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Wednesday, 1 March 2017
Applications of nanotechnology in water and wastewater treatment
Published Date
Water Research 1 August 2013, Vol.47(12):3931–3946,doi:10.1016/j.watres.2012.09.058 Nanotechnology for Water and Wastewater Treatment
Author
Xiaolei Qu
Pedro J.J. Alvarez
Qilin Li,
Department of Civil and Environmental Engineering, Rice University, Houston, TX 77005, USA
Received 13 July 2012. Revised 8 September 2012. Accepted 11 September 2012. Available online 26 March 2013.
Highlights
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Nanotechnology offers opportunities to develop next generation water supply systems.
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Applications of nanomaterials in water treatment and reuse are critically reviewed.
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Nanomaterial properties and mechanisms that enable the applications are discussed.
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Advantages and limitations as well as barriers and research needs are highlighted.
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Challenges include technical hurdles, high cost, and environmental and health risk.
Abstract Providing clean and affordable water to meet human needs is a grand challenge of the 21st century. Worldwide, water supply struggles to keep up with the fast growing demand, which is exacerbated by population growth, global climate change, and water quality deterioration. The need for technological innovation to enable integrated water management cannot be overstated. Nanotechnology holds great potential in advancing water and wastewater treatment to improve treatment efficiency as well as to augment water supply through safe use of unconventional water sources. Here we review recent development in nanotechnology for water and wastewater treatment. The discussion covers candidate nanomaterials, properties and mechanisms that enable the applications, advantages and limitations as compared to existing processes, and barriers and research needs for commercialization. By tracing these technological advances to the physicochemical properties of nanomaterials, the present review outlines the opportunities and limitations to further capitalize on these unique properties for sustainable water management. Graphical abstract
Keywords
Nanotechnology
Nanomaterials
Water and wastewater treatment
Water reuse
Sorption
Membrane processes
Photocatalysis
Disinfection
Microbial control
Sensors
Multifunctional
1 Introduction
Water is the most essential substance for all life on earth and a precious resource for human civilization. Reliable access to clean and affordable water is considered one of the most basic humanitarian goals, and remains a major global challenge for the 21st century.
Our current water supply faces enormous challenges, both old and new. Worldwide, some 780 million people still lack access to improved drinking water sources (WHO, 2012). It is urgent to implement basic water treatment in the affected areas (mainly in developing countries) where water and wastewater infrastructure are often non-existent. In both developing and industrialized countries, human activities play an ever-greater role in exacerbating water scarcity by contaminating natural water sources. The increasingly stringent water quality standards, compounded by emerging contaminants, have brought new scrutiny to the existing water treatment and distribution systems widely established in developed countries. The rapidly growing global population and the improvement of living standard continuously drive up the demand. Moreover, global climate change accentuates the already uneven distribution of fresh water, destabilizing the supply. Growing pressure on water supplies makes using unconventional water sources (e.g., stormwater, contaminated fresh water, brackish water, wastewater and seawater) a new norm, especially in historically water-stressed regions. Furthermore, current water and wastewater treatment technologies and infrastructure are reaching their limit for providing adequate water quality to meet human and environmental needs.
Recent advances in nanotechnology offer leapfrogging opportunities to develop next-generation water supply systems. Our current water treatment, distribution, and discharge practices, which heavily rely on conveyance and centralized systems, are no longer sustainable. The highly efficient, modular, and multifunctional processes enabled by nanotechnology are envisaged to provide high performance, affordable water and wastewater treatment solutions that less rely on large infrastructures (Qu et al., 2013). Nanotechnology-enabled water and wastewater treatment promises to not only overcome major challenges faced by existing treatment technologies, but also to provide new treatment capabilities that could allow economic utilization of unconventional water sources to expand the water supply.
Here, we provide an overview of recent advances in nanotechnologies for water and wastewater treatment. The major applications of nanomaterials are critically reviewed based on their functions in unit operation processes. The barriers for their full-scale application and the research needs for overcoming these barriers are also discussed. The potential impact of nanomaterials on human health and ecosystem as well as any potential interference with treatment processes are beyond the scope of this review and thus will not be detailed addressed here.
2 Current and potential applications for water and wastewater treatment
Nanomaterials are typically defined as materials smaller than 100 nm in at least one dimension. At this scale, materials often possess novel size-dependent properties different from their large counterparts, many of which have been explored for applications in water and wastewater treatment. Some of these applications utilize the smoothly scalable size-dependent properties of nanomaterials which relate to the high specific surface area, such as fast dissolution, high reactivity, and strong sorption. Others take advantage of their discontinuous properties, such as superparamagnetism, localized surface plasmon resonance, and quantum confinement effect. These applications are discussed below based on nanomaterial functions in unit operation processes (Table 1). Most applications discussed below are still in the stage of laboratory research. The pilot-tested or field-tested exceptions will be noted in the text.
Table 1. Current and potential applications of nanotechnology in water and wastewater treatment.
Applications
Representative nanomaterials
Desirable nanomaterial properties
Enabled technologies
Adsorption
Carbon nanotubes
High specific surface area, highly assessable adsorption sites, diverse contaminant-CNT interactions, tunable surface chemistry, easy reuse
Contaminant preconcentration/detection, adsorption of recalcitrant contaminants
Nanoscale metal oxide
High specific surface area, short intraparticle diffusion distance, more adsorption sites, compressible without significant surface area reduction, easy reuse, some are superparamagnetic
Adsorptive media filters, slurry reactors
Nanofibers with core–shell structure
Tailored shell surface chemistry for selective adsorption, reactive core for degradation, short internal diffusion distance
Reactive nano-adsorbents
Membranes and membrane processes
Nano-zeolites
Molecular sieve, hydrophilicity
High permeability thin film nanocomposite membranes
Nano-Ag
Strong and wide-spectrum antimicrobial activity, low toxicity to humans
Adsorption is commonly employed as a polishing step to remove organic and inorganic contaminants in water and wastewater treatment. Efficiency of conventional adsorbents is usually limited by the surface area or active sites, the lack of selectivity, and the adsorption kinetics. Nano-adsorbents offer significant improvement with their extremely high specific surface area and associated sorption sites, short intraparticle diffusion distance, and tunable pore size and surface chemistry.
2.1.1 Carbon based nano-adsorbents
2.1.1.1 Organic removal
CNTs have shown higher efficiency than activated carbon on adsorption of various organic chemicals (Pan and Xing, 2008). Its high adsorption capacity mainly stems from the large specific surface area and the diverse contaminant–CNT interactions. The available surface area for adsorption on individual CNTs is their external surfaces (Yang and Xing, 2010). In the aqueous phase, CNTs form loose bundles/aggregates due to the hydrophobicity of their graphitic surface, reducing the effective surface area. On the other hand, CNT aggregates contain interstitial spaces and grooves, which are high adsorption energy sites for organic molecules (Pan et al., 2008). Although activated carbon possesses comparable measured specific surface area as CNT bundles, it contains a significant number of micropores inaccessible to bulky organic molecules such as many antibiotics and pharmaceuticals (Ji et al., 2009). Thus CNTs have much higher adsorption capacity for some bulky organic molecules because of their larger pores in bundles and more accessible sorption sites.
A major drawback of activated carbon is its low adsorption affinity for low molecular weight polar organic compounds. CNTs strongly adsorb many of these polar organic compounds due to the diverse contaminant–CNT interactions including hydrophobic effect, π–π interactions, hydrogen bonding, covalent bonding, and electrostatic interactions (Yang and Xing, 2010). The π electron rich CNT surface allows π–π interactions with organic molecules with CC bonds or benzene rings, such as polycyclic aromatic hydrocarbons (PAHs) and polar aromatic compounds (Chen et al., 2007; Lin and Xing, 2008). Organic compounds which have –COOH, –OH, –NH2 functional groups could also form hydrogen bond with the graphitic CNT surface which donates electrons (Yang et al., 2008). Electrostatic attraction facilitates the adsorption of positively charged organic chemicals such as some antibiotics at suitable pH (Ji et al., 2009).
2.1.1.2 Heavy metal removal
Oxidized CNTs have high adsorption capacity for metal ions with fast kinetics. The surface functional groups (e.g., carboxyl, hydroxyl, and phenol) of CNTs are the major adsorption sites for metal ions, mainly through electrostatic attraction and chemical bonding (Rao et al., 2007). As a result, surface oxidation can significantly enhance the adsorption capacity of CNTs. Several studies show that CNTs are better adsorbents than activated carbon for heavy metals (e.g., Cu2+, Pb2+, Cd2+, and Zn2+) (Li et al., 2003; Lu et al., 2006) and the adsorption kinetics is fast on CNTs due to the highly accessible adsorption sites and the short intraparticle diffusion distance.
Overall, CNTs may not be a good alternative for activated carbon as wide-spectrum adsorbents. Rather, as their surface chemistry can be tuned to target specific contaminants, they may have unique applications in polishing steps to remove recalcitrant compounds or in pre-concentration of trace organic contaminants for analytical purposes. These applications require small quantity of materials and hence are less sensitive to the material cost.
Produced by exfoliating graphite with strong acids and oxidizers, graphite oxide is a potentially low-cost adsorbent. It was recently reported that sand granules coated with graphite oxide was efficient in removing Hg2+ and a bulky dye molecule (Rhodamine B); its performance was comparable to commercial activated carbon (Gao et al., 2011).
2.1.1.3 Regeneration and reuse
Regeneration is an important factor that determines the cost-effectiveness of adsorbents. Adsorption of metal ions on CNTs can be easily reversed by reducing the solution pH. The metal recovery rate is usually above 90% and often close to 100% at pH < 2 (Li et al., 2005; Lu et al., 2006). Moreover, the adsorption capacity remains relatively stable after regeneration. Lu et al. reported that Zn2+ adsorption capacity of SWNT and MWNT decreased less than 25% after 10 regeneration and reuse cycles, while, that of activated carbon was reduced by more than 50% after one regeneration (Lu et al., 2006). A statistical analysis based on the best-fit regression of Zn2+adsorption capacity and the number of regeneration and reuse cycles suggested that CNT nano-adsorbents can be regenerated and reused up to several hundred times for Zn2+ removal while maintaining reasonable adsorption capacity (Lu et al., 2007).
2.1.2 Metal based nano-adsorbents
Metal oxides such as iron oxide, titanium dioxide and alumina are effective, low cost adsorbents for heavy metals and radionuclides. The sorption is mainly controlled by complexation between dissolved metals and the oxygen in metal oxides (Koeppenkastrop and Decarlo, 1993). It is a two-step process: fast adsorption of metal ions on the external surface, followed by the rate-limiting intraparticle diffusion along the micropore walls (Trivedi and Axe, 2000). Their nanoscale counterparts have higher adsorption capacity and faster kinetics because of the higher specific surface area, shorter intraparticle diffusion distance and larger number of surface reaction sites (i.e., corners, edges, vacancies). For instance, as the particle size of nano-magnetite decreased from 300 to 11 nm, its arsenic adsorption capacity increased more than 100 times (Yean et al., 2005). Much of this observed increase in adsorption was attributed to the increase in specific surface area as the 300-nm and 20-nm magnetite particles have similar surface area normalized arsenic adsorption capacity (∼6 μmol m−2 or 3.6 atoms nm−2) (Auffan et al., 2009, 2008). However, when particle size was reduced to below 20 nm, the specific surface area normalized adsorption capacity increased, with 11-nm magnetite nanoparticles absorbing three times more arsenic (∼18 μmol m−2 or 11 atoms nm−2), suggesting a “nanoscale effect”. This “nanoscale effect” was attributed to the change of magnetite surface structure which creates new adsorption sites (vacancies) (Auffan et al., 2009).
In addition to high adsorption capacity, some iron oxide nanoparticles, e.g., nano-maghemite and nano-magnetite, can be superparamagnetic. Magnetism is highly volume-dependent as it stems from the collective interaction of atomic magnetic dipoles. If the size of a ferro- or ferri-magnet decreases to the critical value (∼40 nm), the magnet changes from multiple domains to single domain with higher magnetic susceptibility (Yavuz et al., 2006). As the size further decreases, magnetic particles become superparamagnetic, losing permanent magnetic moments while responding to an external magnetic field, which allows easy separation and recovery by a low-gradient magnetic field. These magnetic nanoparticles can be either used directly as adsorbents or as the core material in a core–shell nanoparticle structure where the shell provides the desired function while the magnetic core realizes magnetic separation (Fig. 1).
Fig. 1. Multifunctional magnetic nanoparticles. Magnetic nanoparticles are used as the core material in a core–shell nanoparticle structure where the shell provides the desired function while the magnetic core realizes magnetic separation. Silica coating helps functionalization due to the rich silica chemistry.
Metal oxide nanocrystals can be compressed into porous pellets without significantly compromising their surface area when moderate pressure is applied (Lucas et al., 2001). The pore volume and pore size can be controlled by adjusting the consolidation pressure. Thus, they can be applied in forms of both fine powders and porous pellets, which are the likely forms to be used in industry.
Metal based nanomaterials have been explored to remove a variety of heavy metals such as arsenic, lead, mercury, copper, cadmium, chromium, nickel, and have shown great potential to outcompete activated carbon (Sharma et al., 2009). Among them, the application for arsenic removal has attracted much attention. Although a good adsorbent for many organic and inorganic contaminants, activated carbon has limited capacity for arsenic, especially for As(V) (Daus et al., 2004). Several metal oxide nanomaterials including nanosized magnetite and TiO2 have shown arsenic adsorption performance superior to activated carbon (Deliyanni et al., 2003; Mayo et al., 2007). Metal (hydr)oxide nanoparticles also can be impregnated onto the skeleton of activated carbon or other porous materials to achieve simultaneously removal of arsenic and organic co-contaminants, which favors point-of-use (POU) applications (Hristovski et al., 2009a, 2009b).
2.1.2.1 Regeneration and reuse
Metal oxide nano-adsorbents can be easily regenerated by changing solution pH (Sharma et al., 2009). In many cases, the adsorption capacity of metal oxide nano-adsorbents is well maintained after several regeneration and reuse cycles (Hu et al., 2006). However, reduced adsorption capacity after regeneration has also been reported (Deliyanni et al., 2003).
Above all, metal based nano-adsorbents can be produced at relatively low cost. The high adsorption capacity, low cost, easy separation and regeneration make metal based nano-adsorbents technologically and economically advantageous.
2.1.3 Polymeric nano-adsorbents
Dendrimers are tailored adsorbents that are capable of removing both organics and heavy metals. Their interior shells can be hydrophobic for sorption of organic compounds while the exterior branches can be tailored (e.g., hydroxyl- or amine-terminated) for adsorption of heavy metals. The sorption can be based on complexation, electrostatic interactions, hydrophobic effect, and hydrogen bonding (Crooks et al., 2001). A dendrimer-ultrafiltration system was designed to recover metal ions from aqueous solutions (Diallo et al., 2005). The system achieved almost complete removal of Cu2+ ions with initial concentration of 10 ppm and Cu2+ to PAMAM dendrimer-NH2 ratio of 0.2. After adsorption, the metal ion laden dendrimers were recovered by ultrafiltration and regenerated by decreasing pH to 4.
2.1.4 Potential application in water treatment
Nano-adsorbents can be readily integrated into existing treatment processes in slurry reactors or adsorbers. Applied in the powder form, nano-adsorbents in slurry reactors can be highly efficient since all surfaces of the adsorbents are utilized and the mixing greatly facilitates the mass transfer. However, an additional separation unit is required to recover the nanoparticles. Nano-adsorbents can also be used in fixed or fluidized adsorbers in the form of pellets/beads or porous granules loaded with nano-adsorbents. Fixed-bed reactors are usually associated with mass transfer limitations and head loss; but it doesn't need future separation process. Applications of nano-adsorbents for arsenic removal have been commercialized, and their performance and cost have been compared to other commercial adsorbents in pilot tests (Aragon et al., 2007). ArsenXnp is a commercial hybrid ion exchange medium comprising of iron oxide nanoparticles and polymers. ADSORBSIA™ is a nanocrystalline titanium dioxide medium in the form of beads from 0.25 to 1.2 mm in diameter. Both nano-adsorbents were highly efficient in removing arsenic and ArsenXnp required little backwash (Aragon et al., 2007; Sylvester et al., 2007). The estimated treatment cost for ArsenXnp is $0.25∼$0.35/1000 gal if the medium is regenerated, similar to $0.37/1000 gal of Bayoxide E33, a high-performance granular iron oxide adsorbent (Aragon et al., 2007; Westerhoff et al., 2006). ArsenXnp and ADSORBSIA™ have been employed in small to medium scale drinking water treatment systems and were proven to be cost-competitive.
2.2 Membranes and membrane processes
The basic goal of water treatment is to remove undesired constituents from water. Membranes provide a physical barrier for such constituents based on their size, allowing use of unconventional water sources. As the key component of water treatment and reuse, they provide high level of automation, require less land and chemical use, and the modular configuration allows flexible design (Qu et al., 2013).
A major challenge of the membrane technology is the inherent tradeoff between membrane selectivity and permeability. The high energy consumption is an important barrier to the wide application of pressure driven membrane processes. Membrane fouling adds to the energy consumption and the complexity of the process design and operation. Furthermore, it reduces the lifetime of membranes and membrane modules.
The performance of membrane systems is largely decided by the membrane material. Incorporation of functional nanomaterials into membranes offers a great opportunity to improve the membrane permeability, fouling resistance, mechanical and thermal stability, as well as to render new functions for contaminant degradation and self-cleaning.
2.2.1 Nanofiber membranes
Electrospinning is a simple, efficient and inexpensive way to make ultra fine fibers using various materials (e.g., polymers, ceramics, or even metals) (Cloete et al., 2010; Li and Xia, 2004). The resulting nanofibers have high specific surface area and porosity and form nanofiber mats with complex pore structures. The diameter, morphology, composition, secondary structure, and spatial alignment of electrospun nanofibers can be easily manipulated for specific applications (Li and Xia, 2004). Although nanofiber membranes have been commercially employed for air filtration applications, their potential in water treatment is still largely unexploited. Nanofiber membranes can remove micron-sized particles from aqueous phase at a high rejection rate without significant fouling (Ramakrishna et al., 2006). Thus they have been proposed to be used as pretreatment prior to ultrafiltration or reverse osmosis (RO). Functional nanomaterials can be easily doped into the spinning solutions to fabricate nanoparticle impregnated nanofibers or formed in situ (Li and Xia, 2004). The outstanding features and tunable properties make electrospun nanofibers an ideal platform for constructing multifunctional media/membrane filters by either directly using intrinsically multifunctional materials such as TiO2 or by introducing functional materials on the nanofibers. For example, by incorporating ceramic nanomaterials or specific capture agents on the nanofiber scaffold, affinity nanofiber membranes can be designed to remove heavy metals and organic pollutants during filtration.
2.2.2 Nanocomposite membranes
A significant number of studies on membrane nanotechnology have focused on creating synergism or multifunction by adding nanomaterials into polymeric or inorganic membranes. Nanomaterials used for such applications include hydrophilic metal oxide nanoparticles (e.g., Al2O3, TiO2, and zeolite), antimicrobial nanoparticles (e.g., nano-Ag and CNTs), and (photo)catalytic nanomaterials (e.g., bi-metallic nanoparticles, TiO2).
The main goal of adding hydrophilic metal oxide nanoparticles is to reduce fouling by increasing the hydrophilicity of the membrane. The addition of metal oxide nanoparticles including alumina (Maximous et al., 2010), silica (Bottino et al., 2001), zeolite (Pendergast et al., 2010) and TiO2 (Bae and Tak, 2005) to polymeric ultrafiltration membranes has been shown to increase membrane surface hydrophilicity, water permeability, or fouling resistance. These inorganic nanoparticles also help enhance the mechanical and thermal stability of polymeric membranes, reducing the negative impact of compaction and heat on membrane permeability (Ebert et al., 2004; Pendergast et al., 2010).
Antimicrobial nanomaterials such as nano-Ag and CNTs can reduce membrane biofouling. Nano-Ag has been doped or surface grafted on polymeric membranes to inhibit bacterial attachment and biofilm formation (Mauter et al., 2011; Zodrow et al., 2009) on the membrane surface as well as inactivate viruses (De Gusseme et al., 2011). However, its long-term efficacy against membrane biofouling has not been reported. Appropriate replenishment of nano-Ag needs to be addressed for practical application of this technology. CNTs inactivate bacteria upon direct contact (Brady-Estevez et al., 2008). High bacterial inactivation (>90%) has been achieved using polyvinyl-N-carbazole-SWNT nanocomposite at 3 wt% of SWNT (Ahmed et al., 2012). As CNTs are insoluble in water and not consumed, there is no need for replenishment. However, as direct contact is required for inactivation, long term filtration experiments are needed to determine the impact of fouling on the antimicrobial activity of CNTs. Addition of oxidized MWNT at low weight percentage (up to 1.5 wt%) also increases the hydrophilicity and permeability of polysulfone membranes (Choi et al., 2006b).
(Photo)catalytic nanoparticle incorporated membranes (a.k.a. reactive membranes) combine their physical separation function and the reactivity of a catalyst toward contaminant degradation. Much effort has been devoted to develop photocatalytic inorganic membranes consisting of nano-photocatalysts (normally nano-TiO2 or modified nano-TiO2) (Choi et al., 2006a). Metallic/bi-metallic catalyst nanoparticles such as nano zero-valent iron (nZVI) and noble metals supported on nZVI have been incorporated into polymeric membranes for reductive degradation of contaminants, particularly chlorinated compounds (Wu et al., 2005; Wu and Ritchie, 2008). nZVI serves as the electron donor and the noble metals catalyze the reaction.
2.2.3 Thin film nanocomposite (TFN) membranes
Development of TFN membranes mainly focuses on incorporating nanomaterials into the active layer of thin film composite (TFC) membranes via doping in the casting solutions or surface modification. Nanomaterials that have been researched for such applications include nano-zeolites, nano-Ag, nano-TiO2, and CNTs. The impact of nanoparticles on membrane permeability and selectivity depends on the type, size and amount of nanoparticles added.
Nano-zeolites are the most frequently used dopants in TFN and have shown potential in enhancing membrane permeability. The addition of nano-zeolites leads to more permeable, negatively charged, and thicker polyamide active layer (Lind et al., 2009a). One study reported water permeability increased up to 80% over the TFC membrane, with the salt rejection largely maintained (93.9 ± 0.3%) (Jeong et al., 2007). TFN membranes doped with 250 nm nano-zeolites at 0.2 wt% achieved moderately higher permeability and better salt rejection (>99.4%) than commercial RO membranes (Lind et al., 2010). It was hypothesized that the small, hydrophilic pores of nano-zeolites create preferential paths for water. However, water permeability increased even with pore-filled zeolites, although less than the pore-open ones, which could be attributed to defects at the zeolite–polymer interface. Nano-zeolites were also used as carriers for antimicrobial agents such as Ag+, which imparts anti-fouling property to the membrane (Lind et al., 2009b). The zeolite TFN technology has reached the early stage of commercialization. QuantumFlux, a seawater TFN RO membrane, is now commercially available (www.nanoH2O.com).
Incorporation of nano-TiO2 (up to 5 wt%) into the TFC active layer slightly increased the membrane rejection while maintaining the permeability (Lee et al., 2008). When the concentration of nano-TiO2 exceeded 5 wt%, the water flux increased in the cost of reducing rejection, suggesting defect formation in the active layer. Upon UV irradiation, TiO2 can degrade organic contaminants and inactivate microorganisms. This helps reduce organic and biological fouling as well as remove contaminants that are not retained by the membrane. However, the close adjacency between the photocatalyst and the membrane may also lead to detrimental effects on polymeric membrane materials, which needs to be addressed for long-term efficacy (Chin et al., 2006).
CNTs (unaligned) also found their application in TFN membranes due to their antimicrobial activities. Tiraferri et al. covalently bonded SWNTs to a TFC membrane surface (Tiraferri et al., 2011). This approach is advantageous as it uses relatively small amount of the nanomaterial and minimizes perturbation of the active layer. The resulting TFN membrane exhibited moderate anti-bacterial properties (60% inactivation of bacteria attached on the membrane surface in 1 h contact time), potentially reducing or delaying membrane biofouling.
2.2.4 Biologically inspired membranes
Many biological membranes are highly selective and permeable. Aquaporins are protein channels that regulate water flux across cell membranes. Their high selectivity and water permeability makes their use in polymeric membranes an attractive approach to improve membrane performance. Aquaporin-Z from Escherichia coli has been incorporated into amphiphilic triblock-polymer vesicles (Kumar et al., 2007), which exhibit water permeability at least an order of magnitude over the original vesicles with full rejection to glucose, glycerol, salt, and urea. One potential design is to coat aquaporin incorporated lipid bilayers on commercial nanofiltration membranes. On this front, limited success was achieved (Kaufman et al., 2010).
Aligned CNTs have been shown both experimentally and theoretically to provide water permeation much faster than what the Hagen–Poiseuille equation predicts, owing to the atomic smoothness of the nano-sized channel, and the one dimensional single-file ordering of water molecules while passing through the nanotubes (Holt et al., 2006; Hummer et al., 2001). It was predicted that a membrane containing only 0.03% surface area of aligned CNTs will have flux exceeding current commercial seawater RO membranes (Pendergast and Hoek, 2011). However, high rejection for salt and small molecules is challenging for aligned CNT membranes due to the lack of CNTs with uniformly sub-nanometer diameter. Functional group gating at the nanotube opening has been proposed to enhance the selectivity of aligned CNT membranes (Mauter and Elimelech, 2008). By grafting carboxyl functional groups on sub-2-nm CNT openings, 98% rejection of was achieved at low ionic strength by Donnan exclusion (Fornasiero et al., 2008). However, KCl rejection was only 50% at 0.3 mM, and decreased to almost zero at 10 mM. Grafting bulky functional groups at the tube opening could physically exclude salts. However, steric exclusion will significantly reduce membrane permeability (Nednoor et al., 2005). Thus at the current stage, aligned CNT membranes are not capable of desalination. To achieve reliable salt rejection, the CNT diameter must be uniformly smaller than 0.8 nm (Hinds, 2012).
A key barrier for both aquaporin and aligned CNT membranes is the scale-up of the nanomaterial production and membrane fabrication. Large-scale production and purification of aquaporins are very challenging. To date, chemical vapor deposition (CVD) is the most common way to make aligned nanotubes. A continuous high-yield CVD prototype has been designed for producing vertically aligned CNT, paving the way for large-scale production (de Villoria et al., 2011). A post-manufacturing alignment method using magnetic field was also developed (Mauter et al., 2010).
Nanocomposite and TFN membranes have good scalability as they can be fabricated using current industrial manufacturing processes. The high water permeability can reduce the applied pressure or required membrane area and consequently cut cost. This strategy may greatly improve the energy efficiency for treatment of waters with low osmosis pressure, but it may have limited advantage in seawater RO, whose energy consumption is already close to the thermodynamic limit (Elimelech and Phillip, 2011). A recent review ranked current membrane nanotechnologies based on their potential performance enhancement and state of commercial readiness (Pendergast and Hoek, 2011).
2.2.5 Forward osmosis
Forward osmosis (FO) utilizes the osmotic gradient to draw water from a low osmotic pressure solution to a high osmotic pressure one (i.e., the draw solution). The diluted draw solution is then treated by reverse osmosis or thermal processes to generate pure water. FO has two major advantages over the pressure-driven reverse osmosis: it does not require high pressure, and the membrane is less prone to fouling.
The key to FO is to have a draw solute with high osmolality and easily separable from water. Chemicals currently employed for draw solutions include NaCl and ammonia bicarbonate. Therefore, RO or thermal treatment, both energy intensive, is required to recover water from the draw solution. Magnetic nanoparticles were recently explored as a new type of draw solute for its easy separation and reuse. Hydrophilic coating was employed to aid dissolution and increase osmotic pressure. An FO permeate flux higher than 10 L m−2 h−1 was achieved using 0.065 M poly(ethylene glycol) diacid-coated magnetic nanoparticles when deionized water was used as the feed solution (Ge et al., 2011). Magnetic nanoparticles were also applied to recover draw solutes. In a recent study, magnetic nanoparticles (Fe3O4@SiO2) were used to recover Al2(SO4)3 (the draw solute) through flocculation (Liu et al., 2011c).
2.3 (Photo)catalysis
Photocatalytic oxidation is an advanced oxidation process for removal of trace contaminants and microbial pathogens. It is a useful pretreatment for hazardous and non-biodegradable contaminants to enhance their biodegradability. Photocatalysis can also be used as a polishing step to treat recalcitrant organic compounds. The major barrier for its wide application is the slow kinetics due to limited light fluence and photocatalytic activity. Current research focuses on increasing photocatalytic reaction kinetics and photoactivity range (Table 2).
Table 2. TiO2 photocatalyst optimization.
Optimization objectives
Optimization approaches
Optimization mechanisms
Water treatment applications
Enhance photocatalytic reaction kinetics
Size
More surface reactive sites, higher reactant adsorption, lower electron-hole recombination
High performance UV activated photocatalytic reactors
Nanotube morphology
Shorter carrier-diffusion paths in the tube walls, higher reactant mass transfer rate toward tube surface
Low energy cost solar/visible light activated photocatalytic reactors
Anion doping
Band gap narrowing
Dye sensitizer doping
Electron injection
Narrow band-gap semiconductors doping
Electron injection
2.3.1 Nano-photocatalyst optimization
TiO2 is the most widely used semiconductor photocatalyst in water/wastewater treatment owing to its low toxicity, chemical stability, low cost, and abundance as raw material. It generates an electron/hole (e−/h+) pair upon absorbing a UV photon, which later either migrate to the surface and form reactive oxygen species (ROS) or undergo undesired recombination. The photoactivity of nano-TiO2 can be improved by optimizing particle size and shape, reducing e−/h+ recombination by noble metal doping, maximizing reactive facets, and surface treatment to enhance contaminant adsorption.
The size of TiO2 plays an important role in its solid-phase transformation, sorption, and e−/h+ dynamics. Among the crystalline structures of TiO2, rutile is the most stable for particles larger than 35 nm, while anatase, which is more efficient in producing ROS, is the most stable for particles smaller than 11 nm (Fujishima et al., 2008; Zhang and Banfield, 2000). A major cause for the slow reaction kinetics of TiO2photocatalysis is the fast recombination of e− and h+. Decreasing TiO2 particle size lowers volume recombination of e−/h+, and enhances interfacial charge carrier transfer (Zhang et al., 1998). However, when particle size is reduced to several nanometers, surface recombination dominates, decreasing photocatalytic activity. Therefore, the photocatalytic activity of TiO2 has a maximum due to the interplay of the aforementioned mechanisms, which lies in the nanometer range. TiO2 nanotubes were found to be more efficient than TiO2 nanoparticles in decomposition of organic compounds (Macak et al., 2007). The higher photocatalytic activity was attributed to the shorter carrier-diffusion paths in the tube walls and faster mass transfer of reactants toward the nanotube surface.
Noble metal doping can reduce the e−/h+ recombination because the photo-excited electrons tend to migrate to the noble metals with lower Fermi levels while the holes stay in TiO2 (Ni et al., 2007). The photocatalytic activity of TiO2 can also be promoted by creating highly reactive crystallographic facets. Because high-energy {001} facets diminish quickly during crystal growth, anatase TiO2 is usually dominated by the low-energy {101} facets. Using specific capping agent (usually fluoride), the percentage of {001} facets can be increased from less than 10% to up to 89% (Han et al., 2009), substantially enhancing hydroxyl radical production and organic compound decomposition (Han et al., 2009; Murakami et al., 2009). The enhanced activity stems from the strong adsorption of reactants on high-energy facets (Liu et al., 2011b) and the spatial separation of electrons and holes on specific crystal facets (Murakami et al., 2009). The optimal percentage of {001} facets for photocatalysis is still debated (Liu et al., 2011b). Improving contaminant adsorption by modifying photocatalyst surface is another way to enhance photocatalytic activity due to the short life time of ROS. However, little has been done in this area.
Another actively pursued research area is to extend the excitation spectrum of TiO2 to include visible light. The general strategy is doping metal impurities, dye sensitizers, narrow band-gap semiconductors, or anions into nano-TiO2 to form hybrid nanoparticles or nanocomposites (Fujishima et al., 2008; Ni et al., 2007). Metals and anions create impurity energy levels or narrow the band gap; upon visible light excitation, dye sensitizers and narrow band-gap semiconductors inject electrons into TiO2 to initiate the catalytic reactions. Among these methods, anions (especially nitrogen) doping, was considered most cost-effective and feasible for industrial applications (Fujishima et al., 2008), although their stability and long-term efficacy has not been tested. Decreased nitrogen concentration during photocatalysis has been reported (Kitano et al., 2006).
Other than TiO2, WO3 and some fullerene derivatives also have the potential to be used in photocatalytic water treatment. WO3 has a narrower band gap than TiO2, allowing it to be activated by visible light (<450 nm) (Kominami et al., 2001). Pt doping further enhances WO3 reactivity by facilitating multi-electron reduction of O2and improving e−/h+ separation (Kim et al., 2010). Aminofullerenes generate 1O2under visible light irradiation (<550 nm) (Lof et al., 1995) and has been studied to degrade pharmaceutical compounds and inactivate viruses (Lee et al., 2010). Fullerol and C60 encapsulated with poly(N-vinylpyrrolidone) can produce 1O2 and superoxide under UVA light (Brunet et al., 2009). Aminofullerenes are more amenable to immobilization than fullerol and are more effective for disinfection purposes due to their positive charge. 1O2 has lower oxidation potential than hydroxyl radicals produced by TiO2, while it is a more selective ROS and consequently less susceptible to quenching by non-target background organic matter. Fullerenes are currently much more expensive and not as readily available as TiO2.
2.3.2 Potential applications in water treatment
The overall efficiency of a photocatalytic water treatment process strongly depends on the configuration and operation parameters of the photo-reactor. Two configurations are commonly used: slurry reactors and reactors using immobilized TiO2. Various dispersion/recovery or catalyst immobilization techniques are being pursued to maximize its efficiency. Extensive investigation on operating parameters has been carried out with these lab or pilot scale systems. A recent critical review outlines the effects of water quality and a wide range of operating parameters including TiO2 loading, pH, temperature, dissolved oxygen, contaminant type and concentration, light wavelength and intensity (Chong et al., 2010). Readers are referred to this review for details regarding process optimization. A commercial product, Purifics Photo-Cat™ system, has treatment capacity as high as 2 million gallon per day with a small footprint of 678 ft2. Pilot tests showed that the Photo-Cat™ system is highly efficient for removing organics without producing waste streams and it operates with relatively low specific power consumption of about 4 kWh/m3 (Al-Bastaki, 2004; Benotti et al., 2009; Westerhoff et al., 2009). Nano-TiO2 facilitated solar disinfection (SODIS) has been extensively tested and appears to be a feasible option to produce safe drinking water in remote areas of developing countries. The SODIS system can be small scale for one person or scaled up to medium size solar compound parabolic collectors.
Photocatalysis has shown great potential as a low-cost, environmental friendly and sustainable water treatment technology. However, there are several technical challenges for its large scale application, including 1) catalyst optimization to improve quantum yield or to utilize visible light; 2) efficient photocatalytic reactor design and catalyst recovery/immobilization techniques; 3) better reaction selectivity.
Metal oxide nanomaterials such as TiO2 and CeO2 as well as carbon nanotubes have been studied as catalysts in heterogeneous catalytic ozonation processes that provide fast and comparatively complete degradation of organic pollutants. Both radical-mediated and non-radical-mediated reaction pathways have been proposed (Nawrocki and Kasprzyk-Hordern, 2010). The adsorption of ozone and/or pollutants on the catalyst surface plays a critical role in both mechanisms. Nanomaterials have large specific surface area and an easily accessible surface, leading to high catalytic activity. Some nanomaterials were also reported to promote decomposition of ozone into hydroxyl radicals, facilitating degradation process through radical-mediated routes (Orge et al., 2011). For future industrial scale applications, a better understanding of the mechanism of nanomaterial enabled catalytic ozonation is in critical need.
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C bonds or benzene rings, such as polycyclic aromatic hydrocarbons (PAHs) and polar aromatic compounds (Chen et al., 2007; Lin and Xing, 2008). Organic compounds which have –COOH, –OH, –NH2 functional groups could also form hydrogen bond with the graphitic CNT surface which donates electrons (Yang et al., 2008). Electrostatic attraction facilitates the adsorption of positively charged organic chemicals such as some antibiotics at suitable pH (Ji et al., 2009).
was achieved at low ionic strength by Donnan exclusion (Fornasiero et al., 2008). However, KCl rejection was only 50% at 0.3 mM, and decreased to almost zero at 10 mM. Grafting bulky functional groups at the tube opening could physically exclude salts. However, steric exclusion will significantly reduce membrane permeability (Nednoor et al., 2005). Thus at the current stage, aligned CNT membranes are not capable of desalination. To achieve reliable salt rejection, the CNT diameter must be uniformly smaller than 0.8 nm (Hinds, 2012).
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