w a t e r r e s e a r c h 5 8 ( 2 0 1 4 ) 1 7 9 e1 9 7

Available online at www.sciencedirect.com

ScienceDirect journal homepage: www.elsevier.com/locate/watres

Review

Forward osmosis for application in wastewater treatment: A review Kerusha Lutchmiah a,b,*, A.R.D. Verliefde a,c, K. Roest b, L.C. Rietveld a, E.R. Cornelissen b a

Delft University of Technology, Dept. of Water Management, Section Sanitary Engineering, Stevinweg 1, 2628CN Delft, The Netherlands b KWR Watercycle Research Institute, Post Box 1072, 3430 BB Nieuwegein, The Netherlands c Ghent University, Particle and Interfacial Technology Group, Coupure Links 653, 9000 Ghent, Belgium

article info

abstract

Article history:

Research in the field of Forward Osmosis (FO) membrane technology has grown significantly

Received 12 December 2013

over the last 10 years, but its application in the scope of wastewater treatment has been slower.

Received in revised form

Drinking water is becoming an increasingly marginal resource. Substituting drinking water for

14 March 2014

alternate water sources, specifically for use in industrial processes, may alleviate the global

Accepted 17 March 2014

water stress. FO has the potential to sustainably treat wastewater sources and produce high

Available online 4 April 2014

quality water. FO relies on the osmotic pressure difference across the membrane to extract clean water from the feed, however the FO step is still mostly perceived as a “pre-treatment”

Keywords:

process. To prompt FO-wastewater feasibility, the focus lies with new membrane de-

Draw solution

velopments, draw solutions to enhance wastewater treatment and energy recovery, and

Energy generation

operating conditions. Optimisation of these parameters are essential to mitigate fouling,

Forward osmosis

decrease concentration polarisation and increase FO performance; issues all closely related to

Hybrid systems

one another. This review attempts to define the steps still required for FO to reach full-scale

Membrane fouling

potential in wastewater treatment and water reclamation by discussing current novelties,

Wastewater treatment

bottlenecks and future perspectives of FO technology in the wastewater sector. ª 2014 Elsevier Ltd. All rights reserved.

Contents 1. 2.

3.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 180 FO Membranes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182 2.1. Membrane materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 182 2.2. Membrane development . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183 2.3. Membrane orientation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183 Modelling membrane transport . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183 3.1. Concentration polarisation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

* Corresponding author. Delft University of Technology, Dept. of Water Management, Section Sanitary Engineering, Stevinweg 1, 2628CN Delft, The Netherlands. Tel.: þ31 (0)30 60 69 585; fax: þ31 (0)30 60 61 165. E-mail address: [email protected] (K. Lutchmiah). http://dx.doi.org/10.1016/j.watres.2014.03.045 0043-1354/ª 2014 Elsevier Ltd. All rights reserved.

180

4.

5.

6.

7.

8.

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3.2. Membrane structure parameter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186 3.3. Membrane permeability constants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 186 Draw solutions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187 4.1. Effective draw solutions for wastewater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 187 4.2. Recovery systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188 Process conditions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188 5.1. Temperature and pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 188 5.2. Cross-flow velocity and spacer designs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189 5.3. Module configuration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189 Membrane fouling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 189 6.1. Wastewater foulants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 190 6.2. The causes of FO fouling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 190 6.3. Subsequent effects on FO performance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 190 6.3.1. Flux decline . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191 6.3.2. Enhanced rejections . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191 6.4. Fouling detection and cleaning methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 191 FO and other wastewater considerations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192 7.1. Energy aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192 7.2. Micropollutants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193 Concluding remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193 Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 193

Nomenclature A AFM AHA AL B BSA CA cdraw CECP CEOP CF cfeed COD CP CS CTA D DICP DS ECP EDX EfOM EPS FO FS FTIR HF ICP Js Jw Km l MBR

water permeability constant, m/s.Pa atomic force microscopy, e Aldrich Humic Acid, e active layer, e solute permeability coefficient, m/s bovine serum albumin, e cellulose acetate, e concentration of the draw solution, e concentrative external concentration polarisation, e cake enhanced osmotic pressure, e cross-flow, e concentration of the feed solution, e Chemical Oxygen Demand, e concentration polarisation, e corrugated spacer, e cellulose triacetate, e solute diffusion coefficient, m2/s dilutive internal concentration polarisation, e draw solution, e external concentration polarisation, e energy dispersed X-ray, e effluent organic matter, e extracellular polymeric substances, e forward osmosis, e feed side, e Fourier transform infrared spectroscopy, e hollow fibre, e internal concentration polarisation, e salt flux, g/m2 h water flux, l/m2 h mass transfer coefficient, m/s thickness of support layer, mm membrane bioreactor, e

MF NF NOM NPD OMBR PA PAI PAN PBI PDA PEG PES PNF PS PRO R Rg RO S SD SEM SL SWFO T TDS TEP TFC TrOC UF b Dp ε p s s

microfiltration, e nanofiltration, e natural organic matter, e normalised pressure drop osmotic membrane bioreactor, e polyamide, e polyamideeimide, e polyacrylonitrile, e polybenzimidazole, e polydopamine, e poly(ethylene glycol), e polyethersulphone, e plate and frame, e polysulphone, e pressure retarded osmosis, e Rejection of solute, % universal gas constant, L atm K1 mol1 or J/mol K reverse osmosis, e membrane structure parameter, mm Solution-diffusion, e scanning electron microscopy, e support layer, e spiral-wound forward osmosis, e Temperature, K total dissolved solids, e transparent exo-polymer particles thin-film composite, e trace organic contaminant, e ultrafiltration, e van’t Hoff coefficient, e osmotic pressure differential (pDS  pFS) bar porosity of the support layer e osmotic pressure bar reflection coefficient, e tortuosity of the support layer, e

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1.

Introduction

Drinking water is produced mainly from safe water sources, i.e. groundwater, but due to population growth and economic development, exploitation of aquifers and declining groundwater levels have diminished fresh water sources. The unsustainable use of drinking water for purposes other than sustenance, i.e. industrial processes, is therefore of great concern. A possible alternative source is wastewater. Via microfiltration (MF), ultrafiltration (UF), nanofiltration (NF) or reverse osmosis (RO), high quality water can be produced. Important examples of such applications are provided in Table 1. Forward Osmosis (FO), an alternative membrane process, also has the potential to treat wastewater, producing high quality water. FO is a technical term describing the natural phenomenon of osmosis: the transport of water molecules across a semi-permeable membrane. The osmotic pressure difference is the driving force of water transport, as opposed to pressure-driven membrane processes. Despite the availability of over 1000 FO publications since the 18th century (Alsvik and Ha¨gg, 2013; Qin et al., 2012), research in this field has grown significantly since 2005 (Fig. 1). The growing interest was sparked by the commercialisation of membranes specifically tailored to the FO process. The recent completion and operation of the first commercial FO desalination facility (200 m3/day) in Oman, makes FO technology more tangible (2012b). Of the FO literature produced in the last 10 years, approximately 7% employ complex waters. None-the-less, growth in wastewater treatment is steadily increasing. Ultimately, the motivation surrounding FO for the treatment of complex feeds is due to its potential advantages over current technologies: i. FO rejects particles, pathogens and emerging substances (Cartinella et al., 2006; Cath et al., 2005a; York et al., 1999), demonstrates high salt rejections and unlike normal treatment facilities, efficiently removes total dissolved solids (TDS) from complex solutions (Cath et al., 2006; Coday et al., 2014), due to its mean pore radius of 0.25e0.37 nm (Fang et al., 2014; Xie et al., 2012) ii. Due to the lack of high hydraulic pressures, neither the same energy input nor high strength materials are required (Klaysom et al., 2013; Thompson and Nicoll, 2011). In once-through FO systems where recovery is

iii.

iv.

v.

vi.

vii.

unnecessary, FO will be more energy-efficient than RO, enabling applicability in areas where access to electricity is limited (Coday et al., 2014). Extensive pre-treatment systems for FO may be redundant when treating complex feeds, but depends on FO performance and membrane design. In contrast, RO and NF are susceptible to fouling and require pretreatment to promote longevity and reduce costs (Kim et al., 2011; Sutzkover-Gutman and Hasson, 2010). FO has proven excellent operation in terms of durability, reliability and water quality in highly polluted waters, e.g. Hydropack Emergency Supply product (HTI). FO has shown flexibility and applicability due to: a) scalability of the membrane system (Schrier, 2012); b) reduced fouling propensity (Achilli et al., 2009; Lee et al., 2010) and simple cleaning (Hoover et al., 2011; Lutchmiah et al., 2011) compared to RO. FO can be applied for dewatering feeds (Petrotos and Lazarides, 2001; Zhu et al., 2012), useful for effective anaerobic digestion of wastewater, and is simpler, greener and higher in efficacy than traditional dewatering treatments (Chung et al., 2012b). Highly saline streams >83 bar are treatable by FO, not by RO (Hydranautics, 2014).

FO can be engineered and adapted to treat many complex feed types, including: complex industrial streams, i.e. from textile industries, oil and gas well fracturing (Anderson, 1977; Coday et al., 2014; Said, 2011; Votta et al., 1974; Zhang et al., 2014b), landfill leachate (York et al., 1999), nutrient-rich liquid streams (centrate) (Holloway et al., 2007); activated sludge (Cornelissen et al., 2011b, 2008); wastewater effluent from municipal sources (Lutchmiah et al., 2011; Valladares Linares et al., 2013a), simulated wastewaters (Beaudry and Herron, 1997; Cartinella et al., 2006; Cath et al., 2005b,a; Su et al., 2012) and even nuclear wastewaters have been mentioned (Zhao et al., 2012). Shortcomings of FO include the recovery step in closedloop systems, low water fluxes (Seppa¨la¨ and Lampinen, 2004) and reverse solute leakage (Hancock and Cath, 2009). The latter increases operational costs and decreases the driving force. The potential need for wastewater pretreatment is also disadvantageous, but is not exclusive to FO. Furthermore, the incomplete rejection of trace organic contaminants (TrOCs) is still an issue, but depends on the

Table 1 e Important examples of membrane treated wastewaters for reuse. Year

Feed type

1968

municipal wastewater

potable use

1976 2002

wastewater municipal wastewater effluent wastewater municipal wastewater wastewater

groundwater recharge aquifer recharge, potable use potable and industrial use industrial use groundwater recharge

2002 2007 2010

Application

Project/ Company Old Goreangab Water Reclamation Plant Water Factory 21 Torreele Wastewater Treatment Plant NEWater DOW Chemical The Groundwater Replenishment Trial

Location

Membrane treatment

Windhoek, Namibia

UF

Orange County, California Veurne Region, Belgium

RO UF-RO

Bedok and Kranji, Singapore Terneuzen, the Netherlands Perth, Australia

MF-RO UF-RO UF-RO

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Fig. 1 e FO publication growth since 2005 (ACS, ScienceDirect).

employed recovery system (D’Haese et al., 2013). Additionally, saline solutions on either side of the membrane may aggravate concentration polarisation (CP). This review, stimulated by the rapid growth and expansion in water reuse ventures, attempts to define the steps still required for FO to reach full-scale wastewater treatment and water reclamation.

2.

Elimelech, 2008). ICP, however, is still an issue for FO and the main driver for further membrane development. More specifically for wastewater treatment, the FO membrane fouling propensity needs to be addressed more vigilantly. In addition, the ideal FO membrane must allow fast transport of water towards the draw side, with ideally no migration of solutes between the draw and feed solutions, especially in closed-loop applications. Desired FO membrane characteristics for use in wastewater applications thus entail:

FO Membranes

Several recent reviews have focused on FO membrane developments (Alsvik and Ha¨gg, 2013; Klaysom et al., 2013; Wang et al., 2014). This review will only highlight findings important for wastewater applications. The first asymmetric cellulose acetate (CA) RO membranes developed in the 1960’s (Loeb, 1981) were initially intended for FO, however, due to inherent transport limitations were considered ineffective. Other RO membranes too, have not shown convincing results in FO due to hydrophobicity and relatively thick support layers (SLw150 mm) (McCutcheon and Elimelech, 2008). Thick SLs lead to poor performance of osmotically-driven membrane processes, which is mainly related to concentration polarisation (CP). Both internal CP (ICP) and external CP (ECP) exist. CP is caused by a balance between flux, rejection and diffusion, and lowers flux and membrane selectivity (intrinsic membrane selectivity remains unaltered). ICP is exclusive to FO and generally occurs within the porous SL of the membrane, while ECP is present at the surface of the dense active layer (AL). The breakthrough for FO came with the development of thin, FOtailored CTA membranes (w50 mm) by HTI, allowing higher fluxes through reduced ICP (McCutcheon et al., 2005; Mi and

 a dense, ultra-thin, active-separating layer for high solute rejection;  an open, thin (as possible), hydrophilic SL, with high mechanical stability, sustaining long-term operation and reducing ICP;  a high affinity for water (hydrophilicity) for enhanced flux and reduced fouling propensity. Significant advances have been made in these areas, and new, commercial FO membrane modules have hit the market, with more soon to follow (Coday et al., 2014; Klaysom et al., 2013). Dimensions of common spiral-wound FO (SWFO) elements can be found in Table 2.

2.1.

Membrane materials

Different materials are used for FO membranes (Klaysom et al., 2013). The widely-used cellulose triacetate (CTA) membrane (HTI) is highly resistant to chlorine (Lior et al., 2013) and is (mostly) unsusceptible to adsorption of mineral and fatty oils, including petroleum. CTA is less sensitive to thermal, chemical and biological degradation (Mulder, 1996),

w a t e r r e s e a r c h 5 8 ( 2 0 1 4 ) 1 7 9 e1 9 7

Table 2 e The dimensions of common SWFO elements and their effective membrane area. Module

Diameter

Length

Effective membrane areaa (m2) m2

inches mm inches mm 2521 4040 8040 a

2.5 4 8

63.5 101.6 203.2

21 40 40

533.4 1016 1016

0.38e0.53 1.5e3.2 7.5e17.6

Dependent on the number of membrane leaves.

and hydrolysis at alkaline conditions than cellulose, yet hydrolysis in harsh environments, i.e. wastewater, might still be an issue. New generation, commercial thin-film composite (TFC) membranes for FO are reportedly superior than CTA (Klaysom et al., 2013), in terms of permeability, and stability at broader pH ranges (2e12 vs. 3e8 for CTA), whilst still withstanding high pressures required in PRO (2012a). Susceptibility to membrane fouling is also material dependant. Generally, hydrophilic membranes, e.g. CA/CTA are less prone to absorptive fouling than hydrophobic membranes (Thorsen, 2004), making them more suitable for treatment of heavy duty wastewaters. The skin layer of TFCpolyamide (PA) membranes (HTI) was intentionally developed with low contact angles to increase biofouling resistance (2012c). In addition, hydrophobicity of the SL is also important in terms of flux; hydrophilic SLs lead to improved water fluxes and reduced ICP by increasing the wetting of small pores within the SL (McCutcheon and Elimelech, 2008). Alsvik and Ha¨gg (2013) outlined new methods and materials for FO membrane preparation, showing membrane materials to have moved beyond the traditional CTA and TFC-PA/ polysulphone membranes. Novel FO R&D membranes, specific for wastewater treatment, have shown much promise over recent years. Materials now used include hydrophilic CA, with increased fouling resistance (Zhang et al., 2010). New CA fabrication methods allow addition of pore-forming agents and annealing to improve fluxes and salt rejections (Sairam et al., 2011). Double-skinned CA membranes with dense outer skins have also evolved (Su et al., 2012; Wang et al., 2010a; Zhang et al., 2010), showing the potential for wastewater reclamation. Materials like strong and temperature stable thermoplastic polymers have also been used; polybenzimidazole (PBI) is able to self-charge in aqueous solutions, resulting in high salt rejections, high surface hydrophilicity and low fouling tendencies (Wang et al., 2009); polyamide-imides (PAI) provide both cationic and anionic repulsion to salt transfer through the membrane (Setiawan et al., 2011) and can produce NF-like selective layers (Qiu et al., 2012); nanoporous polyethersulphone (PES) with loose finger-like support structures and nano-sized pores allow less ICP (Yu et al., 2011); polysulphone, is mainly used as the SL in TFC membranes (Tiraferri et al., 2011; Wei et al., 2011; Yin Yip et al., 2010), despite its hydrophobicity (limiting fluxes); blends of polymers with polyacrylonitrile (PAN) for TFC-FO membranes show improvements over commercial FO membranes (Bui et al., 2011; Bui and McCutcheon, 2013; Qiu et al., 2011); and hydrophilic polydopamine (PDA) increases fouling resistance (Arena et al., 2011).

2.2.

183

Membrane development

Techniques used in developing membranes strongly influence membrane behaviour and filtration efficiency. Generally, conventional phase inversion is used to produce FO membranes, focusing on the formation of a dense selective layer on top of an asymmetric membrane (Klaysom et al., 2013). Various papers have elaborated on new design techniques and membrane formation mechanisms to optimise specific parameters. These include: 1) tailoring the membrane surface to decrease fouling and improve water fluxes by functionalising the membrane surface and/or embedding functionalised nanoparticles in the polymer (Tiraferri et al., 2012); 2) reengineering the support structure to withstand stress (Alsvik and Ha¨gg, 2013), 3) adding electrospun nanofibres to increase mechanical strength (Hoover et al., 2011) or 4) employing advanced co-extrusion techniques for mechanical stability and high power density (Zhang et al., 2014a). The many materials and fabrication methods tested show the potential for tailoring FO, regardless of the application, to augment membrane performance.

2.3.

Membrane orientation

Most FO membranes have an asymmetric structure with two different layers; an AL and an SL. The AL is generally the dense selective layer, while the porous SL provides the mechanical support. Due to this asymmetry, these FO membranes can be positioned either with the AL facing i) the feed side (AL-FS or FO-mode) or ii) the draw side (AL-DS, RO-mode or PRO-mode). The membrane orientation impacts FO performance (flux and fouling) significantly. Extensive research has reported higher water fluxes in PRO-mode, attributed to less severe concentrative ICP (CICP), but this orientation is also more prone to membrane fouling (Holloway et al., 2007; Tang et al., 2010; Wang et al., 2010b), due to the entrapment of the foulants in the SL, which reduce porosity and enhance ICP. Membrane orientation therefore influences ICP. Fabrication of the ideal FO membrane thus relies on eliminating the ICP problem, mainly caused by salt accumulation in non-ideal feeds (Ramon et al., 2011), by decreasing the SL thickness. If this could be achieved, then ECP would surely become a limiting factor especially in wastewater and seawater applications. Jin et al. (2012b) examined the effects of FO membrane orientation on organic fouling and the rejection of inorganic contaminants, observing more alginate fouling in PRO-mode. In FO-mode, arsenite rejection was enhanced by the alginate fouling, while in the PRO-mode, alginate fouling caused no observable effect on rejection, reportedly due to the counteractive effects of improved sieving and enhanced CICP. Zhao et al. (2011) observed a more dramatic water flux decline in PRO-mode during organic fouling, while the FO-mode also provided higher flux recoveries after cleaning. In many applications, e.g. osmotic membrane bioreactor (OMBR), the PRO-mode is impractical, due to the high fouling environment (Tang et al., 2010). To the best of our knowledge, commercial FO membrane modules are currently only manufactured in FO-mode.

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3.

Modelling membrane transport

In which D is the solute diffusion coefficient. S is given by the product of the SL thickness (l) and tortuosity (s), and is inversely proportional to the porosity (ε).

FO mass transport models, especially for use in wastewater applications, are limited. Of the existing models the majority neglect fouling, mainly because of the complexity and variability of wastewaters. Generally, FO models incorporate ICP and ECP phenomena (Loeb et al., 1997) and are based on solution-diffusion (SD) and convection-diffusion equations (Lee et al., 1981). Flux and rejection characterise membrane performance. Theoretical derivations of the water flux (Jv) have been established for the FO-mode (Loeb et al., 1997; McCutcheon and Elimelech, 2006):

FO membranes with large S values, i.e. thicker and denser membrane supports, result in decreased membrane perfor mance; mainly due to hindered diffusion and an increase in boundary layer thickness. A lower Km can significantly reduce FO permeate flux due to the exponential dependence of CP on Km (Xu et al., 2010). By substituting Eq. (2) into Eq. (1), S can be determined:

  ApDS þ B Jv ¼ Km ln ApFS þ Jv þ B



(1)

In which Km is the mass transfer coefficient, A the pure water permeability constant and B the solute permeability constant. From Eq. (1) it is clear that no linear relationship exists between the osmotic pressure difference (Dp) of the feed (pFS) and draw (pDS) and Jv. This is due to ICP. As mentioned, most FO models do not incorporate fouling, or do so based only on model foulants or synthetic waters, e.g. the salt accumulation model (Xiao et al., 2011) or the SD model for TrOCs (Jin et al., 2011). A better understanding of foulant behaviour in wastewater may lead to more representative models, fouling indications, accurate FO performance predictions and can further assist with membrane tailoring (Wang et al., 2014).

3.1.

Concentration polarisation

ICP is considered a major problem in FO, reducing the water flux and increasing (reverse) solute transport. For CICP (in PROmode), solutes from the feed penetrate the SL by means of convective water flux and direct diffusion (Tan and Ng, 2008b), while the dense AL is less permeable. This can be aggravated by salt leakage from the draw to the feed. In dilutive ICP (DICP) (FO-mode, preferred for wastewater), the flux becomes limited by dilution of the draw solution (DS) in the SL by water flux. Higher water fluxes result in more dilution, and as such, are a self-regulating mechanism consequently lowering fluxes. In FO-mode, this DICP is aggravated further by CECP, where the solutes in the feed are forced against the membrane (Gray et al., 2006; McCutcheon and Elimelech, 2006; Ng et al., 2006; Wang et al., 2007). In most flux models for FO, ECP is assumed to be negligible because of low fluxes and a high mass transfer, however ECP severely impacts feeds with high total dissolved solids (TDS). Simulations have shown ECP to be more prominent when mass-transfer promoting spacers are absent and low cross-flow velocities are used (Gruber et al., 2011).

3.2.

Membrane structure parameter

Recent efforts to improve FO membranes have focused their attention on the membrane structure parameter (S) (Park et al., 2011), which is inversely related to mass transfer through FO membranes:

Km ¼

D S

(2)



s$l ε

(3)

  D Apd;b þ B ln Jv Apa;w þ Jv þ B

(4)

Essentially, thinner, more porous and less tortuous SLs (Song et al., 2011) will have smaller S values and produce higher water fluxes (Yin Yip et al., 2010). S values depend only on membrane structural properties (Table 2). A priori estimation of the S value is however hard to attain due to s, which is intrinsically associated with the microstructure of the SL (Li et al., 2011b). Casting conditions during membrane preparation can optimise the support structure (Mulder, 1996) and decrease S, thereby improving FO membrane performance. Immersion precipitation processes to synthesize the substructure will yield different structures depending on the casting solution (i.e. composition and concentration) and casting conditions (i.e. temperature and humidity). For example, adding a solvent to the coagulation bath is known to delay solvent/non-solvent mixing which leads to thinner and more compact structures. The morphology of the SL (finger-like, sponge-like or scaffoldlike structures) also influences S (Song et al., 2011; Tiraferri et al., 2011; Wei et al., 2011; Yin Yip et al., 2010). By changing annealing conditions and adding fillers during membrane preparation, tortuosity can be influenced to obtain more finger-like pores, which produce low S values.

3.3.

Membrane permeability constants

The constants A and B are necessary parameters for mathematical models of FO. Water passage through dense membranes is generally described and theoretically calculated by Eq. (5) (Cath et al., 2006): Jv ¼ A$s$Dp

(5)

Where s, the reflection coefficient, is 1 when the membrane completely rejects the solute. This is usually assumed in FO studies, due to high rejections (RNaCl: 93e95%) (McCutcheon and Elimelech, 2007; Tan and Ng, 2008b) and is therefore commonly neglected in modelling studies. The actual s depends on the DS used; low s result in low effective driving forces by reverse solute leakage (Su and Chung, 2011; Yong et al., 2012). B relates to solute transport and solute leakage and is dependent on D, its partition coefficient and effective membrane thickness. Despite high rejections, forward and reverse solute diffusion are significant in FO. Forward diffusion occurs when solutes move from the feed (wastewater) into the DS, while reverse diffusion (solute leakage) occurs from the DS into the

w a t e r r e s e a r c h 5 8 ( 2 0 1 4 ) 1 7 9 e1 9 7

feed (Hancock and Cath, 2009). B is solute-dependent and should be minimised to avoid solute leakage. Solute leakage decreases Dp because of (i) a decrease in pDS and (ii) an increase in pFS. Solute leakage impacts DS costs and may also affect the biology in OMBR and FO-wastewater-energy generating systems. A and B are, in practice, often determined in pressurised cross-flow filtration tests. These tests are well adapted to RO, but poorly reflect FO operation. It is questionable whether these tests yield reliable results for FO. A more standardised methodology to test A and B in FO conditions has been recently proposed and tested (Cath et al., 2013). Commercial CTA membranes typically have water permeabilities of 2.54  1.69  1012 m/s.Pa (Table 3), but low values between 0.87 and 1.07  1012 m/s.Pa and high values of 7.1  1012 m/s.Pa were also reported. Additionally, Table 3 shows average B values of 1.27  0.48  107 m/s for NaCl. The large standard deviations for A and B can be attributed to temperature differences (Cornelissen et al., 2008), deviations between small membrane coupons (Phillip et al., 2010) and/or membrane deformation under hydraulic pressure (Blandin et al., 2013; Kim and Elimelech, 2013). In addition, it is possible that the SD model is not ideal for calculating these parameters or that the translation from RO to FO conditions is flawed. In general, the listed membrane materials in Table 3 show NaCl rejections >90%. TFC membranes are typically more water permeable and show large variations in B values (0.02e47.20x10-7 m/s). The future of FO membranes therefore seems to move in the direction of TFC material, as tailoring appears easier and the design versatile.

185

reverse leakages (Ge et al., 2013a). To evaluate performance, the Js/Jv ratio is often regarded (Phillip et al., 2010): Js B ¼ Jv AbRg T

(6)

The DS and related osmotic pressures in the FO process are important factors influencing mass transport and overall process performance. Several DS types (and concentrations) have already been tested in FO (Fig. 2). From Fig. 2, NaCl appears to be the most employed DS (approximately 40% of experiments), due to its high solubility (Cath et al., 2006), but also low cost (Achilli et al., 2010) and relatively high osmotic potential. NaCl has been used as a DS in concentrations between 0.3 and 6 M, but is often used at 0.5 M simulating the osmotic power of seawater and prompting the use of real seawater or RO brine as a DS (Chekli et al., 2012). Seawater is abundant and has frequently been used as a DS (Cath et al., 2010), but contains numerous particles and microorganisms which may foul the reconcentration system (in closed-loop systems) or cause (bio)fouling/ contamination in the FO unit, hindering performance. This increase in fouling potential may be a limitation when using wastewater feeds. In a once-through system for RO desalination though, an FO pre-treatment step can minimise fouling in RO (Cath et al., 2006).

Where b is the van’t Hoff coefficient and Rg the gas constant. Eq. (6) indicates the dependence of the ratio on the membrane transport properties of the AL, but not the SL. Membranes with better selectivities, i.e. higher A and lower B values are thus essential (Klaysom et al., 2013). Additionally, the DS should be non-toxic, easily recoverable in the reconcentration system, and when operated in a bioreactor, should not deteriorate the OMBR, sludge quality or bacterial growth (Bowden et al., 2012). Low to no degradability of the substance should also be considered, unless degradation of the DS (post-FO) is advantageous (Lutchmiah et al., 2014) or irrelevant. The DS price is equally important, unless loss through the FO membrane is nominal and recovery is achieved with minimal energy and costs. Mass transport properties, i.e. diffusivity (D) are also significant. Large molecules diffuse more slowly and therefore leak less through the membrane. Lobo and Quaresma (1989) found that D changed with concentration and salt type. According to the StokeseEinstein relation, D is further influenced by temperature, viscosity of the fluid and particle size. In dilute aqueous solutions, D values of most ions are similar (0.4  109 to 2  109 m2/s, at 25  C), however for the sulphates, MgSO4 and ZnSO4, D is considerably lower, which may explain the lower water fluxes achieved during FO investigations. Wastewater contains many molecules with varying diffusivities, which together with the DS can influence the extent and severity of DICP. The lower the diffusivity, the higher the degree of DICP in the porous SL (Achilli et al., 2010; Hancock and Cath, 2009; Tan and Ng, 2013; Zhao and Zou, 2011). Monovalent ions display higher water fluxes as a result of a higher diffusivity (Cornelissen et al., 2008), however this is accompanied by higher solute leakages, which are ascribed to lower size exclusion and lower electrostatic repulsion of monovalent ions. Higher diffusivity and solute fluxes might simultaneously lower ICP, but the exact effect depends on the operating conditions. Bivalent ions display lower solute leakages as a result of higher steric hindrance and electrostatic repulsions, but also lower water fluxes. PH and temperature should also be considered when choosing a DS. A change in one or both can lead to mineral salt scaling on the membrane surface when scale precursor ions like Ba2þ, Ca2þ, Mg2þ, SO4 2 , and CO3 2 are present (Achilli et al., 2010). Due to the complex ionic matrix of wastewater feeds, Achilli et al. (2010) suggested MgCl2 to be the most ideal DS for wastewater processes in view of its high osmotic efficiency, though material costs are higher than for NaCl. Table 4 summarises the advantages and drawbacks of some DS types considered for wastewater applications.

4.1.

4.2.

4.

Draw solutions

Effective draw solutions for wastewater treatment

A higher osmotic potential than the feed is essential to induce a water flux, but the focus in FO-wastewater treatment for a DS, is on good performance, with high water fluxes and low

Recovery systems

A thermodynamically favourable reconcentration system, in closed-loops, is required to replenish the DS and separate it from the product water. Some methods involve RO,

186

Table 3 e A summary of membrane types used in FO applications and comparisons between their membrane characteristics. Membrane type

Manufacturer

Temperature

Material

Toray

flat-sheet SWC1 SW30-HR (with PET) SW30-HR (No PET) e e e e e e HF HF HF e flat-sheet

Dow

TFC (RO)

Yale NTU NTU NTU NTU NTU NTU NUS NUS NUS Yale

TFC (FO) TFC-1 TFC-2 TFC-3 TFC-4 TFC-5 TFC (#C-FO) TFC (#A-FO) TFC (#B-FO) TFC (PES/SPSf) TFC (PA/PS nanofibre) #1 nanofibre e composite NC-FO #2 NC-FO e

NTU NTU

Structural Mass transfer Rejection parameter coefficient (S) (Km)

[*107 m/s]

e

3.75

2.13

NaCl

e

e

98

(Loeb et al., 1997)

50  1 20  1, 30  1 22  0.5 25 22.5  1.5 23  1 30 20  2 25  0.5 20  1 20  0.5 31  1 22 - 24 23 25  1 25  0.5 20  0.5 30 20  2 22  1 20  1

5.69 3.07 2.17 2.71 3.08 0.87 3.95 7.10 0.98 3.60 1.23 2.21 2.20 2.12 1.89 1.07 e 2.82 20.0 11.91 3.80

e e 1.73 e 1.27 e e e e e 0.73 e 1.70 1.60 1.68 e 0.63 e e e e

(NH4)2CO3 NaCl NaCl (NH4)2CO3 NaCl NaCl NaCl NaCl NaCl NaCl NaCl NaCl/MgSO4 NaCl NaCl NaCl NaCl NaCl NaCl MgSO4 e NaCl

e e e e e e e e 0.60 e 0.48 e 0.40 e e 0.62 e e e e e

e e e e e e e e e e e e AL-FS: 4.2 AL-FS: 5 e e AL-FS: 17.2 e e e e

>95 95e99 e >95 e e e >92 94.1  1.1 e 89.1e96.1 e 90 e >90 92  1.5 e >99 100 99.5 e

(McCutcheon et al., 2006) (Garcia-Castello et al., 2009) (Wang et al., 2010a) (McCutcheon et al., 2005) (Gray et al., 2006) (Holloway et al., 2007) (Tan and Ng 2008a,b) (Cornelissen et al., 2008) (Yin Yip et al., 2010) (Mi and Elimelech, 2010) (Phillip et al., 2010) (Qin et al., 2010) (Tang et al., 2010) (Xiao et al., 2011) (Zhao and Zou, 2011) (Song et al., 2011) (Yong et al., 2012) (Parida and Ng, 2013) (Cornelissen et al., 2008) (Dova et al., 2007) (Mi and Elimelech, 2010)

20  2 25  0.5

5.60 3.55

e e

e NaCl

e 9.58

e e

100 98.9  0.4

(Cornelissen et al., 2008) (Yin Yip et al., 2010)

25  0.5

4.39

e

NaCl

2.16

e

98.3  0.4

(Yin Yip et al., 2010)

25  0.5 23 23 23 23 23 23 23 23 22  0.5 25  0.3

3.18 5.03 5.05 5.28 9.58 3.5 9.60 2.60 6.20 2.14 3.14

e 0.02 0.25 2.60 47.20 0.17 0.61 0.81 0.56 0.31 0.64

NaCl NaCl NaCl NaCl NaCl NaCl NaCl NaCl NaCl NaCl NaCl

0.49 e e e e e 0.55 1.37 0.60 0.24 0.65

e AL-FS: AL-FS: AL-FS: AL-FS: AL-FS: e 1.1 2.5 18.6 e

97.4  0.5 e e e e e e e e 93.5 >96

(Yin Yip et al., 2010) (Xiao et al., 2011) (Xiao et al., 2011) (Xiao et al., 2011) (Xiao et al., 2011) (Xiao et al., 2011) (Chou et al., 2010) (Wang et al., 2010b) (Wang et al., 2010b) (Wang et al., 2012) (Hoover et al., 2013)

25  0.5

4.66

e

NaCl

0.08

e

97  1.0

(Song et al., 2011)

25  0.5

4.52

e

NaCl

0.11

e

97  1.2

(Song et al., 2011)

[e]

[mm]

[*106 m/s]

Reference

[*1012 m/s.Pa]

[ C]

Type with and without support fabric flat-sheet e Double skinned flat-sheet flat-sheet e e flat-sheet e e e flat-sheet Hydrowell Filter Hydrowell Filter e e e Flat-sheet Flat-sheet TS80 DS-11-AG SW30XLE-400i

Draw solute

2.5 2.5 2.5 2.5 2.5

[%]

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CA-3000 (RO) (modified) HTI CA HTI CA NTU CA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA HTI CTA TriSep TFC (NF) General Electric TFC (RO) Dow TFC (RO) (modified) Hydranautics TFC (RO) Dow TFC (RO)

Water Solute permeability (A) permeability (B)

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Fig. 2 e Draw solutions used in FO based on approximately 50% of FO publications. Results are expected to increase in similar ratios when considering all published works.

membrane distillation (MD) and thermal recovery. The chosen recovery system depends on the type of application and solute, the recovery rate required and the energy consumption of the unit. For wastewater treatment, the process in choosing a recovery unit is similar to other FO applications, however

forward solute diffusion needs to be considered. Solutes e.g. sparingly soluble salts and boron, from complex feeds can pass through the membrane to the draw side due to bidirectional diffusion (Hancock and Cath, 2009). This may cause problems for the reconcentration unit, i.e. ions with low

Table 4 e An overview of draw solutions considered in wastewater applications. Draw solution type

Example

Advantages

Inorganic substances

Salts

 High solubility  Low cost  High osmotic pressure potential

Highly soluble zwitterionic substances

Glycine

High charged compounds

EDTA

 High flux, low leakage  Leakage beneficial to biology in subsequent energy-generating units  High water flux  Low reverse leakage  Reconcentration via less energy consuming processes, i.e. NF

Nutrient-rich substances

Fertilisers

 Direct fertigation  No recovery necessary

Readily available sources

Seawater, RO brine

 Abundant source

Thermolytic solutes

Ammonium bicarbonate

 High solubility in water  Recovery by moderate heat

Engineered draw solutions

Magnetic nanoparticles

High osmotic pressuVres at low concentrations  No leakage  Overcomes scaling and  crystallisation issues in MD

Disadvantages

Reference

 Salt leakage may inhibit anaerobic digestion  Clogging/scaling/fouling  ICP  Recovery is not often feasible  Limited storage time due to biodegradation

(Chung et al., 2012a)

 More expensive than common salts  pH dependency  Questionable environmental repercussions  Osmotic equilibrium limits  Dilution of nutrients  TEP fouling,  Seawater: only cost-efficient if applied near coastal areas  Toxic thermolytic product  High diffusive loss  Agglomeration during magnetic separation  Ultrasonication weakens magnetic properties  Viscosity of solution reduces effective driving force and the flux

(Cornelissen et al., 2013, Hau et al., 2014, Yuchun et al., 2012)

(Lutchmiah et al., 2014)

(Phuntsho et al., 2011, Phuntsho et al., 2012) (Cath et al., 2005b, Cath et al., 2010, Logan and Elimelech 2012)

(McCutcheon et al., 2005, Qin et al., 2013) (Ge et al., 2011, Ling et al., 2010)

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solubilities can precipitate and have deleterious effects on the system’s performance. The addition of scale inhibitors to the draw side or filtration of the solution before the reconcentration unit is possible, but will increase operating costs. Furthermore, if solutes, e.g. TrOCs move into the DS, they must be removed in the reconcentration loop. The energy consumption of all adjacent treatment processes must be considered. Using feeds with high organic content could consequently be advantageous in these situations as energy can be extracted and fed to the reconcentration (or other) unit. In doing so, the energy consumption of recovery systems could be reduced. Due to minimised membrane fouling and subsequent reduced costs, the commercialisation of the OMBR is expected in the near future (Zhao et al., 2012). This however also depends on finding a cost-effective DS, e.g. seawater. The use of hybrid systems for wastewater treatment integrated with seawater desalination is interesting to enhance FO and make wastewater recovery commercially feasible. However, as much as these systems are technically feasible, they remain economically and industrially unpractical due to the high energy costs of the recovery unit (Chung et al., 2012a). Furthermore energy balances for many of these integrated systems are still lacking.

5.

Process conditions

5.1.

Temperature and pH

Increasing the temperature in FO processes will increase the water flux, due to the increase in pDS and the decrease in wastewater viscosity (McCutcheon and Elimelech, 2006). A simultaneous increase in solute permeability also occurs, resulting from an increase in the solute diffusion rate through the membrane (Ghiu et al., 2002) and lower ICP. These findings were corroborated for hollow fibre NF membranes (Wang et al., 2009) and CTA-FO membranes (Cornelissen et al., 2008; Ng et al., 2006), additionally showing a 2.4% rise in water flux per  C increment with 0.5 M NaCl (Cornelissen et al., 2008). Li et al. (2011a) and Razmjou et al. (2013) have shown how solar energy can be used to influence fluxes in FO systems. The temperature effect on ICP is also less pronounced in PRO-mode, resulting in higher water fluxes (Gray et al., 2006; McCutcheon and Elimelech, 2006), but also higher salt fluxes. Furthermore, in FO-wastewater applications an increase in temperature will concentrate the wastewater faster, possibly leading to greater fouling of the membrane. Wastewater effluents contain delicate ecosystems that can be adversely affected with increasing temperatures and increasing salt concentrations, however when energy is generated from the organic content (COD) of wastewater (anaerobic digestion), temperatures as high as 35  C may be required. A change in pH might affect certain DSs as well e.g. EDTA or zwitterions. Furthermore, compounds like heavy metals or suspended solids precipitate at high pH ranges, or dissolve (usually) at low pHs. Additionally, energy generation from wastewater requires an optimal pH between 6.5 and 8, although some bacteria can survive extreme pHs.

5.2.

Cross-flow velocity and spacer designs

In FO, lower cross-flow feed velocities may result in higher CECP effects (Loeb et al., 1997; McCutcheon et al., 2006). Hancock and Cath (2009) found ECP to be affected by both DS and FS flows, resulting in maximum water fluxes at higher and equal flow velocities on both sides of the membrane, however other studies observed no change in the water flux under different flow regimes (Qin et al., 2010; Tan and Ng, 2008a). Hancock and Cath (2009) also suggested operating FO processes at low feed and draw cross-flows to minimise reverse solute leakage, however this may reduce water fluxes by increasing ECP and membrane fouling. Spacers effect cross-flow conditions too, influencing mass transfer and encourage mixing, which suppresses the membrane boundary layer and CP. Park and Kim (2013) studied six FO spacer configurations based on geometric characteristics. Submerged-type spacers were found to perform the best and cavity-type spacers the worst, in terms of permeate flux. In spiral-wound configurations feed spacers provide a feed channel between membrane envelopes for the feed water to flow evenly through the membrane element. Channel profiles are dependent on the spacer type, but normally run in an approximate 45 angle to the feed. To allow fouling mitigation in SWFOs, corrugated feed spacers (CS), chevron design (thickness 98 mil), have been developed, although pumping energy is high for these modules to maintain high cross-flow velocities. Thinner CS spacers are being further developed, which should cut the energy requirements by 25e50%, but may increase the effect of fouling in spiral-wound modules as well.

5.3.

Module configuration

Membrane module configurations imply the packing of a membrane into a module to (i) maximise the surface to volume area and (ii) reduce particle deposition by sufficient cross-flow (Degremont, 2011). Selecting a module depends on economic considerations, CP, fouling propensity, ease of fabrication into the module design and suitability of the design for FO-wastewater applications. Configurations interesting for FO-wastewater can be found in Table 5. Spiral-wound FO (SWFO) membranes have developed considerably from a design impractical for the membrane industry (Cath et al., 2006), to becoming a configuration commonly used in FO tests. SWFO elements are preferred if both feed and draw streams are clean. If fouling occurs in SWFOs, cleaning can occur via chemicals, air/water mixing and/or high fluid velocities. Furthermore, improved SWFO membranes should be backwashable to allow more effective fouling removal from the membrane surface and feed channels. For this, Bamaga et al. (2011) suggested short SWFO feed channel lengths and wider spacers. SWFO elements have a lower packing density (typically 20 m2 for 8-inch modules) than SWRO elements (typically 30 m2 for 8-inch). Additionally, SWFO membrane envelopes contain glue lines to create two independent channels inside and outside the envelope, accounting for membrane area loss (Xu et al., 2010). Improved SWFO modules could employ heat welding technology or mechanical methods to seal the membrane edges (Bamaga et al., 2011). Furthermore, the

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Table 5 e Interesting FO module configurations for wastewater treatment. Configuration

Advantages

Disadvantages

Reference

Spiral-wound (SW)

 High packing density  Easy cleaning of fouling deposits

(Bamaga et al., 2011, Cath et al., 2006)

Hollow fibre (HF)

   

Plate and frame (PNF)

 Well-suited to wastewater applications  Less complicated in design Better backwashing  Higher cross-flow velocities

 Limited pressures in membrane envelop  Clogging of spacers  Limited mixing at membrane surface  In open channels the CP film grows undisturbed in the channel  More expensive per m2 of membrane area

Self-supported characteristics Appropriate flow patterns Simplicity of fabrication High packing density

maximum pressure currently allowed in the envelop is 70 kPa (Bamaga et al., 2011); higher pressure differences at the DS side, without damaging the glue lines, are desired. Pilot-scale experience with SWFO modules is still limited, but steadily growing. Hancock et al. (2011) used a SWFO in their pilot-scale study to determine the rejection of TrOCs, showing higher rejections than on bench-scale, while Cornelissen et al. (2011a) found slightly higher fluxes in a 4inch SWFO module (constant DS concentration) compared to lab-scale. Xu et al. (2010) suggested that permeate flow in spiral-wound modules may significantly dilute the bulk DS concentration (if not kept constant), reducing flux performance compared to small coupon tests. Factors like membrane area, feed volume and concentration should be considered in such tests, illustrating the importance of developing scalability studies accurately from lab-scale to full-scale.

6.

Membrane fouling

Membrane fouling concerns the accumulation of solutes and/ or particles on a membrane surface, within membrane pores or feed spacer clogging. This can lead to fouling, scaling or damage of the membrane (Metcalf and Eddy, 2003). FO membranes are generally at a lower risk of membrane fouling due to the lack of hydraulic pressure. Ultimately fouling leads to an additional hydraulic resistance lowering the effective osmotic pressure and water permeability.

6.1.

Wastewater foulants

Major foulants in natural and impaired waters are microorganisms, organic matter and inorganic matter (scaling). Biofouling can be the most limiting factor when employing wastewater, due to the presence of microorganisms and their secretion of extracellular polymeric substances (EPS) to establish biofilm integrity (Flemming et al., 2000; Staudt et al., 2004). Biofouling is influenced by the feed water quality, membrane physico-chemical properties and operating conditions. In an FO-MBR study (Zhang et al., 2012), biofouling did not affect the water permeability much, however the mass transfer coefficient was reduced severely, enhancing ICP. In seawater-FO, silica scaling (Li et al., 2012) or membrane

(Liu et al., 2013, Park and Kim 2013, Setiawan et al., 2011b, 2012, Wang et al., 2007) (Thorsen 2004)

(Bamaga et al., 2011, Gu et al., 2011)

biofouling via transparent exo-polymer particles (TEP) (BarZeev et al., 2009; Valladares Linares et al., 2012) can occur. Organic fouling differs according to the feed water employed. Wastewaters are comprised of effluent organic matter (EfOM), which includes both soluble microbial products and natural organic matter (NOM). NOM has been found to be a serious fouling agent in many membrane processes including FO (Lee et al., 2010; Valladares Linares et al., 2012), therefore mimicking the behaviour of these complex feeds to include all, or the most significant, foulants is very important. Model foulants, i.e. sodium alginate or alginic acid, bovine serum albumin (BSA), Aldrich Humic Acid (AHA) have been used to test the severity of NOM fouling on FO membranes (Liu and Mi, 2012; Mi and Elimelech, 2008; Qi et al., 2012; Tang et al., 2010). Alginate relates to the hydrophilic fraction of EfOM, AHA represents humic acids and BSA the protein portion. Mi and Elimelech (2008) found calcium ions to enhance alginate (humics) fouling, severely declining the flux, compared to alginate fouling in the absence of calcium ions. This humicsecalcium deposition on the membrane surface has been observed using RO brine (Parida and Ng, 2013), showing that calcium-binding can enhance the intermolecular adhesion between foulants, increasing membrane fouling. Bypassing these interactions completely may be difficult, but adding calcium, i.e. as a DS, should be avoided. Zhao et al. (2011) found organic fouling to be more severe and irreversible than inorganic fouling in FO, but scaling may be a greater issue for wastewater, especially at high recoveries. Precipitation can occur with sparingly soluble salts such as calcium carbonate or calcium sulphate (gypsum). Gypsum scaling cannot be controlled by simple pH adjustments, and is also affected by the membrane material, i.e. heterogeneous/surface crystallisation on PA membranes causes severe flux decline (Mi and Elimelech, 2010). In practice, salt precipitation in the form of BaSO4, Mg(OH)2 or silica contribute more to membrane scaling than gypsum.

6.2.

The causes of FO fouling

Many studies have attributed the low fouling propensity in FO to the low flux conditions and lack of applied pressure, however FO fouling works similar to the fouling mechanisms in RO because both processes are governed by chemical and hydrodynamic interactions (Mi and Elimelech, 2008). The

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occurrence of ICP makes FO fouling even more complicated (Tang et al., 2010). An accumulation of solutes (foulants) in the SL is generally caused by reduced porosity and a reduced mass transfer coefficient, enhancing ICP. ICP adversely affects FO flux and fouling at greater DS concentrations and/or greater membrane fluxes, due to the exponential dependence of ICP on the flux level (Tang et al., 2010). More frequent cleaning, i.e. flushing or backwashing may reduce the foulant accumulation in the SL and so alleviate ICP. This may be detrimental to the membrane longevity over time, but could avoid the decrease of flux performance during operation and perhaps prevent the subsequent use of chemicals. Operational parameters like cross-flow velocity can affect the rate and extent of membrane fouling (T.Y. Cath, 2009; Choi et al., 2005). Increasing the cross-flow velocity, prior to cakeforming compaction, can reduce NOM accumulation and hamper growth of the fouling layer in FO membranes. However once the cake layer forms, changes in hydrodynamic conditions barely affect the fouling behaviour (Mi and Elimelech, 2008). Module design also plays a role in fouling. Lab-scale studies may not employ spacers (Hoover et al., 2013), but when applied to full-scale, FO membranes require spacers in both the feed and draw solution channels (Kim and Elimelech, 2012). The selected spacer type, especially in wastewater applications, should promote mixing, decreasing ECP and consequently, the passage of small solutes through the membrane. Additionally, application-specific feed spacer materials, can be fitted to further prevent colloidal and particulate fouling (2012a). The use of a diamond-patterned feed spacer on the feed side dramatically enhances initial FO fluxes (w50% at 0.5 M NaCl and w100% at 2.0 M NaCl) and improves flux stability during fouling (Wang et al., 2010b). However, particle accumulation and biofouling is still found near spacer filaments, due to local hydrodynamic conditions (low shear region) and physical blockage (large particle size), suggesting room for improvement in FO spacer design. The membrane orientation has also been suggested to affect fouling. The FO-mode offers flux stability against both dilution of the bulk DS and membrane fouling, while, the PROmode has shown severe flux reduction after fouling, due to the microstructure difference of the SL, where internal clogging or enhanced ICP can occur (Tang et al., 2010). Other factors influencing the development of a fouling layer on the membrane surface include membrane surface morphology (foulant-membrane interactions), i.e. charge (Boo et al., 2012; Xu et al., 2006) and roughness (Li et al., 2007); membrane surface hydrophobicity; electrostatic attraction (leading to calcium binding); elevated ionic strengths which can lead to an increase in EPS thickness (Orgad et al., 2011); permeation drag, and hydrodynamic shear force. Ultimately, membrane fouling is dependent on the foulant. Therefore it is important to consider the combined interactions of feed water foulants, while still regarding the overall effects of ECP and ICP.

6.3.

Subsequent effects on FO performance

Foulant accumulation adversely affects the quantity (permeate flux) and quality (solute concentration) of the

product water. This can influence membrane performance, i.e. reduce water productivity and/or “permeate” quality; change the normalised pressure drop (NPD); increase energy expenditure and treatment costs, and if severe, cause membrane failure (Goosen et al., 2005).

6.3.1.

Flux decline

Flux decline in FO processes is generally due to a reduction in the driving force, via DS dilution (Cath et al., 2005b), membrane fouling (Holloway et al., 2007) and enhanced CP in the fouling layer (Tan and Ng, 2008b). Reverse salt diffusion too reduces the driving force, but also intensifies cake enhanced osmotic pressure (CEOP) within the fouling layer, elevating osmotic pressures near the membrane surface (feed side) and subsequently reducing Dp and permeate flux (Lee et al., 2010). When affected by fouling, the extent of the flux decline can vary depending on the fouling stage, installation scale and feed type. Membrane fouling/scaling can cause water flux declines of 20% when treating synthetic feeds (COD: 4.5  0.2 g/L) in lab-scale OMBR systems (Achilli et al., 2009) or between 30 and 50% when treating landfill leachate in a fullscale FO system (York et al., 1999).

6.3.2.

Enhanced rejections

Using wastewater effluent, Hancock et al. (2011) found >99% TrOC rejection due to FO membrane fouling, because of interactions between hydrophobic hormones and organic matter. Valladares Linares et al. (2011) found the increase in micropollutant rejection in the presence of the fouling layer due to higher hydrophilicities in fouled FO membranes than clean ones, reducing mass transport capacity, membrane swelling, and the higher negative charge of the membrane surface (related to NOM acids and polysaccharides).

6.4.

Fouling detection and cleaning methods

Immediate fouling detection can ensure the longevity of a membrane and restore membrane performance. Determining the fouling potential of the feed can help predict fouling, however once fouling occurs on the membrane surface, exsitu and off-line methods may be necessary to incorporate future preventative measures. Non-invasive and visual online methods can detect early signs of fouling in real-time, i.e. flux decline, solute rejection and changes in NPD and operating parameters (temperature, feed TDS, permeate flow, recovery). Fig. 3 summarises the fouling detection techniques involved with feeds and FO membrane fouling. Only long-term fouling studies can prove whether FO membranes are resilient to fouling. Short-term vs. long-term studies have varied significantly with regards to fouling severity when treating impaired waters (Cornelissen et al., 2011b; Valladares Linares et al., 2012). However once fouling has been detected, cleaning via physical, chemical or physicochemical methods can be introduced to restore system performance. Depending on the fouling severity, physical cleaning via hydrodynamic modifications (Lee et al., 2010), forward or backward flushing and osmotic backwashing (Coday et al., 2014) might be enough to remove fouling deposits. Higher fouling propensities and fluxes have already been mentioned

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191

Fig. 3 e Methods used in literature to detect the fouling potential of the feed or monitor/analyse fouling on the FO membrane using i) on-line techniques to detect early signs of fouling via changes in operational parameters (flux or rejection) or other non-invasive tools, ii) off-line techniques to determine morphological, chemical and physical properties of the foulant type influencing FO membrane fouling.

for the PRO-mode. This orientation should therefore make flushing and rinsing easier and faster, however using inorganic and organic foulants, Zhao et al. (2011) found the FOmode to provide higher flux recoveries after cleaning than the PRO-mode. When severe, fouling mitigation may require chemical reagents. Reagent selection depends on the feed water, foulant type (Zhao et al., 2011) and membrane material (Lay et al., 2012). Furthermore, these reagents should be able to dissolve and remove most of the deposited materials, without damaging the membrane surface. Precipitation and scaling can be mitigated with anti-scalants and commercial inhibitors, while periodic cleaning with one or more chemicals (acids, bases, oxidising agents, surfactants or chelating agents) might be necessary to maintain long-term process performance. Chemical cleaning typically removes the cake layer, but not foulants in the membrane pores (Holloway et al., 2007). Furthermore, it not only impairs membrane selectivity and shortens membrane life, but also consumes additional energy and produces concentrated waste streams (Ang et al., 2006). Valladares Linares et al. (2013b) assessed various chemical and physical procedures to remove FO membrane fouling, during wastewater-seawater integration. Air scouring proved most effective (w90% flux recovery), with the addition of industrial detergents further increasing recovery; osmotic backwashing proved completely ineffective. Qin et al. (2010) confirmed the potential advantages of fouling reversibility with air scouring employing a pilot OMBR, while a further study by Valladares Linares et al. (2013a) suggested the need

for multiple backwashings to achieve flux recovery, due to ICP aggravation by (seawater) salts. Periodic physico-chemical cleaning or pre-treatment are other options. However, regardless of the cleaning method, irreversible fouling may still occur when treating wastewater, due to the adsorption of biopolymers on the membrane (Valladares Linares et al., 2013b; Valladares Linares et al., 2012).

7.

FO and other wastewater considerations

7.1.

Energy aspects

Employing effective, inexpensive, low energy pre-treatment practices for high TDS solutions may be essential in reducing the potential for membrane fouling, ensuring acceptable performance and protecting the FO membrane (Hancock et al., 2012). Multiple methods are available for the removal of particles, but these do not prevent organic fouling. Furthermore, single-pass membrane processes produce a concentrate stream. In wastewater, the method of disposing this concentrated brine must be deliberated well, since some streams may be difficult to discard and can become a costly element, i.e. leachate management (York et al., 1999). One advantage of applying FO (similar to RO), is that the waste stream becomes concentrated, enabling a smaller flow of waste to be processed. Additionally, the concentration of organic wastewater streams could facilitate renewable energy

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production and nutrient recovery; a goal of the Sewer Mining Concept (Lutchmiah et al., 2011). This could reduce the energy required by the system, e.g. energy produced from the organics can drive the reconcentration unit. Wastewater organics (COD) are directly correlated to the energy value: 1 kg COD ¼ 3.86 kWhtheory (Shizas and Bagley, 2004). In this way energy can be recovered from wastewater. The concept of using FO for energy generation is not new (Devoe, 2005; Lampi et al., 2007) and recent environmental awareness has emphasized the need for sustainable, green technologies. The osmotic Microbial Fuel Cell (OsMFC) (Ge and He, 2012; Ge et al., 2013b; Zhang et al., 2011) is similar to the microbial desalination cell (Cao et al., 2009) desalinating synthetic brackish- and seawater. The OsMFC uses an FO membrane as a barrier between cathode and anode units, extracting water (from wastewater) and generating electricity. These designs, however, require aeration of the cathode chamber, counteracting the benefits of the power generation. Air-cathode OsMFCs have evolved to improve energy recovery (Werner et al., 2013), but reverse salt leakage and fouling remain challenges. Furthermore, energy can be achieved from salinity gradients, where the osmotic pairing of fresh water and seawater is possible. In salinity-driven desalination, pre-dilution can be accomplished by pressure retarded osmosis (PRO) (Post et al., 2007). PRO is a process through which osmotic energy can be harnessed and power generated (Loeb, 1975). Wastewater tested in PRO applications (Chou et al., 2012; Saito et al., 2012; Wang et al., 2012) demonstrated power densities from 4 to 10 W/m2. PRO is growing fast, but is still a young technology requiring more comprehensive studies on a large scale, especially with regards to its environmental impact (Helfer et al., 2014). The operation of FO processes at ambient pressures is interesting as high quality water can be produced using potentially less energy (application-dependant). To produce clean water from wastewater via FO, RO is necessary to recover the monovalent DS, requiring at least 30 bar to operate. That is more than for wastewater treatment via UF-RO. Digestion may thus be the only way to recover energy. Even if magnetic nanoparticles are used, these need to be separated. A permanent magnet is not an option; thermodynamically the same high energy would be required to separate the particles from the magnet. Using a temporary electric/magnetic field will likewise cost more energy than UF-RO, because osmotic pressure will be lost in the FO process due to CP. (NH4)2CO3 as a DS requires waste heat to decompose the ammonium salt, but another source of heating could be more efficient, i.e. natural solar or geothermal energies (Xie et al., 2013a). Solar power has been used for DS regeneration (Schrier, 2012) and increasing dewatering rates (Li et al., 2011a; Razmjou et al., 2013). When used for desalination, the reconcentration unit for FO does not require external pumping energy. The minimum thermodynamic energy required to achieve 50% recovery of fresh water from a 35 g/L TDS solution by RO is 3e5 kWh/m3 (Elimelech and Phillip, 2011; Khaydarov and Khaydarov, 2007). By excluding external pumping energy, the specific power consumption of an FO-solar desalination process can drop to 80% and neutral TrOCs between 40 and 90%. The FO membrane showed R > 90% for most TrOCs at pilot-scale, while rejection at bench-scale was generally lower. Membrane compaction, the fouling layer and optimised hydrodynamic conditions in the pilot-scale system have been suggested to improve rejections. Valladares Linares et al. (2011), spiking secondary wastewater effluent with 13 micropollutants, achieved similar rejections with FO for neutral compounds and >92% rejection for ionic compounds. In FO, rejection of charged TrOCs was found to be governed by electrostatic interaction and size exclusion, while rejection of neutral compounds was dominated by size exclusion (Alturki et al., 2013). Rejection also depends largely on fouling. Valladares Linares et al. (2011) found higher rejections of micropollutants on a fouled membrane than a clean one, with the exception of hydrophilic neutral compounds. Each group of compounds behaves different in the presence of a fouling layer, based on size, charge, polarity, adsorption capacity, mass transport capacity, membrane hydrophilicity after fouling and membrane swelling. By changing the surface properties of FO membranes, the interactions with TrOCs can be limited. Some TrOCs are of more public concern than others. Endocrine disrupting compounds (EDCs) for example, interfere with the endocrine (hormone) system. Cartinella et al. (2006) demonstrated similarly high rejections (77e99%) for estrone and estradiol in FO, but rejection was affected by the feed composition. Adding surfactants to the feed improved rejections, however these may change membrane surface properties e.g. hydrophobicity and charge, which can influence the flux. Boron and arsenic are other TrOCs adversely affecting human health, crop production and aquatic life. Jin et al. (2011) found boron rejections in FO experiments (30e60%) comparable to rejections achieved in RO (40e65%), but boron and arsenic rejections were also affected by the FO membrane orientation (Jin et al., 2012b), due to the degree of fouling. D’Haese et al. (2013), testing 20 TrOCs, found FO to achieve comparable rejection rates to NF, but lower rejections than RO. Most FO studies

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have carried out these investigations with CTA membranes. Using TFC membranes, TrOC rejection has been reported to increase (Jin et al., 2012a; Xie et al., 2013b); the rejection of neutral TrOCs was attributed to the more favourable AL structure of the TFC. Not many stand-alone treatment options can remove TrOCs entirely and regardless of improvements in FO, it may still not completely remove TrOCs either. However a combination of treatment processes (FO hybrids) should be sufficient, if used in a once-through system (D’Haese et al., 2013; Xie et al., 2013a).

8.

Concluding remarks

In this paper, FO has been discussed as a potential candidate for treating wastewater, due to, among other things, its high rejection capacity and low fouling propensity. Ineffective membranes and reverse solute leakage remain the main challenges hindering the growth of FO-wastewater applications. The use of hybrid systems, i.e. FOeMD, FOeRO, FOeNF and OMBR-RO and/or integrated with seawater desalination can enhance FO and make it commercially feasible for wastewater recovery. However overall energy balances for integrated systems are still lacking in order to compare the economic benefits. A better understanding of these concepts will further promote the use of this technology in existing and new applications of wastewater treatment. Moreover, concentrate disposal must be deliberated well, since waste streams containing high concentrations of heavy metals may be difficult to dispose. Energy-efficient pre-treatment of the feed is an option. Additionally, concentration of organics by FO could be advantageous for energy production and nutrient recovery. Higher quality water is in demand, due to the imposition of new and ever-changing water quality standards. Therefore interest in FO technology is growing as a potential, costcompetitive and reliable alternative.

Acknowledgements The authors wish to acknowledge the Dutch Ministry of Economic Affairs, Agriculture and Innovation (Agentschap NL) for their financial support (InnoWator grant: IWA10003).

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Forward osmosis for application in wastewater treatment: a review.

Research in the field of Forward Osmosis (FO) membrane technology has grown significantly over the last 10 years, but its application in the scope of ...
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