CHAPTER 1
Introduction
1.1 Background
Fluorosis caused by high fluoride intake predominantly through drinking water containing F- concentrations > 1 mg.L-1, is a chronic disease manifested by mottling of teeth (dental fluorosis) in mild cases and changes in bone structure (skeletal fluorosis), ossification of tendons and ligaments, and neurological damage in severe cases (Wang and Reardon, 2001; Ghorai and Pant, 2002). Today increasing concern is being expressed that these adverse effects of fluorosis are irreversible.
With the rural population of South Africa in many areas currently using drinking water with high F- concentrations supplied from wells and boreholes, the development of a long term solution for the defluoridation of F- contaminated groundwaters is of critical importance. This would require appropriate water treatment procedures. Appropriate technology must be technically simple, cost-effective, easily transferable, use local resources and must be accessible to the rural community.
The removal of fluoride from water using defluoridation techniques is a common practice world-wide, both in industry and domestically. Current methods of fluoride removal from water include adsorption onto activated alumina, bone char and clay, precipitation with lime, dolomite and aluminium sulfate, the Nalgonda technique (Srimurali et al., 1998), ion exchange (Mohan Rao and Bhaskaran, 1988), and membrane processes such as reverse osmosis, electrodialysis and nanofiltration. These processes are discussed in more detail in Section 2.1.
1.2 The origin and distribution of fluoride in groundwaters
1.2.1 International
High F- concentrations in groundwater are found in many countries around the world, notably the United States of America, Africa, and Asia (Czarnowski et al., 1996; Azbar and Turkman, 2000). The most severe problems associated with high F- waters occur in China (Wang et al., 2002), India (Argawal et al., 2003), Sri Lanka (Disanayake, 1996) and Rift Valley countries in Africa. High fluoride groundwaters have been studied in detail in Africa, in particular Kenya and Tanzania (Moges et al., 1996; Gaciri and Davies, 1999; Chernet et. al., 2002; Mjengera and Mkongo, 2002; Moturi et al., 2002). The abundance of F- in Rift Valley groundwaters is due to the weathering of alkaline volcanic rocks rich in F-. Typical fluoride concentrations of towns in the Rift Valley are between 1 and 33 mg.L-1. High fluoride groundwater is also found in the East Upper Region of Ghana (Apambire et al., 1997). The concentration of fluoride was found to be between 0.11 and 4.60 mg.L-1.
1.2.2 South Africa
High fluoride groundwaters in the Republic of South Africa (see Figure 1.1) have recently been studied. Few of the studies, however, cover the whole of South Africa. Detailed studies were done in Bushveld and Pilanesberg regions (Bond, 1947; Fayazi, 1995; McCaffrey and Willis, 2001). These two areas were seen as areas deserving special investigation because of endemic dental fluorosis.
Fluoride concentrations in rivers flowing through the Kruger National Park were determined by Raubenheimer et al. (1990). All five rivers (Olifants, Luvuhu, Letaba, Sabie and Crocodile rivers) had higher concentrations of F- during the latter part of the dry winter season.
Fayazi (1995) measured the fluoride level in groundwater of the Northern Springbok Flats and attributed high fluoride groundwater to fluoride-bearing minerals, such as fluorite (CaF2). Fluorite is known to occur in the granites of the Bushveld Complex which surround the basins of the Springbok Flats.
Figure 1.1
South African groundwater with F- concentration > 1.5 mg.L-1 (McCaffrey and Willis, 2001)
There is a wide variation in fluoride levels in the natural waters of South Africa. This was evident from an analysis of ca. 3000 boreholes in the Pilanesberg and Western Bushveld area (see Figure 1.2). More than 30% of the boreholes have F- concentrations > 1 mg.L-1. In alkaline waters (pH > 9) F- concentrations up to 30 mg.L-1 were recorded (McCaffrey and Willis, 2001).
0 5 0 1 0 0 1 5 0 2 0 0 2 5 0 3 0 0 3 5 0 4 0 0 4 5 0 5 0 0 0.0 1.0 2 . 0 3.0 4 . 0 5 . 0 6.0 7 . 0 8 . 0 9 . 0 [ F ] / m g . L-1 Frequency Figure 1.2
Distribution of F- concentrations in ca 3000 boreholes in the North West Province.
1.3 Fluorosis
1.3.1 Background
Fluoride has certain physiological properties (Muller et al., 1998; Notcutt and Davies, 1999) of great importance in human health. The role of fluoride in the process of mineralisation of certain tissues is important. At low concentrations fluoride stabilises the skeletal system by increasing the size of apatite crystals and reducing their solubility (Moges et al., 1996). Although beneficial effects can be demonstrated at low concentrations, it has detrimental effects when concentrations exceed the threshold (Li et al., 2001).
The relationship between fluoride and dental caries was first noted in the early part of the 20th century when it was observed that residents of certain areas of U.S.A. developed brown stains on their teeth. In the 1930’s it was observed that the prevalence and severity of this type of mottled enamel was directly related to the amount of fluoride ingested (Dean, 1934).
Endemic fluorosis is known to be global in scope, occurring in all continents and affecting many millions of people. Cases of skeletal fluorosis have been reported all over the world (Hillier et al., 2000). According to a report (Susheela, 2001) from UNICEF, fluorosis is endemic in at least 25 countries across the globe.
The fluorosis problem is most severe in the two most populous countries of the world, China and India (Cao et al., 1997; Fung et al., 1999, Mekonen et al., 2001). Thus for example, in China some 38 million people are reported to suffer from dental fluorosis and 1.7 million from the more severe skeletal fluorosis. In India, 15 states are endemic for fluorosis, with over six million people seriously afflicted (Kailash et al. 1999).
The drinking water standards for fluoride ion stipulated by various authorities are tabulated in Table 1.1. There is uncertainty with regard to the precise concentrations of fluoride in drinking water which are optimal for human health.
Table 1.1: Drinking water standards for fluoride ion prescribed by various authorities (Smet, 1992)
Authority Maximum Fluoride
concentration (mgL-1) WHO (International
Standard)
0.5
US Public Health 0.7-1.2 South African Bureau
of Standards
1.0
1.3.2 Symptoms of fluorosis
1.3.2.1 Skeletal fluorosis
Fluorapatite is an order of magnitute less soluble than hydroxyapatite, the principal mineral constituent of bone. The F- ion aggressively substitutes for the OH- ion, leading to a buildup of F- in bone tissue which eventually lead to skeletal fluorosis.
Early in the development of fluorotic changes in the skeleton, the patient often complains of a vague discomfort in the limbs and the trunk. Pain and stiffness in the back appear next, especially in the lumbar region.
In severe fluorosis, in addition to joint problems, some victims can experience deformation of their bones. The stage at which skeletal fluorosis becomes crippling usually occurs between 30 to 50 years of age in endemic regions. The factors which govern the development of skeletal fluorosis are (a) the prevalence of high levels of fluoride intake, (b) continual exposure to fluoride, (c) strenuous manual labour, (d) poor nutrition and (e) impaired renal function due to disease (Nicolay et al., 1999).
1.3.2.2 Dental fluorosis
Human beings throughout history have suffered from dental fluorosis, but until the 20th century the cause of the condition was unknown. Given the common incidence of high F- groundwaters in the East African Rift Valley (Walvekar and Qureshi, 1982), it is evident that ancient people could have suffered from dental fluorosis.
Mottling of teeth is one of the earliest and most easily recognized symptoms (Choi and Chen, 1979). It is the permanent teeth that are affected the most and
they lose their normal creamy white translucent colour and become rough, opaque and chalky white. Dental fluorosis is a developmental disturbance which increases with time. Therefore primary teeth are less severely affected than the permanent teeth, and those teeth which erupted first (the incisors and first permanent molars) are less affected than those erupting later, the premolars and other permanent molars (Pereira and Moreira, 1999).
The first signs of dental fluorosis (moderate dental fluorosis) are thin white lines running across the entire enamel surface and which can only be seen after drying of the tooth surface (Dean, 1942). With more fluorosis these thin lines become broader, merge and may be clear without the need for drying. At slightly greater severity the tooth surface shows distinct, irregular, opaque or cloudy white areas, caused by increasing porosity of the tooth enamel.
Dental fluorosis has been studied in several parts of South Africa (Carstens et al., 1995; Mothusi, 1995; Louw and Chikte, 1997). The findings by Mothusi (1995) indicated that in North West province, dental fluorosis has resulted in cases of psycological trauma, particularly amongst adolescents. In areas with fluoride levels of drinking water exceeding 3 mg F- L-1, local inhabitants demand that their teeth be extracted and replaced with dentures. Dental fluorosis was also investigated in areas surrounding the Pilanesberg complex (Rudolph et al. 1995). They reported that two of the villages had high F- groundwater. They found that severe dental fluorosis occurred in 28% of the population while 41% had moderate dental fluorosis.
The same results were reported in the study by Du Plessis et al (1995). They found that 39% of school children in the Bloemfontein area had evidence of dental fluorosis. The spatial variations of dental fluorosis in seven villages was studied by Zietsman (1989).
1.4 Ways of solving the problem: defluoridation techniques
The prevention of fluorosis through treatment of drinking water in rural areas is a difficult task because of economical and technological restrictions. Defluoridation of water is the only measure to prevent fluorosis and many different defluoridation methods have been developed (Chaturvedi et al., 1990). However, not many are applicable in the areas where the problems occur. This section gives a brief overview of defluoridation methods (See Chapter 2 for a more detailed discussion).
Defluoridation processes are categorised into four main groups:
Adsorption methods: In these methods sorbents such as bone charcoal, activated alumina, and clay are used in column or batch systems.
Ion exchange methods: These methods require expensive commercial ion exchange resins.
Co-precipitation and contact precipitation methods: These methods coprecipitate F- with for example aluminium sulfate and lime (Nalgonda technique.) or precipitate F- for example with calcium and phosphate compounds.
Membrane processes: These include reverse osmosis, nanofiltration and electrodialysis methods.
Taking into account the realities of the problem as outlined in this Chapter the provision of an affordable and technologically simple solution must obviously lie in empowering the local communities to construct viable defluoridation systems from local and readily available materials.
1.5 Motivation for this study
High fluoride groundwaters have been reported in various parts of South Africa (McCaffrey, 2001) and since groundwater forms a major source of drinking water in rural areas, rural populations are facing a major health problem in these areas.
Fluoride removal using activated alumina and reverse osmosis was proposed as a solution to the problem in South Africa (Schoeman and Steyn, 2000). However, the capital cost for household defluoridation was estimated to be about R 5000 for a 50 L per day unit, using either of the methods. This turned out to be too expensive. There is thus a need to develop low cost methods to remove fluoride from water. The removal of fluoride using locally available clays was studied in many countries where the problem occurs. Adsorption of fluoride onto clay minerals was, however, not studied in South Africa before.
The main objective of this WRC-sponsored research was therefore to study the adsorption of fluoride onto clay sorbents from a mechanistic point of view, and to evaluate South African clays as possible sorbents in defluoridation systems.
In order to achieve this outcome the specific objectives were:
(i) The fundamental study of F- adsorption mechanisms and the derivation of a proper modelling algorithm for fluoride adsorption onto mineral surfaces. This includes the study of the effect of mineralogical and structural characteristics of clays, F- solution chemistry, surface properties of the sorbent, and how this is related to F- adsorption capacity. Chapter 4 discusses adsorption mechanisms and adsorption modelling using the model developed in this study.
(ii) The characterisation of South African clays for their F- adsorption potential and usefulness in defluoridation syste ms. This is the first study of this nature on
(iii) The effect of physical and chemical pre-treatment procedures on enhancing adsorption efficiencies of clays and the mechanistic explanations for these procedures (Chapter 5).
(iv) The development of laboratory scale defluoridation columns to study the efficiency of fluoride removal using different sorbents (Chapter 6).
CHAPTER 2
Literature study
2.1 Defluoridation techniques
The process of removal of fluoride is generally termed defluoridation. A comprehensive search of the literature reveals that fluoride removal techniques fall into three major categories:
Chemical precipitation and coprecipitation Adsorption/Ion exchange
Membrane processes.
The materials studied under each category are tabulated in Table 2.1
Table 2.1: Materials and methods of defluoridation
Method Process/Material Reference
Coprecipitation
(Nalgonda Technique)
Aluminium salts Dahi, 1996
Precipitation Calcium and phosphate compounds
Larsen et al., 1993
Adsorption/Ion exchange
Activated alumina Hao et al., 1986; Schoeman and Steyn, 2000
Fly ash Chaturvedi et al., 1990
Clays Srimurali et al., 1998
Soils Omueti and Jones, 1977
Membrane processes Reverse Osmosis Schoeman and Steyn, 2000 Nanofiltration Lhassani et al., 2001
2.1.1 Precipitation methods
Precipitation methods can be divided into two categories, those based on coprecipitation of adsorbed F- and those based on the precipitation of insoluble fluoride compounds.
2.1.1.1 Methods based on coprecipitation:
Coprecipitation (eg. the Nalgonda Technique) is the process by which aluminium salts (aluminium chloride and aluminium sulphate) is added to F- contaminated drinking waters for treatment (Yang et al., 1999; Yang and Dluhy, 2002). This process is used in three ways.
• A bucket system designed to be used on household scale • Fill and draw plants to be used on community scale
• A waterworks flow system developed for larger communities.
(i) Bucket system
The bucket defluoridation system was first practised for domestic use in Tanzania (Mjengera and Mkongo, 2002). The two chemicals (aluminium chloride and aluminium sulphate) are added simultaneously to the raw water bucket and stirred with a wooden paddle. Lime is added to adjust the pH of water to about 6.7. After addition of the chemicals it is left to settle for about 1 hour.
This process is suitable for a daily routine, where one bucket of water is treated for one day’s water supply of about 20 L. The process produces water with residual F- between 1 and 1.5 mg.L-1 (Dahi et al.; 1996).
(ii) Fill and draw system
This system is also used in Tanzania for the defluoridation of drinking water (Mjengera and Mkongo, 2002). It consists of a cylindrical vessel equipped with a hand operated stirring mechanism. The vessel is filled with raw water and a similar procedure for defluoridation using bucket system is performed. Raw water is pumped onto the tank and the required amounts of alum, lime and bleaching powder are added. The contents are stirred slowly for ten minutes and allowed to settle for two hours. The defluoridated supernatant water is withdrawn and supplied through stand-posts. The settled sludge is discarded.
(iii) Waterworks flow system
A bigger defluoridation system is used for larger communities. This system involves the combined use of alum and lime for the defluoridation process (Dahi, 1996). It consists of several components, namely, reactors, sump well, sludge drying beds, elevated service reservoir, electric room and chemical storehouse. The raw water from the source is pumped to the reaction-cum-sedimentation-tank which is referred to as reactor. A sludge pipe with sluice valve is provided to withdraw the settled sludge once a day.
The Nalgonda technique has been introduced in many countries, e.g. India, Kenya, Senegal and Tanzania. However, the method has a number of disadvantages. These include:
• The treatment efficiency is about 70%, which means the process cannot be used in cases of high fluoride contamination.
• A large dosage of aluminium sulphate, up to 700-1200 mg.L-1 may be needed.
• The adverse health effects of dissolved aluminium species in the treated water.
2.1.1.2 Methods based on F- precipitation with calcium and phosphate compounds:
Many methods of precipitation of fluorides with salts of calcium, aluminium and iron are reported in the literature (Lawler and Williams, 1984; Larsen et al., 1993; Qafas et al., 2002). Precipitation processes are governed by the solubility of a forming salt (Parthasarathy et al., 1986). The most common method of treatment is the precipitation of calcium fluoride using calcium from either lime or calcium chloride.
The fundamental problem that exists using lime arises from the low solubility of the calcium hydroxide. It therefore requires excess of reagent to complete precipitation. The relatively high solubility of the calcium fluoride does not allow a complete removal of F-. An additional difficulty with lime precipitation is the poor settling characteristics of the precipitate.
The lime-based fluoride removal can be improved by using CaCl2-lime mixture. The highly soluble CaCl2 provides more calcium than lime without increasing pH. Fluoride removal by lime and CaCl2-lime costs about the same.
2.1.2 Adsorption methods
Fluoride can be removed by adsorption onto many adsorbent materials. The criteria for selection of suitable sorbents are: cost of the medium and running
costs, ease of operation, adsorption capacity, potential for reuse, number of useful cycles and the possibility of regeneration. Some of the most frequently encountered sorbents a re reviewed in this section.
2.1.2.1 Activated alumina
Activated alumina is a granular form of aluminium oxide (Al2O3) with very high internal surface area, typically in the range of 200-300 m2/g. This high surface area allows the material a very large number of sites where adsorption can occur. It has been widely used for removal of F- from drinking water (Hao et al., 1986; Schoeman and MacLeod, 1987).
The mechanism of F- removal from water is similar to those of a weak base ion exchange resin. Fluoride removal efficiency is excellent (typically > 95%), and is dependent on pH. Fluoride removal capacity is best in the narrow range of pH 5.5 to 6.
Fine (28-48 mesh) particles of activated alumina are typically used for F -removal. The adsorption sites on the activated alumina are also attractive to a number of anions other than F-. The selectivity sequence (Johnston and Heijnen, 2002) of activated alumina in the pH range of 5.5 to 8.5 is:
OH->H2AsO4->Si (OH)3O->HSeO3->F->SO42->CrO42->>HCO3->Cl->NO3->Br->I
-Activated alumina can be regenerated by flushing with a solution of 4% sodium hydroxide which displaces F- from the alumina surface (Schoeman and MacLeod, 1987). This procedure is followed by flushing with acid to re-establish a positive charge on the surface of the alumina.
2.1.2.2 Clays and soils
The first comprehensive study of fluoride adsorption onto minerals and soils was published in 1967 (Bower and Hatcher, 1967). Since the above -mentioned paper was published, several workers studied the adsorption of fluoride. These studies include the use of Ando soils of Kenya ( Zevenbergen et al., 1996), Illinois soils of USA (Omueti and Jones, 1977), Alberta soil (Luther et al., 2996), illite- goethite soils in China (Wang and Reardon, 2001), clay pottery (Chaturvedi et al., 1988; Hauge et al., 1994), fired clay (Bardsen and Bjorvatn, 1995), fired clay chips in Ethiopia (Moges et al., 1996), kaolinite (1997), bentonite and kaolinite (Kau et al., 1998; Srimurali et al., 1998), and fly ash (Chaturvedi et al., 1990).
2.1.2.3 Other sorbents
In addition to activated alumina, clays and soils other materials such as spent bleaching earth, spent catalyst, rare earth oxides, bone charcoal and activated carbon were studied as sorbents for F-. Mahramanlioglu et al (2002) investigated the adsorption of F- using spent bleaching earth. They found that the removal of F- depends on the contact time, pH and adsorbent concentration. Lai and Liu (1996) studied the F- removal from water with spent catalyst. Their findings showed that spent catalyst could be utilized as adsorbent for F- removal. Its adsorption capacity was comparable to that of activated alumina. Raichur and Basu (2001) studied the adsorption of F- onto rare earth oxides. Rare earth oxides showed great potential for F- removal from water. Lu et al (2002) investigated the removal of F- using red mud. The removal of F- using red mud was found to be 82%.
2.1.3 Ion exchange resins
Ion exchange resins are effective in removing F- from water. Mohan Rao and Bhaskaran (1988) studied the removal of F- using ion exchange materials such as sulphonated material from coconut shell, Carbion, Tulsion and Zeocarb 225. From the results, it was evident that Zeocarb 225 has the highest F- removal capacity and sulphonated material of coconut shell has the lowest. It was also indicated that the ion exchange material can be regenerated by aluminium sulphate solution (2-4%). Castel et al (2000) studied the removal of F- by a two way ion exchange cyclic process. This system used two anion exchange columns. The results show that this process can effectively remove fluoride from water.
The use of anion exchange resins for F- removal is not common because of their relatively high costs. The presence of other anions such as chloride and sulphate also presents a major problem when using ion exchange resins for F- removal. Since F- removal is accompanied by sorption of other anions, the sorption capacity is normally lower than 0.5 mg F-.L-1 (Veressinina et al., 2001).
2.1.4Membrane processes
Membrane processes such as reverse osmosis, nanofiltration and electrodialysis are recently developed methods for F- removal from water (Schoeman and Steyn, 2000; Lhassani et al., 2001; Garmes et al., 2002). Not much research has gone underway using these membrane processes. However, the study by Lhassani et al (2001) indicates that F- can be removed using nanofiltration.
2.2 Clay adsorption studies
Bower and Hatcher (1967) indicated that the adsorption of F- onto minerals and soils is accompanied by the release of OH- ions. It was also found that the F -adsorption is concentration dependent and it is described by the Langmuir adsorption isotherm. After this study, many studies on the adsorption of F- using clay minerals and soils were undertaken. Some of the studies involving F -adsorption are discussed below.
2.2.1 Fired clays
Several researchers have studied the removal of F- using fired clays (Hauge et al., 1994; Bardsen and Bjorvatn, 1995; Moges et al., 1996). Hauge et al (1994) studied the defluoridation of drinking water using pottery. The study investigated the effect of temperature on F- adsorption. The results show that clays fired at temperature up to 600°C gave higher F- adsorption. Moges et al (1996) studied the defluoridation of water using fired clay chips in Ethiopia. Their findings indicated that F- adsorption is affected by factors such as initial concentration, mass of adsorbent and the pH of the solution.
2.2.2 Low cost materials
Srimurali et al. (1998) investigated the removal of F- using low cost materials such as kaolinite, bentonite, charfines, lignite and nirmali seeds. Their results show that F- adsorption using nirmali seeds and lignite is low (6 to 8%). The removal of F- by kaolinite is slightly better (18.2%) while charfines and bentonite give higher F- removal capacity of 38 and 46% respectively. Chemical pre-treatment was used to investigate its effect on the removal capacity of these materials. Kau et al. (1998) investigated the adsorption of F- by kaolinite and bentonite. The results show that bentonite was found to have higher F -adsorption than kaolinite. F- removal using China clay was studied Chaturvedi
(Chaturvedi et al., 1988). The results show that low F- concentration, high temperature and acidic pH are factors favouring the adsorption of F-. It was concluded that the alumina constituent of the China clay is responsible for F -adsorption. Chaturvedi et al. (1990) also studied the defluoridation of water by adsorption on fly ash. This study confirms their previous findings that low F -concentration, high temperature and acidic pH favour the adsorption of F-.
2.2.3 Soils
Several studies investigated the adsorption of F- using soils (Omueti and Jones, 1977; Chhabra et al., 1980; Zevenbergen et al., 1996; Bjorvatn et al., 1997; Wang and Reardon, 2001). Chhabra et al. (1980) investigated the effect of varying levels of exchangeable sodium on the adsorption of F- onto sodic soils. It was noted that at equilibrium F- concentration, a decrease in adsorption of F- with increase in the soil exchangeable sodium percentage was observed. Omueti and Jones (1977) studied the adsorption of F- by Illinois soils. They reported that at low concentrations F- adsorption onto soils was described by both Langmuir and Freundlich isotherms. It was also suggested that F- adsorption onto soils was due to the presence of the amorphous aluminium hydroxides. Bjorvatn et al. (1997) studied the defluoridation of water using soil samples from Ethiopia. It was reported that five soil samples from highland areas around Addis Ababa reduced the fluoride content of the water from about 15 to 1 mg.L-1. From this study, it was concluded that the highland soil may be useful for removal of excessive F- from drinking water. Zevenbergen et al. (1996) studied the defluoridation of water using the Ando soil of Kenya. It was concluded that the use of Ando soils appears to be an economical and efficient method for defluoridation of drinking water.
The structure of the clay plays a very important role in determining the charge on the clay surface and type of exchange that can occur with ions in solution. In
charged ions, such as F-. As discussed in detail in Chapter 4, pH plays a dominant role in determining the adsorption capacity as pH modifies the charges on edge positions in phyllosilicates and also those of variably charged minerals such as gibbsite, hematite and goethite. Charges are generally positive under acid conditions and negative in an alkaline environment. The specific pH range for positive and negative surface charge will of course be a function of the pKa values of the metal hydroxides present.
Chemical pre-treatment which includes the use 1% Na2CO3 and 1% HCl was reported to improve the adsorption capacity of clays and soils (Srimurali et al., 1998). In general, it was found that firing and chemical pre-treatment both improve the adsorption capacity of some clays and soils.
Studies on the surface coating of clays and soils were reported (Coleman and Thomas, 1964; Agarwal et al., 2002; Zhuang and Yu, 2002). The coating of clays and soils with Al and Fe hydroxides improves their adsorption capacity.
2.3 South African studies on defluoridation and F
-adsorption
South African studies on defluoridation of drinking water was first started in mid 1980 (Schoeman, 1985). Most of these studies have been concerned with the use of activated alumina for fluoride removal (Schoeman, 1985; Schoeman and Leach, 1986; Schoeman and Steyn, 2000). Schoeman (1985) developed two full scale activated alumina defluoridation plants. These plants reduced the F -content of the underground mine water from 8 to less than 1.5 mg.L-1. This study was followed by a critical evaluation on the performance of the two defluoridation plants (Schoeman and Leach, 1986). In recently published work on this topic, Schoeman and Steyn (2000) evaluated the use of activated alumina and reverse osmosis for the defluoridation of water.
CHAPTER 3
Experimental procedure and methodologies
The procedures described here were developed and used in the study of F -adsorption mechanisms onto mineral substrates (Chapter 4), the characterisation of South African clays for their F- adsorption potential (Chapter 5), and laboratory scale column defluoridator systems discussed in Chapter 6. These methods are discussed in detail in this chapter and will only be referred to in the follo wing chapters.
3.1 Mineralogical characterisation of the clays
3.1.1 Sample preparation
A representative portion of each clay sample was dried overnight at 30°C, after visible pieces of plant material and pebbles were taken out by hand. Each sample was lightly crushed in a swing-disk mill to an arbitrary chosen grain size of <180 µm. Each sample was then dispersed in distilled water using an ultrasonic bath. The clay fraction was separated by using standard centrifugal techniques (USGS, open file report 01-041). Decantation was not an acceptable alternative to centrifugation in this study because normal gravitational methods of particle sedimentation take an inordinate amount of time, and for particles finer than < 0.5 µm Brownian motion interferes with settling (Folk, 1974; Syvitski, 1991). For each sample, three sedimented samples were prepared on glass slides and allowed to dry slowly to produce orientated deposits. One was heat-treated for 30 min at 550ºC, one was left in a desiccator with ethe lene glycol at 40ºC for 24 h, and one was kept in a desiccator for X-raydiffraction.
3.1.2 Identification of clay minerals from X-ray diffraction
Step scans of 4 s per step of 0.05° 2? from 2.50º 2? to 16.00º 2? were done on air-dried, glycolated and heat-treated samples using a Phillips PW 1710 X-ray diffractometer with the following settings:
Tube anode material Cobalt Generator potential 40 kV
Wavelength Ka1 1.78896 Å
Wavelength Ka2 1.79258 Å
Intensity ratio I Ka1/I Ka2 2 Divergence diaphragm 1° Detector diaphragm 0.1 mm
The changes in peak position on swelling and heating were used to confirm mineral identification (Velde, 1995). X-ray powder diffraction analyses were conducted on each sample using side loaded aluminium sample holders to determine the overall mineralogical composition.
3.2 Analytical methods for the determination of F
-3.2.1 Background
Various methods have been reported for the determination of F- in aqueous solutions, such as spectrophotometric eg. the SPADNS-method (Crosby et al., 1968), conductometric (Buffle et al., 1985), complexometric (Pickering, 1986; Saha, 1993), ion chromatographic (IC), and potentiometric or fluoride ion selective electrode (FISE) methods (Van den Hoop et al., 1996; Yuchi et al., 1999; McCaffrey, 2001). From the methods listed above, FISE has been widely used due to its simplicity and short analysis time. Recent developments in ion
chromatographic separation techniques have led to an improvement in the determination of F- in terms of selectivity and sensitivity. Ion chromatography is therefore increasingly used for the determination of various ions and in particular the F- ion. In this study both FISE and IC have been used for the determination of F-.
3.2.2 Reagent and standard solutions
3.2.2.1 Standards
A 1000 mg F-.L-1 sodium fluoride (NaF) stock solution was prepared using deionised water. Standards at a required concentration range were prepared by appropriate dilution of the stock solution.
3.2.2.2 Total ionic strength buffers (TISAB) for FISE determinations
TISAB III (Orion cat No 940911, Thermo Orion) was obtained from Thermo Orion (Beverly, USA). This commercial product consists of a mixture of ammonium chloride (NH4Cl), ammonium acetate (CH3COONH4), CDTA (C14H22N2O8.H2O), and cresol red (C21H18O5S) in water. 2 mL of this buffer was added to the test and standard solutions of 20 mL according to the manufacturer’s recommended procedure for the determination of F-.
TISAB IV was prepared as follows: 84 mL concentrated HCl (36-38%), 242g Tris (hydroxymethyl) aminomethane, and 230g sodium tartrate (Na2C4H4O6•2H2O) were added to 500 mL deionised water, allowed to dissolve, and then made up to 1L in a volumetric flask with deionised water. An equal volume of TISAB IV was added to standards or samples before analysis.
3.2.3 HPIC instrumentation
A Dionex DX-120 ion chromatographic system equipped with a Dionex anion exchange column system (AG 14A + AS14A) was used for the determination of F- concentrations < 0.1 mg.L-1 and also other anions leached from the clay sorbents during adsorption tests. The standard HCO3-/CO32- (1.0 mM HCO3- + 3.5 mM CO32-) eluent recommended for anion determinations with this column was used in all determinations.
3.2.4 FISE instrumentation
Orion F- ion selective electrodes were used for routine determination of F- for concentrations > 0.1 mg.L-1. The fluoride concentration in each of the prepared solutions containing one of the two buffers were measured using a fluoride ion selective electrode (Orion 9609, Orion Research Inc., Beverly, MA) connected to a single junction calomel reference electrode (Orion Model 9609) and pH/ISE meter (Orion Model 520A, Orion Research Inc., Beverly, MA).
3.2.5 The role of total ionic strength buffers in F- determinations
One of the major problems encountered in the analysis of total F- concentrations is the impact of metal ions such as aluminium, iron, magnesium and calcium on the analytical response. Due to binding of F- with these metal ions, only a fraction of the total F- concentration is determined. Elimination of these interferences may be achieved by addition of metal complexing agent. A Total Ionic Strength Adjustment Buffer (TISAB III or TISAB IV) containing complexing agents was therefore added to standards and samples. This buffer ensures:
• that a constant ionic strength is maintained in test solutions and standards. This is important because the F- electrode actually
measures activity and not concentration. The measured result is therefore strongly dependent on the ionic strength of the medium because the activity a, given by the equation below,
a = γF[F-]
will depend on the activity coefficient of the F- ion, γF at the particular ionic strength of the solution.
• that the pH is kept constant at ca. 5.5. This is achieved by an acetic acid buffer in the TISAB III solution and a tartrate buffer in the case of TISAB IV. It is important to maintain the pH of samples and standards above 4 because increasing protonation of F- at pH < 4 could lead to a negative error. The F- sensitive membrane does not respond to F -when it is in its protonated form, HF.
• that F- is released from its metal ion complexes if present. Of particular importance are AlF2+ and AlF2+. TISAB III contains CDTA (cyclohexane diamine tetra acetic acid) that forms a very strong complex with metal ions such as Al3+ replacing the F- in the process. For very high levels of Al3+ and other metals forming complexes with F-, TISAB IV was used. TISAB IV contains hydroxymethyl aminomethane as complexing agent. In the case of low concentration F- determinations a special low level TISAB solution is recommended because TISAB III and IV do not produce accurate results when applied under these conditions. The low level TISAB, however, does not contain a complexing agent such as CDTA and can therefore not compensate for interferences caused by complexation of F- with metals such as Al3+.
FISE offers linear calibration graphs for concentrations larger than 0.1 mg.L-1. Below 0.1 mg.L-1, curvature in the calibration curve increases and it becomes more difficult to compensate for electrode drift and special procedures need to be followed to produce accurate results. These procedures include a 20 min equilibration of the membrane in 0.1 M F- solution prior to calibration and the analysis. Even then the electrode requires more than 10 min to stabilise after each transfer to a new sample or standard. Groundwaters in the study area in most cases have pH values > 7. In alkaline waters the presence of metal cations such as Al3+ that can form complexes with F- is unlikely and the inclusion of CDTA in the TISAB solution is not mandatory. In this study F- determinations were also required in acidic solutions where Al3+ leached from the clay substrates could interfere with the measurement. This would lead to an underestimation in the F- concentration. In this study TISAB III and IV were therefore used as ionic strength adjustment buffers.
3.3 Comparison of the FISE and IC methods
3.3.1 Background
Consistent differences in the determination of F- in natural waters using FISE and HPIC have been reported (McCaffrey, 1994) using the older generation of anion exchange columns and F- membrane electrodes. One of the problems with these columns using the standard HCO3-/CO32- eluent was that the F- peak appeared immediately after the water dip which made the accurate determination of the baseline difficult. The Dionex AG14A + AS14A colum n systems used in this study produced a F- peak well separated from the water dip. To ensure accurate analytical results it was nevertheless decided to perform a limited study to compare the two techniques for F- determination in the test solutions encountered in this study. The aim of this study was to compare the results of F -determination in standard solutions, deionised and tap water, and clay extracts samples as obtained by the two methods
3.3.2 Experimental procedure
The procedure used in this comparative study included the following steps. The two techniques were calibrated using exactly the same set of standards. In case of FISE, however, TISAB IV was added to both samples and standards.
Method:
Spiked samples of F- in deionised water, tap water and solutions obtained from clay extracts were prepared as follows: 25 mL of water was added into a plastic beaker and spiked with 0, 0.1, 0.2, 0.5, 1, 2, 3, 5, 7, 10 mg.L-1 F-. Volumes of water equal to the volume of spiking solution needed for a specific concentration were removed prior to spiking to keep the volume constant. For clay extract preparation, 300 mL of deionised and tap water was added to 10g of clay. The mixture was stirred for 1h, centrifuged for 5 minutes and decanted. 25 mL of extract was added into plastic beakers and spiked in the same way as with deionised and tap water.
3.3.3 Results
3.3.3.1 Comparison of calibration curves
The calibration curves obtained for 0.1, 0.2, 0.5, 1, 2, 5 and 10 mg.L-1 F -standards for the two methods are shown in the Fig. 3.1 and 3.2 below.
Figure 3.1
Fluoride calibration curve using NaF dissolved in MiliQ water and IC.
The calibration curve for ion chromatography shows excellent linearity as indicated by correlation coefficient of 0.9995.
y = -54.317x + 62.173 R2 = 0.9998 0 20 40 60 80 0 0.5 1 1.5 Log [F-] Pot (mV) Figure 3.2
Fluoride calibration curve using NaF dissolved in MilliQ water and FISE.
The calibration curve above shows a sensitivity (response factor) of –54.317 mV per decade and excellent linearity, as indicated by the correlation coefficient of 0.9998. This calibration curve was combined with calibration values for lower F
-concentrations (0.2 and 0.5 mg.L-1 F-) on a graph paper. Determinations of lower F- concentrations were extrapolated from the combined graph.
Table 3.1 shows the results obtained by these two methods and in Fig. 3.3 ISE results using TISAB IV buffer are plotted against F- concentrations obtained by IC. It can be seen that the results are in good agreement. The plot of HPIC versus FISE, which gives a straight line obtained at a slope of 1, indicates that the two methods are comparable. The detection limit for ISE using the linear calibration line section was 0.1 mg.L-1 F- and for IC 0.03 mg.L-1 F-. Because of the higher detection limit for ISE, IC was used for samples where low residual F -concentrations where expected. In general, however, ISE was used because the method is a simple, accurate, quick and economical way to determine F -concentrations.
TABLE 3.1
F- concentrations (mg.L-1) determined by IC and FISE (TISAB IV) ISE IC Known F- conc (mg.L-1) Measured F- (mg.L-1) Measured F- (mg.L-1) 0.2 0.22 0.26 0.5 0.54 0.53 1 0.99 1.24 2 2.10 2.13 5 4.90 5.08 10 10.0 9.93
Figure 3.3
F- concentrations determined by ISE using TISAB IV vs determination by IC
3.3.3.2 Comparison of F- determination in spiked samples of deionised water, tap water and solutions obtained from clay extracts.
(i) Deionised and tap water
Test solutions obtained from spiking deionised water, tap water and aqueous extracts from selected clay adsorbents were analysed in triplicate by both methods. The results of spiking deionised and tap water obtained by the two methods are summarised in Table 3.2. Tap water contained 0.09 mg.L-1 F- and the measured samples were therefore corrected by subtracting this amount.
y = 1.0107x - 0.1041 R2 = 0.9991 0 2 4 6 8 10 0 2 4 6 8 1 0 F-/mg.L-1 determined by ISE F - /mgL -1 determined by IC
TABLE 3.2
F- determined in spiked water samples by IC and ISE using TISAB III and TISAB IV. Concentrations in mg.L-1
ISE TISAB III
Matrix Deionised water Tap water
Spiked F -in mg.L-1 F- in mg.L-1 SD %recovery % error F- in mg.L-1 SD %recovery % error 0.1 0.10 0.001 100.0 0.0 0.09 0.002 90.0 -10.0 0.2 0.20 0.003 100.0 0.0 0.18 0.001 90.0 -10.0 0.5 0.49 0.001 98.0 -2.0 0.49 0.002 98.0 -2.0 1 1.01 0.003 101.0 1.0 0.96 0.003 96.0 -4.0 2 2.01 0.017 100.5 0.5 1.95 0.029 97.5 -2.5 3 2.93 0.016 97.7 -2.3 2.89 0.008 96.3 -3.7 5 5.00 0.014 100.0 0.0 4.92 0.014 98.4 -1.6 7 6.94 0.019 99.1 -0.9 6.81 0.098 97.3 -2.7 10 10.07 0.112 100.7 0.7 9.77 0.028 97.7 -2.3 ISE TISAB IV
Matrix Deionised water Tap water
Spiked F -in mg.L-1 F- in mg.L-1 SD %recovery % error F- in mg.L-1 SD %recovery % error 0.1 0.10 0.001 100 0.0 0.1 0.002 100.0 0.0 0.2 0.20 0.002 100 0.0 0.2 0.002 100.0 0.0 0.5 0.50 0.002 100 0.0 0.5 0.004 100.0 0.0 1 1.01 0.001 101 1.0 0.98 0.001 98.0 -2.0 2 1.98 0.003 99 -1.0 1.99 0.015 99.5 -0.5 3 3.01 0.01 102 0.3 2.96 0.011 98.7 -1.3 5 4.96 0.02 99.2 -0.8 4.95 0.02 99.0 -1.0 7 6.99 0.01 99.9 -0.1 6.96 0.032 99.4 -0.6 10 9.98 0.03 99.9 -0.2 9.98 0.025 99.8 -0.2
ION CHROMATOGRAPHY
Matrix Deionised water Tap water
Spiked F -in mg.L-1 F- in mg.L-1 SD %recovery % error F- in mg.L-1 SD %recovery % error 0.1 0.10 0.002 100.0 0.0 0.09 0.002 90.0 -10.0 0.2 0.21 0.001 105.0 5.0 0.18 0.004 90.0 -10.0 0.5 0.49 0.001 98.0 -2.0 0.53 0.001 106.0 6.0 1 0.99 0.003 99.0 -1.0 0.99 0.002 99.0 -1.0 2 2.03 0.005 101.5 1.5 1.97 0.003 98.5 -1.5 3 3.01 0.002 100.3 0.3 2.96 0.002 98.7 -1.3 5 4.95 0.01 99.0 -1.0 4.97 0.011 99.4 -0.6 7 7.00 0.009 100.0 0.0 7.01 0.02 100.1 0.1 10 10.02 0.021 100.2 0.2 9.95 0.024 99.5 -0.5
It can be seen that the spike recovery values obtained are in the range of 90-100%. Both methods give good and stable fluoride recoveries when using water samples. These results suggest that either TISAB III and TISAB IV were suitable for the accurate determination of fluoride in water samples. Typical chromatograms of water samples spiked with 0.2 mg.F-.L-1 are shown in Figure 3.4 and 3.5 below.
Figure 3.4
A sample chromatogram of spiked deionised water (0.2 mg F-.L-1) determined by ion chromatography. Concentrations (mg.L-1): 1. F- = 0.21, 2. Cl- = 0.02 and 3. SO4
= 0.01.
Figure 3.5
A sample chromatogram of spiked tap water (0.2 mg F-.L-1) determined by ion chromatography. Concentrations (mg.L-1): 1. F- = 0.27, 2. Cl- = 42.56, 3. Br- = 0.02, 4. NO3
= 2.17 and 5. SO4
= 13.54.
In Figures 3.6 and 3.7 below ISE results for spiked deionised and tap water samples using TISAB III and TISAB IV buffers are plotted vs F- concentrations obtained by IC. Both figures show a straight line with correlation coefficient between 0.9999 and 1. The results show a good correlation at F- concentrations
between 0.1 and 10 mg.L-1. Both techniques are therefore capable of analysing correctly in relatively undemanding matrices such as deionised and tap water.
y = 0.9996x - 0.0021 R2 = 0.9999 0 2 4 6 8 10 12 0 2 4 6 8 10 12 F-/mg.L-1 determined by IC F - /mg.L -1 determined by ISE T III T IV Figure 3.6
F- concentrations in spiked tap water samples determined by ISE using TISAB III and IV vs determinations by IC.
The general observation is that application of the ISE using TISAB III and IC procedures leads to lower F- values at low F- concentrations in tap water.
Figure 3.7
F- concentrations in spiked deionised water samples determined by ISE using TISAB III and IV vs determination by IC
y = 0.9971x + 0.0014 R2 = 1 0 2 4 6 8 10 12 0 2 4 6 8 10 12 F-/mg.L-1 determined by IC F - /mg.L -1 determined by ISE T III T IV
(ii) Clay extracts
Extracts were prepared by extracting 10g of clay with 300 mL deionised water and tap water for 1h. Samples were centrifuged for 5 minutes and pH of each sample was measured. Blank values were determined before spiking with F-. Table 3.3 gives a summary of pH of four clay samples used. The clay samples are described in Chapter 5.
TABLE 3.3
pH of clay extracts in deionised and tap water determined after stirring for 1 hour .
Clay samples pH determined for each extract.
Deionised water Tap water
ALU 4.80 5.45
RBM 9.64 8.25
MD2 5.95 7.42
CAL 9.27 8.71
Table 3.3 shows pH values for each sample extract. In case of clay samples (ALU and MD2) the pH of deionised water is lower than that of tap water. For clay samples (RBM and CAL) the pH of deionised water is higher than that of tap water. It is a well known fact that pH plays an important role in the determination of F- in sample matrices.
Accurate determination of F-, however, becomes rather challenging in more complex matrices such as the extracts obtained from adsorption experiments on clay samples. Table 3.4 compares the results obtained by analysing spiked
CAL. The RBM extracts contained 2.58 mg.L-1 F-, MD2 extract contained 0.09 mg.L-1 F-, CAL extract contained 0.55 mg.L-1 F-, ALU extract contained 0.04 mg.L-1 F- and the measured samples were therefore corrected by subtracting this amount.
TABLE 3.4
F- determined in spiked clay extract samples by IC and ISE using TISAB III and TISAB IV. Concentrations in mg.L-1
RBM TISAB III TISAB IV IC
Spiked F -in mg.L-1 F- in mg.L-1 %recovery % error F- in mg.L-1 %recovery % error F- in mg.L-1 %recovery % error 0.1 0.12 120.0 20.0 0.12 120.0 20.0 ND ND ND 0.2 0.3 150.0 50.0 0.22 110.0 10.0 0.74 370 270 0.5 0.49 98.0 -2.0 0.54 108.0 8.0 1.13 226 126 1 0.97 97.0 -3.0 1.10 110.0 10.0 1.47 147 47 2 1.90 95.0 -5.0 2.13 106.5 6.5 2.45 122.5 22.5 3 3.00 100.0 0.0 3.26 108.7 8.7 2.96 98.7 -1.3 5 5.23 104.6 4.6 5.33 106.6 6.6 4.84 96.8 -3.2 7 7.69 109.9 9.9 7.55 107.9 7.9 6.60 94.3 -5.7 10 11.03 110.3 10.3 10.25 102.5 2.5 8.61 86.1 -13.9
ALU TISAB III TISAB IV IC
Spiked F -in mg.L-1 F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error 0.1 0.06 55.0 -45.0 0.09 90.0 -10.0 0.06 60.0 -40.0 0.2 0.08 40.5 -59.5 0.16 80.0 -20.0 0.13 65.0 -35.0 0.5 0.29 58.0 -42.0 0.41 82.0 -18.0 0.35 70.0 -30.0 1 0.60 60.0 -40.0 0.83 83.0 -17.0 0.73 73.0 -27.0 2 1.10 55.0 -45.0 1.70 85.0 -15.0 1.41 70.5 -29.5 3 1.06 35.3 -64.7 2.71 90.3 -9.7 1.93 64.3 -35.7 5 3.84 76.8 -23.2 4.36 87.2 -12.8 4.06 81.2 -18.8
7 5.86 83.7 -16.3 6.30 90.0 -10.0 5.57 79.6 -20.4
10 8.51 85.1 -14.9 8.60 86.0 -14.0 85.6 85.6 -14.4
MD2 TISAB III TISAB IV IC
Spike F -in mg.L-1 F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error 0.1 0.06 60.0 -40.0 0.06 60.0 -40.0 0.05 50.0 -50.0 0.2 0.11 55.0 -45.0 0.11 55.0 -45.0 0.11 55.0 -45.0 0.5 0.23 46.0 -54.0 0.33 66.0 -34.0 0.34 68.0 -32.0 1 0.62 62.0 -38.0 0.72 72.0 -28.0 0.7 70.0 -30.0 2 1.47 73.5 -26.5 1.41 70.5 -29.5 1.59 79.5 -20.5 3 2.40 80.0 -20.0 2.41 80.3 -19.7 2.46 82.0 -18.0 5 4.60 92.0 -8.0 4.41 88.2 -11.8 4.52 90.4 -9.6 7 6.60 94.3 -5.7 6.18 88.3 -11.7 6.35 90.7 -9.3 10 9.90 99.0 -1.0 8.69 86.9 -13.1 8.85 88.5 -11.5
CAL TISAB III TISAB IV IC
Spike F -in mg.L-1 F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error F- in mg.L-1 %recovery %error 0.1 0.10 100.0 0.0 0.08 80.0 -20.0 0.03 30.0 -70.0 0.2 0.13 65.0 -35.0 0.16 80.0 -20.0 0.16 80.0 -20.0 0.5 0.49 98.0 -2.0 0.43 86.0 -14.0 0.40 80.0 -20.0 1 1.00 100.0 0.0 0.62 62.0 -38.0 0.78 78.0 -22.0 2 1.91 95.5 -4.5 1.17 58.5 -41.5 1.30 65.0 -35.0 3 2.88 96.0 -4.0 2.26 75.3 -24.7 2.29 76.3 -23.7 5 5.76 115.2 15.2 4.50 90.0 -10.0 4.76 95.2 -4.8 7 8.02 114.6 14.6 6.00 85.7 -14.3 6.21 88.7 -11.3 10 11.17 111.7 11.7 8.92 89.2 -10.8 8.95 89.5 -10.5
Chromatograms of the extracts were recorded before analysis of the spiked extracts. A sample chromatograph in Figure 3.8 shows the anions present in the
CAL extract. using deionised water. These include anions such as: F-, Ac-, Cl-, NO2-, Br-, NO3-, PO43- and SO42-.
Figure 3.8
Chromatogram for the CAL extract using sodium
carbonate/bicarbonate as eluent and an AS4A anion exchange column. Concentrations (mg.L-1): 2. F- = 0.55, 3. Ac- = 0.04, 4. Cl- = 1.0, 5. NO2 = 0.05, 6. Br- = 0.01, 7. NO3 = 0.03, 8. PO4 = 0.03 and 9. SO4 = 9.63.
A chromatograph for the RBM extract spiked with 0.2 mg F-.L-1 is shown in Figure 3.9. The chromatograph shows excellent resolution where all anions elute away from the water dip and at retention times > 4 min. Unresolved peaks, however, occur between fluoride and acetate. It can be seen that acetate peak elutes next to fluoride, hence the fluoride concentration is not easily determined. Chloride, nitrate and sulphate were both resolved with sharp peaks. Nitrite, bromide and phosphate occurred in lower concentrations.
The accurate determination of F- in these matrices proved to be problematic when using IC. Lower F- concentrations from 0.1 to 2 mg.L-1 were over estimated and higher concentrations underestimated. It should be noted that the presence of anions such as acetate, chloride, nitrite, bromide, nitrate, phosphate and sulphate can interfere with the F- determinations. The presence of polyvalent
cations (Al3+, Fe3+, Mg2+, and Ca2+) may also cause interferences under operating conditions of the IC, as it does with the ISE method. Typical concentrations of these polyvalent cations in the clay extracts were found to be in the range of 1-4 mg.L-1 aluminium, 0-5 mg.L-1 calcium, 0-11 mg.L-1 magnesium and 0-0.5 mg.L-1 iron.
Figure 3.9
A chromatogram for the RBM extract using sodium
carbonate/bicarbonate eluent and an AS4A anion exchange column after spiking with (2 mg.F-.L-1). Concentrations (mg.L-1): 1. F- = 3.41, 2. Ac- = 0.16, 4. Cl- = 0.81, 5. NO2 = 0.09, 6. Br- = 0.01, 7. NO3 = 5.17, 8. PO4 = 0.03 and 9. SO4 = 3.33.
The ion selective electrode method of fluoride determination has found wide spread use. However, when applied to samples in which aluminium concentrations may be high (clay samples) the buffer systems (TISAB buffers) were not effective in combating aluminium interference. Some clay samples contained sufficient Al3+, Fe3+, Ca2+ and Mg2+ to interfere seriously in the determination of fluoride ions when using the fluoride ion selective electrode. In the presence of these metals the spiking method therefore gave erroneous results.
In comparison with IC, the results seem to improve when using ISE. This is because of the availability of TISAB being used. TISAB III contains CDTA (cyclohexane diamine tetra acetic acid) and TISAB IV contains Tris (hydroxymethyl) aminomethane (THA) as complexing agents. The success of the ISE method depends on the effectiveness of the system used to buffer pH and ionic strength and to complex possible interferences such as Al3+, Mg2+, Ca2+ and Fe3+. The F- recovery using these reagents depends on the system under investigation. For instance, both CDTA and THA yield quantitative results of F -analysis in deionised and tap water. In more complex matrices, such as clay extracts, the efficiency of these complexing agents seems to decrease. In case of TISAB III, quantitative F- recoveries were not obtained. The results obtained using ISE with TISAB IV buffer produced the most accurate results and were used as standard analytical procedure in this study.
Chemical interferences are difficult to eliminate completely. Fluoride readily forms complexes with many ions in solution. In order to measure total fluoride present, such complexes need to be destroyed. Polyvalent cations (Al3+, Fe3+, Mg2+, and Ca2+) complexes F- and it becomes necessary to destroy these complexes. Aluminium as was already mentioned presents a significant interference even at levels below 0.2 mg Al / L. Considering stability constants of aluminium fluoride complexes, it is expected that part of the F- in the original sample will be bound by aluminium. Obviously, the added decomplexing agents hydroxymethyl aminomethane for TISAB IV and CDTA (cyclohexane diamine tetra acetic acid) for TISAB III were not able to fully decomplex F- bound by metals. This means the decomplexing agents applied are not able to decomplex F- completely from its aluminium complex form. Thus, for clay extracts which contain aluminium, measured F- concentrations obtained by using ISE method were a fraction of the total F- concentration.
The chemical states and dissociation behaviour of fluoride have been thoroughly investigated. In an aqueous solution, fluoride has three dissolved forms,
depending on the pH: which are HF, H2F2 and F-. Among these dissolved forms, the fluoride ion (F-) is the most dominant form in water with pH values ranging from 3 to 9. However, when tri or tetravalent cations of certain metallic elements such as aluminium or iron exist along with fluoride ions, complexes with fluoride ions are readily formed. Aluminium fluoride is one of the stable complexes which can be formed with fluoride anions. Several different forms of aluminium fluoride can be found with various F-/Al3+ atomic ratios. One aluminium ion (Al3+) in the complex structure is surrounded by a maximum of six octahedrally arranged fluoride ions. This means that the presence of one aluminium ion could form a complex with six free fluoride ions. This would, however, not occur at the concentrations levels encountered in this study.
Therefore, the significant interference by aluminium ions during fluoride determinations using both IC and ISE was caused by the strong affinity of fluoride ions for aluminium ions. Although both decomplexing agents used obviously dissociated aluminium ions from the aluminium fluoride complexes, this process was not complete.
The interference of aluminium in fluoride ion determinations can also be attributed to the formation of insoluble aluminium fluoro-species. Aluminium forms colloidal hydrous oxide particles in the pH range 4-9. This solid can strongly adsorb fluoride, resulting in low fluoride measurements with both IC and ISE.
Anion interferences are particularly problematic when using ion chromatography. These interferences are normally caused by anions with retention times that are close to each other. This could lead to an overlap of one anion to the anion of interest.
For determinations based on anion chromatography, much remains to be learned about aluminium interference effects. With excess of aluminium present, low values seem inevitable unless the aluminium is removed.
3.3.3.3 Effect of acetate on F- determination with IC
The effect of acetate or other organic species extracted from the clays and that elute close to F- was investigated by spiking activated alumina (ALU) extracts with 1 mg.L-1 fluoride and 5 mg.L-1 acetate. Table 3.6 shows results obtained with ISE using TISAB III and IV and IC.
TABLE 3.6
F- determined in ALU extract spiked with 1 mg.L-1 F- and 5 mg.L-1 Ac- by IC and ISE using
TISAB III and IV.
DEIONISED WATER
ALU TISAB III TISAB IV IC
Spiked F -mg.L-1 F- mg.L-1 Recovery (%) Error (%) F- mg.L-1 Recovery (%) Error (%) F- mg.L-1 Recovery (%) Error (%) 1 0.59 60 -40 0.62 62 -38 0.46 46 -54 TAP WATER
TISAB III TISAB IV IC
Spiked F- mg.L-1 F- mg.L-1 Recovery (%) Error (%) F- mg.L-1 Recovery (%) Error (%) F- mg.L-1 Recovery (%) Error (%) 1 0.76 76 -24 0.67 67 -33 0.52 52 -48
The results show lower fluoride concentrations with both IC and ISE. Consistent lower F- concentrations with both methods might also be attributed to the effect of aluminium present in the samples. Ion chromatography studies show that the acetate interferes with fluoride as the two anions elute close to each other. The chromatogram in Figure 3.10, shows the effect of acetate on the fluoride peak.
All anions elute well away from the water dip. From the chromatogram, the interference of acetate with fluoride peak is clear as the two peaks are next to each other. The presence of acetate in the sample results in the underestimation of the fluoride. Therefore, it is difficult to obtain accurate and quantitative fluoride measurements in the presence of acetate. The retention times of other anions also seem to differ when large amounts of acetate were present.
Figure 3.10
Chromatogram showing interference of acetate with the fluoride peak in an activated alumina (ALU) extract spiked with 1 mg.L-1 F- and 5 mg.L-1 acetate. Concentrations (mg.L-1): 1. F- = 0.48, 2. Ac- = 5.63, 3. Cl- = 23.5, 4. NO3 = 0.14, 5. PO4 = 0.03 and 6. SO4 = 0.65. 3.3.4 Conclusions
A good agreement for F- determinations was found for the two methods. The results emphasise the point that the determination of fluoride depends on the matrix characteristics. Consistently lower F- concentrations were found in spiked clay extracts. The presence of Al in the samples resulted in lower F- values using both techniques. Acetate has also proved to be a major interference on fluoride determinations. In the analysis of clay extracts in this study care was taken to develop efficient precleaning procedures to reduce the presence of interfering
3.4 Experimental methods
3.4.1 Determination of elemental composition
3.4.1.1 Solid samples
The major element composition of selected clay samples was determined by XRF by Set Point Technologies.
3.4.1.2 Solutions
Inductively coupled plasma optical emission spectrometry (ICP-OES) was used to determine the elemental composition in solution samples. A method was developed for the determination of the elements: Al, Fe, Mn, Cu, Ca, Mg, Cr, Mn, Ni, Zn, and As using a Varian Liberty 100 ICP-OES spectrometer. A 1000 mg.L-1 Multi IV (Merck SA) multi-element ICP-OES standard was used to prepare calibration standards by appropriate dilution with 2% HNO3.
The analytical wavelengths and experimental conditions used in the ICP-OES measurements are summarised in Table 3.7 and Table 3.8, respectively.
TABLE 3.7
Analytical wavelengths for the determination of metals in solution by ICP-OES Elements Analytical Wavelength (nm) Instrumental detection limit Al 396.153 0.10 Fe 259.940 0.05 Cu 324.754 0.02 Ca 393.366 0.05 Mg 279.553 0.05 Cr 267.716 0.05 Mn 257.610 0.02 Ni 231.604 0.05 Zn 213.856 0.05 As 188.979 0.05 TABLE 3.8
ICP-OES operational conditions
Parameter Setting
Generator power 1.3 kW
Plasma gas flow 15.0 L.min-1
Nebuliser pressure 150 kPa
Sample uptake rate 1 mL.min-1
Integration time 3 s
Viewing height 2-8 mm
PMT voltage 450-650 V
3.4.2 Pretreatment procedures
3.4.2.1 Wash procedure to remove water extractables
All clay samples were extracted by shaking with deionised water in a 1:10 mass(g) to volume(mL) ratio for 10 min, centrifuged at 3900 rpm for 5 min and decanted. Extraction was repeated until clays were free from water extractable ions indicated by the absence of Cl- in ion chromatograms of the extracts. The washed clay was dried overnight at 105 °C, pulverised in an agate mortar and stored in a desiccator.
3.4.2.2 Heat treatment
Clay samples, placed in porcelain crucibles, were heat-treated at the desired temperature in a preheated muffle furnace for 2 h. To study the effect of heating clays at different temperatures, the same procedure was followed through the temperature range 200 to 900oC in steps of 100oC. A temperature of 600 °C was selected for heat treatment. After heat treatment clays were cooled, pulverised and stored in a desiccator.
3.4.2.3 Chemical pretreatment
Clay samples containing F- were chemically treated to remove F- through an ion exchange process with hydroxide. Chemical treatment was also used to activate clay surfaces before F- adsorption. In the chemical pretreatment procedure clays were stirred with 0.1 M NaCO3 for 30 min in a mass(g) to volume (mL) ratio of 1 to 5, centrifuged for 5 min and decanted. The residue was washed with deionised water to remove excess sodium carbonate and again centrifuged and decanted. The residue was then treated with 1% HCl for 30 min in a mass(g) to volume(mL) ratio of 1 to 5 and rinsed until Cl--free. The rinsed solution was tested with
AgNO3 solution initially and finally with HPIC. The residue was dried overnight at 105 °C, pulverised in an agate mortar and stored in a desiccator.
3.4.2.4 Chemical treatment for alkaline clays
Alkaline clays such as clays containing substantial amounts of dolomite or Ca and Mg carbonates, required modification of the chemical pretreatment procedure. The acid treatment was excluded from the procedure described in Section 3.4.2.3, because adding acid to dolomitic clays would cause dissolution of the Ca and Mg carbonates according to the reaction:
CaCO3 + HCl ? Ca2+
+ 2Cl- + H2O + CO2
3.4.2.5 Sequential pretreatment procedure
In the sequential treatment, samples were first heated for 1 h at 600oC and then treated chemically with Na2CO3 and HCl as above.
Table 3.9 summarises the different pretreatment procedures and the notation used in this study.
TABLE 3.9
Notation of pretreatment procedures
Procedure Notation
Removal of water extractables CR
Heat treatment C
Chemical treatment CT
Heat followed by chemical treatment
3.4.2.6 Substrate regeneration procedure
Regeneration of clay substrates was done using the same solutions, 0.1 M Na2CO3 and 1% HCl as for chemical pretreatment.
3.4.3 Ammonium oxalate extraction
The ammonium oxalate extraction (Parfitt, 1989) removes materials without a well-developed crystalline structure and is specifically designed to dissolve Al, Fe and Mn from amorphous soil phases containing these elements. A sample is shaken for 4 h in the dark with ammonium oxalate at pH 3.
0.5 g of clay sample was weighed into a 100 mL polyethylene flask which was wrapped in aluminium foil to protect it from light induced reactions. 50 mL 0.2 M ammonium oxalate at pH 3 was added. The flask was then placed in a shaker for 4 h. After filtration and appropriate dilution the Al, Fe and Mn concentrations were determined in the filtrate by ICP-OES.
3.4.4 Batch adsorption procedures
3.4.4.1 Adsorption at pH 6
Adsorption capacities were determined at pH 6 by shaking 1 g of washed and dried (2 h at 105oC) clay with 50 mL of 10 mg.F-.L-1 NaF solution (C0) for 2 h in polyethylene bottles at 22 °C. The initial pH was adjusted to approximately 6 using NaOH or HCl depending on the acid or base properties of the sample. After equilibration the solutions were centrifuged, the pH measured to ensure that the pH was within ±0.2 pH unit from the target pH, and the residual F- concentration (Ce) determined using ion chromatography or a F- ion selective electrode. The % adsorption was calculated from the residual F- concentration using the equation:
o e O C C C x Adsorption 100 ( ) % = −
To ensure that F- does not adsorb on the inner walls of the adsorption vessels, blank runs were performed. In this procedure a 10 mg.F-.L-1 solution was added to a polyethylene vessel and the F- concentration measured after 2 h and again after 12 h. No reduction in F- concentration was found.
3.4.4.2 Adsorption curves (% adsorption vs pH)
The same method as described in Section 3.4.4.1 was followed for equilibrium solution pH values of about 3, 4, 5, 6, 7, 8, and 9.
3.4.4.3 Competitative adsorption
The effect of other anions on the adsorption of F- was studied by determining the residual F- concentration in the presence of Cl- and SO42-. The same procedure was used as in Section 3.4.4.1.
3.4.4.4 Determination of adsorption kinetics
Adsorption kinetics were determined using the batch adsorption procedure described in Section 3.4.4.1. Residual F- concentrations were, however, measured at the time intervals 10, 20, 30, 45, 60, 120, and 300 min.
3.4.5 Zeta potentials
Zeta potentials were determined by shaking 1g of clay with 50 mL of deionised water. The pH was adjusted to desired values (3-9) using 0.1 M NaOH or HNO3
depending on the acid or base properties of the sample. Zeta potentials were measured using a Malvern ZetaMaster.
3.4.6 Determination of adsorption isotherms
Adsorption isotherms were determined using chemically pretreated clays. 1 g of clay was equilibrated in 45 mL 0.1 M NaClO4 at pH 6 for 14 h. Appropriate amounts of a 1000 mg.L-1 F- stock solution and 0.1 M NaClO4 were then added to the equilibrated samples to prepare 50 mL test solutions with initial concentrations: 2, 5, 10, 20, 50, 100, and 200 mg.F-.L-1. These solutions were equilibrated for 2 h to complete the adsorption process and the residual fluoride concentration and final pH determined.
3.4.7 Determination of ∆pH curves
Clay samples of 1 g each were equilibrated for 14 h in 50 mL 0.1 M NaClO4 solution at pH 3 to 9. After equilibration 0.5 mL of a 1000 mg.L-1 F- stock solution was added so that the initial F- concentration was 10 mg.L-1.
The pH was adjusted with 1M HClO4 or NaOH. The pH was measured at 1h after pH adjustment (pH1h), again after 14h equilibration (pHi), and then at different times after addition of 10 mg.L-1 fluoride (For example pH15 denotes 15 m after addition of F-).
3.4.8 Kinetics of adsorption and pH equilibration
Kinetics of equilibration were determined by weighing 1g of clay and add 47 mL of deionised water and 2.5 mL of 1 M NaClO4. The pH was adjusted with 1 M HClO4 or NaOH. The pH was measured before adding fluoride (pHi). Fluoride solution of 10 mg.L-1 (0.5 mL of 1000 mg.L-1 F-) was added. The pH was then
