TABLE OF CONTENTS

TABLE OF CONTENTS 1  ENHANCEMENT OF DRINKING WATER DISIFECTION METHODS VIA USE OF METAL/ METAL-OXIDE NANOPARTICLES: POTENTIAL AND LIMITATIONS 2  1. INTRODUCTION 2  2. DRINKING WATER CONTAMINANTS 3  2.1. CHEMICAL CONTAMINANS   3  2.2. RADIONUCLEIDES   6  2.3. PATHOGENIC MICROORGANISMS   6  2.3.1. BACTERIA   7  2.3.2. VIRUS   8  2.3.3. PROTOZOA AND HELMINTHES   10  WATER QUALITY TARGETS WITH RESPECT TO MICROBIAL CONTAMINANTS   11  3. HOUSEHOLD WATER TREATMENT 12  4. METHODS USE IN DRINKING WATER DISINFECTION 13  4.2. STRONG OXIDANT-BASED SYSTEMS   14  4.2.1. CHLORINE   15  4.2.2. CHLORINE DIOXIDE   17  4.2.3. CHLORAMINES   18  4.2.4. OZONE  18  4.2.5. HETEROGENEOUS FENTON PROCESS   19  4.1. HEAT AND UV-BASED SYSTEMS   20  4.1.1. HEAT TREATMENT   20  4.1.2. SODIS   20  4.1.3. UV-C   22  4.1.4. HETEROGENEOUS PHOTOCATALYSIS   22  4.1.4.1. TiO2 HETEROGENEOUS PHOTOCATALYSIS   23  4.1.4.2. ZnO HETEROGENEOUS PHOTACATALYSIS   25  4.1.4.3. Fe2O3 HETEROGENEOUS PHOTOCATALYSIS  26  4.3. PHYSICAL SYSTEMS   26  4.3.1. COAGULATION OR PRECIPITATION   26  4.3.2. FILTRATION   27  4.3.2.1. SURFACE FILTRATION   27  4.3.2.1.1. SURFACE FILTER BIOFOULING PREVENTION   29  4.3.2.2. DEPTH FILTRATION   31  4.3.2.2.1. SAND FILTER   31  4.3.2.2.2. ACTIVATED CARBON   32  4.3.2.2.3. PAPER, FIBER AND FABRIC FILTERS   32  4.3.2.2.4. CERAMIC FILTER  33  4.3.2.3. ENHANCEMENT OF BACTERIA REMOVAL IN DEPTH FILTRATION   34  4.3.2.4. ENHANCEMENT OF VIRUS REMOVAL IN DEPTH FILTRATION   35  4.3.2.4.1. DEPOSITION   36  4.3.2.4.2. DLVO FORCES   36  4.3.2.4.3. NON-DLVO FORCES   40  4.3.2.4.4. DLVO THEORY   40  

4.3.2.4.5. VIRUS ADSORPTION AND INACTIVATION   40  4.3.2.4.6. SAND FILTER VIRUS RETENTION ENHANCEMENT   42  4.3.2.4.7. FIBER FILTER VIRUS RETENTION ENHANCEMENT   43  4.3.2.4.8. CERAMIC FILTERS VIRUS RETENTION ENHANCEMENT   45  4.3.2.5. SIMULTANEOUS ENHANCEMENT OF BACTERIA AND VIRUS REMOVAL IN DEPTH FILTRATION  49  4.4. MULTIPLE-BARRIER APPROACHES   52  5. ANALYSIS OF HOUSEHOLD DISNFECTION SYSTEMS 52  6. ANALYSIS OF NANOMATERIALS ENHANCED DISINFECTION 54  6.1. LACK OF UNDERSTANDING OF MICROBIAL CONTROL MECHANISMS   55  6.2. COMPETITIVE ADSORTION   56  6.3. NANOMATERIAL AGGREGATION, RECOVERY AND INMOBILIZATION   56  6.4. PHOTOCATALYTIC ACTIVITY  58  6.4.1. CHARGE CARRIER RECOMBINATION   58  6.4.2. VISIBLE LIGHT ADSORPTION   60  6.5. NANOMATERIAL TOXICITY   60  7. DEVELOPMENT OF NANOTECHNOLOGY ENHANCED POINT OF USE SYTEMS FOR WATER DISINFECTION 64  8. REFERENCES 65  

ENHANCEMENT OF DRINKING WATER DISIFECTION METHODS VIA USE OF METAL/ METAL-OXIDE NANOPARTICLES: POTENTIAL AND LIMITATIONS 1. INTRODUCTION

The availability of potable water has become nowadays a worldwide problem due to the continuous growth in water demand not balanced by an adequate recharge. Additionally, more and more often water sources are suffering from a worsening of their quality due to the indiscriminate discharge of both domestic and industrial effluents without adequate treatments. (F. Macedonio, 2012) (Jonathon Brame, 2011) (Meng Nan Chong, 2010) (S. Malato, 2009) (Nora Savage, 2005) (Chunjian Shi, 2012). Acording to the 2012 update of Progess on drinking water and sanitation published by WHO and UNICEF, over 780 million people are still without access to improved sources of drinking water and 2.5 billion lack improved sanitation. It is also stated that if current trends continue, these numbers will remain unacceptably high in 2015: 605 million people will be without improved drinking water access and 2.4 billion people will lack access to improved sanitation facilities (WHO, 2012). While it is not possible to quantify the proportion of deaths directly due to unsafe drinking water and not attributed to other transmission routes, access to clean drinking water and proper sanitation can provide substantial improvements in health. Consequently assuring reliable access to inexpensive and clean sources of water is within the United Nations Millenium Development goals (Ian Bradley, 2011) (Jonathon Brame, 2011). Addressing these problems calls out

for a tremendous amount of research to be conducted to identify robust new methods of purifying water at lower cost and with less energy, while at the same time minimizing the use of chemicals and impact on the environment (S. Malato, 2009).

2. DRINKING WATER CONTAMINANTS Drinking water does not need to be absolutely pure to be safe. Because water is such a good solvent, pure water containing nothing else is almost impossible to attain. What is required is that drinking water be safe for people in most stages of normal life. Hence, it should contain no harmful concentrations of chemicals, radionuclides or pathogenic microorganisms (NHMRC, NRMM, of Australia, 2011).

2.1. CHEMICAL CONTAMINANS Chemicals contaminants present in water can be arrange by source as: (i) Naturally occurring chemicals, (ii) Chemicals from industrial sources and human residences, (iii) Chemicals from agricultural activities, and (iv) Chemicals used in water treatment or from materials in contact with drinking-water. There are a number of sources of naturally occurring chemicals in drinking water. All natural waters contain a range of inorganic and organic chemicals. The former derive from the rocks and soil through which water flows. The latter derive from the breakdown of plant material or from cyanobacteria and other microorganisms that grow in the water or on sediments (WHO, 2011). Guideline values have been established for arsenic, barium, boron, chromium, fluoride, selenium, uranium and microcystine-LR (WHO, 2011). Arsenides and fluoride are among the best-known, widespread and significant naturally occurring waterborne chemical pollutants: guidelines define their maximum acceptable concentrations as 10 µg/l and 1.5 mg/l respectively. In severe problem areas, increased concentrations of this chemicals in drinking water can lead to skin diseases (e.g. hyperkeratosis), cancer (by arsenic poisoning) or crippling diseases (skeletal fluorosis). These two chemicals alone affect close to hundred million people in developing countries (Maryna Peter-Varbanets, 2010). Cyanobacteria are Gram-negative bacteria that contain chlorophyll (WHO, 2011). Their primary health significance is that many species of cyanobacteria produce toxins, which can be either contained intracellularly, or expressed extracellularly and therefore present in the surrounding water. The two main types of toxin are: (i) Cyclic peptides (microcystins and nodularin). Microcystins cause damage to the liver and are possibly carcinogenic. Furthermore, they are extremely stable chemically and remain potent even after boiling; the half-lives for

breakdown of microcystins in natural water have been shown to range from 5 to 20 days. Nodularin has an identical mode of action to microcystin in animals and is considered to present at least the same risk to human health as microcystin, and (ii) Alkaloids (neurotoxins and cylindrospermopsin). Neurotoxins produced by cyanobacteria include anatoxin A, anatoxin A-S and the saxitoxins. Cylindrospermopsin is a general cytotoxin that blocks protein synthesis. The major pathological effects are damage to the liver, kidneys, lungs, heart, stomach, adrenal glands, the vascular system, and the lymphatic system. Acute clinical symptoms are kidney and liver failure. However, no human deaths have been recorded from ingesting the toxins of cyanobacteria but gastroenteritis may result from drinking water containing toxic species and extended exposure may lead to more serious impacts. (NHMRC, NRMM, of Australia, 2011) (Jungju Lee, 2011) (WHO, 2011). The guideline value for microcystin-LR is 1 µg/l (WHO, 2011). As with other cyanotoxins, a high proportion of microcystins remain intracellular unless cells are lysed or damaged, and can therefore be removed by coagulation and filtration in a conventional treatment plant. It should be noted that using oxidants such as chlorine or ozone to treat water containing cyanobacterial cells, while killing the cells, will also result in the release of free toxin. Free microcystins are readily oxidised by ozone and chlorine and adsorbed by activated carbon. (NHMRC, NRMM, of Australia, 2011). In contrast, other studies have reported that conventional treatment plants have very low removal efficiencies of cyanobacterial toxins (Venkata K. K. Upadhyayula, 2009). Chemicals from industrial sources and human residences can reach drinking water directly from discharges or indirectly from diffuse sources arising from the use and disposal of materials and  products containing the chemicals (WHO, 2011). Pharmaceuticals can be introduced into water sources in sewage by excretion from individuals using these chemicals and, from uncontrolled drug disposal (e.g. discarding drugs into toilets) (WHO, 2011). Most chemicals that may arise from  agricultural activities are pesticides, although their presence will depend on many factors, and not all pesticides are used in all circumstances or climates. Contamination can result from application and subsequent movement following rainfall or from inappropriate disposal methods. (WHO, 2011) (Maryna Peter-Varbanets, 2010). Chemicals used in water treatment and chemicals arising from materials in contact with water may give rise to contaminants in the final water (WHO, 2011).

Some substances are deliberately added to water in the course of treatment (direct additives), some of which may be inadvertently retained in the finished water  (e.g. salts, coagulant residues). Other chemicals may be taken up from contact with surfaces during treatment or distribution. Chlorine disinfectant residual, for example, is a deliberate additive, and its presence confers a benefit. Others, such as disifection byproducts (DBPs), are generated during chemical interactions between disinfectant chemicals and natural organic matter (NOM) in water (WHO, 2011). Natural Organic Matter (NOM) constitutes a diverse group of organic compounds with varying molecular weights and occurs in natural waters due to the decomposition of plant and animal residues. The organic compounds present in NOM can be divided into hydrophobic and hydrophilic fractions. (Venkata K. K. Upadhyayula, 2009) (Anu Matilainen, 2010). The hydrophilic fraction is comprised of carboxylic acids, carbohydrates, and proteins while the hydrophobic fractions comprised of humic and fulvic acids (Venkata K. K. Upadhyayula, 2009). An approach to define hydrophobicity is determination of SUVA value (which is UV254 absorbance divided by the TOC concentration). High SUVA value indicates that the organic matter is composed largely of hydrophobic and high molar mass (HMM) organic material, in comparison a low SUVA value means that water includes mainly organic compounds which are hydrophilic, low molar mass (LMM) and low in charge density (Anu Matilainen, 2010). Hydrophilic portion of NOM is of great concern for drinking water treatment plants because it reacts with strong disinfectants, thereby oxidizing NOM to simple organic acid derivatives. This provides an assimilable form of carbon source for the growth of bacteria (Venkata K. K. Upadhyayula, 2009). While the hydrophobic and high molecular mass NOM, has been generally considered the main DBP precursor in chlorine disinfection (Anu Matilainen, 2010) DBPs have raised a lot of attention in water treatment during the past decades. More than 700 compounds of DBPs have been confirmed, among which trihalomethanes (THMs) and haloacetic acids (HAAs) are the two groups found in highest concentrations and most commonly in drinking water worldwide. Chlorine, ozone, chlorine dioxide and chloramine are the most common disinfectants in use today, and each produces its own suite of chemical DBPs in drinking water. A number of epidemiological studies have suggested an association between water chlorination byproducts and various cancers, but there are insufficient data to determine concentrations at which chlorination by-products might cause an increased risk to human health. Also,

The International Agency for Research on Cancer has reviewed the available data and concluded that there is inadequate evidence to determine the carcinogenicity of chlorinated drinking water to humans (NHMRC, NRMM, of Australia, 2011). Nevertheless, their occurrence in drinking water has been regulated in several countries (Anu Matilainen, 2010). Action to reduce the concentration of disinfection by-products should be encouraged (e.g. by removal of NOM), but disinfection itself must not be compromised: the risk posed by disinfection by-products is considerably smaller than the risk posed by the presence of pathogenic microorganisms (NHMRC, NRMM, of Australia, 2011).

2.2. RADIONUCLEIDES By far the largest proportion of human exposure to radiation comes from natural sources of radiation, including cosmic radiation, or from ingestion or inhalation of radioactive materials. A very low proportion of the total human exposure comes from drinking water. Radiological contamination of drinking water can result from: (i) naturally occurring concentrations of radioactive species (e.g. radionuclides of the thorium and uranium series in drinking water sources), (ii) technological processes involving naturally radioactive materials (e.g. the mining and processing of mineral sands or phosphate fertiliser production) and (iii) manufactured radionuclides, which might enter drinking water supplies from the medical and industrial use of radioactive materials. (NHMRC, NRMM, of Australia, 2011) (WHO, 2011).

2.3. PATHOGENIC MICROORGANISMS The water acting as a passive carrier for pathogens spreads water-borne diseases. These diseases are predominantly due to fecal contamination of water sources and are thus strongly linked to sanitation conditions. Use of such water for drinking and cooking, as well as contact with it and its ingestion during bathing and washing, or even inhalation of small droplets in the form of aerosols, may result in infection. Water-borne deseaes are cause by viruses, bacterias or protozoan parasites (e.g. cholera, typhoid, bacillary dysentery, infectious hepatitis, leptospirosis, giardiasis and gastroenteritis). Water-based diseases are cause by water supporting an essential part of the life cycle of infecting agents (e.g. helminths); such as schistosomiasis, dracunculosis, bilharziosis, philariosis and oncholersosis. Water-related diseases are spread by insects that live in or close to water (e.g. mosquitoes, flies) and include yellow fever, dengue fever, encephalitis, malaria, filariasis, sleeping sickness and onchocerciasis. Finally, washing-water diseases are caused by a lack of adequate quantities of water for the proper

maintenance of personal hygiene (e.g. scabies, trachoma, leprosy, conjunctivitis, salmonellosis, ascariasis, trichuriasis and hookworm) (Maryna Peter-Varbanets, 2010). In relation to water-borne deseaes it is worth mentioning that, the significance of a particular organism in water can vary considerably; for example, a potentially pathogenic organism will not always cause symptomatic disease in a particular individual. The chances of waterborne infections occurring in a community depend on: (i) the concentration of pathogenic organisms in the water, (ii) the virulence of the strain, (iii) the per capita intake of contaminated water, (iv) the infectious dose of the particular pathogen, (v) the susceptibility of individuals and (vi) the incidence of the infection in the community. The occurrence of disease is also related to the relative level of immunity in the community. If, for example, the water supply has been repeatedly contaminated, the community may have become immune to some waterborne pathogens (NHMRC, NRMM, of Australia, 2011).

2.3.1. BACTERIA Bacteria are generally the group of pathogens that is most sensitive to inactivation by disinfection. Some environmental pathogens, such as Legionella and non-tuberculous mycobacteria, can grow in water environments, but enteric bacteria typically do not grow in water and survive for shorter periods than viruses or protozoa (WHO, 2011) (NHMRC, NRMM, of Australia, 2011). Waterborne pathogenic bacteria for wich there is some evidence of health significance related to their ocurrence in drinking water supplies are: Campylobacter jejuni, C. Coli, E. Coli – pathogenic, E. Coli – Enterohaemorrhagic, Francisella tularensis, Leptospira, Salmonela typhi, Other salmoneale, Vibrio cholerae, shigella, Burkholderia pseudomallei, Francisella tularensis, Legionella, Mycobacteria - non tuberculous (WHO, 2011). Campylobacter spp. are Gram-negative, curved spiral rods with a single polar flagellum (WHO, 2011). Campylobacter is an important cause of diarrhoea worldwide, but mortality is low. Compared with other bacterial pathogens, the infective dose is relatively low and can be below 1000 organisms. (WHO, 2011). Wild birds and poultry are the most important reservoirs of Campylobacter. Campylobacter spp, like other bacterial pathogens, can survive for several weeks in unchlorinated tap water (NHMRC, NRMM, of Australia, 2011). Escherichia coli cells are typically rod-shaped, and are about 2.0 µm long and 0.5 µm in diameter. This bacteria is present in large numbers in the normal intestinal flora of humans and animals, where it generally causes no harm. However, in other parts of the body, E. coli can cause serious disease,

such as urinary tract infections and meningitis. A limited number of enteropathogenic strains can cause acute diarrhea with the infective dose being very low (fewer than 100 organisms) (WHO, 2011). E. coli may be transmitted to humans by eating raw or undercooked meats or via foodstuffs or water supplies contaminated with faeces from infected humans or animals (NHMRC, NRMM, of Australia, 2011). Salmonella are motile, Gram-negative bacilli with diameters around 0.7-1.5 µm and lengths from 2-5 µm. S. Typhi causes large and devastating outbreaks of waterborne typhoid. (WHO, 2011). Salmonella are widely distributed in the environment and gain entry into water systems though fecal contamination from livestock, native animals, and incompletely treated waste discharges (NHMRC, NRMM, of Australia, 2011). Shigella is Gram-negative, non-spore-forming, rod-like bacteria. Shigella causes over 2 million infections each year, including about 60000 deaths, mainly in developing countries. The infective dose is low and can be as few as 10–100 organisms (WHO, 2011). Legionellae are Gram-negative, rod-shaped, nonspore- forming bacteria found in a wide range of water environments (WHO, 2011) (NHMRC, NRMM, of Australia, 2011). Mycobacteria are aerobic, rod-shaped bacteria. Non-tuberculous species of Mycobacterium grow slowly in a variety of water environments forming biofilms and some are also pathogenic. The ecology of environmental Mycobacterium spp is poorly understood; however, there is increasing evidence that they can survive and grow in water distribution systems (NHMRC, NRMM, of Australia, 2011). Leptospires are aerobic spirochetes that are typically 0.1 µm in diameter and 5–25 µm in length. (WHO, 2011). Vibrios are small, curved (comma-shaped), Gram-negative bacteria. Toxigenic Vibrio cholerae can cause watery diarrhea (WHO, 2011). Burkholderia pseudomalle is a Gram-negative bacillus commonly found in soil and muddy water, it measures 2-5 µm in length and 0.4-0.8 µm in diameter and is capable of self-propulsion (WHO, 2011).

2.3.2. VIRUS Viruses are among the smallest of all infectious agents. In essence they are molecules of nucleic acid that can enter cells and replicate in them. The viral particle consists of a genome, either ribonucleic acid (RNA) or deoxyribonucleic acid (DNA), surrounded by a protective protein shell, the capsid. Frequently this shell is itself enclosed within an envelope that contains both protein and lipid. Viruses replicate only inside specific host cells, and they are absolutely dependent on the host cell’s synthetic and energy yielding apparatus for producing new viral particles.

(NHMRC, NRMM, of Australia, 2011) (Li, 2010) (Michen, 2010). A total number of 219 virus species may be harmful to humans of which 48 species have the potential to infect humans via the water route worldwide (Michen, 2010). Currently, there are only 8 species of waterborne pathogenic viruses for wich there is some evidence of health significance related to their ocurrence in drinking water supplies namely: Adenoviruses, Astroviruses, Enteroviruses, Hepatitis A virus, Hepatitis E virus, Norovirus, Rotaviruses, Sapoviruses (WHO, 2011). These viruses are all non-enveloped, hydrophilic and with diameters ranging between 20 - 80 nm. Adenoviruses consist of a double-stranded DNA genome in a non-enveloped icosahedral capsid with a diameter of about 80 nm. Adenoviruses cause a wide range of symptomatic infections and can be transmitted by a number of routes including, faecal-oral and inhalation of aerosols (NHMRC, NRMM, of Australia, 2011) (WHO, 2011). Astroviruses consist of a single-stranded RNA genome in a non-enveloped icosahedral capsid with a diameter of about 28 nm (WHO, 2011). Sapoviruses consists of four genera of single-stranded RNA viruses with a non-enveloped capsid (diameter 35–40 nm) (WHO, 2011). Enteroviruses are among the smallest known viruses and consist of a single-stranded RNA genome in a non-enveloped icosahedral capsid with a diameter of 20–30 nm (WHO, 2011). Enteroviruses having a worldwide distribution are one of the most common causes of human infections and can cause a broad range of symptomatic infections (e.g. diarrhea, gastroenteritis and respiratory infections). Transmission is mainly by person-to-person contact and inhalation of aerosols (NHMRC, NRMM, of Australia, 2011). The term hepatitis virus describes a group of viruses that target the liver and cause inflammation. Enteric hepatitis viruses including Hepatitis A and E are transmitted by the faecal-oral route and can be transmitted from contaminated food or water. Hepatitis A is a single-stranded RNA non-enveloped virus. Hepatitis E virus consists of a single-stranded RNA genome in a non-enveloped icosahedral capsid with a diameter of 27–34 nm (WHO, 2011). Hepatitis E is widespread but endemic in regions such as Mexico, Nepal, India, central Asia and parts of Africa (NHMRC, NRMM, of Australia, 2011). Noroviruses are single-stranded RNA non-enveloped viruses, and their major route of transmission is person-to-person by the faecal-oral route (NHMRC, NRMM, of Australia, 2011). Rotaviruses are the most important cause of gastrointestinal infection in children. Approximately, 50–60% of cases of acute gastroenteritis of hospitalized children throughout the world are caused

by human rotaviruses; this infection causes over 600000 deaths each year (WHO, 2011) (Li, 2010) (Leonardo Gutierrez, 2009). Rotaviruses consist of a segmented double-stranded RNA genome in a non-enveloped icosahedral capsid with a diameter of 50–65 nm.  This capsid is surrounded by a second double-layered capsid, giving the virus the appearance of a wheel - hence the name rotavirus.  The diameter of the entire virus is about 80 nm.  The primary mode of transmission is fecal-oral, with inhalation of aerosols also possible. Although infected people excrete large numbers of viral particles, water plays a smaller role than expected in transmission (NHMRC, NRMM, of Australia, 2011).

2.3.3. PROTOZOA AND HELMINTHES Protozoa and helminths are among the most common causes of infection and disease in humans and animals. Water plays an important role in the transmission of some of these pathogens. The control of waterborne transmission presents real challenges, because most of the pathogens produce cysts, oocysts or eggs that are extremely resistant to processes  generally used for the disinfection of water and can survive for long periods in water. However, protozoa and helminthes are of a moderate size (> 2 µm) and can be removed by physical processes (WHO, 2011). The great majority of protozoa in freshwater are natural aquatic organisms of no significance to health. The pathogenic protozoa that may occur in drinking water fall into two functional groups: (i) enteric origin and (ii) environmental origin. Environmental protozoa are responsible for serious cerebral and eye diseases (NHMRC, NRMM, of Australia, 2011). Waterborne pathogenic protozoa for wich there is some evidence of health significance related to their ocurrence in drinking water supplies are: Acanthamoeba, Cryptosporidium hominis/parvum, Cyclospora cayetanensis, Entamoeba histolytica, Giardia intestinalis and Naegleria fowleri. Dracunculus medinensis and Schistosoma are waterborne pathogenic helminths (WHO, 2011). Acanthamoeba are free-living amoebae (10–50 µm) common in aquatic environments.  Under unfavourable conditions, it will develop into a dormant cyst that can withstand extremes of temperature (up to 56 °C), disinfection and desiccation (WHO, 2011). Acanthamoeba species cause both cerebral and corneal disease (NHMRC, NRMM, of Australia, 2011). In recent years, Cryptosporidium has come to be regarded as one of the most important waterborne human pathogens in developed countries. Over 30 outbreaks associated with drinking water have been reported in North America and Britain. Thick-walled oocysts with a diameter of 4–6 µm, shed in faeces are responsible

for transmission. Oocysts are robust and can survive for weeks to months in fresh water. There are a number of species of Cryptosporidium, with C. hominis and C. parvum identified as the main causes of disease in humans via fecal-oral transmision. Waterborne outbreaks of cryptosporidiosis have been attributed to inadequate or faulty treatment and contamination by human or livestock waste (NHMRC, NRMM, of Australia, 2011) (WHO, 2011). Although known as a human parasite for 200 years, Giardia has been regarded seriously as an agent of disease since the 1960s. It has been identified as an important waterborne pathogen, and linked to many outbreaks of illness associated with drinking water, particularly in North America. Giardia has a relatively simple life cycle involving two stages: a flagellate that multiplies in the intestine, and an infective thick-walled cyst (8–12 µm in diameter). Cysts are robust and can survive for weeks to months in fresh water. There are a number of species of Giardia, but human infections (giardiasis) are usually assigned to one, G. intestinalis. Waterborne outbreaks of giardiasis (fecal-oral transmission) have generally been linked to consumption of untreated or unfiltered surface water and contamination with human waste (NHMRC, NRMM, of Australia, 2011) (WHO, 2011). Cyclospora cayetanensis produces thick-walled oocysts of  8–10 µm in diameter (WHO, 2011). Entamoeba histolytica, is the most prevalent intestinal protozoan pathogen worldwide. Entamoeba (diameter 10–60 µm) will develop into a dormant cyst (diameter 10–20 µm) under unfavourable conditions (WHO, 2011). Naegleria fowleri (10–20 µm) feeds on bacteria.  Under adverse conditions, the trophozoite transforms into a circular cyst (7–15 µm); which is resistant to unfavourable conditions (WHO, 2011). Cerebral infection by Naegleria fowleri is strictly waterborne and, although rare, is usually fatal. Since these amoebae are able to colonise piped water supplies, disinfection at the water source may not adequately control them unless the disinfectant pervades the whole distribution system (NHMRC, NRMM, of Australia, 2011).

WATER QUALITY TARGETS WITH RESPECT TO MICROBIAL CONTAMINANTS

Water quality targets are typically not developed for pathogenic microorganisms; monitoring finished water for pathogens is not considered a feasible or cost-effective option because pathogen concentrations equivalent to tolerable levels of risk are typically less than 1 organism per 104 to 105 litres. For the control of microbial hazards, the most frequent form of health-based goal applied is performance targets defined in relation to source water quality. There is insufficient data, and it is not

realistic, to derive performance targets for all potentially waterborne pathogens. The practical approach is to derive targets for reference microorganisms representing groups of pathogens (WHO, 2011). As an example, three reference pathogens along with their significance in drinking water quality assessment are presented bellow. E. coli can be used to assess: (i) source water quality and potential impacts of human and animal waste; (ii) inadequate treatment; and (iii) post-treatment ingress of human and animal waste into distribution systems. The performance target indicates that Escherichia coli should not be detected in any 100 ml sample of drinking water. It also relevant to take into account that E. Coli is not an effective indicator for the presence of enteric protozoa or viruses. There is no performance target for Clostridium perfringens in drinking water. Although, due to their small size (spores - 1 µm) and exceptional resistance to disinfection processes and other unfavourable environmental conditions, C. perfringens spores have been proposed as potential indicators for enteric viruses and protozoa in drinking-water supplies (NHMRC, NRMM, of Australia, 2011). Bacteriophages (also known as phages) are viruses that exclusively use bacteria as their hosts for replication. The physical similarities between coliphages and viruses are particularly useful in validating efficacy of filtration and they can also be used to validate disinfection processes. The performance target indicate that coliphages should not be detected in any 100 mL sample of drinking water. (NHMRC, NRMM, of Australia, 2011). MS2 and φX174 are non-enveloped phages, commonly utilized as model viruses, due to their similarity to the human enteric viruses in shape, size, behavior, and nucleic acid structure. In addition, these microorganisms are usually more resistant to disinfection processes than the majority of pathogenic viruses (Li, 2010) (Leonardo Gutierrez, 2009).

3. HOUSEHOLD WATER TREATMENT The estimates of the population in urban areas with access to a reliable water supply given by the WHO may be set too high. In some cities, water systems draw unsafe water from unprotected or contaminated sources and deliver it to consumers with no or inadequate treatment, despite being classified as safe (Maryna Peter-Varbanets, 2010). This would imply that the number of people without direct access to safe drinking water may be higher than the 0.78 billions mention before. Another problem contributing to the underestimation of the population served by unsafe water is the contamination of water during its distribution to homes via pipes or trucks (Maryna Peter-Varbanets, 2010). Drinking water

recontamination has been observed during the storage and handling due to unhygienic practices in the household. In rural areas of developing and transition countries, investments for centralized systems are often unaffordable given the remote locations and lack of financial resources. In the rare cases where centralized systems are installed, the system often fails to provide drinking water of reliable quality due to unprofessional operation, maintenance and management. Drinking water from a supply network and a central water treatment facility is therefore generally unavailable in rural areas, and typically, water is recovered from surface/ground water sources or rainwater, with no water disinfection before consumption. It can be concluded from these arguments that approaches relying solely on centralized solutions may work in some regions, whereas in many cases structural problems lead to their failure. In this case, promoting alternative water treatment options such as household treatment is often the most feasible way of improving drinking water qualiy. (Maryna Peter-Varbanets, 2010) (Michen, 2010) (WHO, 2011). Household water treatment technologies are also known as point-of-use and point-of-entry water treatment technologies (WHO, 2011). Point-of-use (POU) systems treat only the part of water used for drinking. The minimum requirement for drinking water amounts to about 2 l per person per day, while the maximum for drinking and cooking is 8 l per day. The application of these systems is already widespread: around 19 million people are estimated to use POU water treatment, in addition to the 350 million people who boil their water. Field and laboratory studies indicate that the improvement of water quality through the use of POU technologies results in 30-40% reductions in diarrheal disease (Ian Bradley, 2011). Point-of-entry (POE) systems refer to the treatment of all the water supplied to a household. The treatment capacity is therefore much higher than for POU systems in the order of 100–150 l per person per day (Maryna Peter-Varbanets, 2010) (Nora Savage, 2005).

4. METHODS USE IN DRINKING WATER DISINFECTION The value of pure water supplies has been recognised, for millennia. Hippocrates described an association between water supplies and disease. However, was only by the middle of the 19th century, when Britain was affected by major epidemics of cholera and endemic typhoid, that filtration of river-derived water became a legal requirement in London in 1859. This mesure was the result of the studies carried by John Snow and William Budd, which provide incontrovertible evidence of the role of water

in transmission of these two diseases (NHMRC, NRMM, of Australia, 2011). Microbial contamination has been recognized as the greatest challenge in health risk management in drinking water supply. Consequently, disinfection defined as the removal of pathogenic microorganisms, is the process that has the greatest potential to reduce waterborne disease and thus has the highest impact on drinking water safety (Youwen You, 2005) (S. Malato, 2009). Water disinfecition methods include: (i) strong oxidants-based systems, (ii) heat and UV-based systems, (iii) physical systems, and (iv) combined systems. Most of these methods can be adapted for either large scale or household drinking water disinfection. In addition, different nanotechnological approaches are been tested in order to enhance the efficacy of disinfection methods, for example: enhancement UV-solar disinfection via heterogenous photocalysis using nanosized semiconductor oxides (e.g. TiO2, ZnO), addition of TiO2 and Ag nanoparticles to membrane filters to improve biofouling resistance, enhanced virus rentention or bactericidal capacity of depth filters by surface functionalization using nanosized metals/metal oxides (e.g. ZrO2, Y2O3, Fe2O3, ZVI, MgO, Ag, AgO), and combine approaches such as ozonation with ceramic membrane filtration. Unlike strong oxidants used for disinfection, most of these nanomaterials are relatively inert in water. Therefore, they are not expected to produce disinfection byproducs that compromise drinking water quality, when properly incorporated into disinfection processes. Three main aspects of disinfection methods are discuss: Disinfection mechanisms, effectivity regarding microorganisms inactivation and limitations

4.2. STRONG OXIDANT-BASED SYSTEMS Commonly used disinfection processes are chlorination and chloramination, however ozone and chlorine dioxide are also used. These methods are very effective in killing bacteria and can be reasonably effective in inactivating viruses (depending on type) and many protozoa, including Giardia (NHMRC, NRMM, of Australia, 2011). The efficacy of disinfection by strong oxidants depends greatly on conditions such as contact time, the pH and turbidity, change in pathogen type or concentration and organic matter content, which can shift the balance anywhere between complete removal and complete non-removal of pathogenic microorganisms (Venkata K. K. Upadhyayula, 2009) (S. Malato, 2009) (Maryna Peter-Varbanets, 2010). In order to access the relative efectivity of a certain disinfectant against a test microorganism under specific process conditions the C.t value is determined by multiplying the concentration of

residual disinfectant (in mg/L) by the contact time (in minutes) to achieve a 99% kill. A low C.t value relates to an efective disinfectant. The relative speed of action of typical disinfectants, from most to least effective against microorganisms is ozone, chlorine dioxide, hypochlorous acid, hypochlorite ion, dichloramine and monochloramine. Increased speed of action means a shorter contact time is required, increasing the flexibility of the system. (NHMRC, NRMM, of Australia, 2011)

4.2.1. CHLORINE Chlorine was introduced as a water disinfectant early in the 20th century and still remains the major chemical in use for this purpose around the world. It is a strong disinfectant with excellent bactericidal properties, and is effective at short contact times. In water, chlorine reacts to form hypochlorous acid (HOCl), a very effective disinfectant. The hypochlorous acid dissociates to form hypochlorite ion (OCl–) which is estimated to be 150 to 300 times less effective as a disinfectant than hypochlorous acid. For effective chlorine disinfection, turbidity should be less than one nephelometric turbidity unit and pH should be less than 8 because the relative proportions of HOCl and OCl– in solution depend mainly on pH and, to a lesser extent, on temperature. Lower pH and temperature result in higher proportions of hypochlorous acid. For example, At 0°C and pH 7, 83% exists as HOCl and At 0°C and pH 8.5, only 14% exists as HOCl. Chlorine is suitable for water disinfection at all scales (from POU to Large scale) and commonly used either as calcium hypochlorite powder or sodium hypochlorite liquid (NHMRC, NRMM, of Australia, 2011). This is because these forms of free chlorine are relatively safe to handle, inexpensive and easy to dose (WHO, 2011).

Table1: Characteristics of waterborne patogens for there is some evidence of health significance related to their ocurrence in drinking water supplies and their resistance to chlorine disinfection.

The mechanisms by which chlorine inactivate waterborne pathogenic microorganisms include: (i) impairment of pathogen cellular function by destruction of major constituents (e.g. cell wall), (ii) interference with the pathogen cellular metabolic processes, and (iii) inhibition of pathogen growth by blockage of the synthesis of key cellular constituents (e.g. DNA, coenzymes and cell wall proteins) (Nora Savage, 2005) (S. Malato, 2009). Resistance to disinfection can vary depending on the type of microorganism and within different strains of the same species. In general, it can be said that microorganisms that form spores (or cysts) are more resistant to disinfection than those that do not. The order of resistance of microorganisms to chlorine disinfection is the following: bacteria < viruses < bacterial spores < helminths < protozoa (S. Malato, 2009). Table 1 provides general information of waterborne pathogens and their resistence to chlorine (WHO, 2011) (Michen, 2010). Provided the water has low turbidity 99% of bacteria (e.g. Campylobacter, Salmonella spp, Vibrio cholera, Shigella spp) can be killed with 0.08 mg per minute

per liter at 1-2 °C and neutral pH. In contrast, Legionella and non-tuberculous Mycobacteria resistance to disinfection is enhanced by their ability to survive intracellularly within amoebae (protozoa) and to grow in biofilms (NHMRC, NRMM, of Australia, 2011) (S. Malato, 2009). Viruses such as adenoviruses, enteroviruses, hepatitis viruses, noroviruses and rotaviruses are moderately resistant to chlorine disinfection (99% virus inactivation require 12 mg per minute per liter at 0-5 °C and neutral pH) and persistent in water supply systems (S. Malato, 2009). The most difficult targets for chlorine disinfection are the protozoa (e.g. Acanthamoeba, Cryptosporidium parvum, Cyclospora cayetanensis, Entamoeba histolytica, Giardia intestinalis, Naegleria fowleri), which are infectious at very low concentrations and are moderately - highly persistent in water systems. For example 99% of Giardia inactivation require 230 mg per min per liter at 0-5 °C and neutral pH and Cryptosporidium oocysts will not be killed by concentrations that can be practically used in drinking water. Therefore protozoa in drinking water systems are very hazardous and have to be considered in the choice of treatment (S. Malato, 2009) (Maryna Peter-Varbanets, 2010). Natural water contains inorganic and organic matter that reacts with chlorine. Hence, sufficient disinfectant must be used to satisfy this demand and still provide the required dose for disinfection (NHMRC, NRMM, of Australia, 2011). Reactions of chlorine with naturally occurring organic matter produce chlorinated by-products, which can compromise drinking water quality. Consequently, these compunds has prompted stricter regulations concerning the acceptable levels by the U.S. Environmental Protection Agency (S. Malato, 2009) (Venkata K. K. Upadhyayula, 2009) (Meng Nan Chong, 2010) (Michen, 2010).

4.2.2. CHLORINE DIOXIDE Chlorine dioxide has excellent biocidal activity, though for effective disinfection turbidity should be less than one nephelometric turbidity unit and pH should be less than 8. Chlorine dioxide is a more effective disinfectant than chlorine, although the overall kinetics of bacterial destruction appears to be similar (NHMRC, NRMM, of Australia, 2011). Chlorine dioxide is capable of inactivating most waterborne pathogens, including Cryptosporidium oocysts, with practical doses and contact times (Michen, 2010). Chlorine dioxide is a reactive gas that cannot be easily stored or transported, and must be generated on site. Hence, it’s not a suitable disinfectant for household systems. Futhermore, the oxidative end products of chlorine dioxide disinfection

have been found to be a health hazard (NHMRC, NRMM, of Australia, 2011).

4.2.3. CHLORAMINES Chloramines are formed when chlorine and ammonia are added to water, specifically: monochloramine, dichloramine and nitrogen trichloride (trichloramine). Dichloramine is a stronger disinfectant than monochloramine, but is less stable and has a distinct disagreeable odour. Nitrogen trichloride has an extremely offensive odour and is readily destroyed by sunlight. Though, chloramination has been used for disinfection at all scales (NHMRC, NRMM, of Australia, 2011). The biocidal activity of chloramines is slower than that of free chlorine, for example: Monochloramine (most commoly formed type) requires a contact time of many hours be effective. Chloramines are effective against bacteria but their activity against viruses and protozoa is poor. To achive an effective chloroamination turbidity should be less than one nephelometric turbidity unit and pH should range between 8–9 because low pH (together with a high ratio of chlorine to ammonia) favours the formation of the less stable chloramines (NHMRC, NRMM, of Australia, 2011).

4.2.4. OZONE Ozone is very strong biocide that can be used for disinfection at all scales if electricity is available. In order to achieve effective disinfection turbidity should be less than one nephelometric turbidity unit and the pH should be less than 8 because ozone is unstable above this pH value. Hydrogen peroxide is an effective disinfectant that is not used in largescale plants, but may be used in small-scale or emergency applications. Recently, hydrogen peroxide has been used in conjunction with ozone (peroxone) to provide a more effective oxidising agent (NHMRC, NRMM, of Australia, 2011). Ozonation is very effective in killing bacteria, (i.e. 99% of bacteria can be removed with 0.02 mg of ozone per min per liter at 5 °C and pH 6–7) and reasonably effective in inactivating viruses (depending on type) and many protozoa. For the disinfection of Cryptosporidium, the highest ozone concentration is needed: 40 mg per min per liter at 1 °C (S. Malato, 2009). As for central treatment, ozone for household water treatment must be generated on site. As a result, ozone is not recommended for household water treatment because of the need for a reliable source of electricity to generate it and proper dosing in a small application its relatively high cost (WHO, 2011) (Michen, 2010). Ozone reacts with natural organics to produce lower molecular weight compounds, which are more biodegradable and promote the growth of bacteria. Also, ozonation can lead to the formation of brominated trihalomethanes (THMs). However, the brominated THMs produced in

ozonation usually occur in lower concentrations, and constitute less of a problem, than chlorinated THMs produced by chlorination (NHMRC, NRMM, of Australia, 2011).

4.2.5. HETEROGENEOUS FENTON PROCESS Heterogeneous Fenton process is an advanced Oxidation Processes (AOPs). Although there are different reacting systems, all of them are characterized by the same feature: production of reactive oxygen species (ROS) as free radicals, which are able to oxidize and mineralize almost any organic molecule, yielding CO2 and inorganic ions (S. Malato, 2009) (Grigory Zelmanov, 2008). These processes are also characterized by their not-selective attack, which is a useful attribute for microbial control. In general, as the surface area of the catalyst increase, the generation of reactive oxygen species (ROS) increases in advanced oxidation processes. Hence, catalysts with larger surface area demonstrate stronger antimicrobial activity (Li, 2010) (Tamar Gordon, 2011). Iron-oxide minerals have been used as heterogeneous catalysts for the oxidative degradation of organic compounds through the heterogeneous Fenton reaction in the presence of H2O2 (E. G. Garrido-Ramirez, 2010). Iron oxides, used for water treatment, can be recovered and reused because they are practically insoluble in water. Furthermore, the Fenton-like reaction in the presence of H2O2 does not require strict control of pH as is the case in the homogeneous Fenton process, the drawback is generally diminished oxidation kinetics of the heterogeneous process (S. Malato, 2009). The Fenton-like reaction involves the generation of hydroxyl radicals, by decomposition of hydrogen peroxide on the surface of iron-oxide particles (E. G. Garrido-Ramirez, 2010). Initially hydrogen peroxide (H2O2) and Fe(3+)–OH groups at the iron oxide surface react, froming a surface complex (Fe(2+)O2H). The electronically excited state (Fe(2+)O2H) is deactivated through dissociation of the peroxide radical (O2H), the peroxide radical can immediately react with other compounds (e.g. organic pollutants). The reduced Fe(2+) in the particle surface can react with either hydrogen peroxide, water or dissolved oxygen returning to the initial state, i.e. Fe(3+)-OH (E. G. Garrido-Ramirez, 2010). The efficiency of iron oxide minerals in catalyzing the decomposition of the organic pollutants through the heterogenous Fenton reaction is influenced by several parameters, such as: hydrogen peroxide concentration, surface area and type of the iron oxide, solution pH, ionic strength, and pollutant characteristics. The rate of formation of radicals in iron oxide/hydrogen peroxide systems is proportional

to the product of the concentrations of surface area of the iron oxide and hydrogen peroxide (E. G. Garrido-Ramirez, 2010) (Grigory Zelmanov, 2008). Consequently, the use of iron oxide nanocatalysts is a promising alternative to enhance process effiency. Grygory Zelmanov et. al. used iron(3+) oxide-based nanocatalysts for degradation of organic pollutants. The authors showed that the heterogeneous fenton process using nanocatalysts have improve efficacy and kinetics when compare to homogeneous Fenton and photo-Fenton processes; with no degradation of the nanocatalyst (Grigory Zelmanov, 2008). With regard to oxide type, Fe(3+) oxides (hematite, goethite, lepidocrocite, and ferrihydrite) are catalytically less active than their Fe(2+) counterparts (magnetite and pyrite). At acid pH values, the oxidation process is mainly due to dissolution of iron oxides, promoting the homogeneous Fenton reaction. However, at near neutral and basic pH values, the solubility of iron-oxide minerals decreases, and hence the oxidation of organic compounds is mediated by the heterogeneous Fenton reaction. Under the heterogeneous process conditions, the electrostatic interactions between the catalyst surface and the organic contaminats become important, since surface-adsorbed organic compouds are readly accessible to oxidation (E. G. Garrido-Ramirez, 2010). The potential of heterogeneous (photo)Fenton processes for effective water disinfection using Fe(3+) oxides is supported by five main features: (i) capacity to adsorb waterborne pathogens (e.g. viruses), (ii) capacity to generate reactive oxygen species (ROS) wich are effective for inactivating most pathogens, (iii) capacity to generate charge carries upon solar light radiation, wich enhances ROS generation, (iv) availability and (v) resistance to degradation.

4.1. HEAT AND UV-BASED SYSTEMS 4.1.1. HEAT TREATMENT

Heat treatment, as applied by boiling of contaminated water, is perhaps the oldest way of disinfection and is still the most common means of treating water in the home. Heating water to even 55°C has been shown to kill or inactivate most waterborne pathogenic bacteria, viruses, helminths and protozoa.  A major disadvantage of boiling is its consumption of energy in relation to the availability, cost and sustainability of fuel.  Another problem of boiling is that it provides no residual protection: water can easily be recontaminated after cooling (Maryna Peter-Varbanets, 2010) (Michen, 2010) (WHO, 2011)

4.1.2. SODIS It is widely accepted that solar-UV inactivation of microbial cells occurs through a variety of mechanisms depending on the type of UV. Sunlight

consists mainly of UV-A and hence the main inactivation mechanism is a photooxidative process as well as the generation of reactive oxygen species. ROS generated upon UV irradiation include free radicals (e.g. OH- or O2+) as well as non-radicals (e.g. H2O2). ROS interactions with cells induce oxidation of proteins and membrane damage (S. Malato, 2009). Solar water disinfection (SODIS) is a simple technology for improving the microbiological quality of drinking water by the combined action of the UV-A radiation, oxidative activity associated with dissolved oxygen and heat treatment to destroy pathogenic microorganisms. The SODIS system consists of four basic steps: removing solids from highly turbid (>30 NTU) water by settling or filtration; placing low-turbidity water in clear PET bottles of 1–2 l volume; aerating the water by shaking it in contact with air and exposing the filled, aerated bottles to full sunlight for about 5-6 h.  This system is suitable for treating small volumes of water (<10 l) and has showed diarrheal disease reduction of about 30% (Maryna Peter-Varbanets, 2010) (Mark D. Sobsey, 2008) (WHO, 2011) (S. Malato, 2009) (Michen, 2010). Most gram-negative bacteria including E. Coli are of similar sensitivity to UV-A radiation. Enterobacter cloacae, the species of Shigella, and Salmonella enteritidis are relatively easy to inactivate in comparison to E. coli. In contrast, Salmonella typhimurium, vibrio cholera and Sh. Sonnei are slightly more resistant. Gram-positive bacteria, represented by Enterococci sp. (Enterococcus faecalis) and Bacillus subtilis, are also more difficult to disinfect. In fact, the solar disinfection technique is not effective for Bacillus subtilis spores. Virus show double resistance to solar radiation in comparison to E. Coli and Acanthamoeba Polyphaga is quite resistant to UV-A, even compared with other protozoa (S. Malato, 2009). It is important to emphasize that almost all works on the matter are concentrated in the reduction of a single microorganism, and do not study the real water disinfection in the presence of turbidity, chemical compounds and other microbial species (S. Malato, 2009) Solar disingfection (SODIS) as a drinking water treatment has important advantages to alternative treatments for house hold disinfection: (i) availability in low-income, sun-rich areas, (ii) acceptance due to natural odor and taste of the water, and (iii) sustainability as no chemicals are consumed. Nevertheless, different scientific and technical issues are still under research to overcome the limitations of this technique, such as: (i) the length of time required for microbial inactivation, (ii) the low volume of water generated, (iii) the UV resitance of some relevant

pathogens and (iv) the increase exposure time and disinfection capacity reduction due to high water turbidity (S. Malato, 2009).

4.1.3. UV-C Disinfection using UV-C irradiation (190 nm < λ < 290 nm) has found numerous applications in situations where lack of a disinfectant residual is not important, such as point-of-use disinfection. For POU water disinfection, most systems employ low-pressure mercury arc lamps producing monochromatic UV radiation at a wavelength of 254 nm and disinfection is related to the generation of reactive oxygen species (ROS) in the precense of dissolved oxygen (WHO, 2011) (NHMRC, NRMM, of Australia, 2011). UV-C lamps disinfection is very effective in killing bacteria and reasonably effective in inactivating viruses (depending on type) and many protozoa, including Giardia and Cryptosporidium (S. Malato, 2009). For clear water (turbidity < 1 NTU), 99% bacteria inactivation require UV-C dose is 7 mJ/cm2 and for Cryptosporidium is 5 mJ/cm2. The total number of microorganisms reduced by UV-C is very similar for protozoa and bacteria, because of their similar susceptibility (S. Malato, 2009) (Maryna Peter-Varbanets, 2010). Enteroviruses, Hepatitis viruses, Norovirus and Rotaviruses are sensitive to disinfection using UV-C light. In comparison, Adenoviruses are relatively resistant to UV-C light and the dose required for a 90% (1log) kill is 110mJ/cm2 (NHMRC, NRMM, of Australia, 2011). According to US-EPA the UV dose requirement for 4log removal of viruses is 186 mJ/cm2 (Michael V. Liga, 2011) (Michen, 2010). The main disadvantage of UV-C disinfection is the high dependence on source-water turbidity. In addition, the high doses require for virus inactivation significantly increases the energy demand, which translates into a higher treatment cost. (Michael V. Liga, 2011) (WHO, 2011) (Maryna Peter-Varbanets, 2010) (Michen, 2010). Another limitation is that microorganisms may become viable again in the presence of visible light (photoreactivation) if UV-C treatment is inadequate (NHMRC, NRMM, of Australia, 2011).

4.1.4. HETEROGENEOUS PHOTOCATALYSIS Heterogeneous photocatalysis is an advanced Oxidation Process (AOP). The basic principles of this method are briefly summarized bellow. Semiconductors (e.g., TiO2, ZnO, Fe2O3, WO3, and ZnS) can act as sensitizers for light-induced redox processes due to their electronic structure. Absorption of a photon with an energy hν greater or equal to the bandgap energy Eg generally leads to the formation of an electron/hole pair (S. Malato, 2009) (Qilin Li, 2008) (Meng Nan Chong, 2010)

(Anu Matilainen, 2010). These charge carriers subsequently either recombine and dissipate the input energy as heat, get trapped in metastable surface states, or react with electron donors and acceptors adsorbed on the surface or bound within the electrical double layer (S. Malato, 2009) (Meng Nan Chong, 2010). Many studies have examined the role of experimental variables on the response of microorganisms to photocatalytic treatment. These have included the effects of aeration, pH, photocatalyst type and concentration, light intensity and treatment time. While these experimental parameters have been shown to affect microbial response to photo inactivation, consideration must also be given to the nature of the organism targeted. The response of microorganisms to photocatalytic treatment has been reported to follow the order of protozoa being the most resistant followed by bacterial spores, mycobacteria, viruses, fungi and bacteria (Peter K. J. Robertson, 2012). Furthermore the difference in cell wall complexity between gram positive and gram-negative bacteria has also been shown to affect treatment time. Gram-positive bacteria require a longer treatment time than gram-negative bacteria (Peter K. J. Robertson, 2012). The most important features that anable this process for water disinfection are: (i) the reaction take place at ambient temperature and pressure, (ii) the oxygen necessary for the reaction can be directly obtained from atmosphere and (iii) the energy required can be obtained from the sun (S. Malato, 2009) (Meng Nan Chong, 2010). However, the main issues limiting the wide scale adoption of photocatalysis as a practical water disinfection process are the lack of commercially available visible light absorbing photocatalyst materials and the need for catalyst recovery or immobilization to avoid drinking water contamination (Peter K. J. Robertson, 2012) (Grigory Zelmanov, 2008).

4.1.4.1. TiO2 HETEROGENEOUS PHOTOCATALYSIS During the last decade, titanium dioxide (TiO2) nanoparticles have emerged as promising photocatalysts for water disinfection, because of it’s high phtocatalytic activity, strong resistance to chemical breakdown and photocorrosion, low cost, non-toxicity and capacity to absorb solar-UV light (Nora Savage, 2005) (Qilin Li, 2008) (S. Malato, 2009). In addition, industrial scale water purification systems using TiO2 photocatalysis already exist (e.g. Purifics®), although their main application is organic contaminant degradation instead of disinfection (Nora Savage, 2005) (Qilin Li, 2008). Basically, TiO2 is a broad bandgap semiconductor material that is able to convert energy from light into chemical redox energy. The two most important and

easily synthesized polymorphs of TiO2 are rutile, with a band gap of 3.05 eV (λ = 407 nm), and anatase, with a band gap of 3.26 eV (λ = 381 nm); anatase is more active in photocatalysis although the physical causes are not completely understood so far. Upon irradiation of UV light, two types of photochemical reactions occur on the TiO2 surface, photo-induced redox reaction of the adsorbed substances, and photo-induced hydrophilic conversion of TiO2 itself. In the first reaction, photoelectrons (e-) are generated in the conduction band and holes (h+) are produced in the valence band. The photoelectrons then react with molecular oxygen (O2) to produce superoxide radical anions, and the holes react with water to produce hydroxyl radicals. Finally, these two types of radicals work together to decompose adsorbed organic contaminants (Ying Hui Ngo, 2011). Conversely, in the second reaction, photoelectrons tend to reduce the Ti (4+) cations to the Ti (3+) state, whilst the holes oxidize the O(2−) anions, creating oxygen vacancies. Water molecules that render the surface hydrophilic then occupy oxygen vacancies. The bactericidal properties of TiO2 photocatalysis are commonly related to the first reaction (Peter K. J. Robertson, 2012)(Anu Matilainen, 2010) (Qilin Li, 2008) (Shaily Mahendra, 2009) (Ying Hui Ngo, 2011). While the second reaction can generate hydrofillic repulsion between the TiO2 and pathogenic microorganisms such as non-enveloped viruses, hence reducing adsorption and inactivation capacity of TiO2. TiO2 can kill both Gram-negative and Gram-positive bacteria, although Gram-positive bacteria are less sensitive due to their ability to form spores (seed like bacterial form). The concentration of TiO2 usually required to kill bacteria varies between 100 and 1000 mg/l, depending on the size of the particles and the intensity and wavelength of the light used (Qilin Li, 2008). As mention before solar light has a bactericidal effect, the addition of TiO2 in the presence of solar radiation inactivates microorganisms faster than in the absence of catalyst. (S. Malato, 2009). For example, complete inactivation of fecal coliforms was achieved in 15 min at an initial concentration of 3000 cfu/100 mL by exposing water in TiO2-coated plastic containers to sunlight whereas the same inactivation required 60 min with uncoated containers (Qilin Li, 2008) (Shaily Mahendra, 2009). In addition, microorganisms that are very resistant to UV-A irradiation like Enterobacter cloacae have been successfully inactivated by TiO2 photocatalysis (S. Malato, 2009). However, it has been reported that when TiO2 disinfection (total coliforms and Streptococcus faecalis) is applied to natural water (pH=7.8), no

significant difference between TiO2- photocatalysis and direct solar or UV-C lamp light irradiation is found (S. Malato, 2009). UV-C can kill protozoan cysts such as Giardia and Cryptosporidium, but UV alone is relatively ineffective against viruses unless the contact time and energy output are significantly increased (Shaily Mahendra, 2009). The higher resitance of UV-C of viruses in comparison with both bacteria and protozoa has been attributed to the simpler virus structure, such as greater integrity of the viral capsid and no reliance of viability on chemically fragile enzymes, thus permitting susceptibility only to hydroxyl radicals and not to the less reactive super-oxdie radicals and H2O2 (Min Cho, 2011). However, nano-sized TiO2 has been reported to kill viruses including poliovirus, hepatitis B virus, Herpes simplex virus, and MS2 bacteriophage (Qilin Li, 2008). A recent contribution also demonstrated the potential of TiO2-photocatalysis to inactivate prions (scrapie strain 263K) (S. Malato, 2009). Besides the already mention limitations of UV-based disinfection systems (e.g. turbidity), TiO2-photocatalysis has to deal with other thecnical challenges for widespread application, such as effective catalyst dispersion in order to provide a large surface area available for reaction and catalyst recovery, along with development of catalysts with broader photoactivity range and higher quantum efficiency (e.g. reduced charge carrier recombination).

4.1.4.2. ZnO HETEROGENEOUS PHOTACATALYSIS Among the functional oxides, ZnO is unique because of its twofold semiconducting and piezoelectric properties and its low toxicity render it suitable for personal care applications (e.g UV screening). Further, it is a bactericide effective against both Gram-positive and Gram-negative bacteria (C. Karunakaran V. R., 2011) (Agatino Di Paola, 2012). ZnO (band gap of 3.0 eV) displays n-type conductivity because of its native defects such as zinc interstitials and oxygen vacancies. Silver doping is found to be effective for the fabrication of p-type ZnO (C. Karunakaran V. R., 2011). The suggested mechanism for the antibacterial activity of ZnO is based mainly on photocatalytic formation of reactive oxygen species (ROS) from water and oxygen (Tamar Gordon, 2011) (Nasrin Talebian, 2011). However, Ken Hirota et. al. developed porous ZnO ceramics with sustainable antibacterial activity even in the dark sunshade; the antibacterial activity was attributed to the generation of super-oxide anion (Ken Hirota, 2010). The bacterial surface is negatively charged due to the predomination of the polysaccharides over the

amide. Therefore, bacteria can be easily adsorbed on the ZnO surface (as the surface of other metal oxides with IEPs > 8); wich enables disruption of the bacterial cell membrane by ROS attack (Nasrin Talebian, 2011) (Qilin Li, 2008). Furthermore, adsorbed virus ROS-induced inactivation can be accomplished by viral capsid protein oxidation and/or formation of carbonyl groups on viral capsid proteins (Li, 2010). In many cases, ZnO has been reported to be more efficient than TiO2 but the occurrence of photocorrosion and the susceptibility of ZnO to facile dissolution at extreme pH values, have limited its application in photocatalysis (Qilin Li, 2008) (Agatino Di Paola, 2012). In addition, enhancement of stability against agglomeration, charge carrier separation and visible light adsorption of the catlysts are required for practical application of ZnO in water disinfection (W. G. Xu, 2010).

4.1.4.3. Fe2O3 HETEROGENEOUS PHOTOCATALYSIS Iron oxide nanomaterials can be good photocatalysts absorbing visible light. Compared with commonly applied TiO2, which mainly absorbs UV light with wavelengths of 380 nm (covering only 5% of the solar spectrum), α-Fe2O3 with band-gap of 2.2 eV is a suitable candidate for photocatalysis under visible light. Besids α-Fe2O3 other species of Fe(3+) oxides have been proposed, such as, γ-Fe2O3, α-FeOOH, β-FeOOH and γ-FeOOH, to be used in heterogeneous photocatalysis (Piao Xu, 2012). All mentioned Iron (3+) oxides are capable of producing ROS when irradiated with sunlight. Furthermore, the direct contact of the adsorbed microorganisms (e.g. viruses) with the iron-oxide surface places them directly adjacent to the source of the inactivating oxidant. This physical proximity would be expected to greatly increase inactivation. (Brian M. Pecson, 2012). Nevertheless low photocatalytic activity is frequently encountered because the electron–hole charge recombination on the iron oxide surface, is fast (Piao Xu, 2012) (Agatino Di Paola, 2012). Deposition of noble metals has been employed to address this problem (Piao Xu, 2012) as well as other strategies used for TiO2.

4.3. PHYSICAL SYSTEMS 4.3.1. COAGULATION OR PRECIPITATION

Coagulation or precipitation methods employ natural or chemical materials to coagulate or precipitate suspended particles, including microorganisms, to enhance their sedimentation. These methods may be used along with cloth or fibre media for a straining step to remove the large particles (flocs) formed (WHO, 2011) (NHMRC, NRMM, of Australia, 2011). The coagulation and sedimentation remove a maximum of 90% of bacteria, 70% of viruses and 90% of protozoa (S. Malato, 2009). This technology requires trained personnel to control the process,

which often relies on pH adjustment. This, along with the costs of the coagulant or precipitant, makes this technology expensive for household systems (Michen, 2010)

4.3.2. FILTRATION One great advantage of filtration, if compared to strong oxidant- and UV-based treatments, is the simultaneous and rapid removal of turbidity and microorganisms in a one-stage process. In filtration, two major processes may be distinguished, namely surface filtration and depth filtration. Surface filtration relies to a great extend on the sieving effect, which is the physical straining of particles that are larger than the pore dimension of the filter media. In contrast, depth filtration corresponds to a filter media of increased thickness, and hence increases the path length of contaminants in the filter media. Depth filters are often composed of granular or fibrous materials which further elongate the path length and force contaminants to travel around obstacles where they can be trapped by various transport and surface forces (Michen, 2010). Most household filter technologies operate by gravity flow or by water pressure provided from a piped supply (WHO, 2011). The quantification of retention performance can be obtained either by fractional retention (R) or the log reduction value (LRV) as given in equations (eq. 1) and (eq. 2), respectively. Therefore, the concentration of the component of interest (i) is determined experimentally in the influent water (Ci0) and in the effluent (Ci). (Michen, 2010)

𝑅! = 1 − !!!!!

(eq. 1)

𝐿𝑅𝑉! = −𝑙𝑜𝑔!" 1 − 𝑅! (eq.2) Further relevant parameters that characterise the performance of a filtration process are: the flow rate 𝑉 , which refers to a certain effluent volume (ΔV)

per unit time (Δt); the flux 𝐽 , taking into account the outer filter (or membrane) surface (AFS); and the porosity (ε), which is the volume pore fraction (VP) divided by the total volume of the filtration media (VF) (Michen, 2010).

𝑉 = ∆!∆!

(eq. 3)

𝐽 = !!!"

(eq. 4)

𝜀 = !!!!

(eq. 5) 4.3.2.1. SURFACE FILTRATION

In general, surface filtration processes rely on a driving force that is used to separate two phases. The driving force can be a difference in pressure, concentration, temperature or electric potential. Most surface filtration processes are pressure-driven and employ membrane filters. The separation range of different membrane processes and the corresponding applied pressure ranges for operation

are shown in Fig. 1. Regarding the disifection of drinking water, it is important to assess membrane technologies in relation to waterborne microorganisms such as viruses, bacteria and protozoa. (Michen, 2010)

Figure 1: Separation range of membrane processes in relation to waterborne patogens.

Microfiltration is a standard process for removing particles in drinking water. It is generally assumed that filters with a 0.1 - 0.45 µm pore size retain bacteria. In contrast to this assumption, it has been regularly observed the passage of bacteria through 0.45, 0.22 and even 0.1 µm pore size filters. Bacteria with dimensions of 0.25 to 0.3 µm in diameter and 0.6 to 1.0 µm in length were found to penetrate through 0.22 µm membranes. This phenomenon can be explained due to the availability of some openings in the membrane larger than the nominal pore size, imperfections or defective installation (Michen, 2010). The pore size of ultrafiltration membranes is small enough to ensure high log-removal of microbiological hazards such as Cryptosporidia, Giardia and total bacterial counts. Substantial virus removal can be attained with ultrafiltration membranes since the size of viruses is in the range of the pore dimensions. However, the retention of small species, in particular enteric viruses, may not be complete due to membrane imperfections. Also, the retention of viruses with such membranes might be in part attributed to adsorption processes and thus depend on solution composition and operation conditions. When virus removal needs to be ensured, membranes with smaller pore sizes are preferred which rely on a higher operating pressure and thus higher energy consumption. (Michen, 2010). Nanofiltration and reverse osmosis can be used to

remove all kinds of microorganisms, as well as inorganic contaminants from water (Michen, 2010). The main limitation of membrane systems is membrane fouling (the establishment of a growing filter cake and/or the clogging of pores). Fouling prevention measures for microfiltration and ultrafiltration usually include regular backflushing and cleaning. In case of ultrafiltration, nanofiltration and reverse osmosis, pre-treatment is usually used, and the systems are operated in a cross-flow mode. Such fouling prevention measures require automated process control and regulation (Michen, 2010). Another relevant drawback particularly characteristic of surface filtration techniques for reliable virus removal (i.e. nanofiltration and reverse osmosis) is low flowrates (Markus Wegmann B. M., 2008). Mentioned issues results in increased investment costs and make trained personnel a requierement, thus membrane technology not suited for POU water treatment in less developed nations (Michen, 2010).

4.3.2.1.1. SURFACE FILTER BIOFOULING PREVENTION It is now accepted that microbial populations use cell attachment to solid substrates to survive, forming structured communities called biofilms. Biofilms are defined as biopolymer matrix-enclosed microbial populations adhering to each other and/or surfaces. Consequently, two bacterial living states exist in natural environments: planktonic (free floating) and sessile (attached). While the planktonic growth mode is important for the bacterial spread, biofilms are necessary to allow bacteria to persist and to resist adverse environmental conditions. Although several mechanisms have been postulated to explain the reduced susceptibility of sessile organisms to antimicrobials, it is becoming evident that biofilm resistance is multifactorial. Poor antimicrobial penetration, nutrient limitation, slow growth, adaptive stress responses and the formation of multi-resisting cells are hypothesized to constitute a multi-layered defense. (K. Glinel, 2012). The formation and development of the biofilm follows steps that may vary depending on the biological characteristics of the bacteria present. Nevertheless, four common steps are usually distinguished. In the first step a bacterium is brought into contact with the surface. The second step of biofilm formation consists of adhesion of the bacteria to the surface. Most authors consider that this is the step most influenced by topographical and chemical features of the surface. The third step consists of proliferation of the adherent bacteria and, in most cases, synthesis of the biofilm matrix. Thus the biofilm becomes mature, and is associated with a specific bacterial metabolism and physiology (fourth step). Within this community bacteria interact with each other through membrane organelles

and biochemical signals. This allows, for example, the lysis of biofilm matrix molecules necessary for the release of free bacteria into the surroundings (K. Anselme, 2010). Both titanium dioxide (TiO2) and Silver nanoparticles have been used for fabricating membrane filters in order to prevent biofouling (Jeonghwan Kim, 2010) (Law Yong Ng, 2010). SURFACE FILTER BIOFOULING PREVENTION USING TIO2 Heterogeneous photocatalysis with TiO2 is particularly applied for avoiding and/or remediation of biofouling of membrane filters under UV radiation (Qilin Li, 2008) (Jeonghwan Kim, 2010). The use of TiO2 has encountered various technical problems such as membrane structure deterioration, low photocatalytic activity and loss of deposited TiO2 layer over time (Meng Nan Chong, 2010). In addition, the preparation (e.g. by dipping, slip-casting or electrophoretic deposition) of ceramic membrane derived from TiO2 nanoparticles with out defects has also been found to be more difficult than that of other ceramic materials (Jeonghwan Kim, 2010). SURFACE FILTER BIOFOULING PREVENTION USING SILVER Silver have also been utilized to avoid biofouling in membrane filters, it is suggested that biocidal properties of silver loaded membranes arise due to the release of Ag+ ions. In addition to improvement in the bio-resistance of the membrane, silver decorated membrane filters have been observed to exhibit higher permeability. (T. Pradeep, 2009) (Jeonghwan Kim, 2010). The combine performace of TiO2-Silver membrane has been tested for E. coli inactivation; a 3-4 LRV removal was achieved under UV illumination compared to dark conditions (1-2 LRV). However, the permeate flux decline pattern of membrane was very similar regardless of the removal efficiency of E. coli. (Jeonghwan Kim, 2010). It has also been shown that addiiton of Ag (0.9 wt%) nanoparticles significantly decreased the number of Escherichia coli grown on the membrane surface after filtration, as indicated by the number of colony forming units (CFU) per area of membrane as compared with membranes with no Ag addition (Law Yong Ng, 2010) (Jeonghwan Kim, 2010). When silver nanoparticles are absorbed onto the surface of porous (ceramic) membrane by a weak force such as the Van der Waals attraction, they can be easily washed away when the water flows pass the channels (Jeonghwan Kim, 2010). In addition, silver from the membrane surface is also lost after membrane-bacteria and membrane-virus interactions in ionic form (Law Yong Ng, 2010). Silver lost generates membrane performance to drop over time, due to the lost of the biofouling resistance capacity. Moreover, release of silver nanoparticles can pose the danger of water contamination when membranes

are used for treatment of drinking water (Law Yong Ng, 2010).

4.3.2.2. DEPTH FILTRATION Depth filtration has been established in various techniques, for example non-consolidated beds of sand, diatomaceous earth or activated carbon, fibrous materials in a woven or non-woven matrix or consolidated porous ceramic structures (Michen, 2010). In depth filtration contaminated water is passed through a filter medium with micron-scale open porosity, and retain microbes by a combination of physical and chemical processes, including physical straining, natural die-off, biological action and adsorption. Adsorbtion on the pore walls takes place under the main influence of van derWaals, electrostatic forces and hydrophobic interactions. Because of the relatively coarse porosity (compared to the filtration membranes working on the straining principle), the flowrates are high at low applied pressures, and the system cost is reduced because of the less demanding materials processing required and the absence of high-pressure pumps (Markus Wegmann B. M., 2008) (WHO, 2011). Some depth filters may also use antimicrobial or bacteriostatic surfaces (e.g. silver decorated). Others are biologically active because they develop biofilms able to retain microbes, which often leads to their inactivation and  biodegradation (WHO, 2011).

4.3.2.2.1. SAND FILTER Sand filters remove pathogens and turbidity mainly by biological action and straining. The sand filter itself was designed for water treatment on large scale; however, it has been adapted for use in the home and is known as biosand filtration. The biosand filter (BSF) is an intermittently operated sand filter with top sand layer covered in a biofilm. The biofilm is credited with the enhanced removal of suspended solids and pathogens through mechanical trapping, adsorption, predation, and natural death. Biofilm development reduces flow rate; though, water can be forced through the BSF by gravity alone and consequently these filters do not rely on any energy supply. (Ian Bradley, 2011) (Michen, 2010) (Maryna Peter-Varbanets, 2010) (Mark D. Sobsey, 2008). Field studies suggest biosand filters are able to overcome sustainability obstacles by requiring only one-time purchase and producing sufficient water for daily household use. In addition, this type of filter have been shown to improve water quality and reduce diarrheal disease in rigorous epidemiological studies, and follow-up studies document sustained, effective performance long after implementation, with filter usage rates remaining high years post-implementation. Biosand filter technologies have also shown the potential for large-scale adoption, as they

are use by over 1.5 million people (Mark D. Sobsey, 2008). Laboratory and field studies have documented improved microbiological water quality through the use of the BSF. Due to their small size, protozoan (oo)cysts can challenge removal by sand filtration processes; however, well-designed and properly operated systems can provide removal levels above 99.9% of Giardia cysts and Cryptosporidium oocysts. While, bacterial concentrations are reduced between 70-99.99%, depending on biofilm development and time of sampling (Ian Bradley, 2011) (NHMRC, NRMM, of Australia, 2011). The critical shortcoming of BSFs is they are not effective for removing viruses (Ian Bradley, 2011) (Michen, 2010) (Maryna Peter-Varbanets, 2010) (Mark D. Sobsey, 2008). In natural water conditions of pH 6-8, sand and most viruses are negatively charged, causing a net repulsion and reducing virus removal efficacy by sand filtration. Thus, water contaminated with pathogenic viruses is not yet potable after passing through sand filters (Ian Bradley, 2011).

4.3.2.2.2. ACTIVATED CARBON Activated carbon is a widely used to remove chemical contaminants such as pesticides and cyanobacterial toxins from water (NHMRC, NRMM, of Australia, 2011). POU systems exclusively comprising powdered activated carbon (PAC) filters tested for removal of bacterial pathogens from raw water had the problem of leakage of microorganisms in the effluent water. Unlike chemical contaminants, most biological adsorbates (protozoa, bacteria and virus) have sizes larger than the pore size of the microporous activated carbon. This size difference makes majority of pore surface area inaccessible to microorganisms thereby limiting their removal. (Venkata K. K. Upadhyayula, 2009). Activated carbon virus removal from water has been shown range from 9 to 75% (Michen, 2010).  Consequently, these filters are not recomended for microorganism control.

4.3.2.2.3. PAPER, FIBER AND FABRIC FILTERS Paper, fiber or fabric filters may be applied for household water treatment. They can be effective in the removal of larger water-borne pathogens such as schistosomes and Faciola species, guinea worm (Dracunculus medinensis) larvae, and bacterial pathogens associated with (can survive within) relatively large copepods, such as the bacterium Vibrio cholerae (WHO, 2011). However, these filters are not recommended for the general disinfection of household water because their pores are too large to significantly retain viruses, bacteria and smaller protozoan parasites (Maryna Peter-Varbanets, 2010) (WHO, 2011). More recently nanofibers are been used to fabricate surface filters. The microfiltration fiber filters (see section 4.3.2.1)

are of particular interest as they show high microorganisms removal capacity at low transmembrane pressure (<300 kPa) (Meng Nan Chong, 2010).

4.3.2.2.4. CERAMIC FILTER Porous ceramic (i.e. fired clay) media are used to filter large microorganisms (e.g. Protozoa) from drinking water by size exclusion (Mark D. Sobsey, 2008). While, smaller size microorganisms (e.g. bacteria and viruses) are retain by a combination of the previously mention principles active on depth filters. As sand filters, field studies suggest ceramic filters are able to overcome sustainability obstacles by requiring only one-time purchase, producing sufficient water for daily household use, and achieving large scale adoption and continued use (Mark D. Sobsey, 2008). In addition, ceramic water filters provide the highest prevention of diarrhoeal diseases (Michen, 2010). Field studies in which ceramic water filters were distributed to groups in communities in Bolivia, Colombia and Cambodia showed about a 50% reduction in diarrhoea when compared to the control groups. Alternative routes of infection and/or passage of microorganisms might explain this partial illness reduction (Michen, 2010) (Venkata K. K. Upadhyayula, 2009). Ceramic filters safely remove large-sized microorganisms of the group protozoa like Cryptosporidium parvum oocyst, Giardia lamblia, Giardia muris, Tetrahymeny and Paramecium in excess of 99.99 % (US-EPA regulations stablish a minimum removal of 99.9% for giardia cysts). Removal of bacteria appears scattered among studies. When filters were subjected to tests in which relatively small volumes of water were passed through the filter, the bacteria removal could be considered high (up to > 99.9999%) and most often no single bacterium could be detected in the effluent. However, when filters were challenged with bacteria for a longer period, the retention performance in most studies dropped down to values ranging from 50 to 99.9% (Michen, 2010). Indicating that they do not constitute an absolute physical barrier for these microbes. Viruses are the least studied microorganisms and showed large scatter in removal ranging from 90% up to 99.999%. Viruses are not expected to be retained by straining in these filters and removal may be attributed to adsorption mechanisms (Michen, 2010). As for biosand filters the main drawback of ceramic depth filters its virus retention. Hence, the challenge is thus to develop a filter medium which will attract and trap virus particles at its surface and which possesses the largest possible adsorber area (Markus Wegmann B. M., 2008). Furthermore, bacteria removal capacity should also be improved.

4.3.2.3. ENHANCEMENT OF BACTERIA REMOVAL IN DEPTH FILTRATION

Bacteria removal capacity of depth filters (without bioactive layers) has been improve mainly by addition of bactericidal materials. Silver-based nanoparticles has been intensively investigated because they effectively eliminate bacteria at relatively low concentrations; concentrations that may be considered not toxic for human cells (Qilin Li, 2008) (Robert Prucek, 2011). Silver is broad spectrum and toxic (to varying degrees) to numerous strains of bacteria, fungi, algae, and possibly some viruses. In addition, silver is reasonably effective at penetrating biofilms, which has been a drawback identify for other antibacterial materials (Duncan, 2011) (Wenning Shen, 2011) (Amparo Llorens, 2012). However, bacterial strains that manifest or develop silver-resistance are known (Duncan, 2011). Recently, it was showed that the bactericidal property of silver is related to its valence states, the higher the valence state the stronger the bactericidal property (Wenning Shen, 2011). This is probably because surface oxidation and ionic exchange are both mandatory to achieve the release of silver ions (Amparo Llorens, 2012) (Ying Hui Ngo, 2011). Based on reported data the bactericidal capacity of silver nanomaterials are ordered in the following sequence: AgO > Ag2O > Ag. Therefore, AgO is promising antimicrobial material. Though the oxide low decomposition temperature (280°C) may limit its use in depth filters. Several antibacterial mechanisms of silver have been postulated based on the release of silver ions, such as: (i) interaction with cytoplasmatic components such as ribosomes inhibiting the expression of enzymes, (ii) prevention of DNA replication, (iii) protein inactivation via interation with sulfur hindering the respiratory chain and the mechanisms related to energy production in the cell and (iv) deregulation of the homeostatic equilibrium through induction of a massive proton (Na, H, Ca) leakage through the membrane. In adition, ROS (i.e. oxygen superoxide) formation can take place at the silver surface (Qilin Li, 2008) (T. Pradeep, 2009) (Law Yong Ng, 2010) (Jean-Yves Bottero, 2011) (Ying Hui Ngo, 2011) (Amparo Llorens, 2012). Similarly as for bacteria the antiviral properties of silver ions involve the interaction with viral DNA and thiol groups in proteins (Qilin Li, 2008) (Law Yong Ng, 2010). Silver nanoparticles have been found to be potent agents against numerous species of bacteria, including: E. coli, Enterococcus faecalis, Staphylococcus aureus and epidermidis, Vibrio cholerae, Pseudomonas, Shigella flexneri, Bacillus, Proteus mirabilis, Salmonella enterica Typhimurium,

Micrococcus luteus, Listeria monocytogenes and Klebsiella pneumoniae. Silver nanoparticles are also toxic to fungi: Candida albicans, Aspergillus niger, Trichophyton mentagrophytes, and algae (Nora Savage, 2005) (Duncan, 2011) (Wenning Shen, 2011) (Jean-Yves Bottero, 2011). However, silver is a weak biocide that shows limited inactivation capacity, since viruses, protozoan cysts and bacterial spores, are not inactivated at silver concentrations employed for point-of-use drinking water treatment (Michen, 2010) (Li, 2010) (NHMRC, NRMM, of Australia, 2011) Silver nanoparticles are used as antimicrobial agent in over 100 consumer products, among these commercial home water purification systems such as Aquapure®, Kinetico®, and QSI-Nano®, which are reported to remove 99.99% (4LRV) of pathogenic bacteria, viruses and protozoan cysts (Qilin Li, 2008) (Shaily Mahendra, 2009). Nevertheless, it has been indicated that for example silver pre-coated activated carbon filters, used for additional drinking water treatment at the point of use have only a limited operating life and are relatively high cost (Maryna Peter-Varbanets, 2010). In addition, the potential toxicity of Ag nanoparticles to humans is still a matter of considerable debate. Consequently, strong binding forces between Ag nanoparticles and substrate are desirable to reduce possible exposure (Ying Hui Ngo, 2011).

4.3.2.4. ENHANCEMENT OF VIRUS REMOVAL IN DEPTH FILTRATION

Viruses are the smallest and exhibit the highest mobility among pathogens and hence are more difficult to remove by physical processes such as sedimentation and filtration than bacteria and protozoan cysts (WHO, 2011) (Youwen You, 2005) (Brian M. Pecson, 2012) (Li, 2010) (Chunjian Shi, 2012). Filtration can reduce viral particles concentrations by 90% or more depending on effectiveness of operation and pore size. Though, it is generally only in complex multistage treatments that viruses can be removed with efficiency in excess of 99.9% (Markus Wegmann B. M., 2008) (NHMRC, NRMM, of Australia, 2011). Therefore, removal of viruses from water using filters with pore sizes larger than the viral dimension represents a particular technological challenge. In addition, to ensure effective disinfection by depth filtration, viruses must either be inactivated or irreversibly attached to correctly inmovilized adsorbent material across the water quality conditions that might be encountered (Brian M. Pecson, 2012). In order to be suitable for practical applications, a depth filter must exhibit a flowrate of 60 l/h at an operating pressure of 3 bar. Furthermore, as a prerequisite to being certified as a true virus

filter by the United States Environmental Protection Agency (U.S. EPA), the unit must remove in excess of 99.99% of viral particles (25 nm diameter) from feed water containing 107 plaque-forming units (PFU) per liter between pH 5 and 9 (Markus Wegmann B. M., 2008).

4.3.2.4.1. DEPOSITION The deposition of contaminants generally involves two steps, namely transport and attachment. Transport of submicron (Brownian) particles to a solid surface in a porous media is controlled by convection and diffusion, while that of larger (non-Brownian) particles is control by forces arising from gravity and fluid drag. Thus, transport depends on particle size and flow rate. On the other hand interfacial forces influence the attachment efficiency factor, which is a ratio describing the fraction of adhering particles to the filter media. An attachment efficiency of unity implies that every collision leads to permanent attachment. This condition should be achieved when surface properties are adjusted to eliminate repulsion. Repulsive forces can reduce the attachment efficiency by many orders of magnitude, such that practically no deposition occurs. Hence, the attachment (or adsorption) of particles is dominated by colloidal interactions in the vicinity of the filter surface, where various long- and short-range forces between the adsorbent and the adsorbate determine either a net attractive or repulsive interaction. (Michen, 2010)

4.3.2.4.2. DLVO FORCES Named after Derjaguin, Landau, Verwey and Overbeek, in the DLVO theory interaction between two surfaces is assumed to be composed of two main contributions the Van der Waals force and the electrostatic double layer force. The Van der Waals attractive interactions between bodies of material, arising from the interaction of oscillating dipoles in the interatomic bonds of each body, manifest themselves in various aspects of behaviour ranging from the determination of surface energies, and consequently wetting behaviour, to the stability of colloidal suspensions. The Hamaker constant (AH) is a specific quantity used to represent these interactions for a given system. AH can be determined either by force measurements between the bodies (using for example an Atomic Force Microscope) or by calculations based on physical properties of the materials (Michen, 2010). The electrostatic double layer force is originated by the material surface charge. At the interface between a material and the environment, the surface Gibbs energy differs when compared to the interior conditions of the bulk. In order to minimise the free energy, atoms or ions at the surface may undergo reconstruction (i.e. rearrangement of surface

atoms/ions) or relaxation (i.e. contraction or expansion of atoms/ions relative to their surface plane). Also, the adsorption of foreign atoms or ions can minimise the surface free energy, leading to a new surface structure (Michen, 2010) When solids, such as metal oxides (MO) are in contact with water, the surface of the solid adsorbs water molecules to minimise the surface energy. The protonation and deprotonation of –MOH surface groups leads to an electrically charged surface that depends on the degree of ionisation (proton transfer) and consequently on the pH of the environment. The pH-value at which the Z-potential switches its sign is referred to as the isoelectric point (IEP) and is a characteristic parameter of the material in equilibrium with its environmental water chemistry. At pH values below the IEP, the surface is positively charged, indicating an excess of surface protons, whereas above the IEP the surface is negatively charged (Michen, 2010). Although particles in aqueous media bear an electric charge, the electric state of a surface depends on the spatial distribution of free charges in its neighbourhood. Theoretical models of charged surface-water interfaces were first introduced with the pioneering work of Gouy, Chapman and Stern. Such models are combined in the term electrical double layer (EDL). In Gouy-Chapman-Stern model, the surface hydroxyl groups determine the 0-plane as sketched in Fig. 2a. Charged species may be strongly bound at the α-plane within the Stern layer or weakly bonded in the β-plane outside the Stern layer. The δ-plane opens up the diffuse layer where the localisation of charges is balanced by electrostatic and thermal forces. Thereby, the electrical potential (U), decays exponentially with distance from the δ-plane (Michen, 2010). The development of U as a function of distance (x) from the MO surface is sketched in Figure 2b. The sketch also presents the thickness of the EDL, which is defined as the diffuse layer. The diffuse layer, also known as the inverse Debye length, (1/κ), is the distance at which the Stern potential UStern has fallen to the value 1/e (Euler’s number). κ is the Debye-Hückel parameter (Michen, 2010). When two surfaces approach each other the EDL is disturbed and the force rising from such perturbation is known as the electrical double layer force.

Figure 2: (a) Electric double layer scheme. (b) Electrical potential as a function of distance.

ISOELECTRIC POINT OF METAL OXIDES Kosmulski listed IEP values reported in the literature, and have discussed the differences reported in IEPs (Kosmulski, The pH-dependent surface charge and the points of zero charge, 2002) (Kosmulski, pH-dependent surface charging and points of zero charge II. Update, 2004) (Kosmulski, pH-dependent surface charging and points of zero charge. IV. Update, 2009) (Kosmulski, The pH-dependent surface charging and points of zero charge V. Update, 2011). Table 2 present IEP and IEP range of selected metal oxides.

Table 2: Isoelectric point of selected metal oxides.

ISOELECTRIC POINT OF VIRUSES Viruses as well as metal oxides possess a surface charge in aqueus media. In the case of non-enveloped viruses, the functional groups of the coat protein determine to a great extent the net surface charge of the virus. This electrical charge has an impact on the mobility and sorption behavior of the viral particles. Accordingly, the IEP is a crucial value that controls virus deposition onto solid surfaces (Michen, 2010) (Li, 2010).

Figure 3: Summary of the results of the review of virus IEPs performed by Michen and Graule (B. Michen, 2010).

In table 1 (see section 4.2.1) the IEPs of waterborne pathogenic viruses for wich there is some evidence of health significance related to their ocurrence in drinking water supplies according to (WHO, 2011) are included. The Gaussian function fit in figure 3 reveals a mean value of 5.0 and a standard deviation of 1.3. Despite the statistical relevance of the data presented in figure 3 the following issues should be

taken into account when IEP values from literature are used to explain virus adsorption behaviour: (i) IEP values reported in literature showed a deviation of up to 4.8 pH units in the IEP of a single virus species, (ii) the co-existence of two IEPs within a single enterovirus stock has been confirmed and (iii) localized regions on the virus surface can vary widely in charge, as the surface is made up of multiple protein subunits (Michen, 2010) (Li, 2010).

4.3.2.4.3. NON-DLVO FORCES The hydrophobic attraction force originates when polar water molecules are squeezed between two hydrophobic surfaces, since contact between water molecules and the surface is essentially avoided the hydrophobic surfaces have a preference to associate with each other. Surfaces can be characterised in terms of their hydrophobicity by measuring the water-contact angle (θ) on a solid substrate, whereby a larger θ refers to a material of hydrophobic character. Common materials are divided into wetting (θ < 90°) and non-wetting (θ > 90°) surfaces. However,  hydrophobic interactions become effective between surfaces with θ > 65° (Michen, 2010). Hydration (or structural) monotonically repulsive or oscillatory forces can arise from the way water molecules (or other solvent molecules) order themselves at the solid-liquid interface and how such a structure responds to a neighbouring surface (Hans-Jurgen Butt, 2005) (Michen, 2010). A hydrogen bond (0.2-0.4 nm) is an attractive interaction between a proton donor (A) and a proton acceptor (B). Atoms A and B are usually the highly electronegative F, O, N, S, for which dipole moments are large. The bond energy (or enthalpy) ranges from 0.8 to 170 kJ/mol. Mentioned NON-DLVO interactions excluding hydrophobic attraction are considered short-range forces (relevant at surface separations < 10 nm).

4.3.2.4.4. DLVO THEORY DLVO potential energy calculations can be used to describe the balance of electrostatic repulsion and van der Waals attraction forces as a function of virus − surface separation distance. Thereby, the sum of the van der Waals interaction energy (ΦVDW) and the electrostatic double layer interaction energy (ΦEDL) gives the total interaction energy (ΦDLVO) (Michen, 2010). However, DLVO-theory has failed to give a unifying description. Non-DLVO interactions have been included in some models and are known as extended DLVO theories (X-DLVO) to improve the description. Such non-DLVO contributions to the total interaction energy can be simply included in ΦXDLVO calculations by summation of additional terms (Michen, 2010) (Li, 2010).

4.3.2.4.5. VIRUS ADSORPTION AND INACTIVATION

The removal of microorganisms by depth filtration is restricted to larger-sized pathogens such as protozoa and several bacteria. The additional removal of viruses by adsorption is currently the concern of studies that have aimed towards a low-cost single stage filtration process with the potential for POU applications. It has been shown that viruses often possess a negative surface (capsid) charge at pH values occurring in natural water sources (figure 3), although a smaller fraction of viruses may also carry a positive surface charge (Markus Wegmann B. M., 2008) (Michen, 2010). In the case of non-enveloped hydrophilic viruses (relevant in water disinfection) DLVO interactions lead the virus adsortion process to a large extent (other forces surface forces may also contribute) and rentention is improved wherever virus and adsorbent carry oppositely charged surfaces (Markus Wegmann B. M., 2008) (Michen, 2010). A filter able to capture positively and negatively charged viruses must provide a heterogeneous surface that bears surface charges of both signs. Thereby, an electropositive adsorbent is commonly introduced into an electronegative filter media in order to enhance retention of negatively charged viruses (Michen, 2010). The adsorption of Poliovirus (non-enveloped) to various powdered surfaces: SiO2, Fe2O3, MnO2, CuO, Al2O3, Al and Si has been studied. Results led the authors to propose that DLVO theory should be regarded as a principal approach with which the adsorption of viruses on various inorganic surfaces can be described. Other contributions such as, hydrophobic interactions and hydrogen bonding appeared to be of secondary importance. It has also been stated that hydrophobic interactions can be expected to play a major role in the adsorption of lipid-containing viruses (envelope viruses) and their adsorptive behaviour could be expected to be quite different to that of the more hydrophilic viruses (non-enveloped virus). Some studies also suggest hydrogen bonding as the dominant force in virus adsorption to polymers and inorganic surfaces (Michen, 2010). In relation to Fe(3+) oxides it has been stated that the adsorption capacity rises from both electrostatic interactions and chemisorption. As a result, adsorption may take place on a neutral surface or even with the same charge as the adsorbing species. Both of the above interactions provide iron oxides with a high sorption affinity toward viruses and cyanobacterial toxins (microcystin-LR) (Li, 2010) (Leonardo Gutierrez, 2009) (Jungju Lee, 2011). The inactivation of viruses (i.e. render a virus into a uninfectious state) on surfaces has ben found to be specific to certain types of viruses, although the underlying principles remain unclear (Michen, 2010).

A non-enveloped virus can be inactivated by the alteration or removal of its nucleic acid and/or its proteins. Alteration of a surface protein might prevent a virus from attaching to or entering the cell, while alteration of internal virus proteins can destroy properties, such as enzyme activities, essential for the replication of the virus (Li, 2010). Hence, inactivation by adsorption has been related to alteration of the unique and energy-stable pattern of virus proteins by strong surface interactions (Li, 2010). Joseph N. Ryan et. al. proposed a surface inactivation mechanism based on electrostatic interactions between virus structure and hematite-coated quartz. The mechanism is as follows: (i) virus attachment to Fe(III)-OH, favored by electrostatic attraction; (ii) inactivation at the surface, caused by the strong attachment force; (3) release of non-infective or disintegrated viruses to solution (partial) (Joseph N. Ryan, 2002). Virus inactivation has also been reported by adsorption onto Al, Zn, Mg, Fe, Fe3O4, CuO and MnO2, whereas viruses were not inactivated after adsorption onto Fe2O3 and SiO2 (Brian M. Pecson, 2012) (Michen, 2010).

4.3.2.4.6. SAND FILTER VIRUS RETENTION ENHANCEMENT

Viruses generally have a low affinity for sand (composed to a large degree by SiO2), and the degree of removal is dependent on flow rates, pH, the presence of organic matter and particularly on the concentration of divalent cations. Since SiO2 has an IEP < 4 (see table 2) and thus bears a negative surface charge, as do most viruses, in natural waters, the low affinity of viruses for SiO2 based materials is in agreement with repulsive forces originating from electrostatic interactions (Michen, 2010). Investigations of bacteriophages deposition kinetics in packed beds of quartz have shown low retention (< 1 LRV) at pH 5 and 0.01 M NaCl. In comparison, retention experiments conducted at pH equal to the virus IEP (3.5 for MS2), revealed an increased removal around 3 LRV at 0.01 M NaCl (Michen, 2010). It has also been showed that IEP controlls adsorption and transport through sand/clay soils of five different spherical bacteriophages. These results underline the major role of electrostatic forces in virus retention (Michen, 2010). In comparison with pure sand, it has been shown that sand coated with metal oxides and hydroxides can remove significantly more viruses. These metal oxide coatings improve virus rentention by altering the sand surface charge (Mark D. Sobsey, 2008) (Markus Wegmann B. M., 2008) (Li, 2010) (Brian M. Pecson, 2012). Retention studies of bacteriophages MS2 and φX174 as well as the human pathogenic Aichi virus in columns filled with sand, goethite coated sand and aluminium oxide coated; showed little retention (< 11

%) of all viruses on clean sand but recorded a significant improvement of 4 to 6 LRV with coated sands at pH 7.5. Among the viruses, adhesion on goethite- and aluminium oxide-coated sands followed the order of MS2 > Aichi > φX174. These results were consistent with adhesion force measurements using an atomic force microscope which, revealed average forces of 19, 9 and 0.5 nN between MS2 and aluminium oxide coated sand, goethite coated sand and clean sand, respectively. Strong electrostatic, and, to a lesser extent, hydrophobic interactions were proposed to be responsible for the significant improvement in virus removal on the modified sands. X-DLVO calculations provided qualitative correct predictions on the adsorption trends observed. (Michen, 2010). Jie Zhuang et. al. also studied the removal of MS2 (IEP = 3.9) and φX174 (IEP = 6.6) viruses on goethite-coated sand. In terms of their attachment and inactivation on goethite, about 90% of the MS2 particles removed by goethite were inactivated at pH 7.5. In comparison, 74% of the goethite-bound ϕX174 particles lost their infectivity at pH 7.5. In addition it was showed that the presence of phosphate in solution reduced virus attachment and appeared to protect the viruses from being inactivated. Conversely, presence of carbonate in solution increased virus attachment and inactivation (Jie Zhuang, 2008). Brian M. Pecson et. al. studied the adsorption and inactivation of MS2 and φX174 using Fe (3+) oxide coated sand. Authors reported that uncoated sand caused no discernible adsorption, the iron oxide coating transformed the sand into a matrix capable of adsorbing large quantities of both viruses (>109 PFU/g). Adsorption alone (i.e. in the dark), however, had no impact on virus infectivity. Consequently, the release of adsorbed viruses reinserted infective viruses back into solution (Brian M. Pecson, 2012). In addition, Viruses that were not adsorbed onto the iron oxide coated sand showed low levels of inactivation after 8 h of sunlight exposure (<1 LRV). Inactivation was much greater if the viruses were adsorbed onto the coated sand before exposure to sunlight. Thus, the key factors leading to enhanced MS2 inactivation were (i) adsorption onto iron coated sand and (ii) exposure to sunlight. Compared to MS2, φX174 exhibited a much lower extent of photoinactivation when adsorbed onto coated sand (Brian M. Pecson, 2012).

4.3.2.4.7. FIBER FILTER VIRUS RETENTION ENHANCEMENT

Filters composed of fibreglass have been coated with a cationic polymer (PEI) and with hydroxides of Al, Fe, Mg and Ca in order modify surface charge. These filters were found to adsorb a greater percentage of coliphages and enteroviruses than the

untreated filters (Michen, 2010) (Markus Wegmann B. M., 2008). More recently nano-sized fibers, composed of boehmite (γ-AlO(OH)), applied on glass fibre filter material have been shown to remove MS2 phages up to 99% at neutral pH. The fibres are approximately 2 nm wide and 200 to 300 nm in length with an IEP of 9.4 and a specific surface area up to 500 m2/g. The technology is already on the market as the so-called Disruptor® from Ahlstrom (Michen, 2010). Iron oxide and zinc oxide nano-sized particles has also been used to enhace virus retention and inactivation of non-woven glass fiber substrate. Authors stated that this substrate provides large surface area and enables the assemble of filters with low-pressure drop compared to other substrates (Li, 2010). Leonardo Gutierrez et. al. evaluated the use of glass fiber coated with hematite nanoparticles (3-20 nm and IEP = 6.9) for virus (rotavirus and MS2) removal from water. Control experiments, performed for MS2 virus using fiberglass substrate with no coating of iron oxide nanoparticles, showed no removal of viruses. In comparison, iron oxide coated filters showed fast adsorption kinetics and high adsorption capacity for MS2 (IEP = 3.6) and rotavirus (IEP = 4.5) in the pH range 6 – 9 under non-competing conditions (Leonardo Gutierrez, 2009) (Li, 2010). Only a small fraction of infectious RV (0.5 – 2.1%) was successfully recover in comparison to MS2 (44 – 45.8%), possibly due to irreversible adsorption or structural damage to the capsid of rotavirus (RV) when interacting with hematite. (Leonardo Gutierrez, 2009). The observations that MS2 and RV exhibited different behavior upon adsorption onto iron oxide suggest that the inactivation of a given viral particle might depend on the proteins of its capsids or the robustness of its structure. (Leonardo Gutierrez, 2009) (Li, 2010). Obtained results under non-competing conditions indicated that both viruses adsorbed to hematite nanoparticles. Nevertheless, in natural water systems many competitors for adsorption sites are found. (Leonardo Gutierrez, 2009). Adsorption capacity of hematite-coated fiberglass was reduced from 6LRV to 2LRV units when NOM and bicarbonate ions were present in solution in comparison to non-competing conditions (Leonardo Gutierrez, 2009). Zinc (3.5 nm and IEP = 9) coated glass fiber filters offer an increase in capacity by orders of magnitude over iron oxide coated filters. Compared to iron oxide nanoparticles, zinc oxide nanoparticles didn‘t show an improvement in inactivation kinetics but inactivation capacities did increase by two orders of magnitude from 99% to 99.99%. The viricidal effect of ZnO was mainly attributed to generation of

reactive oxygen species (ROS), which can cause permanent damages to MS2 viruses and to a less extent to its higher IEP. In addition, zinc oxide nanoparticles have higher affinity to viruses than the iron oxide nanoparticles in presence of competing species. Although zinc oxide dissolves in aqueous media the dissolution rate is low (0.2 mg/l over one month – EPA secondary limit 5 mg/l). Consequently this approach appears to be efective to remove viruses from drinking water (Li, 2010).

4.3.2.4.8. CERAMIC FILTERS VIRUS RETENTION ENHANCEMENT

Among porous ceramics filters fired clay are commonly used for water treartment at the point of use (Mark D. Sobsey, 2008). Although, renew interest has been observed in diatomite based filters. The various species of diatomites have different morphologies and pore sizes, which range from the nano to the micro scale. Therefore, this material shows a high surface area and porosity, making it suitable for use in filtration with the potential to combine high virus retention with high flowrates and low cost (Michen, 2010) (Wenning Shen, 2011) (Markus Wegmann B. M., 2008). Since diatomite (IEP<2) is a silica-based material it electrostatically repels most viruses. In order to promote the electrostatic adsorption of viruses, the material can be modified with coatings that impart a positive surface potential between pH 5 and 9. Besides possessing a positive surface charge, the coating material (adsorbent) needs to fulfil additional requirements, such as a low solubility in water, and it should not be harmful to humans as leaching may cause serious health effects (Markus Wegmann B. M., 2008) (Michen, 2010). Surface charge modification of diatomite based filters has been achieved both with the application of organic polyelectrolytes and of metallic hydroxides of aluminum, calcium, iron and magnesium and virus removal efficiencies on the order of 99% have been demonstrated (Michen, 2010) (Markus Wegmann B. M., 2008). Also, small amounts of silver (or copper) have been added into diatomite-based depth filters to inhibit the growth of bacteria within the filter (Michen, 2010). More recently, Wenning Shen et. al. deposit ultrafine AgO powders (8.5 wt%) on porous diatomite ceramics in order to avoid this shortcoming. These filters showed bacteria inactivation capacities against S. aureus and E. coli of 99.9%. (Wenning Shen, 2011) Markus Wegmann et. al. modified diatomite based filters with zirconia and yttria to enhance virus retention. The characteristics of diatomite-based filters are summarized in Table 3 (Markus Wegmann B. M., 2008). A different alternative used by Michen is the introduction of the sorbent material (MgO) prior

to sintering. Considering production requirements, the adsorbent material needs to be stable at temperatures above 1000°C and should be low-cost (Michen, 2010). Base filter characteristics reported by Michen are similar to those in table 3, however mean pore size was 3.2 µm.

Table 3: Diatomite-based filter characteristics

Markus Wegmann et. al. modify the surface charge of diatomite based filters using a colloidal dispersion of Zr(OH)x nanoparticles (5-10 nm). The filter elements were immersed (dip-coated) for 2h in the pristine Zr(OH)x suspension which was allowed to permeate the porous medium by capillary forces alone. After dipcoating, the filters were dried and heat treated to fix the Zr(OH)x coating in place on the filter medium surface (Markus Wegmann B. M., 2008). After heat treatment at 250 °C, the coated material possesses a specific surface area of 25.5 m2/g and an IEP of 9. When challenged with MS2 phages, filters with this Zr(OH)x coating generally retain 7LRV of the virus-like particles in the pH range 5–9. The problem with this coating is that it is washed out of the filter in use due to insufficient consolidation of the Zr(OH)x particles during heat treatment. The coating is rendered physically stable by heat treatment at 400 °C; however, the specific surface area of the coated filter is reduced to 12.5 m2/g. MS2 retention in these filters drops off in the pH range 8–9, and considering that the IEP of the coated material after calcination at 400 °C lies at pH 5.5, this result indicates that electrostatic interactions are the predominant mechanism promoting the adsorption of virus-particles on the Zr(OH)x coating. (Markus Wegmann B. M., 2008). In a later work Markus Wegman et. al. use nanosized (10 nm) hydrated yttrium oxide as the coating material. The porous filter elements were dip-coated as in the previous work (basic procedure). After coating was dried, filters were heat-treated to consolidate the coating and obtain the desired Y2O3 phase (Markus Wegmann B. M., 2008). Afer heat treatment it was observed that the nanometric particles migrated towards the inner and outer surfaces of the tube-shaped diatomite-based filters during drying, resulting in a U-shaped concentration

profile of the coating material across the tubewall (see figure 5). Since this phenomenon leads to blockage of porosity in the near surface regions of the filter and thereby reduces the flowrate possible at a given operating pressure, various approaches were considered to flatten the concentration profile as far as possible. (Markus Wegmann B. M., 2008). In order to flatten the Y2O3 concentration profile immobilizing the nanoparticles by gelling the suspension was investigated. Despite the pronounced gelling effect observed upon cooling the suspensions containing agarose, the yttrium profiles showed little improvement relative to those obtained using the basic procedure. The gel formation using direct coagulation casting (DCC) process successfully prevented colloid particle migration during drying. However, measurements of the specific surface areas of these coated filters yielded values around 2 m2/g, representing absolutely no improvement over the virgin diatomite filter (2.2 m2/g (Markus Wegmann B. M., 2008). Another approach considered for prevention particle migration after infiltration was freeze-drying. This procedure yielded fairly flat yttrium profiles across the filter wall and acceptable specific surface area was achieved (15 m2/g). Despite the promising results, the procedural complexity and the associated limited throughput of the process precluded it from being used as the standard infiltration procedure (Markus Wegmann B. M., 2008). The solution ultimately chosen was the following: Instead of infiltrating filter elements with standard dimensions as previously, elements with a slightly larger outer diameter were used and drying from the inner surfaces was prevented, thereby forcing water diffusion and particle migration to occur primarily towards the outer surface. By subsequently machining the outer diameter of the filter tube down to the standard dimension, elements were produced which exhibited a largely uniform coating concentration across the tube wall. (Markus Wegmann B. M., 2008). However, this procedure has the clear disadvantage of wasting high cost nanoparticles.Two immediate problems were identified in previously presented studies, specifically (i) a limited virus capacity/lifetime of the filters, and (ii) washing out of the coatings during operation.

Figure 5: Yttrium concentration profile

Michen used MgO nanoparticles as virus sorbent material. This material was chosen because it has high thermal stability and it is readily available at a relatively low cost. In addition MgO has the highest IEP reported. (Michen, 2010). However, MgO is soluble and reactive in water. Diatomite-based depth filters containing 10, 15, 20 and 30 wt% of MgO were produced. It was found that, with an increasing amount of adsorbent, the filters were more susceptible to forming cracks during drying and firing. In the case where 30 wt% of MgO was incorporated, no single usable filter was obtained. Hence, only modifications with 10, 15 and 20 wt% of MgO could be used for filtration experiments. However, it was stated that such filters showed cracks in the longitudinal direction with lengths up to 50 mm (Michen, 2010). The introduction of MgO particles led to an increase of the specific surface area and caused a reduction in the mean pore size of the filters with increasing amounts of adsorbent. (Michen, 2010). A smaller pore size resulted in the reduction of permeability as evidenced by the flux measurements. In contrast, the porosity slightly increased with higher loads of MgO and reflects the increased occurrence of cracks (Michen, 2010). The ζ-potential measurements of MgO modified filters showed a significant change towards less negative z values compared to base filters. However, the modified filters still possessed a net negative charge between -10 and -50 mV in the investigated pH range. Diferent mineral compounds were formed during the sintering process of the DE, clay and MgO mixture, namely SiO2, MgO, Mg2SiO4 and MgSiO3. (Michen, 2010)

The modified filters were tested in soft water at pH 5, 7 and 9 to assess their removal efficiency in regard to the bacteriophage MS2. It has been shown that the introduction of MgO into depth filters can significantly improve the retention of bacteriophages MS2 by more than 4 LRV compared to the reference filter. After the filter has been in operation, the virus retention efficiency was improved. The improve bacteriophage removal in the MgO modified filters, was accompanied by a change in filter surface characteristics associated with water-based reactions. The changes involve the formation of magnesium oxyhydroxides originating from the bare MgO and/or from the generated compounds Mg2SiO4/MgSiO3 during sintering and a morphological change of MgO (Michen, 2010). Since the mentioned changes influence the filter performance during operation, such filters were not recommended for drinking water production due to significant variations in virus removal. Moreover, the filter may alter the pH in the effluent to values above pH 9, which does not fulfil drinking water regulations (Michen, 2010). The experimental results on virus removal were verified with DLVO and X-DLVO calculations. These models supported the understanding of virus retention in depth filters. However, these models were only partly capable of explaining retention behavior, probably because they do not account for virus inactivation (Michen, 2010).

4.3.2.5. SIMULTANEOUS ENHANCEMENT OF BACTERIA AND VIRUS REMOVAL IN DEPTH FILTRATION

Nanoparticles that are very sensitive to oxidation or reduction (e.g. ZVI or CeO2) can cause important effects when they interact with microorganisms; thus they can be use to enhance sumultaneusly bacteria and virus removal of depth filters. For example, it has been showed that 50% mortality of E. coli occurred for concentrations of CeO2 > 3 mg/L when the nanoparticles are in close contact with the cell membrane. The toxicity of CeO2 is associated with the reduction of 30% of surface CeO2 to Ce2O3, as a consequence of this reduction a chain reaction resulting in lipid peroxidation within the liposaccharid layer on the cell membrane is initiated (Jean-Yves Bottero, 2011). This mechanism may also be effective for inactivation of enveloped and non-enveloped viruses in presence of CeO2 (IEP = 7.3). Iron is the most ubiquitous of the transition metals and the fourth most plentiful element in the Earth's crust. The extreme reactivity has traditionally made iron nanoparticles difficult to study and inconvenient for practical applications (Law Yong Ng, 2010). However, among the nano-absorbents popular today, zero valent iron is the most utilized, primarily because it covers the broadest range of environmental contaminants: organic pollutants,

pesticides, arsenic, nitrate and heavy metals (T. Pradeep, 2009) (R. A. Crane, 2012) (Nora Savage, 2005). The mechanism of reactivity for ZVI is similar to the mechanism of corrosion (T. Pradeep, 2009). From a simple perspective, corrosion is the degradation of a metal to a low-energy state stimulated by the enviroment. Corrosion reactions and corrosion product(s) are responsible for the reductive transformation and/or physical removal of exposed species to ZVI. Surface precipitated iron oxide/hydroxide is initially porous, thus the material can develop a “core–shell” structure during the early stages of reaction capable to induced both sorption (at the oxide/hydroxide) and chemical reduction (at the metallic iron oxide/hydroxide interface) simultaneously. However, as the reaction progresses the material’s porosity decreases and direct Fe(0)–H2O/O2 and Fe(0)–contaminant interactions are no longer possible. It should also be noted that because the kinetics of the initial stages of Fe(0) oxidation are so rapid, ZVI that is introduced to an environmental system will already have a film of surface oxide develop directly after synthesis. Therefore, it should be understood that although this nanomaterial (ZVI) is referred to as being metallic, each particle exists in natural conditions with a thin but encapsulating layer of surface oxide (R. A. Crane, 2012). Nowadyas, the capacity of ZVI in microorganism control is under investigation. Melanie Auffan et. al. access the biocidal capacity of ZVI, magnetite and maghemite towards E.coli as a function of the oxidation state. Under the experimental conditions used by the authors the zeta potential is on the order of 10-20 mV for the three iron-based nanoparticles (γ-Fe2O3, Fe3O4 and ZVI), hence the affinity of the three materials for the bacterial wall is elevated, and the aggregation state is similar. Therefore, the difference in the toxicity response cannot be related to these parameters and the redox changes between Fe(0)/Fe(2+)/Fe(3+) were then directly related to the increase cytotoxicity observed γ-Fe2O3 < Fe3O4 < ZVI. Authors stated that the main source of toxicity of iron-based nanoparticles towards E. coli was related to ROS generation during oxydation of reduced iron species (Melanie Auffan, 2008) (Satinder K. Brar, 2010). Zhi Qiang Li et. al. showed that nanosized ZVI (aprox. 28% Fe(0) content ) was less toxic to E. coli exposed under aerobic conditions (0.8 LRV inactivation) compared to exposure under anaerobic conditions (5 LRV inactivation). The lower observed toxicity during exposure under aerobic conditions was related to the formation of Fe(3+) oxides, compared to corrosion under anaerobic conditions wich favors Fe3O4 (magnetite) formation. This

formation of Fe(3+) oxides can at the particle surface inhibit electron transfer due its lower conductivity when compared to magnetite (Zhi Qiang Li, 2010). Other authors have shown that the bactericidal activity of ZVI in the presence of dissolved oxygen is comparable to that of silver nanoparticles, although with reduced lifetime (Chang Halee, 2008). Youwen You et. al. performed the first demonstration of virus (φX174 and MS2) removal from water by zerovalent iron. After desorption of both adsorbed viruses recoveries were low (0.1%), indicating that the phages were either irreversibly adsorbed or rendered non-infectious. Futhermore when ZVI was added to a flow-through sand filter, over 5 LRV removal of both viruses could be achieved within an effective contact time with iron of 20 min (Youwen You, 2005). The removal was attributed to adsorption of these phages to iron oxides and hydroxides formed via corrosion of ZVI. Chunjian Shi et. al. further assess the feasibility of ZVI for drinking water treatment, by evaluating the ability of ZVI to remove pathogenic viruses in treatment plant water samples. Clean sand alone achieved limited or no removal of viruses, conversely when the sand column was amended with 15% ZVI, removal efficiencies of 4.5–6 LRV were achieved for all three viruses (MS2, φX174, and AiV) regardless of the water used (Chunjian Shi, 2012). Ian Bradley et. al. study the potential advancement in household water treatment technologies by amending the biosand filters with iron. Results showed that, both MS2 and rotavirus were treated to US-EPA standards for virus removal (see section 4.3.2.4.8), by adsorption to positively charged iron oxides generated by the corrosion of ZVI. Authors stated that duration of effective virus removal by ZVI-amended biosand filtration depends on both source water conditions and quantity and composition of iron material added (Ian Bradley, 2011). Based on obtain results it was concluded that ZVI can be used for removal of bacterias and viruses from water. Although important issues need to be futher studied such as: long-term performace, impact of pH and disolved oxygen on virus retention, effects of competitive adsorption on virus removal efficacy, effectivity agains protozoa, and the mechanisms reponsable for virus inactivation (Youwen You, 2005) (Ian Bradley, 2011). ZVI with particle size at the nano-scale exhibits superior activity because of their larger surface area and higher reactivity. However, similar to other nano-materials, this ultra-fine powder has a strong tendency to agglomerate into larger particles, resulting in an adverse effect on performance; another disadvantage of this material is the recovery after usage. Using a support material (e.g. clay minerals) for nano-sized ZVI is one

possible way to solve this problem (Ray L. Frost, 2010).

4.4. MULTIPLE-BARRIER APPROACHES These are any of the above technologies used together, either simultaneously or sequentially, for water treatment. These combination treatments include coagulation plus disinfection, depth filtration plus disinfection or depth filtration plus surface filtration, among others. Some are  commercial single-use chemical products in the form of granules, powders or tablets containing a chemical coagulant, such as an iron or aluminium salt, and a disinfectant, such as chlorine. When added to water, these chemicals coagulate impurities to promote rapid and efficient sedimentation and also deliver the chemical disinfectant (e.g. free chlorine) to inactivate microbes (Mark D. Sobsey, 2008) (WHO, 2011). Extensive reductions of bacteria, viruses and parasites (protozoa and helminthes) were reported, and the costs of treatment were estimated to be relatively low (US$ 0.01 /l). However, the socio-cultural acceptance of disinfection with chlorine-containing reagents tended to be low in some cases, due to taste and odor problems. Moreover, if insufficient time intervals are applied for reaction and sedimentation, the effectiveness of these methods is low (Maryna Peter-Varbanets, 2010). Other combined treatment technologies are physical devices that include two or more stages of treatment, such as depth or surface filters to physically remove microbes and either strong oxidant or UV radiation to kill any remaining microbes not removed by the filter. (WHO, 2011). Another alternative is the use of iron oxide ceramic membranes to degrade organic foulants via heterogeneous Fenton reactions in precense of strong oxidants (see section 4.2.5). Recent studies have demonstrated that iron oxide-coated ceramic membrane improves degradation of organic foulants such as NOM by ozonation, remarkably reducing membrane fouling (Jeonghwan Kim, 2010). Combined ozonation-filtration setups required further testing for disinfection purpuses.

5. ANALYSIS OF HOUSEHOLD DISNFECTION SYSTEMS In principle, all centralized technologies can be applied in to household treatment of drinking water. For the smallest scale of systems, the POU systems, some specific technologies and systems have been developed, described and evaluated for household use on the basis of several performance criteria. Besides efficiency in improving the microbiological quality of the water and the system costs, these performance criteria include the ease of use of the system, its environmental sustainability, sociocultural acceptability and required maintenance (Maryna Peter-Varbanets, 2010). Any POU water-treatment technology without consideration of these criteria is

unlikely to be either successful or sustainable. Moreover, systems for decentralized applications should preferably be independent of utilities such as electricity or tap pressure; hence energy source is another relevant performace parameter (Maryna Peter-Varbanets, 2010) (Mark D. Sobsey, 2008). The data included in table 4 was taken from (Maryna Peter-Varbanets, 2010) (Mark D. Sobsey, 2008) and (WHO, 2011). The investment costs listed in Table 4 include the costs needed to buy, deliver and install the system. The operational costs include the costs of reagents, energy and servicing if needed, as well as maintenance and replacement of parts of the system. The World Health Organization (WHO) categorizes costs for POU systems as low, medium and high on a worldwide basis that includes the poorest people. The categories for annual household cost estimates in US dollars are less than US$ 10 for low, US$ 10–100 for moderate and >US$ 100 for high. Performance is access using the base lines of microbial retention reported by World Health Organization (WHO, 2011): (++) the baseline is equal or above 4 LRV, (+) the baseline is bellow 4 LRV and above 1 LRV and (-) the base line is 1 LRV or bellow. Ease of use as evaluated as follows (++) daily operation is limited to filling in of raw water and collection of treated water and (+) requires additional (time consuming) operations which, however, may be performed by unskilled person with no or little training (Maryna Peter-Varbanets, 2010). Sustainability of the system is (+) if system may be produced locally from locally available materials, with limited use of chemicals and non-renewable energy and (-) if the system requires chemicals or non-renewable energy sources for daily operation (Maryna Peter-Varbanets, 2010). Social acceptability can be ranked as follows: (++) application is based on tradition or it is already in use; (+) available studies showed good social acceptance and (+/-) available studies are contradictory, or the results depend on the region studied (Mark D. Sobsey, 2008) (Maryna Peter-Varbanets, 2010).

Table 4: Features of descentralized systems for drinking water disinfection.

Beside the taste and odour problems and the generation of DBP’s, chlorine disinfection is ineffective for inactivating oocysts of the waterborne protozoan Cryptosporidium; in addition it requires low turbidity and pH control. Although technologies such as ozone, UV-irradiation or membrane processes can provide safe drinking water, these techniques become impractical for use in less developed nations due to an absent infrastructure and lack of trained personnel as well as high energy consumption, all leading to high costs. Solar disinfection requires low turbidity raw water and a considerable amount of time to obtain safe drinking water. Slow sand filtration is cheap and may be completely built from locally available materials, however inadequate maintenance is often a main reason of system malfunction. Flocculation must usually be combined with a filtration step in order to produce safe drinking water. While ceramic filter depth filtration alone fails when small sized pathogens, such as viruses, are to be removed (provide filter structure integrity is maintain). Table 4 shows that an ‘‘ideal’’ solution, namely one whose maintenance is limited and able to produce safe drinking water at low cost, does not yet exist for household systems (Maryna Peter-Varbanets, 2010). However if depth filters can be modified to enhance bacteria and virus retention, kinetics of solar disinfection can improve, or nontoxic antibacterial materials can be used to coat vessels for safe storage of disinfected water (e.g. boiling achvies 6 LRV for all pathogenic microorganisms), Harmful health effects related to microbial contamination of drinking water could be effectively reduced with reduced impact on the environment.

6. ANALYSIS OF NANOMATERIALS ENHANCED DISINFECTION Nowadays, nanomaterials have been used to enhance effectivity and/or kinetics water disinfection

methods based on photocatalytic and filtration processes and to confer antibacterial properties to packing materials (e.g. PET water bottles). Table 5 presents a summary of caracterisitcs of different tecniques and nanomaterials currently studied in order to enhance available water disinfection methods that that have the potential to be adapted to systems for household water treatment (POU and POE).

Table 5: Summary of characteristics of diferent techniques and nanomaterials used in order to enhace disinfection methods

In addition to the already recognized mechanisms and systems limitations, five main aspecst has been identified particularly that limit nanosized materials use for drinking water disinfection, namely: (i) lack of understanding of microbial control mechanisms, (ii) competitive adsortion, (iii) nanomaterial aggregation, recovery and inmobilization, (iv) photocatalytic activity and (v) nanoparticle toxicity.

6.1. LACK OF UNDERSTANDING OF MICROBIAL CONTROL MECHANISMS Detail study of microorganism control mechansms both attachement (sorption) and inactivation mechanisms need to be further study. For example in relation to virus adsortion, only the systematic study of all possible surface interactions (electrostatic, van der Waals, hydrophobic, steric, hydrogen bonds and so on) acting between the virus and adsorbent material may finally lead to the understanding this process (Michen, 2010). Positively charged metal oxides and hydroxides seem to be promising materials that are capable of removing non-enveloped viruses with a rather hydrophilic character. In contrast, viruses having an envelope (or a more complex morphology) tend to be hydrophobic in nature; thus nanostructured surfaces, which favour hydrophobic attraction and show increased surface area, may be beneficial in the retention of type of viruses (Michen, 2010). Future research should focus in particular on inactivation

processes on surfaces, since this would exclude potential problems of filters working on the adsorption principle, such as breakthrough upon filter capacity exhaustion or virus desorption during use (Michen, 2010). Particularly, the antimicrobial capacity and life span of highly reactive materials such as ZVI and CeO2, wich have shown capacity for simultaneous removal of viruses and bacteria. Never the less toxicity of ceria should be considered (see section 6.5)

6.2. COMPETITIVE ADSORTION In relation to competitive adsorption in water disifection further research is needed to address issues such as catalyst/sorbent fouling, impact on disinfection efficacy and kinetics, and need for catalyst/sorbent regeneration to ensure the sustainability of these new nanomaterials enhanced water disinfection appoaches. (Michael V. Liga, 2011). Virus adsorption capacity of positively charged oxides and hydroxides is impared by the occurrence of competitive adsorption of other species. For example, humic acid (NOM component) is known to (a) cause physical fouling of finely porous filter media, (b) possess surface charge characteristics similar to those of virus particles and (c) exhibit higher mobility when compared to viruses. The latter points (i.e. b and c) are critical, as it means that competitive adsorption will occur, depleting the filter of adsorption sites and reducing its capacity for virus adsorption (Markus Wegmann B. M., 2008) (Leonardo Gutierrez, 2009). For example, Markus Wegmann et. al. showed that in the presence of 1 mg/l humic acid, the 4 LRV limit is reached after significantly less water has been pumped through the coated filter (150 l) compared to the amount filtered when the tap water is free of humic acid (500 l) (Markus Wegmann B. M., 2008). Leonardo Gutierrez et. al. indicated that the adsorption capacity of hematite coated fiber glass was reduced down 4 LRV units when NOM and bicarbonate ions were present in solution in comparison to non-competing conditions (Leonardo Gutierrez, 2009). In addition, Zhi Qiang Li et. al. showed that the adsorption of NON to ZVI nanopartiples resulted in less than a 0.2 LRV reduction in viable E. coli cells, while a 5 LRV reduction was observed for bare ZVI. Competitive adsorption does not preclude the use of depth filters coated with electropositive oxides as POU or POE devices for virus removal from pre-treated water. However, it greatly impairs their effectivity as stand alone solution in disinfection of source water.

6.3. NANOMATERIAL AGGREGATION, RECOVERY AND INMOBILIZATION Nanoparticle aggregation negates most of the benefits derived from the use of nanosized materials. Hence if stable nanoparticle dispersions are required

an efficient separation process should be implemented downstream to retain the nanoparticles (S. Malato, 2009) (Meng Nan Chong, 2010) (Peter K. J. Robertson, 2012). Nanosorbents or nanocatalysts used in water treatment can be retain by membrane filters, more recently the magnetic properties of nanozised magnetite has been employ to facilitate the recovery of these nanosized materials and to prevent their agglomeration. The most highlighted work to date regarding this topic has been the removal of arsenic from drinking water using magnetite as sorbent and low-magnetic field induced recovery. However, functionalization of Fe3O4 surfaces has also led to the preparation of promising adsorbents for other heavy metals such as copper, lead and mercury that can be readly recovered by magnetic separation (T. Pradeep, 2009) (Piao Xu, 2012) (Tamar Gordon, 2011). On the other hand immobilization of nanomaterials onto inert supports can reduce (onto larger size particles) or eliminates (onto filter surfaces) the need for separation and nanoparticle agglomeration. However, the active surface area will be afected since inmobilization onto a support entails a reduction in reactive surface atoms; as a concequence the interaction between catalyst/adsorbent and the target microorganism is decreased (S. Malato, 2009) (Anu Matilainen, 2010) (Yalei Zhang H. G., 2011). Since the specific surface area of the support largely determines the active surface area of immobilized catayst/sorbent; the use of nanofibers, nanowires or nanorods as high surface area inmobilizers is been currently explored. Similarly, clays are receiving increasing attention as supports for nanosized catalysts and adsorbents of owing to their high adsorption capacity and cost effectiveness (Meng Nan Chong, 2010) (Yalei Zhang H. G., 2011). For example, TiO2 crystals has been deposited on different types of clay minerals, which includes bentonite, sepiolite (Elina Manova, 2010), montmorillonite (Robert Kun, 2006), semectites (Elina Manova, 2010), vermiculites (Elina Manova, 2010), palygorkite (D. Papoulis, 2010) (Lili Zhang, 2011) (L. Bouna, 2011), hallosyte (D. Papoulis, 2010) and kaolinite (Yalei Zhang H. G., 2011) (K. Koci, 2011). Other catalyst has also been inmobilized onto clay minerals, W. G. Xu et. al. synthesize sepiolite-ZnO composites (W. G. Xu, 2010). An additional benefit of anchoring nanoparticles on a suitable substrate is reducing the environmental risk asociated to nanoparticles release. Nevertheless, there are two main questions related to anchoring or inmobilization: (1) how strong is the adhesion between the nanoparticles and the substrate and (2) how the interaction between substrate and nanoparticles affects their functionality. Coating

washing-out has been commonly observed in surface modified depth filters. For example, Markus Wegmann ascribed material loss to insufficient sintering, and thus to the weak bonding of coating particles within themselves and to the underlying substrate (Markus Wegmann B. M., 2008). Higher sintering temperatures avoid the washing-out phenomenon. This result is, however, not without its drawbacks: The specific surface area of the filters drops dramatically. Furthermore, the MS2 filtration tests confirmed that retention was severely impaired in filters sintered at high temperature, though it was unclear whether this was due only to the loss in capacity (i.e. reduced surface area), or if a shift in the coating IEP also occured (Markus Wegmann B. M., 2008). Indicating that question (1) entails a compromise between strong adhesion and large active surface area.

6.4. PHOTOCATALYTIC ACTIVITY The strong attenuation of UV light by water (particularly natural waters contaminated by microorganisms) reduces the efficiency and hence increases the energy cost of photocatalytic water treatment. In contrast, visible light is less attenuated and a greater portion of the sun’s spectrum could be utilized in a solar photocalytic process (Peter K. J. Robertson, 2012). For example, solar TiO2 photocatalysis process efficiency can be considered linearly dependent on the energy flux but only 5% of the whole solar spectrum is available for TiO2 band gap. A realistic assumption of solar collector efficiency of 75% and 1% for the catalyst means 0.04% original solar photons are efficiently used in the photocatalytic process (S. Malato, 2009). In addition, quantum efficiency is defined by the competition between charge–carrier recombination, trapping and interfacial charge transfer. Therefore, the photocatalytic activity of metal oxide semiconductors can be enhanced via sunlight photoactivation, retarding electron-hole recombination and improving trapping and transfer of charge carriers to the interface (Nora Savage, 2005) (Shaily Mahendra, 2009) (C. Karunakaran G. A., 2011) (Qilin Li, 2008) (Yalei Zhanga, 2011).

6.4.1. CHARGE CARRIER RECOMBINATION Reduction in carrier recombination may be attained through sensitization, and/or through preparation of a mixed phase system (Ranjith G. Nair, 2011). Noble metals (e.g. Ag, Ni, Cu, Pt, Rh, and Pd) with Fermi level lower than TiO2 catalyst have also been deposited on the TiO2 surface for enhanced charge separation. These metals were reported to enhance electron transfer, but require good knowledge on the optimal deposited amount during the fabrication process (Meng Nan Chong, 2010). Since a high ammount might reduce photon absorption and/or

become electron–hole recombination centers, negatively affecting photocatalysis efficiency (S. Malato, 2009). Among noble metals, silver; have received much attention due to twofold capacity to enhance charge carrier separation and solar light activation. Ag(0) is believed to enhance photoactivity by facilitating electron–hole separation by forming the Schottky barrier at the TiO2-Ag interface and visible light absorption by silver surface plasmons is thought to induce electron transfer to TiO2 (Qilin Li, 2008) (Ranjith G. Nair, 2011) (Zhigang Xiong, 2011) (C. Karunakaran V. R., 2011). Recently, it was demonstrated that doping TiO2 with silver greatly improved photocatalytic inactivation of bacteria (E. coli) and viruses (MS2) (Michael V. Liga, 2011). It has also been reported that deposited Ag(0) increased adsorption of molecular oxygen with the latter enhancing production of reactive oxygen species (ROS) with concomitant increase in bacterial/virus photodegradation (Zhigang Xiong, 2011). In addition, leaching of antimicrobial Ag(+) can also contribute to bacteria inactivation (Michael V. Liga, 2011). Rare earth metal ions (La, Ce, Er, Pr, Gd, Nd and Sm) has also been used to enhance the photocatalytic activity of TiO2. Among rare earth metal ions, Gd ions were found to be most effective in enhancing the photocatalytic activity due to their higher ability to transfer charge carriers to the interface (S. Malato, 2009). TiO2 catalysts with mixed phases of anatase and rutile shows enhanced photocatalytic activity because of the suppressed recombination of photogenerated electrons and holes via traping. (Yalei Zhang H. G., 2011). In addition to anatase and rutile, TiO2 can form the brookite phase, which has been the least studied in many aspects of its properties due to the difficulty in synthesis. Recently, it was reported that the anatase–brookite composite nanocrystals (formed during the synthesis TiO2/kaolinite composites) showed higher photocatalytic activity than anatase and anatase-rutile (Degussa P25) (Yalei Zhang H. G., 2011). Other stidues reported mixed phase systems that combined TiO2 with CuO. Copper oxide nanoparticles change the energy level and work as electron acceptor that favor the electrons transfer, enhancing both photocatalytic and antibacterial activity of TiO2 (Ayman Yousef, 2012). Due to its narrow bandgap, Fe2O3 can be applied as a sensitizer of TiO2 photocatalyst. Electrons in the valence bands of TiO2 are driven into Fe2O3 due to formation of the built-in field in Fe2O3–TiO2 heterojunction. The charge transport between the valance bands of Fe2O3 and TiO2 is regarded as an effective process to promote photocatalytic activity, since it results in an increase in the electron–hole

recombination time (Piao Xu, 2012). In another work C. Karunakaran et. al. synthesized mixed phase ZnO–TiO2 nanocomposite capable of effectively the inactivation of E. coli under visible light (C. Karunakaran G. A., 2011).

6.4.2. VISIBLE LIGHT ADSORPTION Although noble metals coupling could be efficient in prolonging the surface charge separation, more economical transition metals (e.g. Cr, Fe, Ni, V) or non-metals (e.g. N, C, F, S) doping are usually prefer for industrial application due to their cost-effectiveness (Meng Nan Chong, 2010). Transition and non-metals doping, incorporates ions into the TiO2 crystal lattice. Such incorporation introduces impurity in the bandgap of TiO2 thus, reducing the photonic energy requirements (Meng Nan Chong, 2010) (S. Malato, 2009) (Ranjith G. Nair, 2011). For example, high temperature diffusion of iron cations has been used to obtain Fe-doped TiO2 coatings with enhanced photoactivity from TiO2 coatings deposited onto stainless steel (S. Malato, 2009). More recently, the use of non-metal dopants has been shown to improve the photoactivity into the visible spectrum and feasibility of TiO2 catalysts for industrial application (Meng Nan Chong, 2010). Unlike transion metal ions (cations), anions are less likely to form recombination centers and, therefore, are more effective to enhance the photocatalytic activity (S. Malato, 2009). N- and S-doped TiO2 have been reported to be effective for disinfection (E. coli) under visible light (λ > 400 nm) (S. Malato, 2009) (C. Karunakaran G. A., 2011). As an interesting example coupling previously mention mechanisms used to obtaian TiO2-based catalysts with superior properties is the work of Ranjith G. Nair et. al., they synthesized vanadium-doped titania sensitized by silver. The cathalyst exhibited a remarkable enhancement in the visible light photocatalytic activity in inactivating Escherichia coli. The enhanced photoactivity in inactivating the bacteria was attributed to (i) the dopant (vanadium) induced a marked red-shift in the absorption spectrum compared to base TiO2, (ii) the synergistic effect of Ag assisted charge separation as well as the surface plasmon enhanced absorption, and (iii) the charge traps generated in the mixed phase catalyst (Anatase-Rutile) (Ranjith G. Nair, 2011).

6.5. NANOMATERIAL TOXICITY Nanomaterials shows great promise in water disinection. However, some engineered nanomaterials could also become water pollutants that threaten human and ecosystem health (Qilin Li, 2008) (Jonathon Brame, 2011) (Jeonghwan Kim, 2010). The chemical nature and the physicochemical reactions occurring at the nanoparticle/cell interface

are the key to the toxic biological effects of nanoparticles. In particular, nanoparticles can affect biological targets through generation of reactive oxygen species (ROS). It is known that mineral surfaces may promote the formation of ROS via dissolution (e.g. Ag, AgO ZnO) and reduction of oxygen by electron transfer reactions (e.g. TiO2, Fe(3+) oxides, ZnO, Fe, Fe(2+) oxides CeO2) (Melanie Auffan, 2008). Nevertheless, available information is insufficient to determine the highest allowable concentration of a particular nanomaterial in drinking water. (Qilin Li, 2008) TiO2 particles (>100 nm) are known to be harmless to humans and animals. Although nano-scale TiO2 was classified recently as a possible carcinogen (if inhaled) by International Agency for Research on Cancer, its potential ingestion via water is not expected to be a major concern, as reflected by its use in toothpastes (Qilin Li, 2008) (Jean-Yves Bottero, 2011). Zinc is an essential trace element found in virtually all food and potable water in the form of salts or organic complexes. The diet is normally the principal source of zinc and the daily requirement for adult men is 15–20 mg/day (7.5 – 10 mg/l), in agreement the US-EPA guideline for zinc in drinking water is 5 mg/l. The toxic effect of ZnO nanoparticles has been mainly attributed to dissolved metal ions: at concentrations up to 10 mg/l ZnO (bulk and nano) is almost fully dissolved and bioavailable at pH between 7.5-8.2 (I. Blinova, 2010). Though, Xuan Li, observed that zinc oxide nanoparticles with diameter of 3.5 nm cause significant cytotoxic response at 3.75 mg/l and genotoxitc (chromosome damage) response at 35 mg/l in Chinese hamster ovary (CHO) cells (Li, 2010). Iron is found in natural fresh waters at levels ranging from 0.5 to 50 mg/l (not of health concern at these levels). Iron may also be present in drinking water as a result of the use of iron coagulants or the corrosion of steel and cast iron pipes during water distribution. Estimates of the minimum daily requirement for iron range from about 10 to 50 mg/day (5 – 25 mg/l). Due to the physiological relevance of iron in cell metabolism, iron-based nanoparticles were initially considered to be non-cytotoxic. It has, however, been recognized that the small size of Iron-based nanoparticles might pose an additional hazard as they can reach high local concentrations within the cells, making them more difficult to be cleared from the body (Stefan J. Soenen, 2011). The primary mechanism for cellular damage related to nZVI are the oxidation reactions, where redox cycling and the generation of reactive oxygen species (ROS) from reduced Fe within a cell can cause lipid peroxidation and damage to internal structures such as DNA. This result also suggests

that iron oxides/hydroxides produced by ZVI nanoparticle corrosion present significantly less risk, because their inherent nano-toxicity is lower (R. A. Crane, 2012). The greatest risk that ZVI nanoparticles pose to humans is most obviously during handling, but by using the material in liquid suspension, any immediate human danger is removed. (R. A. Crane, 2012). CeO2 nanoparticle cause significant toxicity at low concentrations (Jean-Yves Bottero, 2011). CeO2 is genotoxic at environmental concentrations (< 1 mg/L). The tocixity of CeO2 is related to redox phenomena that result in the production of ROS and the homeostatic unbalance for a series of ions Na+, Ca2+, K+ in the cell wall (Jean-Yves Bottero, 2011). Silver occurs naturally in water souerces, mainly in the form of insoluble and immobile oxides, sulfides and salts at concentration typically bellow 5 µg/l. However, levels in drinking-water treated with silver ions for disinfection may be above 50 µg/l (WHO, 2011). These is not a mayor concern because only a small percentage of silver ions is absorbed; retention rates in humans and laboratory animals range between 0% and 10% and silver levels, up to 100 µg/l, can be tolerated without risk to human health. However the Europen union safety regulations indicate that the presence of silver ions in food matrices should be limited to 50 µgAg (+)/kg food and the US FDA approved the use of silver nitrate as an antimicrobial in bottledwater, with a concentration not exceeding 17 µg Ag(+)/l of water (Duncan, 2011) (Amparo Llorens, 2012). The effect of increased use of nanoscale silver on human health and the environment is uncertian (Duncan, 2011) (Amparo Llorens, 2012). It is, for example, still unclear to what extent the biochemical pathways which facilitate processing of silver ions apply to Ag nanoparticles, to what extent Ag nanoparticles pass through the intestinal lining intact or are dissolved into silver ions in the highly acidic environment of the stomach, and to what extent Ag nanoparticles can pass through natural biological barriers such as the blood–brain barrier, the placenta or into breast milk. It is also crucial to note that regardless of one’s interpretation of the available body of literature, there have been almost no attempts to study the cumulative effects of chronic Ag nanoparticle exposure, and systematic investigations of the relationship between particle characteristics (size, shape, surface charge, etc.) and toxicity have yet to be performed. (Duncan, 2011) In vitro toxicological studies have shown that Ag nanoparticles may not be benign to isolated mammalian cells (Amparo Llorens, 2012) (Duncan, 2011). Human lung fibroblasts and glioblastoma cells exposed to these particles exhibit reduced ATP

content, increased ROS production, damaged mitochondria, DNA damage and chromosomal aberrations in a dose-dependent manner compared to controls, suggesting that Ag nanoparticles have the potential to be cytotoxic, genotoxic, antiproliferative and possibly carcinogenic. Exposure of spermatogonial mouse stem cells to 15 nm Ag nanoparticles at low (10 µg/l) levels in vitro results in cellular morphological changes and mitochondrial damage, etc., and thus Ag nanoparticles may represent a threat to male reproductive health under some conditions (Duncan, 2011). It has been stated that Ag nanoparticles are safe for skin contact at concentrations up to 6.25 mg/l; however this dosage may be expected to be highly Ag particle-size and -shape dependent, so the absolute usefulness of this mass-based dosage metric is debatable. Despite reports of Ag nanoparticle cytotoxicity, some other studies have arrived at contrary conclusions: one group of researchers found that epidermal cells are unaffected by antimicrobial-relevant concentrations of Ag nanoparticles and several groups have determined that Ag nanoparticles contained in bone-cements cause no observable toxic effects on human osteoblasts under the tested conditions (Duncan, 2011) In relation with environmental aspects, the potential for nanosilver to adversely affect beneficial bacteria in the environment still unclear, there is also a risk that use of silver nanoparticles will lead to the development of resistance among harmful bacteria. Additionally, release of nanosilver can increase silver concentrations in water sources leading to adverse effects, such as bioaccumulation in fish and killing of aquatic life (Satinder K. Brar, 2010). For example, zebra fish embryos show a dose-response relationship (5-100 mg/l) associated with mortality and delayed hatching in the presence of Ag(0) nanoparticles between 5 and 20 nm. The mean lethal concentration varies from 25 to 50 mg/l and depends on the development stage of the embryos. This toxicity appears to be specific to Ag(0) nanoparticles and not of silver ions (Jean-Yves Bottero, 2011). Cytotoxicity and genotoxicity has also been reported in fish cells due to silver nanoparticles accumulating in the gill tissue (Amparo Llorens, 2012). Antimicrobial nanoparticles made of natural materials can be used in place of those containing heavy metals (e.g. nanosilver). Naturally occurring chitin and certain peptides have been long recognized for their antimicrobial properties. It is only recently that these materials have been engineered into nanoparticles. They are promising for low-cost and low-tech disinfection applications. The antibacterial mechanism of natural peptides, such as silk moth cecropins, is the formation of nano-scale channels in

bacterial cell membranes, which causes osmotic collapse. For chitosan several antimicrobial mechanisms have been proposed. One of these mechanisms involves positively charged chitosan particles interacting with negatively charged bacterial cell membranes, causing an increase in membrane permeability and eventually rupture and leakage of intracellular components (Qilin Li, 2008).

7. DEVELOPMENT OF NANOTECHNOLOGY ENHANCED POINT OF USE SYTEMS FOR WATER DISINFECTION

Finally is relevant to mention the product development procedure used T. Pradeep in order to commercialize the first nanotechnology-based point-of-use divice for pesticide removal from drinking water in India, as model of how nanomaterials enhanced POU disinfection systems can be concieved. The first phase of the product development programme was to extend the applicability of materials for pollutans of relevance to local drinking water. The second phase was to adapt the technology to the needs of domestic customers, i.e. online, point-of-use drinking water purifier. An important aspect of the adaptation was to remove pollutans using supported nanoparticles to prevent contamination of drinking water with nanoparticles. This necessitated the study of nanoparticles incorporated on suitable supports. It is important to understand the need for a suitable support, i.e. non-reactivity between support and nanomaterials, availability of large nanomaterial surface area for adsorption, extremely low solubility of support in drinking water and non-toxicity. This phase should also involve the development of a pre-treatment mechanism for contaminated water when required (T. Pradeep, 2009). The third phase was to develop a commercial scale production plant for the systems. The largescale production of nanomaterials or products functionalized with nanomaterials is usually challenged by issues such as precise control over the processes, commercial viability of the process, availability of sophisticated instrumentation for production, quality tests and workforce development for continuous production. A wet route is recomended, because it offers a larger degree of control over production processes. The fourth phase was to develop a suitable mechanism for environmentally responsible disposal of residuals and filter regeneration strategy as well of materials recovery from the spent filters. This entails study of cost-effective regeneration and/or separation of pollutants adsorbed to the filter surface. In addition, the product development involved various levels of field trials, and incorporation of the inputs from the field into the manufacturing process. Product commercialization involved the development of suitable audio-visual aids to inform the public of

various aspects of the technology (T. Pradeep, 2009).

8. REFERENCES  

Nora Savage, M. S. (2005). Nanomaterials and water purification: Opportunities and challenges. Journal of Nanoparticle research , 7, 331-342. S. Malato, P. F.-I. (2009). Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catalysis today , 147, 1-59. Jonathon Brame, Q. L. (2011). Nanotechnology enabled water tretment and reuse: Emerging opportunities and challengesfor developing countries. Trends in Food Science & Technology , 22, 618-624. F. Macedonio, E. D. (2012). Efficient technologies for worldwide clean water supply. Chemical Engineering and Processing , 51, 2-17. Meng Nan Chong, B. J. (2010). Recent developments in photocatalytic water treatment technology: A review. Water Research , 44, 2997-3027. Li, X. (2010). Applications for nanomaterials in critical technologies. University of Illinois. Urbana-Champaign: ProQuest LLC. T. Pradeep, A. (2009). Noble metal nanoparticles for water purification: A critical review. Thin Solid Films , 517, 6441-6478. Venkata K. K. Upadhyayula, S. D. (2009). Application of carbon nanotube technology for removal of contaminants in drinking water: A review. Science of the Total Environment , 408, 1-13. Jungju Lee, H. W. (2011). Adsorption of microcystin-Lr onto iron oxide nanoparticles. Colloids and Surfaces A: Physicochem. Eng. Aspects , 372, 94-100. Anu Matilainen, M. S. (2010). Removal of natural organic matter from drinking water by advanced oxidation processes. Chemosphere , 80, 351-365. E. G. Garrido-Ramirez, B. K. (2010). Clays and oxide minerals as catalysts and nanocatalysts in Fenton-like reactions: A review. Applied Clay Science , 47, 182-192. Peter K. J. Robertson, J. M. (2012). Removal of microorganisms and their chemical metabolites from water using semiconductor photocatalysis. Journal of Hazardous Materials , 211-212, 161-171. R. A. Crane, T. B. (2012). Nanoscale zero-valent iron: Future prospects for an emerging water treatment technology. Journal of Hazardous Materials , 211-212, 112-125.

Satinder K. Brar, M. V. (2010). Engineered nanoparticles in wastewater and wastewater sludge - Evidence and Impacts. Waste Management , 30, 504-520. Grigory Zelmanov, R. S. (2008). Iron(3) oxide-based nanoparticles as catalysts in advanced organic aqueous oxidation. Water Research , 42, 492-498. Melanie Auffan, W. A.-a.-Y. (2008). Relation between the redox state of iron-based nanoparticles and their cytotoxicity toward Eschericha Coli. Environ. Sci. Technol. , 42, 6730-6735. Piao Xu, G. M. (2012). Use of iron oxide nanomaterials in wastewater treatment: A review. Science of the Total Environment , 424, 1-10. Shaily Mahendra, Q. L. (2009). Nanotechnology-enabled water disinfection and microbial control: Merits and limitations. In N. Savage, Nanotechnology applications for clean water (pp. 157-166). Houston, TX: William Andrew Inc. Ying Hui Ngo, D. L. (2011). Paper surfaces functionalized by nanoparticles. Advances in Colloid and Interface Science , 163, 23-38. Agatino Di Paola, E. G.-L. (2012). Survey of photocatalytic materials for environmental remediation. Journal of Hazardous Materials , 211-212, 3-29. Qilin Li, S. M. (2008). Antimicrobial nanomaterials for water disinfection and microbial control: Potential applications and implications. Water Research , 42, 4591-4602. Law Yong Ng, A. W. (2010). Polymeric membranes incorporated with metal/metal oxide nanoparticles: A comprehensive review. Desalination , xxx-xxx. Duncan, T. V. (2011). Applications of nanotechnology in food packing and food safety: Barrier materials, antimicrobials and sensors. Journal of Colloid and interface Science , 363, 1-24. Stefan J. Soenen, P. R.-G.-M. (2011). Cellular toxicity of inorganic nanoparticles: Common aspects and guidelines for improved nanotoxicity evaluation. Nanotoday , 6, 446-465. Jean-Yves Bottero, M. A. (2011). Manufactured metal and metal-oxide nanoparticles: Properties and perturbing mechanisms of their biological activity in ecosystems. Geosciences , 343, 168-176. I. Blinova, A. I. (2010). Ecotoxicity of nanoparticles of CuO and ZnO in natural water. Environmental Pollution , 158, 41-47. Amparo Llorens, E. L. (2012). Metallic-based micro and nanocomposites in food contact

materials and active food packing. Trends in Food Science & Technology , 24, 19-29. Jeonghwan Kim, B. V. (2010). The use of nanoparticles in polymeric and ceramic membrane structures: Review of manufacturing procedures and performance improvement for water treatment. Environmental Pollution , 158, 2335-2349. K. Anselme, P. D. (2010). The interaction of cells and bacteria with surfaces structured at the nanometre scale. Acta Biomaterialia , 6, 3824-3846. K. Glinel, P. T. (2012). Antibacterial surfaces developed from bio-inspired approaches. Acta Biomaterialia , 8, 1670-1684. Chang Halee, J. Y. (2008). Bactericidal effect of zero-valent iron nanoparticles on Escherichia Coli. Environ. Sci. Technol. , 42, 4927-4933. Zhi Qiang Li, K. G. (2010). Adsorbed polymer and NOM limits adhesion and toxicity of nano scale zerovalent iron to E. coli. Environ. Sci. Technol. , 44, 3462-3467. Robert Prucek, J. T. (2011). The targeted antibacterial and antifungal properties of magnetic nanocomposite of iron oxide and silver nanoparticles. Biomaterials , 32, 4704-4713. Joseph N. Ryan, R. W. (2002). Field and laboratory investigations of inactivation of viruses (PRD1 and MS2) attached to iron oxide-coated sand. Environ. Sci. Technol. , 36, 2403-2413. Youwen You, J. H. (2005). Removal and inactivation of waterborne viruses using zerovalent iron. Environ. Sci. Technol. , 39, 9263-9269. Leonardo Gutierrez, X. L. (2009). Adsorption of rotavirus and baceriophage MS2 using glass fiber coated with hematite nanoparticles. Water Research , 43, 5198-5208. Ian Bradley, A. S. (2011). Iron oxide amended biosand filters for virus removal. Water Research , 45, 4501-4510. Brian M. Pecson, L. D. (2012). Photoinactivation of virus on iron-oxide coated sand: Enhancing inactivation in sunlit waters. Water Research , 46, 1763-1770. Chunjian Shi, J. W. (2012). Removal of viruses and bacteriophages from drinking water using zro-valent iron. Separation and Purification Technology , 84, 72-78. H. A. Foster, D. W. (2010). Antimicrobial activity of titania/silver and titania/copper films prepared by CVD. Journal of Photochemistry A: Chemistry , 216, 283-289. Ranjith G. Nair, J. K. (2011). Enhanced visible light photoctalytic disinfection of gram negative, pathogenic Escherichia Coli bacteria with Ag/TiV

oxide nanoparticles. Colloids and Surfaces B: Biointerfaces , 86, 7-13. Zhigang Xiong, J. M. (2011). Silver-modified mesoporous TiO2 photocatalyst for water purification. Water Research , 45, 2095-2103. Lei Liu, Z. L. (2012). Concurrent filtration and solar photocatalytic disinfection/degradation using high-performance Ag/TiO2 nanofiber membrane. Water Research , 46, 1101-1112. Ayman Yousef, N. A.-D.-h. (2012). Inactivation of pathogenic Klebsiella pneumoniae by CuO/TiO2 nanofibers: A multifunctional nanomaterial via one-step electrospining. Ceramics International , xxx-xxx. Min Cho, E. L.-H. (2011). Inactivation and surface interactions of MS2 bacteriophage in TiO2 photoelectrocatalytic reactor. Water research , 45, 2104-2110. Michael V. Liga, E. L. (2011). Virus inactivation by silver doped titanium dioxide nanoparticles for drinking water treatment. Water Research , 45, 535-544. Ken Hirota, M. S. (2010). Preparation of zinc oxide ceramics with a sustainable antibacterial activity under dark conditions. Ceramics International , 36, 497-506. A. Khalil, M. A. (2011). Augmented photocatalytic activity of palladium incorporated ZnO nanoparticles in the disinfection of Escherichia Coli microorganism from water. Applied catalysis A: General , 402, 162-167. Nasrin Talebian, M. R. (2011). Enhanced antibacterial performance of hybrid semiconductor nanomaterials: ZnO/SnO2 nanocomposite thin films. Applied Surface Science , 258, 547-555. C. Karunakaran, V. R. (2011). Optical, electrical, photocatalytic and bactericidal properties of microwave synthesized nanocrystalline Ag-ZnO and ZnO. Solid State Sciences , 13, 923-928. C. Karunakaran, G. A. (2011). Preparation and characterization of ZnO-TiO2 nanocomposite for photocatalytic disinfection of bacteria and detoxification of cyanide under visible light. Materials Research Bulletin , 46, 1586-1592. Tamar Gordon, B. P. (2011). Shynthesis and characterization of zinc/iron oxide composite nanoparticles and their antibacterial properties. Colloids and Surfaces A: Physicochem. Eng. Aspects , 374, 1-8. Peng Yuan, M. F. (2009). Montmorillonite-supported magnetite nanoparticles for temoval of hexavalent chromium [Cr(VI)] from aqueous solutions. Journal of Hazardous materials , 166, 821-829.

Shuzhen Li, P. W. (2010). Synthesis and characterization of organo-montmorillonite supported iron nanoparticles. Applied Clay Science , 50, 330-336. Ray L. Frost, Y. X. (2010). Synthesis, characterization of palygorskite supported zero-valent iron and its aplication for methylene blue adsorption. Journal of Colloid and Interface Science , 341, 153-161. S-M. Lee, C. L. (2012). Iron oxide nano-particles-immobilized-sand material in the treatment of Cu(II), Cd(II) and Pb(II) contaminated waste waters. Chemical Engineering Journal , xxx-xxx. Slavica Lazarevic, I. J.-C. (2012). Removal of Co2+ ions from aqueous solutions using iron-functionalized sepiolite. Chemical Engineering and Processing , 55, 40-47. M. Randelovic, M. P. (2012). Synthesis od composite by application of mixed Fe, Mg (hydro)oxides coatings onto bentonite - A use for the removal of Pb(II) from water. Journal of Hazardous Materials , 199-200, 367-374. Robert Kun, K. M. (2006). Synthesis and structural and photocatalytic properties of TiO2/montmorillonite nanocomposites. Applied Clay Science , 32, 99-110. Elina Manova, P. A.-L.-H. (2010). New titania-clay nanostructured porous materials. Microporous and Mesoporous Materials , 131, 252-260. D. Papoulis, S. K.-K. (2010). Palygorskite- and Halloysite-TiO2 nanocomposites: Synthesis and photocatalytic activity. Applied Clay Science , 50, 118-124. K. Koci, V. M. (2011). Comparison of pure TiO2 and kaolinite/TiO2 composites as catalysts for CO2 photocatalytic reduction. Caalysis Today , 161, 105-109. Lili Zhang, J. L. (2011). Palygorskite and SnO2-TiO2 for the photodegradation of phenol. Applied Clay Science , 51, 68-73. Yalei Zhang, Y. G. (2011). Stable TiO2/rectorite: Preparation, characterization and photocatalytic activity. Applaid Clay Science , 51, 335-340. L. Bouna, B. R.-C. (2011). Synthesis, characterization and photocatalytic activity of TiO2 supported natural palygorskite microfibers. Applied Clay Science , 52, 301-311. W. G. Xu, S. F. (2010). Photocatalytic degradation in aqueous solution using quantum-sized ZnO particles supported on sepiolite. Journal of Colloid and Interface Science , 351, 210-216. Jonas Tokarsky, P. C. (2010). Adhesion of silver nanoparticles on the clay substrates: Modeling

and experiment. Applied Surface Science , 256, 2841-2848. Wenning Shen, L. F. (2011). Ultrafine silver(II) oxide particles decorated porous ceramic composites for water treatment. Chemical Engineering Journal , 175, 592-599. Jaroslaw Drelich, B. L.-Y. (2011). Vermiculite decorated with copper nanoparticles: Novel antibacterial hybrid material. Applied Surface Science , 257, 9435-9443. Markus Wegmann, B. M. (2008). Modification of ceramic microfilters with colloidal zirconia to promote the adsorption of viruses from water. Water Research , 42, 1726-1734. Markus Wegmann, B. M. (2008). Nanostructured surface modification of microporous ceramics for efficient virus filtration. Journal of the European Ceramic Society , 28, 1603-1612. Mark D. Sobsey, C. S. (2008). Point of use household drinking water filtration: A practical, effective solution for providing sustained access to safe drinking water in the developing world. Environ. Sci. Technol. , 42, 4261-4267. Jie Zhuang, Y. J. (2008). Interactions between viruses and goethite during saturated flow: Effects of solution pH, carbonate, and phosphate. Journal of Contaminant Hydrology , 98, 15-21. Michen, B. (2010). Virus removal in ceramic depth filters: The electrostatic enhanced adsorption approach. Technischen Universitat Bergakademie Freiberg, Freiberg. Maryna Peter-Varbanets, C. Z. (2010). Descentralized systems for potable water and the potential of membrane technology. Water Research , 43, 245-265. WHO. (2012). Progress on drinking water and sanitation. World Health Organization, Unicef. WHO. (2011). Guidelines for drinking-water quality, fourth edition. World Health Organization. Kosmulski, M. (2002). The pH-dependent surface charge and the points of zero charge. Journal of Colloid and Interface Science , 253, 77-87. Kosmulski, M. (2004). pH-dependent surface charging and points of zero charge II. Update. Journal of Colloid and Interface Science , 275, 214-224. Kosmulski, M. (2009). pH-dependent surface charging and points of zero charge. IV. Update. Journal of Colloid and Interfase Science , 337, 439-448. Kosmulski, M. (2011). The pH-dependent surface charging and points of zero charge V. Update. Journal of Colloid and Interface Science , 353, 1-15. NHMRC, NRMM, of Australia. (2011). NHMRC, NRMMC (2011) Australian drinking water

guidelines paper 6 national water quality management strategy. Canberra. Yalei Zhang, H. G. (2011). A novel mixed-phase TiO2/kaolinite composites and their photocatalytic activity for degradation of organic contaminants. Chemical Engineering Journal , 172, 936-943. B. Michen, T. G. (2010). Isoelectric points of viruses. Journal of Applied Microbiology (109), 388-397. Hans-Jurgen Butt, B. C. (2005). Force measurements with the atomic force microscope: Technique, interpretation and applications. Surface Science Reports , 59, 1-152.