porous hydroxyapatite. Structure and conductive
982
properties, J. Hazard. Mater. B114 (2004) 41-44.
pt
984
[52] X. Fan, D.J. Parker, M.D. Smith, Adsorption kinetics of fluoride on low cost materials, Water Res. 37 (2003) 4929-4937.
Ac ce
983
ed
980
985
[53] H. Wang, J. Chen, Y. Cai, J. Ji, L. Liu, H.H. Teng, Defluoridation of drinking water by
986
Mg/Al hydrotalcite-like compounds and their calcined products, Appl. Clay Sci. 35
987
(2007) 59-66.
988
[54] L. Lv, J. He, M. Wei, D.G. Evans, X. Duan, Factors influencing the removal of fluoride
989
from aqueous solution by calcined Mg-Al-CO3 layered double hydroxides, J. Hazard.
990
Mater. B133 (2006) 119-128.
41
Page 41 of 62
991
[55] D.P. Das, J. Das, K. Parida, Physicochemical characterization and adsorption behaviour
992
of calcined Zn/Al hydrotalcite-like compound (HTlc) towards removal of fluoride from
993
aqueous solution, J. Colloid Interface Sci. 261 (2003) 213-220. [56] C. Diaz-Nava, M. Solache-Rios, M.T. Olguin, Sorption of fluoride ions from aqueous
995
solutions and well drinking water by thermally treated hydrotalcite, Sep. Sci. Technol.
996
38 (2003) 131-147.
cation synthetic resin, Sep. Sci. Technol. 37 (2002) 89-103.
cr
998
[57] Y. Ku, H. Chiou, W. Wang, The removal of fluoride ion from aqueous solution by a
us
997
ip t
994
[58] K.M. Popat, P.S. Anand, B.D. Dasare, Selective removal of fluoride ions from water by
1000
the aluminium form of the aminomethylphosphonic acid-type ion exchanger, React.
1001
Polym. 23 (1994) 23-32.
[59] A.A.M. Daifullah, S.M. Yakout, S.A. Elreefy, Adsorption of fluoride in aqueous
M
1002
an
999
solutions using KMnO4-modified activated carbon derived from steam pyrolysis of rice
1004
straw, J. Hazard. Mater. 147 (2007) 633-643.
1007
fluoride ions onto carbonaceous materials, J. Colloid Interface Sci. 275 (2004) 35-39.
pt
1006
[60] I. Abe, S. Iwasaki, T. Tokimoto, N. Kawasaki, T. Nakamura, S. Tanada, Adsorption of
[61] Y.H. Li, S. Wang, X. Zhang, J. Wei, C. Xu, Z. Luan, D. Wu, B. Wei, Removal of
Ac ce
1005
ed
1003
1008
fluoride from water by carbon nanotube supported alumina, Environ. Technol. 24 (2003)
1009
391-398.
1010 1011 1012 1013
[62] P.D. Nemade, A.V. Rao, B.J. Alappat, Removal of fluorides from water using low cost adsorbents, Water Sci. Tech. Water Supply 2 (2002) 311-317. [63] N. Viswanathan, S. Meenakshi, Selective fluoride adsorption by a hydrotalcite/chitosan composite, Appl. Clay Sci. 48 (2010) 607-611.
1014
[64] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Fluoride sorption by nano-
1015
hydroxyapatite/chitin composite, J. Hazard Mater. 172 (2009) 147-151. 42
Page 42 of 62
1016
[65] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Uptake of fluoride by nano-
1017
hydroxyapatite/chitosan, a bioinorganic composite, Bioresource Technol. 99 (2008)
1018
8226-8230.
1019
[66] W. Ma, F. Ya, M. Han, R. Wang, Characteristics of equilibrium, kinetics studies for adsorption of fluoride on magnetic-chitosan particle, J. Hazard. Mater. 143 (2007) 296-
1021
302.
[67] S.P. Kamble, S. Jagtap, N.K. Labhsetwar, D. Thakare, S. Godfrey, S. Devotta, S.S.
cr
1022
ip t
1020
Rayalu, Defluoridation of drinking water using chitin, chitonsan and lanthanum-
1024
modified chitosan, Chem. Eng. J. 129 (2007) 173-180.
us
1023
[68] M.G. Sujana, H.K. Pradhan, S. Anand, Studies on sorption of some geomaterials for
1026
fluoride removal from aqueous solutions, J. Hazard. Mater. 161 (2009) 120-125.
1027
[69] M. Agarwal, K. Rai, R. Shrivastav, S. Dass, Defluoridation of water using amended
1031 1032 1033
M
ed
1030
[70] M. Srimurali, A. Pragathi, J. Karthikeyan, A study on removal of fluorides from drinking water by adsorption onto low-cost materials, Environ. Pollut. 99 (1998) 285-289. [71] M. Islam, R. Patel, Thermal activation of basic oxygen furnace slag and evaluation of its
pt
1029
clay, J. Cleaner Production 11 (2003) 439-444.
fluoride removal efficiency, Chem. Eng. J. 169 (2011) 68-77.
Ac ce
1028
an
1025
[72] A. Tor, N. Danaoglu, G. Arslan, Y. Cengeloglu, Removal of fluoride from water by
1034
using granular red mud: Batch and column studies, J. Hazard. Mater. 164 (2009) 271-
1035
278.
1036 1037 1038 1039
[73] V.K. Gupta, I. Ali, V.K. Saini, Defluoridation of wastewaters using waste carbon slurry, Water Res. 41 (2007) 3307-3316. [74] R. Piekos, S. Paslawska, Fluoride uptake characteristics of fly ash, Fluoride 32 (1999) 14-19.
43
Page 43 of 62
1040 1041 1042
[75] Y.D. Lai, J.C. Liu, Fluoride removal from water with spent catalyst, Sep. Sci. Technol. 31 (1996) 2791-2803. [76] Y. Tang, X. Guanc, J. Wang, N. Gaob, M.R. Mcphaild, C.C. Chusuei, Fluoride adsorption onto granular ferric hydroxide: effects of ionic strength, pH, surface loading,
1044
and major co-existing anions, J. Hazard. Mater. 171 (2009) 774-779.
ip t
1043
[77] Y. Zhou, R.J. Haynes, Sorption of heavy metals by inorganic and organic components of
1046
solid wastes: Significance to use of wastes as low-cost adsorbents and immobilising
1047
agents, Crit. Rev. Environ. Sci. Technol. 40 (2011) 909-977.
us
1048
cr
1045
[78] N. Chubar, Physico-chemical treatment of micropollutants: Adsorption and ion exchange, in: J. Virkutyte, R.S. Varma, V. Jegatheesan, (Eds.), Treatment of
1050
Micropollutants in Water and Wastewater, IWA publishing, London, 2010, pp. 165-203. [79] W.H. Zou, R.P. Han, Z.Z. Chen, J. Shi, H.M. Liu, Characterization and properties of
M
1051
an
1049
manganese oxide coated zeolite as adsorbent for removal of copper (II) and lead (II) ions
1053
from solution, J. Chem. Eng. Data 51 (2006) 534-541.
1056 1057 1058 1059 1060 1061 1062 1063
the accumulation of arsenic in lake sediments, Water Res. 9 (1985) 1029-1032.
pt
1055
[80] T. Takamatsu, M. Kawashima, M. Koyama, The role of Mn2+ rich manganese oxide in
[81] E.N. Peleka, E.A. Deliyanni, Adsorptive removal of phosphates from aqueous solutions,
Ac ce
1054
ed
1052
Desalination 245 (2009) 357-371. [82] K. Goh, T. Lim, Z. Dong, Application of layered double hydroxides for removal of oxyanions: A review, Water Res. 42 (2008) 1343-1368. [83] K. Vaaramma, J. Lehto, Removal of metals and anions from drinking water by ion exchange, Desalination 155 (2003) 157-170. [84] S. Wang, Y. Peng, Natural zeolites as effective adsorbents in water and wastewater treatment, Chem. Eng. J. 156 (2010) 11-24.
44
Page 44 of 62
1064 1065 1066 1067
[85] C. Diaz-Nava, M.T. Olguin, M. Solache-Rios, Water defluoridation by Mexican heulandite-clinoptilolite, Sep. Sci. Technol. 37 (2002) 3109-3128. [86] P. Chingombe, B. Saha, R.J. Wakeman, Surface modification and characterisation of a coal-based activated carbon, Carbon 43 (2005) 3132-3143. [87] C.Y. Yin, M.K. Aroua, W.M.A.W. Daud, Review of modifications of activated carbon
1069
for enhancing contaminant uptakes from aqueous solutions, Sep. Purif. Technol. 52
1070
(2007) 403-415.
1074 1075
cr
us
1073
Purif. Technol. 70 (2010) 329-337.
[89] A. Bhatnagar, M. Sillanpää, A review of emerging adsorbents for nitrate removal from water. Chem. Eng. J. 168 (2011) 493-504.
an
1072
[88] R. Mahmudov, C.P. Huang, Perchlorate removal by activated carbon adsorption. Sep.
[90] C.S. Sundaram, N. Viswanathan, S. Meenakshi, Defluoridation chemistry of synthetic
M
1071
ip t
1068
hydroxyapatite at nano scale: equilibrium and kinetic studies, J. Hazard Mater. 155
1077
(2008) 206-215.
ed
1076
[91] A. Bhatnagar, V.J.R. Vilar, C.M.S. Botelho, R.A.R. Boaventura, A review of the use of
1079
red mud as adsorbent for the removal of toxic pollutants from water and wastewater,
1080
Environ. Technol. 32 (2011) 231-249.
1082 1083 1084
Ac ce
1081
pt
1078
[92] D. Mohan, C.U. Pittman Jr, Arsenic removal from water/wastewater using adsorbents-A critical review. J. Hazard. Mater. 142 (2007) 1-53. [93] H. Genc-Fuhrman, H. Bregnhoj, D. McConchie, Arsenate removal from water using sand-red mud columns, Water Res. 39 (2005) 2944-2954.
1085
[94] C. Zhu, Z. Luan, Y. Wang, X. Shan, Removal of cadmium from aqueous solutions by
1086
adsorption on granular red mud (GRM), Sep. Purif. Technol. 57 (2007) 161-169.
45
Page 45 of 62
1087
[95] I.A.M. Yunusa, P. Loganathan, S.P. Nissanka, V. Manoharan, M.D. Burchett, C.G.
1088
Skilbert, D. Eamus, Application of coal fly ash in agriculture: a strategic perspective,
1089
Crit. Rev. Environ. Sci. Technol.42 (2012) 559-600. [96] W. Nigussie, F. Zewge, B.S. Chandravanshi, Removal of excess fluoride from water
1091
using waste residue from alum manufacturing process, J. Hazard. Mater. 147 (2007)
1092
954-963.
ip t
1090
[97] R.H. Masue, T.A. Loeppert, T.A. Kramer, Arsenate arsenite adsorption and desorption
1094
behaviour on coprecipitated aluminium:iron hydroxides, Environ. Sci. Technol. 41
1095
(2007) 837-842.
us
cr
1093
[98] L. Yan, Y. Xu, H. Yu, X. Xin, Q. Wei, B. Du, Adsorption of phosphate from aqueous
1097
solution by hydroxyl-aluminum, hydroxyl-iron and hydroxyl-iron-aluminium pillared
1098
bentonites, J. Hazard. Mater. 179 (2010) 244-250.
1102
M
ed
1101
1989.
[100] N.Y. Mezenner, A. Bensmaili, Kinetics and thermodynamic study of phosphate adsorption on iron hydroxide-eggshell waste, Chem. Eng. J. 147 (2009) 87-96.
pt
1100
[99] S.D. Faust, O.M. Aly, Adsorption Processes for Water Treatment, Butterworths, Boston,
Ac ce
1099
an
1096
46
Page 46 of 62
ip t cr
an
Initial (I), pH; Adsorption (Ads) capacity (mg/g) and Temperat- Equilibrium other results. Maximum (max) o (E) ure ( C) concentration (mg/L); Adsorbent concentration (g/L); column ht, height; d, diameter; FL, flow rate Metal oxides and hydroxides 3-12; I, 1-100; Max ads pH 3-8, then decreased with 10, 25 2 increased pH to 12; Langmuir ads max (pH 6-7) 10oC, 3.68; 25oC, 5.97
Best kinetic model to fit data;
Best equilibrium model to fit data
Reference
Pseudo-first order, Bangham model Pseudo-first, pseudosecond, diffusion
Langmuir
[7]
Langmuir
[29]
Pseudo-first, pseudosecond, diffusion
Langmuir
[36]
Equilibrium (equil)
M
Adsorption method: batch (B), column (C); Water type: synthetic (S), wastewater (W)
B; S
Manganese oxide coated alumina
B, C; S
Manganese oxide coated alumina
B, C; S
B, 3-12; 25 C, 5.2; 25
Ac c
Granular ferric hydroxide
ep te
d
Adsorbent
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents
B, 3-12; 30 C, 7; 30
B, I, 6.242.1; C, I 5; 140; 50cm ht, 2.4 cm d, FL 2.39 m3/m2h B, I, 2.5-30 C, 3.56; 0.5x0.028m d, 0.3mht, FL 2.19 m3/m2h
B, Max ads pH 4-6, decreased with increased pH from 6 to 12. Langmuir ads max 7.1 at pH 5.2 C, breakthrough point 669 bed volume B, max ads pH 4-7, lowest at pH 12; Langmuir ads max 2.85 at pH 7 C, Bed saturation F concentration 1.25 g/L 47
Page 47 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
C, 3.56; 0.5x0.028m d, 0.3mht, FL 2.19 m3/m2h B, I, 2.5-14; 4-40 C, 5; 200 550 mm ht, 51mm d, FL 20,30mL/min -; 12.5
Pseudo-first, pseudosecond, diffusion
Langmuir
[36]
Surface adsorption, diffusion
Langmuir and Freundlich
[47]
Max ads pH 5-7 Langmuir ads max 16.3 at pH 6
Pseudosecond order
[49]
First 5 min most F ads; 40 min equil -
Langmuir and Freundlich Langmuir -
[51]
Langmuir
[52]
us
B, max ads pH 4-7, lowest at pH 12; Langmuir ads max 1.08 at pH 7
B, I, 2.5-30
C, Bed saturation F concentration 0.47 g/L max ads pH 7 Langmuir ads max 2.41 at pH 7
B; S
Rare earth oxides
B; S
3-11; 29
E 1-30; 1-8
Max ads pH 6-6.5 Langmuir ads max 196 at pH 6.5
Hydroxyapatite (HA)
B; S
I, 19-19000; 10
Hydroxyapatite (HA)
B; S
-; Room temp 6; Room temp
At initial F concentration of ≤190 mg/L, 60-80% F removed by porous HA and 30-35% by crystalline HA >90% F adsorption; Langmuir ads max 4.54 compared to fluorspar 1.79, activated quartz 1.16, calcite 0.39, quartz 0.19
M
Activated alumina
B, 4-10; Room temp C, 7; room temp 4-11; 30
96% removal
d
ep te
B, C; S
B, 3-12; 30 C, 7; 30
Ac c
Activated alumina
B, C; S
an
Activated alumina
I;2.5 x 10-5-6 x 10-2; 17
Pseudosecond order
[50]
48
Page 48 of 62
ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
7; 25
B; S
5-10; 30
Zn/Al LDH calcined (450oC) Na/Mg/Al LDH calcined (500oC)
B; S
4-10; 30
B; S, well water
S: 5,7,9; Room temp W:8.4,8.5 Room temp
us
B; S
an
Mg/Al LDH calcined and uncalcined Mg/Al LDH calcined (200-800oC)
B, I, 10.4-104 Ads max pH 4.7-7.3, Langmuir ads max 0.1 3.8 at pH 7 C, I, 72; 4; 9 cm ht, 1 cm Column ads capacity 3.4 d, FL 0.5 mL/min Layered double hydroxides (LDH) E, 0-70; Calcination increased F ads (130oC 2.5 optimum). Ads capacity 35 at equilibrium F concentration 70 mg/L E, 0-250; Max ads pH 6.0; Langmuir ads max 213 1 at pH 6. Ads highest for calcination temp 500oC. Ads Mg/Al LDH>Ni/Al, Zn/Al LDHs, Mg/Al molar ratio 2:1 best E, 2-60; Max ads pH 6; Langmuir ads max 17 at 1 pH 6. Max ads 13.43 for adsorbent dose 0.2g/L, F concentration 10 mg/L, pH 6 S: I, 5; Distribution coefficient between solid 10 and solution, S: pH 5, 3501; pH 7, W: I, 5.9, 6.9; 1047; pH 9, 1155 10 W: 414-437
M
B 4-9; 20 C 7.5; 20
Ac c
Pseudo-firast order
Langmuir
[43]
adsorption reached max at 15 min -
-
[53]
Freundlich
[54]
ads reached equil in 4 h
Langmuir
[55]
80-97% F removal in 1h
-
[56]
d
B; W C; W
ep te
La impregnated silica gel
49
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ip t cr
Chelating resin (CR), anion exchange resin (AER) Metals loaded Amberlite resin
B; S, field water
3-11; 30
B, C; Field water
B: 1-8; 30 C: 6.0; 30
Al-Amberlite resin
B, C; S
B: E, 2-13; 10 C: I, 16; 2; 0.8 mm d, 50 mm ht, FL,1 mL/min I, 2-10; 20
ep te
B: E, 0-60; 1.6 C: I, 7.9; 2; 8 mm d, FL 6 mL/h -; B: 4-9.1; 30 C: 5.5, 6.7 C: I, 40; 30 20-50; 2cm d, 16cm ht, FL 280700 mL/h
Ac c
an
B: 1-10; 25 C: 7; 25
M
B, C; S
Ion exchange resins and fibres B: ads max at pH 7 for TUA and pH 910 for amberlite (A). At pH 7, ads capacity of TUA three times that of A. Langmuir max for TUA at pH 7, 61. C: column ads capacity 50 pH had no significant effect on ads; Langmuir ads max at pH 7: CR 1.3, AER 1.5. At low F concentration CR removed 30% more F than AER. Field water: CR had higher F removal B: Langmuir ads max at pH 7, resin with La, Ce, Al, 25; Fe, 49; Y, 19. pH for ads max: Ce 4-7, Fe 3, Al 5-9, L 3-7 C: column ads capacity for La resin at pH 6, 20. Bed saturation 300 bed vol. B: max ads at pH 4-7, low ads pH 9, Langmuir ads max 4.6 at pH 4 C: column ads capacity at pH 5.5 and 6.7 were 1.13 and 0.72 (FL 460 mL/h)
d
Thio-urea modified Amberlite (TUA)
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
Ads reached equil in 30 min
Langmuir
[34]
Pseudosecond order
Langmuir
[30]
Ads equil time: La 2 h; Ce 1 h; Fe, Al, Y > 5 h
Langmuir
[39]
Pseudosecond order
Langmuir
[57]
50
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ip t cr
B: 2-6; 25 C: -; 25
B: E, 0-60; 4 C: I, 10; 1 1.5cm d, 5cm ht, FL 2.5-3.5 mL/min E, 0-100; 5
B, Tap water
2-9; 30
Al, La, Zr loaded natural zeolites (Z) Al loaded four synthetic zeolites Al, La loaded synthetic zeolites
B; S, tap water (T)
-; 30
B; S
2-11; 25
S: I, 1-20; 2 T: I, 2.9; 6 I, 5-80; 0.05-0.2
B; S, field water (F)
S: 3.5-9; 20-40 F: 7.4; 30
S: I, 10-80; 2 F: I, 3.3, 4; 1-4
ads max at pH 2.6-6.9. Ads capacity 30 for 5 g adsorbent at E = 100 mg/L Zeolites S: Langmuir ads max : Zr (Z) 3.4-4.1, La(Z) 2.4-2.6, Al(Z) 2.0-2.4 T: F removal(%): Zr(Z) 91.1, La(Z) 90.4, Al(Z) 89.7 Ads max pH 4-8; F ads at equilibrium concentration 40 mg F/L and pH 4-6 was 6-16 S: ads max at pH 6-9; Langmuir ads max 20oC, Al(Z) 34, La(Z) 45.At 40 mg/L equilibrium conc., ads capacity 16 vs 8 for Al(Z) vs La(Z). (F): F conc reduction more by Al(Z) than La(Z)
ep te
d
Al-chelating porous anion exchanger
Ac c
B: ads max at pH 3.0 and decreased with increased pH to 6; Freundlich adsorption constant PO43- > AsO43- > FC: column ads capacity for PO43, AsO43, F- 156, 96, 45, respectively
an
B, C; S
M
Ion exchange fibre
us
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d) Pseudosecond order
Freundlich
[31]
1 h equil time (95% adsorption)
-
[58]
-
Langmuir
[23]
Elovich
RedlichPeterson
[41]
-
RedlichPeterson
[40]
Equil reached in 5 min
51
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ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
4.6-9.2; 25
E, 0-20
B; S
3-11; 25
E, 0-16
B; S
3-11; 25
-
B; S
2-12; -
I, 4-30; -
B; S
us
B; S
ep te
3-7; 25
Carbon materials Ads max at pH 2, then decreased with pH. Langmuir ads max at 25oC, 15.9. Ads decreased with temperature increase Ads capacity of bone char increased with pH. Freundlich adsorption constant highest for bone char and lowest for carbon black Ads max at pH 7; ads capacity at equilibrium concentration 10 mg F/L for activated carbon, γ Al2O3, and ACNT: 0.32, 3.7, and 4.1, respectively Ads max at pH 6-9; ads max of 9.6 at pH 6 and calcined temp 450oC for 0.5 mg/L adsorbent and initial concentration 6 mg F/L Ads decreased from pH 2 to pH 12; Fish bone charcoal had the highest F ads
an
E, 5-20; -
M
2-10; 25, 45, 55
d
B; S
Ac c
KMnO4 modified activated carbon Bone char, activated carbon, carbon black Aligned carbon nanotubes (ACNT) Alumina loaded CNT (calcined, 250-1050oC) Wood, animal , fish bone activated charcoal Al impregnated activated carbon
I, 0.5-15 E, 0-6.5; -
Ads decreased from pH 3 to 7. Calcining at 300oC gave the highest ads among 300-1000oC. Langmuir ads max 1.07, plain carbon 0.49
Pseudosecond order
LangmuirFreundlich
[59]
-
Freundlich
[60]
-
Freundlich
[12]
Equil reached at 20 h
Freundlich
[61]
Most of F removed at 2 h
Langmuir, Freundlich
[62]
-
Langmuir, Freundlich
[42]
52
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ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
Chitin/nanohydroxyapatite composite Chitosan/nano hydroxyapatite composite
B; S
3-11; 30
I, 6-12; 2
B; S
3-11; 30
I, 9-15; 2
Magnetic (Fe)-chitosan
B; S
S: 5-9; 30
I, 5-140; 1
La loaded chitosan
B; S, field water (F)
S: 5-9; 30 F: 7; 30
S: E, 0-15; F: I, 10.2; -
Natural chitosan Zr loaded collogen
B; S B; S
2-10; 20 3.5-11; 30
-
Natural materials Ads max at pH 3, decreased to pH 11. At pH 7, 10 mg F/L, 2 g adsorbent/L, composite ads 1.3; LDH alone 1.0, chitosan alone 0.05; Langmuir ads max 1.9 for composite Ads max at pH 3, decreased with increased pH to pH 11; Langmuir ads max 8.4 at pH 7 Ads max at pH 3, decreased with increased pH to pH 11; At pH 7, 10 mg F/L, adsorbent 5 mg/L, composite ads capacity 1.56, hydroxyapatite 1.30, chitosan 0.05; Langmuir ads max 2.04 pH no significant effect. Ads higher than activated alumina. Langmuir twosite max ads 24
us
I, 9-15; 2
an
3-11; 30
M
B; S
Ac c
ep te
d
Chitosan/ LDH composite
I, 19-95; 1
S: Ads max pH 6.7; max ads 5.5 at pH 6.7 and equilibrium conc.15 mg F/L F: ads capacity 1 compared to 2 for distilled water at equilibrium conc. 8 mg F/L Max ads at pH 6; Langmuir ads max 1.39 at pH 6 Max ads pH 5-8, drastic decrease from pH 9 onwards. Langmuir ads max 2.18 at pH 5-8
PseudoLangmuir second order, diffusion
[63]
PseudoLangmuir second order, Freundlich diffusion PseudoLangmuir second order, diffusion
[64]
Langmuir Equil at 90 min; Pseudo- (one and second order two sites)’ Bradley Pseudo-first order, diffusion
[66]
Equil at 5 min Equil at 500 min
[22]
Langmuir, Freundlich Langmuir
[65]
[67]
[13]
53
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ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
B; S, field water
Al, Fe, Ca loaded silty clay (C)
B; S
S: E, 5-40; 2 F: I, 10.25; 14 I, 10; 100
Soil with high Fe oxide (calcined 500oC)
B, C; S
S: 2.5-10; 32 F: 7.75; 32 Al, Ca-C, C alone: 7-8; Fe-C 2.7-3.0; 30 B: 7; 25 C: 7; 25
Volcanic ash soil
B; S
2.5-9.5; 32
I, 5; 1-8
Increased pH, decreased ads, Be: 46 to 29%, Ka: 38 to 5% for 2g/L adsorbent
ep te
B: max ads 1 at equilibrium concentration 60 mg/L
-; 18
B:I, 4.8-95; 33 C: I, 4; 47; 50 mL column, FL 0.2 mL/min I, 0-50; 40
Ac c
Bentonite B; (Be), S kaolinite (Ka)
pH 2-5, > 90% ads, then decreased with increased pH to pH 12 (17%); Langmuir ads max 21.3 at pH 5-7 S: ads max at pH 5-6 and 3-5; Langmuir ads max 12-15 at pH 5 F: 3 repeated stages of ads reduced F concentration to <0.5 mg/L % F removal: C alone 25, Al-C 78-94’ Ca-C 32-52, Fe-C 94-98
us
E: 0-50; 4
an
2-12; 29
M
B; S
d
La loaded cross-linked gelatin Laterites with Ni
C: F concentration reached 1 mg/L after 120 pore volumes. Column ads capacity 0.15 At equilibrium concentration of 19 mg/L, ads capacity was 2.9 Langmuir ads max 5.5
Equil at 40 min; pseudofirst order Pseudo-first order, diffusion
Langmuir
[44]
Langmuir
[68]
-
-
[69]
-
Freundlich
[38]
Equil in ~2h Langmuir at low conc, >24 h at high conc -
[21]
[70]
54
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ip t cr
Table 1. Characteristics of adsorptive removal of fluoride from water by different adsorbents (cont.d)
us
B; S C; S
B:2.5-7.3; 25 C: 4.7; 25
B; S
1-10; -
B: I, 5-150; 2.5 C: I, 5;10, 0.635 cm2 area,15cm ht, FL 2 mL/min I, 100-1000; 1-8.4
B; S C; W
B:2- 11; 25 C: -; 25
Fly ash (class F, 9.1% CaO)
C; S
Fly ash (2.22 % CaO) Alum sludge (calcined and uncalcined)
B; S B; S
10.1; 20
2-9.5; 30 3.5-8.8; 32
Industrial by-products Ads increased from 21 to 93% as pH increased from 2 to 7, then nearly constant; Langmuir ads max at 25oC, 4.6 and 45oC, 8.1 B: ads max at pH 4.7; Langmuir ads max 8.92 at pH 4.7 C: column total ads capacity 2.05 (0.64 by batch trial for initial F, 5 mg/L
an
I, 1-50; 5
M
2-10; 25
Ads max at pH 5.5; ads capacity A 4.8, UA 1.0 at equilibrium concentration 20 mg/L; Langmuir ads max A 6.3, UA 3.1at pH 5.5 B: ads max at pH 7.58; Langmuir ads max 4.3 at pH 7.5
d
ep te
Activated red mud (A), unactivated (UA) Waste carbon slurry 450oC (activated)
B; S
Ac c
Basic oxygen furnace slag (BOFS) heated1000oC Granular red mud
I, 1-11;1 C: I, 11; 0.5; 0.9 cm2 area, 3.1 cm ht, FL 1.5 mL/min I, 0-100; 450; 40 cm ht, 4.5 cm d, FL 2 mL/h E, 0-3; 20 E, 0-15; 0.5-16
C: breakthrough column ads capacity 4.16 F concentration in effluent reached 0 mg/L after 120-168 h Ads max at pH 6.5; Langmuir ads max 20 at pH 6.5 Ads max at pH 6; calcined higher ads capacity than uncalcined; Langmuir ads max 5.39 at pH 6
Equil at 35 min; pseudofirst order
Langmuir
[71]
Equil at 6 h; pseudosecond order
Freundlich, RedlichPeterson
[72]
Equil at 2 h
Langmuir
[4]
Equil at 1 h; Pseudo-first order
RedlichPeterson
[73]
-
-
[74]
pseudo-first order pseudo-first order; diffusion
Langmuir
[46]
-
[48]
55
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ip t cr
E, 0-40; 10
Ads max (1.6 for initial F 20 mg/L, pseudo-first adsorbent 10 g/L) at pH 2, then continue order; 70% to decrease. Max ads 28 ads in 10min
-
[75]
d
M
an
us
2-9.5; 25
ep te
B; S
Ac c
Spent catalyst
56
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ΔH0 (kJ/mol)
ΔS0 (J/mol/0C)
Reference
22.7
8.4
[71]
8.23
0.1
[63]
Granular ferric hydroxide Geomaterials
-39.2 25 -5.8 to -15.9 32 to 62
7.34
15.8
[7]
-0.1 to -30
-3 to -10
Nanohydroxyapatite/ chitin composite Nanohydroxyapatite/ chitosan composite
-7.1 30
-7.0 40
-6.8 50
11.8
-3.8 30
-3.4 40
-3.4 50
8.6
Modified activated carbon Waste carbon slurry
-3.6 -0.9 25 45 -25.4 -26.5 25 35
Chelating ion exchange resin
-5.7 30
Calcined Zn/Al LDH
-25.0 -25.6 30 40
Fly ash
-1.3 30
-1.6 40
-26.3 50
-2.0 50
us
[64] [65]
-48.6
-15.1
[59]
7.3
11
[73]
-2.4
4.2
[30]
-5.7
6.4
[55]
6.5-6.9
17.1
[46]
an
-5.6 50
1.5
[68]
1.6
M
0.9 55 -27.6 45
ed
-5.5 40
Ac ce
Basic oxygen furnace slag LDH/chitosan composite
pt
Adsorbent
ip t
ΔG0 (kJ/mol) ----------------------Temperature (oC) -0.38 -0.72 -1.94 25 35 45 -6.81 -6.74 -6.72 30 40 50
cr
Table 2. Thermodynamic parameters for fluoride adsorption by different adsorbents
57
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Table 3. Fluoride desorption and adsorbent regeneration Desorption/regeneration reagent (BV, bed volume; d and ht, column internal diameter and height; FL, flow rate
Results
Reference
Amberlite modified with thio-urea (ATU) (B) and (C)
(B) 100 g F-loaded ATU shaken 40 min with 5 mL HNO3, H2SO4, HCl, NaOH (0.01, 0.001 M), 25oC. (C) 2 g ATU, 0.8 mm d, 50 mm ht, FL 1 mL/min, 25oC
(B) maximum desorption [34] (99.9%) with 0.01 M HCl. Next best 0.01 M H2SO4 (77.6%). (C) maximum recovery 92.3% (0.01 M HCl), 35 mL volume used
Manganese oxide coated alumina (MOCA) (B)
0.25 g MOCA adsorbed with F, 50 85% desorption at pH 12. Little mL water with pH adjusted to 3.0- F desorbed at pH < 10 13.5, shaken 24 h, 25oC
Granular red mud (GRM) (C)
10 g GRM, 0.635 cm2 cross sectional area, 10 cm ht, 10 mL 0.2 M NaOH, FL 1 ml/min, 25oC
MOCA and activated alumina (AA) (C)
0.028 m d, 0.1, 0.2 and 0.3 m ht, 2.5% NaOH. 3 cycles of regeneration/adsorption
Al treated zeolite (B)
0.1 g F loaded adsorbent, 50 mL acetic acid, water, NaHCO3, NaOH shaken for 1 d. 4 desorption steps. 25oC
AA (C)
50.8 mm d, 550 mm ht, repeated cycles of regeneration/reactivation
ip t
Adsorbent (batch (B), column (C) method used)
us
cr
[29]
[72]
94.6%F desorbed for MOCA, 91%F for AA in Ist cycle. 2nd cycle no reduction in adsorption, 3rd cycle 5% reduction
[36]
NaHCO3 most effective (63%desorption), Acetic acid least effective (5%). 4 zeolites with NaHCO3 desorption, 1st step 61-67%, total 75-95%
[41]
Loss of 2% F uptake capacity after 5 cycles of regeneration
[47]
Ac ce
pt
ed
M
an
Adsorption (mg/g) decreased from 2.05 to 0.82 in 4 regeneration/adsorption cycles. %desorption from 87 to 46
Zr loaded collagen fibre (B)
0.1 g, 100 mL water with pH adjusted with NaOH, HNO3, 0.5 h shaking, 30oC
pH < 9 very little desorption, pH [13] 11.5 97% desorption
La loaded gelatin (B)
1st wash 1 M NaOH, then water wash or acid wash with 1:1 HNO3 to neutral pH
After adsorption/regeneration process for 3 times, maximum adsorption capacity reduced from 98.5 to 82.3%
[44]
Calcined Zn/Al LDH (B)
NaOH (0.001-0.04 M), 1 g adsorbent treated with maximum F/L of NaOH for 6 h, 30oC.
F desorption increased from 5.12 to 8.55 mg/g (100%). Desorption increased with NaOH concentration
[55]
Ion exchange fibre (B), (C)
(B) 0.1 g adsorbent F loaded, mixed with 25 mL 0.1 M NaCl, NaOH, HCl for 2 h, 25oC. (C) 1.5 cm d, 5
(B) Desorption 20% for HCl, 60% for NaCl, 80% for NaOH. (C) 100% desorbed with 5
[31]
58
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mLNaOH
Rare earth oxides (B)
0.2 g adsorbent, 100 mL water, pH adjusted, shaken for 30 min. 29oC
pH<6, F desorbed to ~ 0. pH ~ 12 >95%. Regeneration decreased adsorption 98 to 91%
[50]
Activated alum sludge (B)
F loaded sludge 4g/L, water pH adjusted, 4 h shaking, 30oC
pH 2-7 desorption ~ 0. > pH 7 desorption increased, highest at pH 12
[48]
La loaded silica gel (C)
Adsorbent loaded with F eluted with dilute NaOH (pH 8.5). 1 cm d, 9 cm h, FL 0.5 mL/min, 20oC.
Column was regenerated
[43])
Ac ce
pt
ed
M
an
us
cr
ip t
cm ht, eluted with 0.5 M NaOH, FL 1 mL/min, 25oC
59
Page 59 of 62
List of figures
ip t
Fig. 1. Common technologies for defluoridation of drinking water [8,22,23]
Ac ce
pt
ed
M
an
us
cr
Fig. 2. Mechanisms of F adsorption (, Adsorbent; Me, multivalent metallic cation)
60
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Adsorption/Ion Exchange
! ! Most widely used ! ! Medium cost ! ! Low effectiveness; cannot remove F below 5 mg/L because of high solubility product of CaF2; need secondary treatment ! ! Large amounts of chemicals required ! ! Precise control of chemicals additions (frequent testing of feed and treated water) ! ! Costs of chemicals, chemical storage and feeding system ! ! Large volumes of waste sludge; disposal problem ! ! Acid neutralisation of treated water required ! ! Toxic chemicals left in treated water (AlF complexes, SO4)
! ! Most widely used ! ! Can be costly (especially ionexchange resins), but can use low-cost adsorbent (including certain waste materials) ! ! Effective even at low F concentration ! ! Simplicity and flexibility of design ! ! Ease of operation ! ! No waste production ! ! Low selectivity against some anions for adsorption/all anions for ion exchange competing ions ! ! Frequent adsorbent regeneration or replacement required ! ! Granular adsorbent better for good hydraulic flow ! ! Effective, mostly at pH < 7 for adsorption
cr
us
an M
ed
Reverse osmosis
ip t
Precipitation/coagulation
Ac ce
pt
! ! Excellent removal ! ! Very high capital cost. Very high operational (energy) cost ! ! No chemicals required ! ! No waste production ! ! No ion selectively; beneficial nutrients and other contaminants removed together with F ! ! Some membranes pH sensitive ! ! F concentrated residue disposal problem ! ! Water wasted ! ! Clogging, scaling and fouling problems
Electrodialysis
! ! Excellent removal ! ! High capital cost. High operational (energy) cost ! ! No chemicals required ! ! No waste production ! ! No ion selectively; beneficial nutrients and other contaminants removed together with F ! ! Skilled labour required ! ! Polarization problem
Fig. 1.
61
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(b)
Hydrogen Bonding
ip t
Ion Exchange
an
us
cr
(a)
Ligand Exchange
(d)
Adsorbent surface chemical modification
Ac ce
pt
ed
M
(c)
Fig 2.
62
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