pasteur.epa.gov · web viewreverse osmosis forward osmosis hybrid technologies membrane bioreactors...
TRANSCRIPT
Membrane Processes for the Separation of Potential Emerging
Pollutants
Suhas P. Dharupaneedia, Sanna Kotrappanavar Natarajb, Mallikarjuna Nadagoudac,
Kakarla Raghava Reddyd, Shyam S. Shuklae, and Tejraj M. Aminabhavie*
a*Department of Chemistry, St. Joseph’s College, Langford Road, Bengaluru 560 027, India.
bCentre for Nano and Material Sciences, Jain University, Jain Global Campus, Kanakapura,
Ramanagaram, Bangalore - 562112, India.
cU.S. Environmental Protection Agency, ORD, NRMRL, WSD, WRRB, Cincinnati, Ohio
45268, United States.
dSchool of Chemical and Biomolecular Engineering, The University of Sydney, Sydney,
Australia 2006.
eDepartment of Chemistry and Biochemistry, Lamar University, Beaumont, Texas 77710,
United States.
___________________________________________________________________________
________________________________
*Corresponding author.
E-mail addresses: [email protected] (T. M. Aminabhavi)
1
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
Contents
1. Introduction
2. Potential emerging pollutants (PEPs)
3. Membrane-based separation processes (MBSPs)
3.1 Microfiltration
3.2 Ultrafiltration
3.3 Nanofiltration
3.4 Reverse osmosis
3.5 Forward osmosis
3.6 Hybrid technologies
3.6.1 Membrane bioreactors (MBR)
3.6.2 Photocatalytic membranes/reactors (PMs/PMRs)
4. Conclusions
5. References
2
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
ABSTRACT
The potential emerging pollutants (PEPs) including hazardous chemicals, toxic
metals, bio-wastes, etc., pose a severe threat to the human health, hygiene and ecology by
way of polluting the environment and water sources. These pollutants may have been
originated from the industrial effluent discharges from chemical factories, pharmaceutical,
food and metal processing industries. When these PEPs mix in water sources, they pollute the
water, thereby disturbing the benign environment. Considerable efforts have been made to
alleviate the environmental pollution, but the crisis still exists due to the non-availability of
appropriate methods of treatment. Innumerable methods have been developed for the
treatment of effluents to separate the toxic chemicals/metals. Of these, membrane-based
separation processes (MBSPs) employed to address the separation of toxic PEPs are proven
to be quite effective as compared to conventional techniques to produce clean water from the
waste streams at an affordable cost, using minimum energy. These methods have been widely
explored to achieve an efficient separation of PEPs. Among the many MBSPs used,
microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), reverse osmosis (RO), forward
osmosis (FO) and some judicially chosen hybrid technologies are widely employed. This
review attempts to address the advantages and application potential of MBSPs over the
conventional effluent treatment methods. The data compiled from various laboratories over
the past decade are critically discussed and the review provides in-depth analysis as well as
plausible solutions to the environmental pollution issues.
Keywords
Emerging pollutants, wastewater treatment, MBSPs, separation, polymers
3
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
1. Introduction
Of the World's total availability of water, nearly 71% of Earth's surface is covered by
saltwater and people living with the remaining 29% need fresh water to sustain life. The
World Health Organization (WHO) estimates that there are more than 1 billion people who
cannot get clean water. Health monitoring authorities report that ~4000 children die every
day due to waterborne diseases as result of water pollution around the world [1]. Therefore,
the world is facing severe drinking water crisis due to environmental hazards. The available
water sources are polluted due to human activities and discharge of effluents from various
industries. Even though many emerging water treatment technologies have been widely
adopted to treat the polluted water, but all of these methods are not highly successful at large
scale applications.
Of the many treatment methods, membrane-based separation processes (MBSPs) have
become more popular in recent days. If proper treatment methodologies are not developed or
adopted, then there will be more chaos on severe health issues due to water contamination
with waterborne pathogens due to increased discharge of potential emerging pollutants
(PEPs) (toxic chemicals, pharmaceuticals, heavy metals, fertilizers, sludge, endocrine
disrupters, etc.,) into water sources [2-4]. Thus, there is an increasing awareness that calls for
urgent intervention via technology innovation to avoid the already strained good water
supply, human health, and hygiene. Even though distillation-based technologies dominate the
industrial scale water treatment approaches, efforts are being made to develop economical,
energy efficient and straightforward MBSPs to provide clean water for many human activities
[5, 6].
PEPs are mostly originated from the many industrial processes and the effluents
discharged at the source require clear identification, separation, and disposal as otherwise
these wastes pose serious problems to water quality and ecology at large [7]. The United
States Environmental Protection Agency (USEPA) categorized the PEPs as hazardous
4
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
materials that lack regulatory standards [8]. The PEP-contaminated wastewaters usually
follow many vicious pathways [9, 10] as typically displayed in Figure 1, finally reaching the
water sources used for human consumption. Therefore, handling of PEPs is a severe problem,
because the majority of conventional wastewater treatment technologies have repeatedly
proven to be inefficient to eliminate even the trace amounts of toxic components [10]. In case
of sewage sludge and soils, PEPs may directly diffuse from the waste streams to reach the
groundwater, making the treatment methods much more difficult.
Fig. 1.Typical sources of PEPs. Reprinted with permission from Reference [8], copyright (2016) Elsevier.
Conventional water treatment methods such as adsorption, bio-oxidation, coagulation,
sedimentation, and filtration, even in hybrid combinations such as chlorination and UV
radiation, have been widely used in the literature; but most of these approaches are found to
be inadequate for an effective water treatment [11]. Therefore, water purification becomes
expensive and energy-intensive. In order to meet these challenges, innovative treatment
methods have been developed from time-to-time and used for water treatment. Of all these
methods (physical, chemical, biological) used, the widely explored MBSPs have provided
distinct advantages, including high water quality with a high rate of recovery and low
maintenance costs. Several newer membranes developed could improve separation
efficiencies. This has prompted extensive research efforts to adopt advanced MBSPs in water
treatment tasks [12, 13].
5
1
2
3
4
5
6
7
8
9
1011
12
13
14
15
16
17
18
19
20
21
22
Given the urgencies mentioned above, the present review attempts to critically
address the issues related to the separation of PEPs using microfiltration (MF), ultrafiltration
(UF), nanofiltration (NF), reverse osmosis (RO), and forward osmosis (FO). Some of the
hybrid technologies such as membrane bioreactors (MBRs) and photocatalytic membrane
reactors (PMRs) have also been developed. This review will also briefly summarize the the
types of membrane materials, their pore characteristics, operating methods to control the
filtration processes, and performance testing compared to other technologies. However,
desalination and related technologies, including thermal and solar distillation, will not be
considered in this review as these aspects are covered in previously published reviews [14-
16].
2. Potential environmental pollutants (PEPs)
Fig. 2. The yearly production of PEPs with environmental impact (million tons)
indicators [re-drawn from http://ec.europa.eu].
Understanding the effect of PEPs on human health and ecology in trace quantities is a
formidable task that requires an extensive study in terms of human biology, chemical nature
6
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
of the PEPs as well as the surrounding atmospheric parameters. Studies in the literature
suggests that >70% of PEPs are environmentally hazardous and toxic to human health [17].
Several organic chemicals and metals have potential applications as raw materials in
developing industrial products, but when these are discharged in the open atmosphere without
proper treatment, they pose severe threat to the ecology.
Figure 2 displays the data available on the production of such PEPs (in million tons)
during the period of one decade (2004-2013) and their environmental impact assessment.
These are displayed in three categories: (i) total production of PEPs, (ii) environmentally
harmful PEPs (causing superficial damage), and (iii) PEPs with severe environmental impact.
It can be seen that the magnitude of PEPs produced remains constant year-by-year, which
might increase in the near future. But we do not have adequate treatment technologies to
alleviate these pollutants. Among the several categories of PEP sources, pharmaceuticals
have received the most significant attention [18, 19]. The bio-active PEPs include drugs,
biologics, diagnostic agents, nutraceuticals, fragrances, sunscreen agents, etc. Such bioactive
metabolites turn into complex mixtures of PEPs when present in an aquatic environment.
Table 1: Concentration of various potential emerging pollutants in effluent and its percentage
removal by waste water treatment plants (WWTP).
Type of PEPs Compound Concentration in
effluent (µg/L)
Removal rate %
by WWTP
Phthalates
Diethyl phthalate 19.64 96.5
Dibutyl phthalate 12.44 95.8
Benzyl butyl phthalate 9.17 92.4
Di-(-2-ethyl hexyl) phthalate 39.68 90.2
Di methyl phthalate 2.07 71
Psyco-
stimulants
Caffeine 56.63 96.9
Paraxanthin 2.07 71
Desinfectants Triclosan 0.85 76.8
Cosmetics Galaxolide 4.28 76.2
Tonalide 0.87 76.2
Diuretics
Furosemide 0.41 59.8
Hydrochloro thiazide 2.51 53.2
Diatrioate 3.3 0.2
Metoprolol 1.53 55.8
Propanolol 0.19 48.5
7
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
β-Blockers
Sotalol 1.66 52.6
Analgesics and
anti-
inflammateries
Ibuprofen 13.4 74.2
Ketoprofen 0.48 31.1
Ketorolac 0.41 44
Clofibric acid 0.21 39.1
Antiepileptics
Antipyrin 0.04 32.5
Codein 2.86 32.5
Diclofenac 1.04 34.6
Antibiotics
Doxycyclin 0.65 35.4
Norfloxacin 0.11 54.3
Sulfamethoxazole 0.32 17.5
Trimethoprim 0.43 1.4WWTP/STP=Waste Water Treatment Plants ٭
The dioctyl phthalate, also known as DOP, is commonly used as an ingredient in
personal care products, food packaging materials, blood containers, and tubings. These PEPs
are to be tackled effectively [21]. Polychlorinated biphenyls (PCBs) are another group of
PEPs that generally exist in the fatty tissues in humans and are frequently discharged in the
environment. These are generally deposited in sediments due to their limited solubility in
water [22]. Polyaromatic hydrocarbons (PAHs) are listed as hazardous PEPs on a priority
basis by the USEPA. Bisphenol-A is used as a raw material for polycarbonate-based
healthcare plastics products. Deblonde et al. [20] carried out a survey on the efficiency of
WWTPs in the removal of PEPs. The concentration of PEPs in effluent water ranged from as
low as 0.007 to as high as 56.63 g/L, but the removal rate ranged from 0 – 97 %. Phthalates
showed > 90 % removal, while for antibiotics it varied from 50 to 71 %. Analgesics, anti-
inflamatory and beta-blockers are the most resistant to degradation with a value of 30-40 %.
The removal efficiency of pharmaceuticals such as tetracycline, codeine was alarmingly low (
> 10 %) (see Table 1). Dyeing of fabrics in textile mills consumes enormous amounts of
water of which >98% will end up as wastewater containing PEPs. Majority of textile
wastewater will combine with the urban wastewater, sewage and surface waters to end up
contaminating the groundwater, surface water sources as well as the soil [23, 24]. Organic
dyes and metal traces used for dye preparation will further contaminate food that may directly
affect the human health.
8
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
3. Membrane-based separation processes (MBSPs)
MBSPs are the type of phase changing technologies with a variety of applications
used to remove PEPs from wastewater sources. Membranes are produced from different types
of materials including polymers, ceramics, zeolites etc., which give rise to specific filtering
features depending upon the surface charge, pore size and hydrophobicity. MBSPs, can be
classified as MF, UF, NF, RO, and FO that utilize different types of membranes depending
on their pore sizes and morphologies as well as specific separation needs. Figure 3 shows a
schematic representation of different MBSPs. The mode of separation in each of these
processes varies from solution-diffusion to molecular diffusion to size exclusive principle
[26].
MF membranes have larger pore sizes (0.1 to 5 µm) than UF membranes, which typically
reject materials in size range of 0.1—10 µm. The UF regime with pore sizes 0.01 to 0.1 µm
reject colloidal particles, macromolecules, biopolymers, and viruses whose sizes range from
0.01 to 0.2 µm based on the size exclusive principle. Commercially, UF is the most widely
used method for wastewater treatment, water remediation, recovery of surfactants in
industrial cleaning, food processing, protein separation, etc. The UF membranes are generally
fabricated using cellulose derivatives, inorganic materials such as TiO2, Al2O3, ZrO, etc in
addition to typical polymers such as poly(acrylonitrile) (PAN), poly(sulfone amide) (PSA),
poly(ether sulfone) (PES), poly(vinylidene fluoride) (PVDF), etc. [27].
9
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
Fig. 3. Schematics of MBSP spectrum including process name, size range and potential
solute rejected over the prescribed range of pores. Reprinted with permission from Reference
[28], copyright (2016) Royal Society of Chemistry.
NF membranes reject PEPs in the size range of 0.001—0.01 µm, which includes most of
the organics, biomacromolecules, and a variety of metallic salts (beyond divalent salts). The
performance of NF falls between RO and UF. Compared to NF, UF and MF membranes, RO
membranes are non-porous and are prepared from dense polymers with voids, free volume
channels or pore sizes ranging from ~0.0001 to 0.001 µm. Hence, RO membranes separate
low molecular weight minerals, including metal ions as well as PEPs. The transport
mechanism in RO membranes is by molecular diffusion through statistically distributed (a
theoretical distribution with finite mean and variance) free volumes [26]. The most common
applications of RO are in the treatment of pulp and paper mill effluents to produce potable
water [29, 30].
In most MBSPs, the choice of precursor and the method of membrane preparation
depends on the desired application. General parameters such as water flux, high PEP
10
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
rejection, engineering design, stability under stress, chemical resistance and operating
temperature as well as pressures applied govern the choice of polymers. The performance of
membranes is mainly dependent on the pore structure, pore size distribution and morphology
of the polymer chain network in the membrane matrix [31-35]. Except in some MF processes,
inorganic membranes have been tested for many applications. Ceramic or zeolite powders
have been used in limited applications for membrane preparations due to their high costs. In
the last decade, nanocomposite membranes that contain various types of nanoparticles (NPs)
dispersed in a polymer matrix have emerged as high-performance RO and NF membranes
that are successfully commercialized for separation applications [36-40].
In all the MBSP modules developed and tested, flat sheet, hollow fiber, tubular, and spiral
wound, flat sheet membranes have been employed in domestic as well as industrial
wastewater treatment. Basic membrane structure and nature of the feed (pH), ionic strength,
solute size, pressure, temperature, and solute concentration govern the PEP rejection along
with the mass transfer across the membrane in NF and RO processes [41, 42].
RO technology has been used ever since Loeb and Sourirajan [43, 44] discovered
the first popular cellulose-based membranes several decades ago. RO has issues such as high
energy requirement, membrane fouling, and concentration polarization. This has led
researchers to develop osmotically driven forward osmosis (FO), in which the engineered
osmotic gradient across the membrane plays an important role in mass transportation and
separation [45, 46]. Since then FO has become more suitable and energy efficient for treating
the feed with a high fouling tendency (e.g., landfill leachate), which may not be economical
by other pressure-driven MBSP like RO. The stand-alone FO process has some niche
applications, such as fertilizer dilution and fruit juice concentration. Initially, FO has been
treated as an efficient pre-treatment step for subsequent processes in which purified water can
be recovered from diluted draw solution [45-47].
3.1 Microfiltration (MF)
As discussed earlier, MF has large pore size ranges and is mainly used to remove the
colloidal particles, dyes, organic matter and other high molecular weight soluble PEPs from
the waste streams. MF can be efficiently used to complex industrial PEPs along with other
hybrid combination of techniques. MF has been widely used to treat domestic wastewater
usually containing hormones such as estrone (E1), 17β-estradiol (E2), and 17α-
ethynylestradiol (EE2); E1 and E2 are classified as serious category PEPs in water that can
11
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
damage the human endocrine even if present in trace concentrations [48, 49]. Baby and
personal care products packed in polycarbonate casing utilize bisphenol A (BPA) as one of
the monomers. Studies have pointed out that BPA is a hazardous and environmental
endocrine disrupter [50]. Han et al. [51] employed a bench scale crossflow MF system to
separate PEPs such as estrogens from wastewater using a series of membranes prepared from
PES, cellulose acetate (CA), nitrocellulose, polyester, regenerated cellulose, and polyamide-
66 (PA). With a concentration of 0.2 µM solution containing estrogens, the PA membranes
with a pore size of 0.2 µm exhibited a sorption capacity of 81 L.m -2 (0.44 µg.cm-2) for E1,
150 L.m-2 (0.82 µg.cm-2) for E2, 208 Lm-2 (1.23 µg.cm-2) for EE2, and 69 L.m-2 (0.32 µg.cm-2)
for BPA. Surface adsorption of some of these PEPs at higher concentrations severely affected
the membrane performance, causing membrane scaling. However, the presence of organic
matter in the feed significantly affected the flux in polyamide membranes. PEPs showed
consistent interaction with polyamide membranes via H-bonding, thereby showing an
efficient removal of PEPs from the feed solutions.
Health and safety regulatory authorities are continually observing and documenting
the risk associated with the presence of micro-PEPs, such as progesterone. There have been
several adsorption techniques developed using activated carbon and biopolymers to separate
such micro-PEPs from the wastewater sources [52]. However, these methods face many
shortcomings owing to high-pressure drops, clogging, slow mass transfer and lack of
appropriate methods to recover adsorbing materials. To overcome these difficulties, Ragab et
al. [53] employed a spiral-wound configuration using zeolite imidazolate metal-organic
framework (ZIF-8) nanoparticles incorporated into poly(tetrafluoroethylene) (PTFE) double
layer polymeric membranes. This system provides shorter bed height and larger pores for
convective flow. Here, the ZIF-8 nanoparticle-induced MF membrane showed high flux
along with the rejection of as high as 95% for PEPs (hormones) at low operating pressures.
Such high rejection efficiency was due to hormones chemically interacting via H-bonding
with the high surface area of ZIF-8.
The PhACs such as diclofenac and ibuprofen when present in concentration ranges of
0.14 – 1.48 µg/L and 160 – 169 µg/L, respectively [54, 55] were degraded using UV/TiO2
photocatalysis [56]; the photocatalyst was finally separated from the contaminated
wastewater to obtain disinfected potable water. For this treatment, a hybrid MF was also
found to be useful for separating PEPs and simultaneous recycling of the photocatalyst TiO2.
On the other hand, Fischer et al. [57] used in-situ TiO2 synthesis to develop composite
12
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
membranes of PES with PVDF polymers using titanium tetra-isopropoxide (TTIP)
hydrolysis. These photoactive TiO2-induced composite membranes could remove degraded
methylene blue, diclofenac and ibuprofen efficiently. Further, the high concentration PEPs,
such as carbamazepine, diclofenac, atenolol, azithromycin erythromycin and pesticides (162–
240 ng/L) were removed from wastewater in a treatment plant to the extent of ~98% using
hybrid MF-RO process [58].
A review by Liu and Wong [59] discussed the importance of the techniques used to
remove PEPs such as antibiotics, supplements, drugs, cosmetics and other personal care
products (PPCPs). In continuation of such studies, Wang et al. [60] developed the CNT-
loaded membranes for efficient removal (~95%) of triclosan, acetaminophen, and ibuprofen
from the aquatic environments. This study suggested the capabilities of CNT loaded
composite membranes in removing PhACs due to favorable CNT-PhACs interaction.
3.2 Ultrafiltration (UF)
Unlike the MF process, UF has a solute rejection regime of above 2 kD molecular
weight for the macro-PEPs. Considering many of the PEPs that fall into the category of
macromolecular range, UF has been successful to treat such PEPs from wastewaters [61]. For
macro-PEPs in wastewater in the presence of other organics, UF alone cannot produce
ultrapure water. For such situations, the UF with commercial NF range membranes were used
to achieve the desired levels of water purity by removing PEP-contaminated secondary
effluents [62]. This study included the removal of eleven PEPs dissolved in municipal
secondary effluents using combined UF and NF processes. Majority of PEPs included in this
study were pharmaceuticals and pesticides. Several other studies suggested that hybrid
processes in which RO membrane showed severe fouling were combined with coagulation
and disk filtration to reduce fouling. Chon et al. [63] developed a large-scale water
reclamation unit comprising of a combination of coagulation and disk filtration (CC–DF)
along with UF/RO membranes to remove PEPs such as atenolol, carbamazepine, caffeine,
and sulfamethoxazole. However, this hybrid process was not effective to remove PEPs, but
RO membranes achieved high removal efficiency. Interestingly, the negatively charged PEPs
were retained efficiently by the tight membranes compared to the neutral pollutants. In all
these studies, membranes were washed by desorbing PEPs from the surface of UF and RO
membranes.
13
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
Acero et al. [64] used different micelles, such as sodium dodecyl sulfate(SDS), Triton
X-100 (TX-100), Tween 20 (TW-20), cetylpyridinium chloride (CPC) and
cetyltrimethylammonium bromide (CTAB) to improve the UF performance in separating
PhACs (acetaminophen, metoprolol, caffeine, antipyrine, sulfamethoxazole, flumequine,
ketorolac, atrazine, isoproturon, 2-hydroxybiphenyl and diclofenac). The authors observed
that CPC and CTAB, cationic micelles, could remove the negatively charged and hydrophilic
PhACs. Among all the micelles, CPC showed the optimum performance of up to 95%. This
type of micellar-enhanced UF can be suggested as a cost-effective alternative to separate
PEPs.
Activated carbon can be used as a pre-treatment as well as a post-UF adsorption step
for the retention of PEPs. Therefore, powder-activated carbon-UF or granular-activated
carbon-UF was designed for removing low molecular weight PEPs, which otherwise would
be difficult to remove using UF alone [65]. Acero et al. [66] used activated carbon and a UF
step-wise method to remove eleven PEPs (acetaminophen, metoprolol, caffeine, antipyrine,
sulfamethoxazole, flumequine, ketorolac, atrazine, isoproturon, 2-hydroxyphenyl and
diclofenac) in a UF treatment plant. Interestingly, low PhACs dose of 10–50 mg/L was
adequate to remove the PEPs from wastewater. Benitez et al. [67] used a UF-NF hybrid
process to separate four PhACs (amoxicillin, naproxen, metoprolol, and phenacetin) from the
secondary effluents. In both UF and NF, permeate fl fro were greatly influenced by
membrane morphology, applied pressure, and operating temperature. The retention
coefficients in UF membranes were higher for naproxen than metoprolol with the lowest
being phenacet. However, for commercial scale membranes, the trend was highest for
amoxicillin and lowest for phenacetin, probably due to tight pore size structures, leading to
direct rejection and electrostatic repulsion associated with the membranes. In this study,
except phenacetin, NF achieved the highest retention of 80% for PhACs.
Boleda et al. [68] considered UF-RO processes in a comparative assessment of other
existing treatment techniques for removing twenty-nine PhACs. Drinking water treatment
plants (DWTP) with an effluent comprising oxychlorination, floc, and sand filtration were
compared in two parallel treatment approaches. On the other hand, UF in combination with
RO and chlorination was designed to study the separation effects. Results showed that
advanced treatment processes were efficient over the conventional treatment in eliminating
PhACs up to 94%. This was attributed to carbon filtration, which is suitable for removing
conventional pollutants, but not the PEPs.
14
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
Bagastyo et al. [69] reported the electrochemical oxidation (ECO) process to
eliminate PEPs from RO concentrate streams. A moderate to high salinity feed was removed
efficiently using the ECO process [70]. In continuation of these efforts, Urtiaga et al. [71]
attempted to remove PhACs using a point-of-use hybrid UF-RO system. For major PEPs such
as atenolol, bezafibrate, caffeine, and diclofenac, UF was useful up to 20%. The
concentrations of PEPs in permeate varied between 4 to 44 ng/L (see Fig.4 for process
details). ECO of RO reject with diamond electrodes have decreased the total PEP content
from 149 µg/L down to 10 µg/L. Authors concluded that at high electro-oxidation intensity,
PEPs’ concentration was reduced drastically.
Fig. 4. Summary of micro-PEPs removal during (a) UF and (b) RO treatment. Reprinted with
permission from Reference [71], copyright (2013) Elsevier.
Wray et al. [72] studied the impact of surface stress on membrane fouling control
during removal of organic PEPs using an UF treatment plant for sixteen different PhACs and
EDCs. The results indicated that retention was dependent on specific water matrix in which
increased retention was achieved with higher concentrations of organic matter. In a control
study, it was shown that contaminant scale formation might not have acted as a secondary
selective barrier for retaining macromolecules and hydrophobic micro-PEPs. Under higher
shear stress conditions, a lower fouling propensity compound retention was improved for
water matrices with higher concentrations of organic matter and biopolymers. The
interactions between organic micro-PEPs, mainly hydrophobic, neutral compounds and
biopolymers in solution, induced enhanced retention in all these cases. This study proved the
value of implementing air scouring as a UF fouling control approach.
Some of the WWTP technologies have often shown limited success about PEPs
degradation as a whole [73]. Often, UF fails to achieve the required standards, and hence,
various hybrid technologies have been designed and tested, including AOPs. Acero et al.
15
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
[74] used AOP techniques for removing series of PEPs, such as 1-H-benzotriazole,
chlorophene, and nortriptyline, dissolved in different streams. Here, pre-treated wastewater
was fed into UF, and both permeate, and retentate was later treated using ozone or chlorine
separately; UF step removed the PEP traces with the exception of 1-H-benzotriazole.
Chlorination and ozonation were also effective in reducing PEPs in the concentrated stream,
providing effluent that may be recycled using activated sludge treatment in WWTPs. Thus,
following both the treatment methods showed promising results for the removal of selected
micro-PEPs.
Recently, ultrasound-treated AOPs have also proven promising for removing priority
organic PEPs from wastewater and solid wastes [75]. Ultrasonication works for both
biodegradable and non-biodegradable/refractory organic compounds. Cailean et al. [76]
employed a hybrid ultrasonication-UF process for the removal of 4-chlorophenol (4-CP). The
results suggested that homogeneous Sono-Fenton processes could degrade 4-CP in less than
1h, which was dependent upon the amplitude, power density, and other operating parameters.
The Sono-Fenton process was efficient up to 45% to reduce the organic load to UF process.
Thus, the process is more efficient in the case when the concentration of 4-CP is high in
aqueous effluents.
Ultrasound (US) techniques have also provided an alternative technology to control
membrane fouling, but the ultrasonic parameters affecting the efficiency of membrane
cleaning have not been fully explored. Naddeo et al. [77] studied the US-assisted UF with
varying ultra-sonic frequencies for cleaning UF membrane surfaces during PEP removal.
Here, ultrasonic field drastically reduced the membrane fouling even at a lower frequency (35
kHz). On the other hand, PhACs, such as diclofenac, carbamazepine, and amoxicillin, were
difficult to remove from WWTPs [78]. Secondes et al. [79] also employed the hybrid process
of UF, adsorption and US irradiation concurrently to reject PhAC containing PEPs. The UF
alone was ineffective to remove the PEPs even up to 10%. Addition of activated carbon
retained 99%, and the use of US further increased the removal capacity to almost 100%.
Also, all three PEPs were removed up to 99.5% at 35 kHz ultrasound frequency.
3.3 Nanofiltration (NF)
NF has a much tighter pore regime than UF membranes, thus adding several benefits,
such as divalent salts and textile wastewater recovery. Neira-Ruízet et al. [80] performed a
case study on the untreated wastewater from agricultural and urban wastes. This study was
conducted in order to remove five PEPs viz., carbamazepine, BPA, triclosan, butyl benzyl
16
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
phthalate, and 4-nonylphenol using a commercial NF-270 under 800 kPa pressure. In these
studies, for preventing scaling, a pre-treatment step was applied. Lopen-Muñoz et al. [81]
employed NF for the removal of sulfamethoxazole, diclofenac sodium, hydrochlorothiazide,
4-acetamidoantipyrine, nicotine and ranitidine hydrochloride using two commercial NF-90
and NF-270 membranes. Interestingly, at low pressures, the rejection of PEPs varied almost
linearly. Steric hindrance and dynamic interaction between the solute PEPs and membrane
interface were responsible for achieving the rejection. Solute retention by NF 90 was quite
high (>95%), whereas for NF-270, it was low, ranging from 75% (for nicotine) to 95% (for
ranitinide hydrochloride).
Polyamide membrane exhibits surface adsorption property leading to poor separation
performance [82]. Using surface adsorption as a parameter, Semião et al. [82] studied the
adsorption and retention of estrone and estradiol using polysulfone, polyester and polyamide
membranes. Among all the membranes tested, polyamide NF membranes showed the highest
hormone adsorption. Selective layer morphology and pore size were found to be critical in all
surface adsorption and retention of PEPs. The size of the pore in association with steric
exclusion and pH of the medium were also crucial for hormone surface adsorption.
Interestingly, at pH 7, high solute and membrane interaction (compared to pH 11) was
attributed to electrostatic repulsive effect of the solute from the membrane surface. In another
study, the presence of hormone-type moieties and tert butyl phenol in secondary wastewater
was studied using NF-270 [83]. The authors used a commercial membrane in which PEPs
were retained up to 90%. Here, also steric exclusion was central to the separation of
hormone-mimicking compounds.
The PhAC elimination using the traditional treatment processes have shown limited
success [84]. Kim et al. [85] used grafted polyamide membranes with methacrylic acid cross-
linked with ethylene diamine (ED) to separate BPA, ibuprofen and salicylic acid. In this
study, BPA showed a high rejection of 95%, whereas pristine membranes showed 74%. Also,
rejection of ED-modified membrane for ibuprofen and salicylic acid was slightly lower than
those of metallic acid modified membranes. Interestingly, succinic acid membranes recovered
their electro-negative surface that helped to retain all the PEPs in the concentrate. In another
study, Sun et al. [86] used NF hollow fiber membrane having a charged surface for efficient
removal of cyclophosphamide. The membrane was fabricated using hyperbranched
polyethyleneimine (PEI) as a cross-linking agent onto polyamide-imide backbone. The
17
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
spongy-like porous membrane support provided a high flux with high structural stability for
water permeation even at elevated pressures.
Nghiem et al. [87] evaluated the retention of non-ionizable carbamazepine in the
presence of ionizable PhACs, sulfamethoxazole and ibuprofen using NF. It was observed that
retention of carbamazepine in the concentrate was independent of feed chemistry. PhAC
retention increased when the compound was transformed from a neutral to a negatively
charged compound with an increase in its pKa value. Therefore, retention of the negatively
charged sulfamethoxazole and ibuprofen was elevated as a result of increased ionic strengths.
Along with the MBSPs, adsorptive treatment [88] and AOPs [89] have also been
investigated to eliminate PEPs from wastewater. Liu and coworkers [90] performed the
feasibility study of removing antibiotics, namely norfloxacin (NOR), ofloxacin(OFL),
roxithromycin (ROX) and azithromycin, from wastewater treatment plant. The separation
scheme is depicted in Fig. 5. High rejections up to 98% were obtained in all the NF
experiments. The UV/O3 process achieved excellent removal efficiencies up to 87%,
dissolved organic carbon (DOC) of 40%, an increased BOD/COD ratio of 4.6 times, and a
reduction of acute toxicity up to 58%. Overall, NF efficiently removed all the antibiotics from
the WWTP effluents, but UV/O3 was able to further eliminate antibiotics in the NF
concentrate. Therefore, zero discharge of micro-PEPs from WWTPs was achieved using the
proposed hybrid system.
WWTP
Fig. 5. Hybrid NF combined with adsorption/AOP used for wastewater treatment. Redrawn
with permission from Reference [90], copyright (2014) Elsevier.
18
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
Developing novel membranes for purification of active pharmaceutical ingredients
(APIs) from the classes of genotoxic impurities is a challenging task [91, 92]. Martínez et al.
[91] used PES-NF membranes for the recovery of 1-(5-bromo-fur-2-il)-2- bromo-2-
nitroethane in a cross-flow NF configuration with retentions of >80%. The most prominent
results in terms of flux of ethanol were obtained from membranes prepared with 25% and
75% of PES along with a commercial NF-270 membrane.
Recently, the solvent filtration membrane process, also known as organic solvent NF (OSNF)
or organophilic NF (ONF), is an emerging technology for organic-based effluent treatment
[92]. Székely et al. [92] used OSNF to remove genotoxic impurities from a total of nine API
model feeds having macrolides and amides. The study focused only on understanding the
potentials of replacing stage extractions in traditional purification techniques in which typical
API recoveries reached ~80%. From an environmental view-point, the NF filtration was
solvent intensive. Thus, for economic reasons, process integration was recommended for
solvent recovery. The recently published review [12] on NF membrane covers several aspects
of such applications, especially in wastewater treatment and desalination.
Székely et al. [93] evaluated the feasibility of various methods for the removal of 1,3-
diisopropylurea (IPU). The use of OSNF at a dilution ratio of 3 was optimum to achieve 90%
removal of IPU with as low as ~2.5% loss of the model API. A novel IPU selective
molecularly-imprinted polymer (MIP) was used to remove the trace amounts of IPU, thereby
achieving 83% removal for a feed of 100 ppm concentration in a single step. The
combination of OSNF with diafiltration (DF) at a dilution ratio of 3:1 MIP showed a
reduction of IPU from 100 mg IPU/g of API to 2 mg IPU/g.
Ahmad et al. [94] used four types of commercial NF membranes (NF-90, NF-270,
NF-200 and DK) of ~200 MWCO to separate pesticides, dimethoate and atrazine from the
contaminated water sources. Among all the membranes tested, the commercial NF-90
showed the highest rejection, while NF-270 showed higher mass transport. On the other hand,
NF-90 showed the most significant potential for acetaminophens retention from aqueous
media.
Perfluro octane sulfonates (PFOS), a new class of PEPs with fluorinated alkane
sulfonates, are widely used in surfactants, coating materials, fire retardants, lubricants, metal
plating solutions and polymer additives [95]. PFOS are persistent, bio-accumulative, and
toxic even at trace concentrations [96]. In this regard, methods were developed for an
19
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
effective removal of trace amounts of PEPs to ensure PFOS-free drinking water. One of the
notable efforts was made by Zhao et al., [97] and the authors used NF-270 membranes to
remove PFOS from the simulated surface water containing calcium ions. The results showed
that increase of calcium chloride concentration enhanced the PFOS rejection from 94% to
99% for a feed as low as 100 ppb.
Even though chlorination has been widely used as a disinfectant in wastewater
treatment, recent studies have reported that chlorination of organic matter in freshwater
resulted in the formation of disinfection by-products. Trihalomethanes (THM) is a by-product
belonging to a new PEP class. Uyak et al. [98] enhanced the retention efficiencies of THM
using NF-200 membranes that removed more of THM than that of NF DS5 membrane,
suggesting that NF is one of the best available technologies for removing THM. Even though
moderately high rejection values were observed for the majority of organic micro-PEPs in
NF/RO, trace amounts of PEP could still be found in the permeate. Hence, it is important to
judge the efficiency of these systems in removing such PEPs. To verify this, Verliefde et al.
[99] made an essential contribution by prioritizing these contaminants and the efficiency of
NF in their separation as displayed in Table 2. NF was proved to be effective in removing
larger PEPs, smaller hydrophiles and charged micro-PEPs. Rejection of PEPs by NF was
qualitatively predicted, and the results were compared with the literature data to understand
their removal efficiency.
Table 2: Qualitative rejection prediction based on octane-water partition coefficient (log Kow) and experimental retention values for the prioritized PEPs. Reprinted with permission from Reference [99], copyright (2007) Elsevier.
PEPs Molar mass (g/mol)
Rejection (%)
Hormones
17 β-Estradiol 272 85-100
17 α-Ethinylestradiol
296 n.a.
Estrone 270 60-90
Progesterone 314 90-100
Testosterone 288 80
Industrial chemicals
20
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
212223
Bisphenol A 228 70
p-Dimethyl phthalate
194 65-80
p-Diethyl phthalate 222 65-80
Nonyl phenol 220 70-90
*MTBE 88 89.6
PFOS 152 97
Pesticides
Atrazine 216 68-98
Simazine 202 75-93
Chloropyrifos 350 > 99
Pharmaceuticals
Primidone 218 72-87
Carbamazepin 236 93
Ibuprofen 206 30-95
MTBE-Methyl tertiary butyl ether; PFOS- Perfluro octane sulfonates
It is still the subject of debate as to why RO is a preferred process given that NF
demonstrated comparable removal efficiencies or even better results regarding the operation
and maintenance costs. Yangali-Quintanilla et al. [100] established that NF is an efficient
technique for the removal of organic PEPs compared to RO. Thus, the removal of neutral
compounds such as dioxane was achieved from ~82% to 85% for both NF and RO. However,
the removal of ionic compounds was more than 97% for both NF and RO processes.
Zeng et al. [101] developed dopamine (DA)-modified halloysite nanotubes
(HNT)/PVDF blends by functionalizing HNTs with DA and blending with PVDF. These
membranes were tested for removing direct red-28 (DR-28), direct yellow-4 (DY-4) and
direct blue-14 (DB-14) dyes. The blend membranes increased the water flux by about 80%
compared to its nascent counterpart. The modified membranes showed dye rejection by 86%
for DR-28, 85% for DY-4 and 94% for DB-14. In efforts to separate micro-PEPs from
wastewater and in drinking water sources, Ilyas et al. [102] recently developed weak
polyelectrolyte multilayer (PEM)-based hollow fiber NF membranes. Notably, the PEMs
consisting of weak polyelectrolytes, such as poly(allylamine hydrochloride) (PAH) and
21
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
poly(acrylic acid) (PAA) were coated in a layered fashion onto UF support to obtain PEM-
based NF membranes. These membranes were further used to investigate the retention of
varying size (200 – 400 g/mol) micro-PEPs, which were charged and hydrophilic. The micro-
PEPs included for the study were: atenolol, sulfamethoxazole, naproxen, atrazine, and
bezafibrate. The membranes prepared at a pH of 6 showed high a retention up to 80%.
In a recent study by Wang et al. [103], multi-walled carbon nanotube (MWCNT)-
based membranes were prepared and used for the removal of primary effluent PEPs from the
source water. In order to increase the membrane performance and to overcome the
interference of biopolymers and humic acid, the authors have used coagulation as a pre-
treatment. The method was used for removing aminophen, caffeine, triclosan, and
carbendazim to the extent of 11-34 %.
3.4 Reverse osmosis (RO)
In recent years, there has been a growing interest to incorporate hybrid RO/NF
membrane processes for the treatment of sewage and industrial wastewater. Boleda et al.
[104] studied the feasibility of RO in eliminating certain drugs and metabolites from the
secondary treated wastes in a pilot plant. Three different commercial membranes (LE, BW30,
and XFR) were used to study the rejection of a range of PEPs, but no significant data on
rejection rates were observed.
Ozaki et al. [105] studied the retention efficiencies of thirteen pharmaceuticals and
personal care products (PPCPs) and five EDCs by simultaneous adsorption, size exclusion,
and diffusion methods. Size exclusion was probably a dominant phenomenon in the tight NF
membrane. Besides, molecules having higer Kow (octane-water partition coefficient) values
were adsorbed onto the membrane surface and pores to transport across via diffusion
mechanisms. Rejection of PPCP and EDCs were improved when the solution pH was higher
than the solute pKa values, suggesting that electrostatic repulsion played a vital role in the
rejection of dissociated solutes. Therefore, the degree of dissociation in organic micro-PEPs
at the desired pH values is essential for separation studies.
Kegel et al. [106] studied drinking water treatment plants equipped with RO and
activated carbon filters for the removal of several micro-PEPs. The rejection rates for
hydrophilic and small MW solutes, such as nitrosodimethyl amine, dioxane, and 2-
methylisoborneol were quite low. The solute removal by activated carbon filtration was
22
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
robust. Removal of antibiotics using both NF and RO from a simulated wastewater
manufacturing plant was investigated by Košutić et al. [107]. Rejection of antibiotics by both
the membranes was >98%. On the other hand, low MWCO NF membrane retention for
smaller antibiotics was ineffective, which varied from 0.517% to 0.976% for RO and tight
NF, respectively.
The contamination of N-nitrosodimethylamine (NDMA), a member of a family of
extremely potent carcinogen N-nitrosoamine, in potable water occurs as a result of
disinfection carried out on wastewaters [108]. In 2002, the Health Services Department in
California established 10 ng/L as the critical level of NDMA in drinking water [109].
Plumlee et al. [110] described a method for NDMA detection with a high extraction
efficiency using solid phase extraction (SPE). Later, authorities have detected NDMA using
the above method in the secondary effluent of the Orange county sanitation district in
California to the level of 20–59 ng/L. Even though the initial tests using MF on wastewater
effluent were ineffective to reduce NDMA concentrations, but the secondary effluent
treatment using RO with TFC membranes typically resulted in NDMA rejection of ~50–65%.
Khazaali et al. [111] studied the removal of BPA using low-pressure RO. A critical
range of pressure (408 -476 kPa) was effective for BPA retention. By changing the pH from 8
to 10, BPA rejection decreased, but when BPA was ionized, the interaction between the ions
caused higher rejection. At more elevated feed concentrations, the effect of concentration
polarization was more significant, and BPA concentration in the permeate was elevated. In
any case, a maximum of 87% BPA rejection was obtained at 50 mg/L feed concentration.
Cyclophosphamide (CP) is one of the commonly used drugs in chemotherapy, which
adversely affects living organisms if present in water. The rejection of CP in feed water using
NF, RO and MBR was investigated by Wang et al. [112]. According the results, RO was
effective in CP retention up to 90%, but the rejection of CP was 20–40% in NF. On the other
hand, for MBR effluent treatment, CP rejection rate by NF was enhanced, suggesting that
both MBR-RO and MBR-NF hybrid systems are promising for the treatment of real
wastewater containing CP.
Al-Rifai et al. [113] evaluated a range of micro-PEPs at different processing points in
a water recycling plant. The removal efficiencies of eleven PhACs and two EDCs were
examined using MF and RO processes. It was found that salicylic acid was abundant in
WWTP effluent (11–38 µg/L), followed by BPA (6 to 23 µg/L). Further, the concentration of
all PEPs decreased drastically from primary to secondary treatment. The significant retention
23
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
efficiencies in recycled water was >97%, resulting in product water concentrations of < 0.1
µg/L for most of the PEPs (apart from BPA (0.5 µg/L)). In this study, even though >0.5 µg/L
was measured in the product water for BPA, but its presence is a serious concern and is a
challenging task to engineer the RO systems for water recycling. In the first of its kind
parallel study, AOP was combined with a RO as an alternative approach to remove PhACs
from the organic matter effluent and other inorganic constituents. In this regard, Abdelmelek
et al. [114] examined the removal of PhACs using RO and further treated the retentate using
AOP. The degradation was monitored by excitation-emission matrix spectroscopy, where
•OH radical associated with proteins in RO retentate suggested efficient removal of PhACs.
The results from AOP treatment also revealed that MBSP has efficiently removed the PPCPs
from the effluent even in the presence of both organic and inorganic constituents.
A new class of PEPs known as β-blockers was identified that can cause severe risks to
human health. Among these, some are commonly used drugs, such as metoprolol and
propranolol, classified as potentially toxic to aquatic organisms. Benner et al. [115] studied
the effect of ozonation for the mitigation of β-blockers. Second-order rate constants on four
different β-blockers, including acebutolol and propranolol, were investigated by applying
ozone and •OH radical techniques. The ozonation of RO brine effluents was sufficient to
eliminate the β-blockers. However, tests on chlorinated and non-chlorinated WWTP effluent
showed increased ozone stability, but a decrease in •OH radical exposure, proving the
effectiveness of RO for the removal of β-blockers.
The treatment of sewage wastewater has been a challenging task using the
conventional methods. However, realizing that oxidation processes can reject the organics
from the contaminated streams, James et al. [116] developed a hybrid advanced oxidation
reverse osmosis (AOP-RO) method to treat PEPs from the secondary municipal wastewaters.
Using this method, >99% of PEPs and endocrine disrupting chemicals (EDCs) were removed
successfully in a pilot-scale experiment. Interestingly, for the EDC removal such as N-
nitrosodimethylamine (NDMA), the H2O2 dose was crucial. Further, Alonso et al. [117] used
a pilot scale commercial spiral wound membrane to remove antibiotic (ciprofloxacin) from
water with a high ionic strength using RO, which could remove ciprofloxacin up to > 90%.
3.5 Forward osmosis (FO)
In the recent past, forward osmosis (FO) has attracted as a promising technique for
water treatment and desalination. The method utilizes osmotic gradients artificially created by
24
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
high concentration draw solution across the membrane. In contrast to other MBSPs, FO is
less vulnerable to fouling because it is used in wastewater treatments and food industry as
well as to concentrate biomacromolecules, recover valuable metals, remove toxic metals, and
fill leachates. FO is an osmotic concentration process operating at zero hydrostatic pressure,
providing sustainable water treatment solution, but the method has not yet been applied for
large-scale applications except seawater desalination [118]; moreover, the lack of suitable FO
membranes restricted its commercial exploitation. The essential components required for
efficient FO process are: (i) membranes that are less prone to susceptibility to internal
concentration, (ii) efficient draw solutions and (iii) effective draw solution recovery process.
Aquaporins are the naturally occurring water channels in proteins that can be used
efficiently as semi-permeable water pathways in FO processes [119]. Aquaporins embedding
vesicles in TFC (<200 nm) membranes were deposited onto a porous polysulfone flat sheet
support [120] to act as FO channels. Novel membranes made of aquaporin have shown >90%
rejection for urea with a pure water permeation rate of 10 L/m2.h against 2M NaCl as
dissolved salt. Hancock et al. [121] studied the rejection of PEPs using FO in comparison
with a hybrid FO-RO system both at lab and pilot scales [see Fig. 6]. In both these systems,
more than thirty compounds of different kinds, including non-ionic, hydrophobic, negatively
and positively charged species were analyzed. The rejection of non-ionic compounds in RO
increased with increasing MW of PEPs. The RO process showed better rejection than FO for
BPA. On the other hand, FO showed better for % rejection rate (>99%) for methylparaben,
oxybenzone, amitriptyline, and triclosan compared to RO.
Fig. 6. Schematic representation of FO-RO hybrid system. Reprinted with permission from
Reference [122], copyright (2014) Elsevier.
25
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
Linares et al. [123] investigated the differences in the rejection of thirteen different
PEPs, which were detected by the FO in wastewater sources that had five hydrophilic
nonionic, three hydrophobic nonionic and four hydrophilic ionic micro-PEPs. The fouled FO
membrane secondary wastewater effluent (SWWE) was used as the feed. The resulting
effluent was desalinated at low pressure with the RO membrane. In standalone FO, the
rejection of hydrophilic neutral compounds varied between 49% and 85%, while for
hydrophobic neutrals, the rejection ranged from 40% to 88%. For ionic PEPs, rejections
ranged between 93% and 97%. Alturki et al. [124] tested different commercial cellulose
acetate (CA)-based membranes for FO, pressure retarded osmosis (PRO) and RO modes.
Compared to other NF membranes of similar MWCO, such as the commercial cellulose
triacetate (CTA) membrane, CA membranes showed higher water flux with better PEPs
retention. In RO mode, electrostatic interactions played a significant role in the retention of
electro-active PEPs. In FO and PRO modes, the retention of active PEPs was governed by the
electrostatic interaction between the membrane matrix and solute, while the rejection of
neutral compounds was dominated by size exclusion in which case retention was higher for
PEPs with high MW. In all the cases, retention of neutral PEPs was higher in FO compared to
RO.
Haloacetic acids (HAAs) are the well-known disinfection by-products (DBPs) that are
present with the highest concentrations in chlorinated or chloraminated sewage treatment
plant effluent. Trichloroacetic acid (TCAA) concentration can be as high as 471 μg/L in
chlorinated wastewater effluent [125]. In this study, a rapid detection method for analyzing
HAAs in wastewater effluents was investigated by ultra-performance liquid chromatography-
electrospray ionization tandem mass spectrometry (UPLC-MS/MS) method. Kong et al.
[126] studied the rejection of HAAs using FO attached with reverse draw solute permeation
experiments. The retention ratio for each HAA increased with an increased draw solute (DS)
concentration for the active layer facing the feed water (AL-FW) orientation. The rejection
rates for all HAAs were more than 95% for AL-FW orientation, but ranged from 74% to 89%
for the active layer facing the draw solution (AL-DS) orientation in 1 M NaCl draw solution
and the reverse draw solute flux of AL-FW orientation was lower compared to AL-DS
orientation.
Primary fine chemicals processing industrial effluents contain organic PEPs such as
phenols, aniline, and nitrobenzene [127], which can penetrate barriers of existing treatment
techniques, making them ineffective [128]. In this study, eight commercially available
activated carbons were studied for the removal of organic micro-PEPs, and their removal
26
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
efficiencies were discussed. Among the other MBSPs, RO is preferred for treatment of
effluent containing organic micro-PEPs, but the high operating pressure in RO often adds to
the cost of producing water and the membranes usually experience fouling problems[129].
In order to address the above-mentioned issues, Cui et al. [130] compared the
efficiency of FO in (i) lab-scale FO membranes under both FO and RO modes and (ii)
commercially available RO membranes under the RO mode for the removal of organic PEPs.
The lab scale fabricated FO membranes that had a thin-film polyamide layer onto the ultra-
porous support. The rejections of these TFC membranes to all three organic PEPs (aniline) in
FO were between 72% and 90%. Further, Xie et al. [131] studied the effects of pH and
membrane orientation on permeate flux and PEP retention for carbamazepine and
sulfamethoxazole using FO and pressure retarded osmosis (PRO) processes. It was observed
that permeate flux was lower in FO mode than in PRO mode. The retention of neutral PEP,
carbamazepine, was pH independent in both the operation modes. However, the retention of
carbamazepine was lower in PRO mode than in FO. Authors suggested steric barrier as the
probable cause for such separation patterns for neutral carbamazepine in FO.
Cartinella and co-workers [132] studied the removal of estrone and 17β-estradiol
using direct contact membrane distillation (DCMD) and FO. The DCMD showed >99% for
hormone, >99.9% for urea, and >99% for ammonia rejections at a constant flux. On the other
hand, FO removed estrone and estradiol equally, but hormone rejection was affected by the
initial feed concentrations. The environmental impact due to toxic load of olive mill
wastewater is estimated to be more severe than the municipal sewage. The FO was also used
in tackling the issues related to treatment of oil contaminated wastewater from the olive oil
extraction industries. The quantity of olive mill wastewater generated is typically ~5 m3/ton,
leaving high COD (220 g/L) in wastewater. The high variability of feed composition and
presence of antibacterial phenolic compounds makes oil-contaminated waste difficult to treat
[133]. Gebreyohannes et al. [134] used a single step FO plant to purify olive mill
wastewater against 3.7M MgCl2 draw solution with >98% PEP rejection, including
biophenols. However, MBR-based pre-treatment before FO was reduced by 92%, resulting in
30% flux enhancement. Recovery was achieved up to 95% pure water permeability with CTA
membranes.
Han et al. [135] developed a hybrid process of FO–coagulation/flocculation (CF) for
treating textile wastewater. FO was used for spontaneous recovery of water from wastewater
via FO-CF. The FO–CF hybrid system exhibited unique advantages of high water flux and
27
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
recovery rate using composite FO membranes with an initial water flux of 36.0 L/m2.h and
dye rejection of 99.9% against 2 M NaCl as the draw solution. At high flux rates,
12.0 L/m2.h, a 90% water recovery rate from wastewater was achieved. However, trace
PhACs from wastewater using FO concentrate remains an unsolved issue. To address this
problem, Liu et al. [136] designed an integrated FO system with electrochemical oxidation
(ECO) referred to as FO with ECO method. A synergistic effect can be observed in the
hybrid method wherein, antibiotic rejections by FO were increased due to the degradation of
antibiotics, while ECO was improved in the combined method (see Fig. 7). Results
demonstrated that the hybrid method exhibited excellent rejection of antibiotics up to 98% by
degrading >99% after 3 h of operation.
Fig. 7. Schematic representation of FOwEO process showing enhanced rejection and
elimination of antibiotics simultaneously. Reprinted with permission from Reference [136],
copyright (2015) Elsevier.
Recently, FO was evaluated both in bench scale and pilot scale applications for
landfill leachate treatment containing a range of PEPs [137]. Here, a new method of
combining activated sludge in hybrid combination with FO for wastewater treatment had
many advantages over other methods. The high rejection was achieved through FO in
retaining small PEPs in a biological reactor, thus significantly enhancing their retention time
in the reactor. A recent report on short-term bench scale studies [138] indicated that OMBR
offered well-designed treatment solution to produce high-quality water.
3.6 Hybrid technologies
3.6.1 Membrane bioreactors (MBR)
28
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
Except for membrane modules and aeration steps, MBR treatment is very much similar to
conventional activated sludge (CAS) techniques in which biodegradation and separation takes
place simultaneously. Radmanovic et al. [139] used both MBR and CAS methods to assess
the PhACs removal efficiencies at different operating conditions. Among them, MBR showed
higher PhAC-PEPs removal (~80%) than CAS technique. The membrane part of MBR
typically comprise MF regime (0.4 μm) pores since the size of organic flocs to be separated
from wastewater are around 10–100 μm. Even in some cases, low MWCO membranes have
been used to improve the efficiency of MBRs [140], but the use of low, tight membranes
increase energy consumption since their permeability is lower than the loose MF membranes
and requires high operating pressures. In another attempt, Petrovic et al. [141], studied the
removal of PhACs using flat sheet (0.4 μm pore size) as well as hollow fiber (0.05 μm pore
size) modules. Here, PEP removal in flat sheet was better than the hollow fiber membrane
due to higher surface area and lower MWCO of UF offered by the hollow fiber module.
The use of MBR units in paper mills have shown exceptional rate of PEPs removal
efficiencies when combined with MF. Also, dimensions of tubular membranes had an
influence on the quality of permeate in which 8-mm tubular PVDF membrane installed
outside the bioreactor drastically reduced the COD. The MBR-treated effluent was re-
circulated with no detrimental effect on product quality, which saved fresh intake water of a
paper mill for bleaching process by 80% and discharges by 50% [142].
In a pilot scale hybrid combination of MBR with NF, Li et al. [143] treated
textile wastewater containing COD, organic PEPs, color, and turbidity to achieve >90%
removal efficiency with a simultaneous water recovery. The NF membrane showed
considerable fouling due to presence of protein-like substances and a small amount of humic
acid (650-6,000 Da). In another study, performance of a commercial side
stream membrane bioreactor (SSMBR) and submerged membrane bioreactor (SMBR) was
studied for the treatment of textile wastewater [144]. The SSMBR showed COD removal by
>90%, with a color rejection of 20-90%.
Boonyaroj et al., [145] used a two-stage MBR system under extended sludge age
condition to enrich nitrifying bacteria. During MBR operation, organic removal efficiencies
exceeded 90%, while phenolic PEPs such as BPA and 4-methyl-2,6-di-tert-butyl phenol
(BHT) were removed to the extent of 65% and 75%, respectively. Furthermore, BPA and
BHT were biodegraded to the extent of 88% and 75%, respectively by enriched nitrifying
sludge. A recent review by Besha et al. [146] addressed PEP separation by activated a sludge
29
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
and MBR processes. The author discussed RO, FO and MBRs methods along with their
advantages for wastewater treatment. In addition, the review also provided an overview of
PEPs on microbial activities.
The sludge retention time (SRT) and liquid flux rate control MBR process. Komesli et
al. [147] investigated the removal efficiencies of five EDCs (diltiazem, acetaminophen,
progesterone, estrogen, and carbamazepine) using one full scale and one pilot scale MBR
plant considering the effect of SRT. Diltiazem was completely removed, but its removal was
not achieved in pilot scale applications to the extent of >85%. However, carbamazepine was
not removed in both the plants, while the removal of progesterone and estrogen remained the
same in both the plants. The difference in performance in full scale MBR and pilot MBR is
attributed to the occurrence of vibrations on membrane surfaces in full scale applications,
which helped to remove surface foulants.
Ragavan et al. [148] studied the fate and removal behaviour of 12 antibiotics
belonging to 5 different classes in an osmotic membrane bioreactor (OMBR). The FO
showed >75% rejection for all the antibiotics. Interestingly some antibiotics like
ciprofloxacin and roxithromycin showed biodegradation as a significant removal pathway,
while ofloxacin and roxithromycin showed the highest biosorption onto activated sludge.
Over all, OMBR was an effective method to treat antibiotics.
Park et al. [149] studied the effect of addition of two coagulants viz., polyaluminium
chloride (PACl) and chitosan into MBR system for the removal of pharmaceuticals and
PPCPs. By adding coagulants membrane permeability was increased remarkably 2.3 and 2.8
times for PACl and chitosan, respectively. Chitosan had little effect as a coagulant in PPCPs
removal, but PACl showed an increase in membrane efficiency up to 17-23 %. Overall,
combination of MBR with coagulation has reduced membrane fouling and increased
operation time.
3.6.2 Photocatalytic membranes/reactors (PMs/PMRs)
Photocatalytic membranes (PMs) and photocatalytic membrane reactors (PMRs) have
been of recent trends that have provided greater synergistic advantages for PEPs removal;
these methods can be combined with MBSPs. In such a configuration, suspended
photocatalysts can mineralize organics to minimize fouling and enhance membrane efficiency
[150]. PMRs comprise of (i) TiO2 powder suspended in the reactors and (ii) reactors with
immobilized TiO2 on a substrate material (e.g., glass, quartz, mesoporous materials, stainless
30
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
steel or polymers) [151]. In the latter, location of where the photocatalyst is impregnated onto
the support has some drawbacks. These drawbacks restrict mass transfer and block the active
surface, resulting in lesser access to irradiation and reaction mixture as well as the possibility
of catalyst deactivation. The studies have been reported in both configurations of PMRs,
depending on the type of membrane modules. Among these, submerged membrane
photoreactors have been successfully used for obtaining high-quality water, as depicted in
Fig. 8 [152].
Fig. 8. Submerged membrane photocatalytic reactor. Reprinted with permission from
Reference [152], copyright (2006) Elsevier.
The design of self-cleaning membranes for eco-friendly separation operations have
met with limited success. The PMRs that simultaneously separate and mineralize organic
PEPs in the feed stream have the better potential as self-cleaning membranes. The TiO2-based
PMs have shown advantages such as anti-fouling ability due to photocatalytic degradation of
foulants and confining PEPs within photocatalytic chamber. Photocatalysts also can offer
high flexibility to suit various membrane modules for industrial applications. Moosemiller et
al. [153] performed concurrent membrane separation and TiO2-based photocatalytic
oxidation processes simultaneously by using γ-Al2O3, and TiO2 supported ceramic
membranes and found them to be more stable [154]. Compared to other semiconducting
metal oxides, TiO2 received much greater attention due to its excellent characteristics to
photodegrade organic PEPs present in contaminated water in the presence of UV light
irradiation. Thus, TiO2/polymer composite photocatalytic membranes have multi-functional
31
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
properties such as the ability to photocatalytically remove the PEPs fouling the membrane
surface, the ability of high disinfection to produce clean water from wastewater at a high
membrane-flux with low fouling potential. Such composite membranes have excellent
aqueous stability for efficient and large-scale water-treatment applications.
Recently, TiO2 nanostructures (0D to 3D) were functionalized with various polymeric
membranes through different strategies and investigated their removal efficiency of various
environmental PEPs under different light sources (see Table 4). The higher removal
efficiency of photocatalytic composite membrane is due to a synergistic effect between
polymeric membrane and nano-TiO2, well-designed morphological structures, excellent
stability of composite membrane, high catalytic activity and recyclability.
Table 3. The list of recent TiO2-based hybrid photocatalytic membranes used for the removal
of various environmental pollutants.
Target pollutant Membrane type Light source % removal
Ref.
AlkalineMB PEO Xe-mercury 150 W - [165]
MB PAN-PDA Vis. Light(300 W Xenon)
90.00 [170]
MB PVDF/ZnO Visible light 80.00 [172]
MB PVDF-TrFE UV light 100.00 [184]MB PVA/bentonite No light 94.00 [180]
Tetrazine PVDF-TrFE Sun light 77.77 [166]
AcidicAcid orange 7 Quartz fiber UV light - [174]
Acid orange 7 PEI UV light 90.40 [175]
Acid orange 7 PVA/bentonite No light 85.00 [180]
Phenolic
Phenol PSF 50 W Halogen 74.00 [181]
BPA PVDF UV light 85.00 [173]
BPA PSF Visible light 92.30 [177]
2,4-dichlorophenol PES Visible light 63.74 [176]
32
1
2
3
4
5
6
7
8
9
10
11
12
Azo compound
Congo-red PVC UV light 95.00 [167]
FluorescentEosin yellow PSF Visible light 92.00 [168]
Miscellaneous
AT-POME PVDF UV light 67.30 [169]
Toluene PI UV light 74.96 [171]Oil-water separation PVA UV light 99.00 oil
rejection[178]
Cr (VI) Chitosan Visible light 54.00 at pH 4
[179]
CO2 reduction PVDF UV light 19.80 mol/Gcatal
yst/hour
[182]
Bromate Bromate UV light >90.00 [183]]
Ceramic membranes are prepared by electrospinning, coating, hot-pressing, sol-gel,
hydrothermal synthesis-filtration, grafting, electrochemical deposition, and other methods
followed by etching and anodizing TiO2 film onto suitable supports [155]. In addition,
photocatalytic efficiency of TiO2-PMs was improved via doping with WO3 such as silicon
(Si). Doping TiO2 with Si is an efficient way to enhance photocatalytic capability, thermal
and mechanical stability, quantum-size effect and surface wettability of the photocatalyst
[156]. Further, Ag-doped TiO2 exhibited improved photocatalytic efficiency and bactericidal
capability [157]. The TiO2 doped with tungsten showed enhanced visible light absorption by
narrowing the energy band gap, thereby increasing the possibility for solar photocatalysis
[158]. Thus, doping with tungsten or combining with WO3 imparted better band gap
reduction, leading to better photochemical degradation. Several PhACs and their active
metabolites have been treated efficiently by both PMRs and PMs. Molinari et al. [159]
utilized the combined polycrystalline TiO2 and NF processes with different membranes for
photocatalytic degradation of PhACs like furosemide, ranitidine (hydrochloride), ofloxacine,
phenazone, naproxen, carbamazepine and clofibric acid in a PMR process.
In the past, PMs have been studied for removing humic acids [155], textile dyes
[160], and bacterial disinfection [161], but limited studies are available on the remediation of
PhACs. One of the most relevant studies on the removal of PhACs by PMs was reported
33
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
[162], wherein TiO2 nanofibers integrated with a stainless steel filter were used for complete
removal of PEPs such as antihistamine and cimetidine drugs through photocatalytic process.
Electrospun TiO2 nanofibers were integrated into stainless steel filter through hot pressing
with a PVDF nanofiber as an interlayer or binder with >99% removal of PEPs. Here, the
presence of suitable binding layer has reduced the contact between PhACs and TiO2 layer,
offering increased flux. The composite membranes are highly suitable for photocatalytic
materials in the energy level, and they have recently applied for the removal of organic PEPs
[163, 164] due to rapid charge separation in electron-transfer process. Different metal oxides
and carbon materials have been coupled with polymer membrane supports by numerous
researchers and studied photocatalytic composite membranes for the removal of various
environmental pollutants as the results presented in Table 4.
Table 4. The list of recent different metal oxides and nanocarbons-based composite
photocatalytic membranes used for removal of various environmental pollutants.
Membrane type Target PEPs Light source Removal (%)
Ref.
ZnO/PAN MO UV light 99 [185]
ZnO/CA-PU RR 11RO 84
UV light 9890
[186]
ZnO/PES MB UV light 70 [187]
Halloysite/PVDF DR 28DY 4DB 4
Vis light 86.58593.7
[188]
WO3/membrane-coated stainless meshes
MB UV light 99.9 [189]
H4SiW12O40/CA TetracyclineMB
300 W mercury lamp
63.894.6
[190]
Ag3PO4/PAN MB 200W mercury light
98 [191]
CuMn2O4/ceramic membrane Benzophenone-3
UV light 81.1 [192]
SiO2on SiC substrate MB UV light 72 [193]GO-TiO2/PVDF BSA UV light 92.5 [194]GO/s-PBC MB
MOUV light 88
70[195]
rGO-g-C3N4/CA-PDA MB Vis light 99.8 [196]
GO/Triethanolamine-TiO2 Congo Red UV light 68 [197]
34
1
2
3
4
5
6
7
8
9
10
11
12
13
g-C3N4/Carbon fabric NOx Vis light 64 [198]
Bi2O2CO3-MoS2/Carbon fabric NO Vis light 68 [199]
g-C3N4-CNTs/Al2O3 membrane Phenol Vis light 94 [200]
4. Concluding remarks
In recent years, the presence of PEPs in water sources has created adverse effects on
human health, hygine and ecology. This situation has created continued pressure on
developing newer water treatment technologies since natural attenuation and conventional
treatment processes are not suitable to remove all the PEPs. An overview of the state-of-the-
art technologies based on MBSPs that are avilable to remove the PEPs in water is undertaken.
As our literature survey, the PEPs are originated in many different groups such as synthetic
chemicals, pharmaceutically active compounds, naturally occurring constituents and
biological species such as microorganisms. Among the many conventional methods of their
treatment, MBSPs, predominantly NF and RO techniques have proven suitable for removing
some of the PEPs. However, research efforts on the use of MF and UF technologies are
somewhat moderate. The recently developed hybrid MBR and FO techniques in combination
with such methods as AOPs and PMs/PMRs are quite popular in the present scenario. It is
realized that despite several inherent limitations, the membrane-based processes are quite
effective to achieve the removal efficiencies compared to other conventional water treatment
technologies. The main trends in this field are highlighted along with the recommendations
for further improvements/developments in the current status and knowledge gaps as well as
future directions.
Acknowledgment
Dr. Suhas thanks to the management of St. Joseph's College, Bangalore for awarding "Seed
Money for Research". Professors T.M. Aminabhavi and Shyam S. Shukla are thankful to
Lamar University, Beaumont, Texas for financial support of this study.
Disclosure Statement
No competing financial interests exist.
35
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
Disclaimer
The U.S. Environmental Protection Agency, through its Office of Research and
Development, funded and managed, or partially funded and collaborated in, the research
described herein. It has been subjected to the Agency's administrative review and has been
approved for external publication. Any opinions expressed in this paper are those of the
author(s) and do not necessarily reflect the views of the Agency; therefore, no official
endorsement should be inferred. Any mention of trade names or commercial products does
not constitute endorsement or recommendation for use.
Abbreviations
PEO – Poly(ethylene oxide)
MB – Methylene bluePVDF - Polyvinylidene difluoride (PVDF)
PSF - Polysulfone
PVDF-TrFE - poly[(vinylidenefluoride-co-trifluoroethylene]
PVC – Polyvinyl chloride
AT-POME - Aerobically treated palm oil mill effluent
PAN-PDA - Polyacrylonitrile-polydopamine
PI - Polyimide
PVDF-TrFE - Poly(vinylidenefluoride-trifluoroethylene)
BPA – Bisphenol A
PEI - Poly(ether imide)
PES - Poly(ether sulfone)
PVA - Polyvinylalcohol
AO 7 - Acid organge 74
O-CP - Orthochlorophenol
M-NP - M-Nitrophenol
MO - Methyl organge
CA-PU - Cellulose acetate-Polyurethane
RR 11 - Reactive red
RO 84 - Reactive orange
DR 28 - Direct red
DY 4 - Direct yellow
DB 4 - Direct blue
GO - Graphene oxide
BSA - Bovine serum albumin
36
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
s-PBC - Sulfonated pentablock copolymer
PDA - Polydopamine
GQDs - Graphene quantum dots
g-C3N4 - Graphitic carbon nitride
NOx - Nitrogen oxides
5. References
[1] Progress on sanitation and drinking water: 2015 update and MDG assessment
https://www.wssinfo.org/fileadmin/user_upload/resources/JMP-Update-report-
2015_English.pdf . Accessed on 05-05-2017, ISBN 978 92 4 150914 5, UNICEF and World
Health Organization 2015.
[2] T. Deblonde, C. Cossu-Leguille, P. Hartemann, Emerging pollutants in wastewater: A
review of the literature, Int. J. Hyg. Environ. Health. 214 (2011) 442–448.
[3] M. Stuart, D. Lapworth, E. Crane, A. Hart, Review of risk from potential emerging
contaminants in UK groundwater, Sci. Total Environ. 416 (2012) 1–21.
[4] S. Nasseri, S. Ebrahimi, M. Abtahi, R. Saeedi, Synthesis and characterization of
polysulfone/graphene oxide nano-composite membranes for removal of bisphenol A from
water, J. Environmental Management 205 (2018) 174-182.
[5] J.D. Ortego, T.M. Aminabhavi, S.F. Harlapur, R.H. Balundgi, A review of polymeric
geosynthetics used in hazardous waste facilities, J. Hazard. Mater. 42 (1995) 115–156.
[6] M. Sairam, S.K. Nataraj, T.M. Aminabhavi, S. Roy, C.D. Madhusoodana, Polyaniline
membranes for separation and purification of gases, liquids, and electrolyte solutions, Sep.
Purif. Rev. 35 (2006) 249–283.
[7] N.S. Thomaidis, A.G. Asimakopoulos, A.A. Bletsou, Emerging contaminants: a tutorial
mini-review, Global NEST J. 14 (2012) 72–79.
[8] M. Barbosa, N.F.F. Moreira, A.R. Ribeiro, M.F.R. Pereira, A.M.T. Silva, Occurrence and
removal of organic micropollutants: An overview of the watch list of EU Decision 2015/495,
Water Res. 94 (2016) 257–279.
[9]S. Mompelat, B. Le Bot, O. Thomas, Occurrence and fate of pharmaceutical products and
by-products from resource to drinking water, Environ. Int. 35 (2009) 803–814.
[10] J.O. Tijani, O.O. Fatoba, L.F. Petrik, A review of pharmaceuticals and endocrine-
disrupting compounds: Sources, effects, removal, and detections, Water. Air. Soil Pollut. 224
(2013) 1-29.
37
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[11] O.A.H. Jones, N. Voulvoulis, J.N. Lester, Human Pharmaceuticals in wastewater
treatment processes, Crit. Rev. Environ. Sci. Technol. 35 (2005) 401–427.
[12] A.W. Mohammad, Y.H. Teow, W.L. Ang, Y.T. Chung, D.L. Oatley-Radcliffe, N. Hilal,
Nanofiltration membranes review: Recent advances and future prospects, Desalination. 356
(2015) 226–254.
[13] P. Krzeminski, L. Leverette, S. Malamis, E. Katsou, Membrane bioreactors – a review
on recent developments in energy reduction, fouling control, novel configurations, LCA and
market prospects, J. Memb. Sci. 527 (2016) 207-227.
[14] N. Ghaffour, J. Bundschuh, H. Mahmoudi, M.F.A. Goosen, Renewable energy-driven
desalination technologies: A comprehensive review on challenges and potential applications
of integrated systems, Desalination. 356 (2015) 94–114.
[15] L.F. Greenlee, D.F. Lawler, B.D. Freeman, B. Marrot, P. Moulin, Reverse osmosis
desalination: Water sources, technology, and today’s challenges, Water Res. 43 (2009) 2317-
2348.
[16] O. Rodriguez, J.M. Peralta-Hernandez, A. Goonetilleke, E.R. Bandala, Treatment
technologies for emerging contaminants in water: A review, Chem. Eng. J. (2017)
DOI:10.1016/j.cej.2017.04.106.
[17]Production of environmentally harmful chemicals, by environmental impact class
EUROSTAT.See http://ec.europa.eu/eurostat/tgm/table.do?tab=table&init=1&
%20plugin=1&language=en&pcode=ten00011 accessed on 26.08.2016.
[18] J. Wang, S. Wang, Removal of pharmaceuticals and personal care products (PPCPs)
from wastewater: A review, J. Environ. Manage. 182 (2016) 620–640.
[19] A. Mendoza, J. Aceña, S. Pérez, M. López de Alda, D. Barceló, A. Gil, Y. Valcárcel,
Pharmaceuticals and iodinated contrast media in a hospital wastewater: A case study to
analyse their presence and characterise their environmental risk and hazard, Environ. Res.
140 (2015) 225–241.
[20] T. Deblonde, C. Cossu-Leguille, P. Hartemann, Emerging pollutants in wastewater: A
review of the literature, Int. J. Hyg. Environ. Health. 214 (2011) 442–448.
[21] M. Blanchard, M.J. Teil, D. Ollivon, L. Legenti, M. Chevreuil, Polycyclic aromatic
hydrocarbons and polychlorobiphenyls in wastewaters and sewage sludges from the Paris
area (France), Environ. Res. 95 (2004) 184–197.
[22] J. Sánchez-Avila, J. Bonet, G. Velasco, S. Lacorte, Determination and occurrence of
phthalates, alkylphenols, bisphenol A, PBDEs, PCBs and PAHs in an industrial sewage grid
38
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
discharging to a municipal wastewater treatment plant, Sci. Total Environ. 407 (2009) 4157–
4167.
[23] S.D. Kim, J. Cho, I.S. Kim, B.J. Vanderford, S.A. Snyder, Occurrence and removal of
pharmaceuticals and endocrine disruptors in South Korean surface, drinking, and waste
waters, Water Res. 41 (2007) 1013–1021.
[24] T.D. Minh, B. Y. Lee, M.T.N. Le, Methanol-dispersed of ternary
Fe3O4@γ-APS/graphene oxide-based nanohybrid for novel removal of benzotriazole from
aqueous solution, J. Environmental Management 209 (2018) 452-461.
[25] J.N. Chakraborty, Waste-water problem in textile industry, Fundamentals and Practices
in Colouration of Textiles; J.N. Chakraborthy (Ed.),Woodhead Publishing, India
Pvt.Ltd.,New Delhi, 2010 pp. 381-408.
[26] P. Munk, T.M. Aminabhavi, Introduction to macromolecular science, Wiley, New York,
2nd Ed., 2002.
[27] J. Wang, Z. Bai, Fe-based catalysts for heterogeneous catalytic ozonation of emerging
contaminants in water and wastewater, Chem. Eng. J. 312 (2016) 79-98.
[28] A. Lee, J.W. Elam, S.B. Darling, Membrane materials for water purification: design,
development, and application, Environ. Sci. Water Res. Technol. 2 (2016) 17–42.
[29] S.K. Nataraj, K.M. Hosamani, T.M. Aminabhavi, Distillery wastewater treatment by the
membrane-based nanofiltration and reverse osmosis processes, Water Res. 40 (2006) 2349–
2356.
[30] S.K. Nataraj, K.M. Hosamani, T.M. Aminabhavi, Nanofiltration and reverse osmosis
thin film composite membrane module for the removal of dye and salts from the simulated
mixtures, Desalination. 249 (2009) 12–17.
[31] D.P. Suhas, A. V Raghu, H.M. Jeong, T.M. Aminabhavi, Graphene-loaded sodium
alginate nanocomposite membranes with enhanced isopropanol dehydration performance via
a pervaporation technique, RSC Adv. 3 (2013) 17120-17130.
[32] S.P. Dharupaneedi, R.V. Anjanapura, J.M. Han, T.M. Aminabhavi, Functionalized
graphene sheets embedded in chitosan nanocomposite membranes for ethanol and
isopropanol dehydration via pervaporation, Ind. Eng. Chem. Res. 53 (2014) 14474-14484.
[33] D. P. Suhas, T. M. Aminabhavi, A. V. Raghu, Mixed matrix membranes of H-ZSM5
loaded poly(vinyl alcohol) used in pervaporation dehydration of alcohols: Influence of
Silica/Alumina ratio, Polym. Eng. Sci. 54 (2014) 1774-1782.
39
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
[34] V.T. Magalad, G.S. Gokavi, K.V.S.N. Raju, T.M. Aminabhavi, Mixed matrix blend
membranes of poly(vinyl alcohol)–poly(vinyl pyrrolidone) loaded with phosphomolybdic
acid used in pervaporation dehydration of ethanol, J. Memb. Sci. 354 (2010) 150–161.
[35] V.T. Magalad, G.S. Gokavi, M.N. Nadagouda, T.M. Aminabhavi, Pervaporation
separation of water - ethanol mixtures using organic - inorganic nanocomposite membranes,
J. Phys. Chem. C. 115 (2011) 14731–14744.
[36] S.K. Nataraj, K.S. Yang, T.M. Aminabhavi, Polyacrylonitrile-based nanofibers - A state-
of-the-art review, Prog. Polym. Sci. 37 (2012) 487–513.
[37] S.K. Nataraj, B.H. Kim, M-d. Cruz, J. Ferraris, T.M. Aminabhavi, K.S. Yang, Free
standing thin webs of porous carbon nanofibers of polyacrylonitrile containing iron-oxide by
electrospinning, Carbon Letters, 63 (2009) 218-220.
[38] M. Mandegari, H. Fashandi, Untapped potentials of
acrylonitrile-butadiene-styrene/polyurethane (ABS/PU) blend membrane to purify dye
wastewater, J. Environmental Management 197 (2017) 464-475.
[39] T.M. Aminabhavi, N. Mallikarjuna, Polymeric membranes: Polymeric nanocomposites;
Barrier properties and membrane application, Polym. News. 29 (2004)193-195.
[40] R.W. Baker, Membrane technology and applications, second ed., John Wiley and Sons,
Chichester, 2004.
[41] A.R.D. Verliefde, E.R. Cornelissen, S.G.J. Heijman, E.M.V. Hoek, G.L. Amy, B. Van
Der Bruggen, J.C. Van-Dijk, Influence of solute-membrane affinity on rejection of uncharged
organic solutes by nanofiltration membranes, Environ. Sci. Technol. 43 (2009) 2400–2406.
[42] V. Yangali-Quintanilla, A. Sadmani, M. McConville, M. Kennedy, G. Amy, Rejection
of pharmaceutically active compounds and endocrine disrupting compounds by clean and
fouled nanofiltration membranes, Water Res. 43 (2009) 2349–2362.
[43] S. Loeb, S. Sourirajan, Sea water demineralization by means of an osmotic membrane,
in: Saline Water Conversion—II, Advances in Chemistry, American Chemical Society,
Washington, DC, 1963, pp. 117–132.
[44] S. Loeb, S. Sourirajan,High flow porous membranes for separating water from saline
solutions, US 3133132 A, 1964.
[45] S. Zhao, L. Zou, C.Y. Tang, D. Mulcahy, Recent developments in forward osmosis:
Opportunities and challenges, J. Memb. Sci. 396 (2012) 1–21.
[46] A. Achilli, T.Y. Cath, E.A. Marchand, A.E. Childress, The forward osmosis membrane
bioreactor: A low fouling alternative to MBR processes, Desalination. 239 (2009) 10–21.
40
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[47] B. Mi, M. Elimelech, Organic fouling of forward osmosis membranes: Fouling
reversibility and cleaning without chemical reagents, J. Memb. Sci. 348 (2010) 337–345.
[48] R. Owen, S. Jobling, Environmental science: The hidden costs of flexible fertility,
Nature. 485 (2012) 441–441.
[49] M. Huerta-Fontela, M.T. Galceran, F. Ventura, Occurrence and removal of
pharmaceuticals and hormones through drinking water treatment, Water Res. 45 (2011)
1432–1442.
[50] U.S. Food and Drug Administration (U.S. FDA), 2014. Bisphenol A (BPA): Use in Food
Contact Application.Avilable at:
https://www.fda.gov/newsevents/publichealthfocus/ucm064437.htm#summary accessed on
07.05.17.
[51] J. Han, S. Meng, Y. Dong, J. Hu, W. Gao, Capturing hormones and bisphenol A from
water via sustained hydrogen bond driven sorption in polyamide microfiltration membranes,
Water Res. 47 (2013) 197–208.
[52] J. Rivera-Utrilla, M. Sánchez-Polo, V. Gómez-Serrano, P.M. Álvarez, M.C.M. Alvim-
Ferraz, J.M. Dias, Activated carbon modifications to enhance its water treatment applications.
An overview, J. Hazard. Mater. 187 (2011) 1–23.
[53] D. Ragab, H.G. Gomaa, R. Sabouni, M. Salem, M. Ren, J. Zhu, Micropollutants removal
from water using microfiltration membrane modified with ZIF-8 metal organic frameworks
(MOFs), Chem. Eng. J. 300 (2016) 273–279.
[54] K. Kümmerer, Pharmaceuticals in the Environment, Annual Rev. Environ.Res. 35
(2010) 57-75.
[55] Y. Zhang, S.U. Geißen, C. Gal, Carbamazepine and diclofenac: Removal in wastewater
treatment plants and occurrence in water bodies, Chemosphere. 73 (2008) 1151–1161.
[56] D. Gerrity, S. Snyder, Wastewater and Drinking Treatment Technologies in:
B.W.Brooks, D.B. Huggett (Eds.) Human Pharmaceuticals in the Environment - Current and
FuturePerspectives, Springer, New York, 2012, pp. 225.
[57] K. Fischer, M. Grimm, J. Meyers, C. Dietrich, R. Gläser, A. Schulze, Photoactive
microfiltration membranes via directed synthesis of TiO2 nanoparticles on the polymer
surface for removal of drugs from water, J. Memb. Sci. 478 (2015) 49–57.
[58] S. Rodriguez-Mozaz, M. Ricart, M. Kock-Schulmeyer, H. Guasch, C. Bonnineau, L.
Proia, M.L. de-Alda, S. Sabater, D. Barcelo, Pharmaceuticals and pesticides in reclaimed
water: Efficiency assessment of a microfiltration-reverse osmosis (MF-RO) pilot plant, J.
Hazard. Mater. 282 (2015) 165–173.
41
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[59] J.L. Liu, M.H. Wong, Pharmaceuticals and personal care products (PPCPs): A review on
environmental contamination in China, Environ. Int. 59 (2013) 208–224.
[60] Y. Wang, J. Zhu, H. Huang, H.-H. Cho, Carbon nanotube composite membranes for
microfiltration of pharmaceuticals and personal care products: Capabilities and potential
mechanisms, J. Memb. Sci. 479 (2015) 165–174.
[61] J.M. Laîné, D. Vial, P. Moulart, Status after 10 years of operation — overview of UF
technology today, Desalination131 (2000) 17-25.
[62] J.L. Acero, F.J. Benitez, F. Teva, A.I. Leal, Retention of emerging micropollutants from
UP water and a municipal secondary effluent by ultrafiltration and nanofiltration, Chem. Eng.
J. 163 (2010) 264–272.
[63] K. Chon, J. Cho, H.K. Shon, A pilot-scale hybrid municipal wastewater reclamation
system using combined coagulation and disk filtration, ultrafiltration, and reverse osmosis:
Removal of nutrients and micropollutants, and characterization of membrane foulants,
Bioresour. Technol. 141 (2013) 109–116.
[64] J.L. Acero, F.J. Benitez, F.J. Real, F. Teva, Removal of emerging contaminants from
secondary effluents by micellar-enhanced ultrafiltration, Sep. Purif. Technol. 181 (2017)
123–131.
[65] M. Campinas, M.J. Rosa, Assessing PAC contribution to the NOM fouling control in
PAC/UF systems, Water Res. 44 (2010) 1636–1644.
[66] J.L. Acero, F.J. Benitez, F.J. Real, F. Teva, Coupling of adsorption, coagulation, and
ultrafiltration processes for the removal of emerging contaminants in a secondary effluent,
Chem. Eng. J. 210 (2012) 1–8.
[67] F.J. Benitez, J.L. Acero, F.J. Real, G. Roldán, E. Rodriguez, Ultrafiltration and
nanofiltration membranes applied to the removal of the pharmaceuticals amoxicillin,
naproxen, metoprolol and phenacetin from water, J. Chem. Technol. Biotechnol. 86 (2011)
858–866.
[68] M.R. Boleda, M.T. Galceran, F. Ventura, Behavior of pharmaceuticals and drugs of
abuse in a drinking water treatment plant (DWTP) using combined conventional and
ultrafiltration and reverse osmosis (UF/RO) treatments, Environ. Pollut. 159 (2011) 1584–
1591.
[69] A.Y. Bagastyo, J. Keller, Y. Poussade, D.J. Batstone, Characterisation and removal of
recalcitrants in reverse osmosis concentrates from water reclamation plants, Water Res. 45
(2011) 2415–2427.
42
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[70] A. Anglada, A. Urtiaga, I. Ortiz, Contributions of electrochemical oxidation to waste-
water treatment: Fundamentals and review of applications, J. Chem. Technol. Biotechnol. 84
(2009) 1747–1755.
[71] A.M. Urtiaga, G. Pérez, R. Ibáñez, I. Ortiz, Removal of pharmaceuticals from a WWTP
seondary effluent by ultrafiltration/reverse osmosis followed by electrochemical oxidation of
the RO concentrate, Desalination. 331 (2013) 26–34.
[72] H.E. Wray, R.C. Andrews, P.R. Bérubé, Surface shear stress and retention of emerging
contaminants during ultrafiltration for drinking water treatment, Sep. Purif. Technol. 122
(2014) 183–191.
[73] A. Jelic, M. Gros, A. Ginebreda, R. Cespedes-Sánchez, F. Ventura, M. Petrovic, D.
Barcelo, Occurrence, partition and removal of pharmaceuticals in sewage water and sludge
during wastewater treatment, Water Res. 45 (2011) 1165–1176.
[74] J.L. Acero, F.J. Benitez, F.J. Real, E. Rodriguez, Elimination of selected emerging
contaminants by the combination of membrane filtration and chemical oxidation processes,
Water. Air. Soil Pollut. 226 (139) (2015) 1-14.
[75] G. Sekaran, S. Karthikeyan, C. Evvie, R. Boopathy, P. Maharaja, Oxidation of refractory
organics by heterogeneous Fenton to reduce organic load in tannery wastewater, Clean
Technol. Environ. Policy. 15 (2013) 245–253.
[76] D. Cailean, C. Teodosiu, A. Friedl, Integrated Sono-Fenton ultrafiltration process for 4-
chlorophenol removal from aqueous effluents: Assessment of operational parameters (Part 1),
Clean Technol. Environ. Policy. 16 (2014) 1145–1160.
[77] V. Naddeo, L. Borea, V. Belgiorno, Sonochemical control of fouling formation in
membrane ultrafiltration of wastewater: Effect of ultrasonic frequency, J. Water Process Eng.
8 (2015) 92–97.
[78] B.I. Escher, R. Baumgartner, M. Koller, K. Treyer, J. Lienert, C.S. Mc-Ardell,
Environmental toxicology and risk assessment of pharmaceuticals from hospital wastewater,
Water Res. 45 (2011) 75–92.
[79] M.F.N. Secondes, V. Naddeo, V. Belgiorno, F. Ballesteros Jr., Removal of emerging
contaminants by simultaneous application of membrane ultrafiltration, activated carbon
adsorption, and ultrasound irradiation, J. Hazard. Mater. 264 (2014) 342–349.
[80] F. Neira-Ruíz, A.A. Arévalo, J.C.D. Álvarez, B.J. Cisneros, Operating conditions and
membrane selection for the removal of conventional and emerging pollutants from spring
water using nanofiltration technology: the Tula Valley case, Desalin. Water Treat. 42 (2012)
117–124.
43
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[81] M.J. López-Muñoz, A. Sotto, J.M. Arsuaga, Nanofiltration removal of pharmaceutically
active compounds, Desalin. Water Treat. 42 (2012) 138–143.
[82] A.J.C. Semião, M. Foucher, A.I. Schäfer, Removal of adsorbing estrogenic
micropollutants by nanofiltration membranes: Part B-Model development, J. Memb. Sci. 431
(2013) 257–266.
[83] L.D. Nghiem, A.I. Schäfer, M. Elimelech, Nanofiltration of hormone mimicking trace
organic contaminants, Sep. Sci. Technol. 40 (2005) 2633–2649.
[84] I. Khouni, B. Marrot, P. Moulin, R. Ben Amar, Decolourization of the reconstituted
textile effluent by different process treatments: Enzymatic catalysis, coagulation/flocculation
and nanofiltration processes, Desalination. 268 (2011) 27–37.
[85] J-H. Kim, P-K. Park, C-H. Lee, H-H. Kwon, Surface modification of nanofiltration
membranes to improve the removal of organic micro-pollutants (EDCs and PhACs) in
drinking water treatment: Graft polymerization and cross-linking followed by functional
group substitution, J. Memb. Sci. 321 (2008) 190–198.
[86] S.P. Sun, T.A. Hatton, T-S. Chung, Hyperbranched polyethyleneimine induced cross-
linking of polyamide-imide nanofiltration hollow fiber membranes for effective removal of
ciprofloxacin, Environ. Sci. Technol. 45 (2011) 4003–4009.
[87] L.D. Nghiem, A.I. Schäfer, M. Elimelech, Role of electrostatic interactions in the
retention of pharmaceutically active contaminants by a loose nanofiltration membrane, J.
Memb. Sci. 286 (2006) 52–59.
[88] C. Djilani, R. Zaghdoudi, A. Modarressi, M. Rogalski, F. Djazi, A. Lallam, Elimination
of organic micropollutants by adsorption on activated carbon prepared from agricultural
waste, Chem. Eng. J. 189–190 (2012) 203–212.
[89] W.T.M. Audenaert, D. Vandierendonck, S.W.H. Van Hulle, I. Nopens, Comparison of
ozone and HO induced conversion of effluent organic matter (EfOM) using ozonation and
UV/H2O2 treatment, Water Res. 47 (2013) 2387–2398.
[90] P. Liu, H. Zhang, Y. Feng, F. Yang, J. Zhang, Removal of trace antibiotics from
wastewater: A systematic study of nanofiltration combined with ozone-based advanced
oxidation processes, Chem. Eng. J. 240 (2014) 211–220.
[91] M.B. Martínez, B. Van der Bruggen, Z.R. Negrin, P.L. Alconero, Separation of a high-
value pharmaceutical compound from waste ethanol by nanofiltration, J. Ind. Eng. Chem. 18
(2012) 1635–1641.
44
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
[92] G. Székely, J. Bandarra, W. Heggie, B. Sellergren, F.C. Ferreira, Organic solvent
nanofiltration: A platform for removal of genotoxins from active pharmaceutical ingredients,
J. Memb. Sci. 381 (2011) 21–33.
[93] G. Székely, J. Bandarra, W. Heggie, B. Sellergren, F.C. Ferreira, A hybrid approach to
reach stringent low genotoxic impurity contents in active pharmaceutical ingredients:
Combining molecularly imprinted polymers and organic solvent nanofiltration for removal of
1,3-diisopropylurea, Sep. Purif. Technol. 86 (2012) 79–87.
[94] A.L. Ahmad, L.S. Tan, S.R.A. Shukor, Dimethoate and atrazine retention from aqueous
solution by nanofiltration membranes, J. Hazard. Mater. 151 (2008) 71–77.
[95] S. Deng, Q. Yu, J. Huang, G. Yu, Removal of perfluorooctane sulfonate from
wastewater by anion exchange resins: Effects of resin properties and solution chemistry,
Water Res. 44 (2010) 5188–5195.
[96] Y. Zhou, B. Wen, Z. Pei, G. Chen, J. Lv, J. Fang, X. Shan, S. Zhang, Coadsorption of
copper and perfluorooctane sulfonate onto multi-walled carbon nanotubes, Chem. Eng. J. 203
(2012) 148–157.
[97] C. Zhao, J. Zhang, G. He, T. Wang, D. Hou, Z. Luan, Perfluorooctane sulfonate removal
by nanofiltration membrane the role of calcium ions, Chem. Eng. J. 233 (2013) 224–232.
[98] V. Uyak, I. Koyuncu, I. Oktem, M. Cakmakci, I. Toroz, Removal of trihalomethanes
from drinking water by nanofiltration membranes, J. Hazard. Mater. 152 (2008) 789–794.
[99] A. Verliefde, E. Cornelissen, G. Amy, B. Van der Bruggen, H. van Dijk, Priority organic
micropollutants in water sources in Flanders and the Netherlands and assessment of removal
possibilities with nanofiltration, Environ. Pollut. 146 (2007) 281–289.
[100] V. Yangali-Quintanilla, S.K. Maeng, T. Fujioka, M. Kennedy, G. Amy, Proposing
nanofiltration as acceptable barrier for organic contaminants in water reuse, J. Memb. Sci.
362 (2010) 334–345.
[101] G. Zeng, Z. Ye, Y. He, X. Yang, H. Shi, Z. Ye, Y. He, X. Yang, H. Shi, Z. Feng,
Application of dopamine-modified halloysite nanotubes/PVDF blend membranes for direct
dyes removal from wastewater, Chem. Eng. J. 323 (2017) 572-583.
[102] S. Ilyas, S.M.F. Abtahi, N. Akkilic, H.D.W. Roesink, W.M. De-Vos, Weak
polyelectrolyte multilayers as tunable separation layers for micro-pollutant removal by
hollow fibernanofiltration membranes, J. Membr. Sci. (2017) DOI:
10.1016/j.memsci.2017.05.027
45
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
[103] Y. Wang, H. Huang, X. Wei, Influence of wastewater percoagulation on adsorptive
filtration of pharmaceuticals and personal care products by carbon nanotube membranes,
Chem. Eng. J. 333 (2018) 66-75.
[104] M.R. Boleda, K. Majamaa, P. Aerts, V. Gomez, M.T. Galceran, F. Ventura, Removal
of drugs of abuse from municipal wastewater using reverse osmosis membranes, Desalin.
Water Treat. 21 (2010) 122–130.
[105] H. Ozaki, N. Ikejima, Y. Shimizu, K. Fukami, S. Taniguchi, R. Takanami, R.R. Giri, S.
Matsui, Rejection of pharmaceuticals and personal care products (PPCPs) and endocrine
disrupting chemicals (EDCs) by low pressure reverse osmosis membranes, Water Sci.
Technol. 58 (2008) 73–81.
[106] F.S. Kegel, B.M. Rietman, A.R.D. Verliefde, Reverse osmosis followed by activated
carbon filtration for efficient removal of organic micropollutants from river bank filtrate,
Water Sci. Technol. 61 (2010) 2603–2610.
[107] K. Košutić, D. Dolar, D. Ašperger, B. Kunst, Removal of antibiotics from a model
wastewater by RO/NF membranes, Sep. Purif. Technol. 53 (2007) 244–249.
[108] W.A. Mitch, J.O. Sharp, R.R. Trussell, R.L. Valentine, L. Alvarez-Cohen, D.L. Sedlak,
N-nitrosodimethylamine (NDMA) as a drinking water contaminant: A review, Environ. Eng.
Sci. 20 (2003) 389–404.
[109] Office of research and development (ORD), National centure for environmental
assessment (U.S. EPA), 2002. Avilable at: http://www.epa.gov/ngispgm3/iris/search.htm
accessed on 08.05.17.
[110] M.H. Plumlee, M. López-Mesas, A. Heidlberger, K.P. Ishida, M. Reinhard, N-
nitrosodimethylamine (NDMA) removal by reverse osmosis and UV treatment and analysis
via LC-MS/MS, Water Res. 42 (2008) 347–355.
[111] F. Khazaali, A. Kargari, M. Rokhsaran, Application of low-pressure reverse osmosis
for effective recovery of bisphenol A from aqueous wastes, Desalin. Water Treat. 52 (2014)
7543–7551.
[112] L. Wang, C. Albasi, V. Faucet-Marquis, A. Pfohl-Leszkowicz, C. Dorandeu, B.
Marion, C. Causserand, Cyclophosphamide removal from water by nanofiltration and reverse
osmosis membrane, Water Res. 43 (2009) 4115–4122.
[113] J.H. Al-Rifai, H. Khabbaz, A.I. Schäfer, Removal of pharmaceuticals and endocrine
disrupting compounds in a water recycling process using reverse osmosis systems, Sep. Purif.
Technol. 77 (2011) 60–67.
46
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[114] S.B. Abdelmelek, J. Greaves, K.P. Ishida, W.J. Cooper, W. Song, Removal of
pharmaceutical and personal care products from reverse osmosis retentate using advanced
oxidation processes, Environ. Sci. Technol. 45 (2011) 3665–3671.
[115] J. Benner, E. Salhi, T. Ternes, U. von-Gunten, Ozonation of reverse osmosis
concentrate: Kinetics and efficiency of beta blocker oxidation, Water Res. 42 (2008) 3003–
3012.
[116] C.P. James, E. Germain, S. Judd, Micropollutant removal by advanced oxidation of
microfiltered secondary effluent for water reuse, Sep. Purif. Technol. 127 (2014) 77–83.
[117] J. J. S. Alonso, N. El Kori, N. M._an-Martel, B. D. Río-Gamero, Removal of
ciprofloxacin from seawater by reverse osmosis, J. Environ. Manage. 217 (2018) 337-345.
[118] T.Y. Cath, A.E. Childress, M. Elimelech, Forward osmosis: Principles, applications,
and recent developments, J. Memb. Sci. 281 (2006) 70–87.
[119] C.Y. Tang, Y. Zhao, R. Wang, C. Hélix-Nielsen, A.G. Fane, Desalination by
biomimetic aquaporin membranes: Review of status and prospects, Desalination. 308 (2013)
34–40.
[120] Y. Zhao, C. Qiu, X. Li, A. Vararattanavech, W. Shen, J. Torres, C. Hélix-Nielsen, R.
Wang, X. Hu, A.G. Fane, C.Y. Tang, Synthesis of robust and high-performance aquaporin-
based biomimetic membranes by interfacial polymerization-membrane preparation and RO
performance characterization, J. Memb. Sci. 423–424 (2012) 422–428.
[121] N.T. Hancock, P. Xu, D.M. Heil, C. Bellona, T.Y. Cath, Comprehensive bench- and
pilot-scale investigation of trace organic compounds rejection by forward osmosis, Environ.
Sci. Technol. 45 (2011) 8483–8490.
[122] H. Luo, Q. Wang, T.C. Zhang, T. Tao, A. Zhou, L. Chen, X. Bie, A review on the
recovery methods of draw solutes in forward osmosis, J. Water Process Eng. 4 (2014) 212–
223.
[123] R.V. Linares, V. Yangali-Quintanilla, Z. Li, G. Amy, Rejection of micropollutants by
clean and fouled forward osmosis membrane, Water Res. 45 (2011) 6737–6744.
[124] A.A. Alturki, J.A. McDonald, S.J. Khan, W.E. Price, L.D. Nghiem, M. Elimelech,
Removal of trace organic contaminants by the forward osmosis process, Sep. Purif. Technol.
103 (2013) 258–266.
[125] J. Duan, W. Li, P. Sun, Q. Lai, D. Mulacahy, S. Guo, Rapid determination of nine
haloacetic acids in wastewater effluents using ultra-performance liquid chromatography-
tandem mass spectrometry, Anal. Lett. 46 (2013) 569–588.
47
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[126] F-xin Kong, H-wei Yang, X-mao Wang, Y.F. Xie, Rejection of nine haloacetic acids
and coupled reverse draw solute permeation in forward osmosis, Desalination. 341 (2014) 1–
9.
[127] S. Yang, Y. Sun, H. Yang, J. Wan, Catalytic wet air oxidation of phenol, nitrobenzene
aniline over multi-walled carbon nanotube as catalysts, Front. Environ. Sci. Eng., 9 (2015)
436–443.
[128] F. Zietzschmann, J. Altmann, A.S. Ruhl, U. Dünnbier, I. Dommisch, A. Sperlich, F.
Meinel, M. Jekel, Estimating organic micro-pollutant removal potential of activated carbons
using UV absorption and carbon characteristics, Water Res. 56 (2014) 48–55.
[129] T.S. Chung, X. Li, R.C. Ong, Q. Ge, H. Wang, G. Han, Emerging forward osmosis
(FO) technologies and challenges ahead for clean water and clean energy applications, Curr.
Opin. Chem. Eng. 1 (2012) 246–257.
[130] Y. Cui, X.Y. Liu, T-S. Chung, M. Weber, C. Staudt, C. Maletzko, Removal of organic
micro-pollutants (phenol, aniline and nitrobenzene) via forward osmosis (FO) process:
Evaluation of FO as an alternative method to reverse osmosis (RO), Water Res. 91 (2016)
104–114.
[131] M. Xie, W.E. Price, L.D. Nghiem, Rejection of pharmaceutically active compounds by
forward osmosis: Role of solution pH and membrane orientation, Sep. Purif. Technol. 93
(2012) 107–114.
[132] J.L. Cartinella, T.Y. Cath, M.T. Flynn, G.C. Miller, K.W.Hunter Jr., A.E. Childress,
Removal of natural steroid hormones from wastewater using membrane contactor processes,
Environ. Sci. Technol. 40 (2006) 7381-7386.
[133] C. Russo, A new membrane process for the selective fractionation and total recovery of
polyphenols, water and organic substances from vegetation waters (VW), J. Memb. Sci. 288
(2007) 239–246.
[134] A.Y. Gebreyohannes, E. Curcio, T. Poerio, R. Mazzei, G. Di Profio, E. Drioli, L.
Giorno, Treatment of olive mill wastewater by forward osmosis, Sep. Purif. Technol. 147
(2015) 292–302.
[135] G. Han, C-Z. Liang, T-S. Chung, M. Weber, C. Staudt, C. Maletzko, Combination of
forward osmosis (FO) process with coagulation/flocculation (CF) for potential treatment of
textile wastewater, Water Res. 91 (2016) 361–370.
[136] P. Liu, H. Zhang, Y. Feng, C. Shen, F. Yang, Integrating electrochemical oxidation into
forward osmosis process for removal of trace antibiotics in wastewater, J. Hazard. Mater. 296
(2015) 248–255.
48
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[137] T.Y. Cath, N.T. Hancock, C.D. Lundin, C. Hoppe-Jones, J.E. Drewes, A multi-barrier
osmotic dilution process for simultaneous desalination and purification of impaired water, J.
Memb. Sci. 362 (2010) 417–426.
[138] J.J. Qin, M.H. Oo, G. Tao, E.R. Cornelissen, C.J. Ruiken, K.F. de Korte, L.P. Wessels,
K.A. Kekre, Optimization of operating conditions in forward osmosis for osmotic membrane
bioreactor, Open Chem. Eng. J. 3 (2009) 27–32.
[139] J. Radjenovic, M. Petrovic, D. Barceló, Analysis of pharmaceuticals in wastewater and
removal using a membrane bioreactor, Anal. Bioanal. Chem. 387 (2007) 1365–1377.
[140] F. Zaviska, P. Drogui, A. Grasmick, A. Azais, M. Héran, Nanofiltration membrane
bioreactor for removing pharmaceutical compounds, J. Memb. Sci. 429 (2013) 121–129.
[141] M. Petrovic, M.J.L. de Alda, S. Diaz-Cruz, C. Postigo, J. Radjenovic, M. Gros, D.
Barcelo, Fate and removal of pharmaceuticals and illicit drugs in conventional and membrane
bioreactor wastewater treatment plants and by riverbank filtration., Philos. Trans. R. Soc. A.
367 (2009) 3979–4003.
[142] B. Hepp, L. Joore, H. Schonewille, H. Futselaar, Membrane filtration: A sustainable
water treatment technology within modern papermaking concepts, Papers Technol. 46 (2005)
41-48.
[143] K. Li, C. Jiang, J. Wang, Y. Wei, The color removal and fate of organic pollutants in a
pilotscale MBR-NF combined process treating textile wastewater with high water recovery,
Water Sci. Technol. 73 (2016) 1426–1433.
[144] S.I. Bouhadjar, S.A. Deowan, F. Galiano, A. Figoli, J. Hoinkis, M. Djennad,
Performance of commercial membranes in a side-stream and submerged membrane
bioreactor for model textile wastewater treatment, Desalin. Water Treat. 3994 (2015) 1–11.
[145] V. Boonyaroj, C. Chiemchaisri, W. Chiemchaisri, K. Yamamoto, Enhanced
biodegradation of phenolic compounds in landfill leachate by enriched nitrifying membrane
bioreactor sludge, J. Hazard. Mater. A 323 (2017) 311-318.
[146] A.T. Besha, A.Y. Gebreyohannes, R.A. Tufa, D.N. Bekele, E. Curcio, L. Giorno,
Removal of emerging micropollutants by activated sludge process and membrane bioreactors
and the effects of micropollutants on membrane fouling: A review, J. Env. Chem. Eng. (2017)
DOI: 10.1016/j.jece.2017.04.027
[147] O.T. Komesli, M. Muz, M.S. Ak, S. Bakirdere, C.F. Gokcay, Comparison of EDCs
removal in full and pilot scale membrane bioreactor plants: Effect of flux rate on EDCs
removal in short SRT, J. Environ. Manage. 203 (2017) 847-852.
49
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[148] D.S.S. Raghavan, G. Qiu, Y-P. Ting, Fate and removal of selected antibiotics in an
osmotic membrane bioreactor, Chem. Eng. J. (2017), doi: 10.1016/j.cej.2017.10.026.
[149] J. Park, N. Yamashita, H. Tanaka, Membrane fouling control and enhanced removal of
pharmaceuticals and personal care products by coagulation-MBR, Chemosphere (2018), doi:
10.1016/j.chemosphere.2018.01.063.
[150] J. Grzechulska, A.W. Morawski, Photocatalytic labyrinth flow reactor with
immobilized P25 TiO2 bed for removal of phenol from water, Appl. Catal. B Environ. 46
(2003) 415–419.
[151] R. Molinari, L. Palmisano, E. Drioli, M. Schiavello, Studies on various reactor
configurations for coupling photocatalysis and membrane processes in water purification, J.
Membr. Sci. 206 (2002) 399–415.
[152] J. Fu, M. Ji, Z. Wang, L. Jin, D. An, A new submerged membrane photocatalysis
reactor (SMPR) for fulvic acid removal using a nano-structured photocatalyst, J. Hazard.
Mater. B 131 (2006) 238–242.
[153] M.D. Moosemiller, C.G. Hill, M.A. Anderson, Physicochemical Properties of
Supported γ-Al 2 O3 and TiO2 Ceramic Membranes, Sep. Sci. Technol. 24 (1989) 641–657.
[154] H. Zhang, X. Quan, S. Chen, H. Zhao, Fabrication and characterization of silica/titania
nanotubes composite membrane with photocatalytic capability, Environ. Sci. Technol. 40
(2006) 6104–6109.
[155] M.J. Benotti, B.D. Stanford, E.C. Wert, S.A. Snyder, Evaluation of a photocatalytic
reactor membrane pilot system for the removal of pharmaceuticals and endocrine disrupting
compounds from water, Water Res. 43 (2009) 1513–1522.
[156] Y. Su, S. Chen, X. Quan, H. Zhao, Y. Zhang, A silicon-doped TiO2 nanotube arrays
electrode with enhanced photoelectrocatalytic activity, Appl. Surf. Sci. 255 (2008) 2167–
2172.
[157] K.K. Akurati, A. Vital, J.P. Dellemann, K. Michalow, T. Graule, D. Ferri, A. Baiker,
Flame-made WO3/TiO2 nanoparticles: Relation between surface acidity, structure and
photocatalytic activity, Appl. Catal. B Environ. 79 (2008) 53–62.
[158] K.A. Michalow, A. Heel, A. Vital, M. Amberg, G. Fortunato, K. Kowalski, T.J. Graule,
M. Rekas, Effect of thermal treatment on the photocatalytic activity in visible light of TiO2-
W flame spray synthesised nanopowders, Top. Catal. 52 (2009) 1051–1059.
[159] R. Molinari, F. Pirillo, V. Loddo, L. Palmisano, Heterogeneous photocatalytic
degradation of pharmaceuticals in water by using polycrystalline TiO2 and a nanofiltration
membrane reactor, Catal. Today. 118 (2006) 205–213.
50
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[160] D.M. de-Araújo, P. Cañizares, C.A. Martínez-Huitle, M.A. Rodrigo, Electrochemical
conversion/combustion of a model organic pollutant on BDD anode: Role of sp3/sp2 ratio,
Electrochem. Commun. 47 (2014) 37–40.
[161] R.A. Damodar, S-J. You, H-H. Chou, Study the self cleaning, antibacterial and
photocatalytic properties of TiO2 entrapped PVDF membranes, J. Hazard. Mater. 172 (2009)
1321–1328.
[162] S. Ramasundaram, H.N. Yoo, K.G. Song, J. Lee, K.J. Choi, S.W. Hong, Titanium
dioxide nanofibers integrated stainless steel filter for photocatalytic degradation of
pharmaceutical compounds, J. Hazard. Mater. 258–259 (2013) 124–132.
[163] N. Nasrollahi, V. Vatanpour, S. Aber, N.M. Mahmoodi, Preparation and
characterization of a novel polyethersulfone (PES) ultrafiltration membrane modified with a
CuO/ZnO nanocomposite to improve permeability and antifouling properties, Separation and
Purification Technology, 192 (2018) 369-382.
[164] P. Karaolia, I.M. Kordatou, E. Hapeshi, C. Drosou, Y. Bertakis, D. Christofilos, G.S.
Armatas, L. Sygellou, T. Schwartz, N.P. Xekoukoulotakis, D.F. Kassinos, Removal of
antibiotics, antibiotic-resistant bacteria and their associated genes by graphene-based TiO2
composite photocatalysts under solar radiation in urban wastewaters, Appl. Catalysis B:
Environmental 224 (2018) 810-824.
[165] M. Coto, S. C. Troughton, J. Duan, R. V. Kumar, T. W. Clyne, Development and
assessment of photo-catalytic membranes for water purification using solar radiation, Applied
Surface Sci. 433 (2018) 101-107.
[166] L. Aoudjit, P.M. Martins, S. Madjene, D.Y. Petrovykh, S.L. Mendez, Photocatalytic
reusable membranes for the effective degradation of tartrazine with a solar photoreactor, J.
Hazardous Materials 344 (2018) 408-416.
[167] R. Ahmad, J.K. Kim, J.H. Kim, J. Kim, Effect of polymer template on structure and
membrane fouling of TiO2/Al2O3 composite membranes for wastewater treatment, J.
Industrial and Eng. Chem. 57 (2018) 55-63.
[168] A.T. Kuvarega, N. Khumalo, D. Dlamini, B.B. Mamba, Polysulfone/N,Pd co-doped
TiO2 composite membranes for photocatalytic dye degradation, Separation and Purification
Tech. 191 (2018) 122-133.
[169] M.N. Subramanian, P.S. Goh, W.J. Lau, B.C. Ng, A.F. Ismail, AT-POME colour
removal through photocatalytic submerged filtration using antifouling PVDF-TNT
nanocomposite membrane, Separation and Purification Tech. 191 (2018) 266-275.
51
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
[170] Y. Shi, D. Yang, Y. Li, J. Qu, Z. Yu, Fabrication of PAN@TiO2/Ag nanofibrous
membrane with high visible light response and satisfactory recyclability for dye
photocatalytic degradation, Appl. Surface. Sci. 426 (2017) 622-629.
[171] G. Yang, J. Su, Y. Guo, H. Xu, Q. Ke, Fabrication of TiO2/PI composite nanofibrous
membrane with enhanced photocatalytic activity and mechanical property via simultaneous
electrospinning, J. Materials Sci. 52 (2017) 5404-5416.
[172] N. Li, Y. Tian, J. Zhang, Z. Sun, J. Zhao, J. Zhang, W. Zuo, Precisely-controlled
modification of PVDF membranes with 3D TiO2/ZnO nanolayer: enhanced anti-fouling
performance by changing hydrophilicity and photocatalysis under visible light irradiation, J.
Membrane Sci. 528 (2017) 359-368.
[173] N.A.M. Jaafar, A.F. Ismail, M. A. Mohamed, M.A. Rahman, M.H.D. Othman, W.J.
Lau, N. Yusof, Preparation and performance of PVDF-based nanocomposite membrane
consisting of TiO2 nanofibers for organic pollutant decomposition in wastewater under UV
irradiation, Desalination 391 (2016) 89-97.
[174] M.H. Fraile, R. Liang, M.J. Arlos, R.X. He, P. Peng, M.R. Servos, Y.N. Zhou,
Concurrent photocatalytic and filtration processes using doped TiO2 coated quartz fiber
membranes in a photocatalytic membrane reactor, Chem. Eng. J. 220 (2017) 531-540.
[175] D.K. Wang, M. Elma, J. Motuzas, W.C. Hou, D.R. Lopez, T. Zhang, X. Zhang,
Physicochemical and photocatalytic properties of carbonaceous char and titania composite
hollow fibers for wastewater treatment, Carbon 109 (2016) 182-191.
[176] S.N. Hoseini, A.K. Pirzaman, M.A. Aroon, A.E. Pirbazari, Photocatalytic degradation
of 2,4-dichlorophenol by Co-doped TiO2 (Co/TiO2) nanoparticles and Co/TiO2 containing
mixed matrix membranes, J. Water Process Eng. 17 (2017) 124-134.
[177] Q. Wang, C. Yang, G. Zhang, L. Hu, P. Wang, Photocatalytic Fe-doped TiO2/PSF
composite UF membranes: Characterization and performance on BPA removal under visible-
light irradiation, Chem. Eng. J. 319 (2017) 39-47.
[178] E. Karimi, A. Raisi, A. Aroujalian, TiO2-induced photo-cross-linked electrospun
polyvinyl alcohol nanofibers microfiltration membranes, Polymer 99 (2016) 642-653.
[179] L. Wang, C. Zhang, F. Gao, G. Mailhot, G. Pan, Algae decorated TiO2/Ag hybrid
nanofiber membrane with enhanced photocatalytic activity for Cr(VI) removal under visible
light, Chem Eng. J. 314 (2017) 622-630.
[180] A. Bouazizi, M. Breida, B. Achiou, M. Ouammou, J. I. Calvo, A. Aaddane, S. A.
Younssi, Removal of dyes by a new nano–TiO2 ultrafiltration membrane deposited on low-
cost support prepared from natural Moroccan bentonite, Appl. Clay Sci. 149 (2017) 127-135.
52
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[181] R. Mukherjee, S. De, Preparation of polysulfone titanium di oxide mixed matrix hollow
fiber membrane and elimination of long term fouling by in situ photoexcitation during
filtration of phenolic compounds, Chem. Eng. J. 30 (2016) 773-785.
[182] J.Z.Y. Tan, N.M. Nursam, F. Xia, Y.B. Truong, I.L. Kyratzis, X. Wang, R.A. Caruso,
Electrospun PVDF–TiO2 with tuneable TiO2 crystal phases: synthesis and application in
photocatalytic redox reactions, J. Materials Chem. A 2017, 5, 641-648.
[183] K.Y.A. Lin, C.H. Lin, S.Y. Chen, H. Yang, Enhanced photocatalytic reduction of
concentrated bromate in the presence of alcohols, Chem. Eng. J. 303 (2016) 596-603.
[184] P.M. Martins, R. Miranda, J. Marques, C.J. Tavares, G. Botelho, S.L. Mendez,
Comparative efficiency of TiO2 nanoparticles in suspension vs. immobilization into P(VDF–
TrFE) porous membranes, RSC Adv. 6 (2016) 12708-12716.
[185] N.D. Tissera, R.N. Wijesena, C.S. Sandaruwan, R.M. de Silva, A. Alwis, K.M.N. Silva,
Photocatalytic activity of ZnO nanoparticle encapsulated poly(acrylonitrile) nanofibers, Mat.
Chem. Phys. 204 (2018) 195-206.
[186] A. Rajeswari, S. Vismaiya, A. Pius, Preparation, characterization of nano ZnO-blended
cellulose acetate-polyurethane membrane for photocatalytic degradation of dyes from water,
Chem. Eng. J. 313 (2017) 928-937.
[187] G. Ognibene, D.A. Cristaldi, R. Fiorenza, I. Blanco, G. Cicala, S. Scire, M.E. Fragala,
Photoactivity of hierarchically nanostructured ZnO–PES fibre mats for water treatments,
RSC Adv. 6 (2016) 42778-42785.
[188] G. Zeng, Z. Ye, Y. He, X. He, X. Yang, J. Ma, H. Shi, Z. Feng, Application of
dopamine-modified halloysite nanotubes/PVDF blend membranes for direct dyes removal
from wastewater, Chem. Eng. J. 323 (2017) 572-583.
[189] M.A. Gondal, M.S. Sadullah, T.F. Oahtan, M.A. Dastageer, U. Baig, G.H. Mckinley,
Fabrication and Wettability Study of WO3 Coated Photocatalytic Membrane for Oil-Water
Separation: A Comparative Study with ZnO Coated Membrane, Scientific Reports 7 (2017)
1686.
[190] W. Li, T. Li, G. Li, L. An, F. Li, Z. Zhang, Electrospun H4SiW12O40/cellulose acetate
composite nanofibrous membrane for photocatalytic degradation of tetracycline and methyl
orange with different mechanism, Carbohydrate Polymers 168 (2017) 153-162.
[191] G. Panthi, S.J. Park, S.H. Chae, T.W. Kim, H.J. Chung, S.T. Hong, M. Park, H.Y. Kim,
Immobilization of Ag3PO4 nanoparticles on electrospun PAN nanofibers via surface
oximation: Bifunctional composite membrane with enhanced photocatalytic and
antimicrobial activities, J. Ind. Eng. Chem 45 (2017) 277-286.
53
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
[192] Y. Guo, Z. Song, B. Xu, Y. Li, F. Qi, J.P. Croue, D. Yuan, A novel catalytic ceramic
membrane fabricated with CuMn2O4 particles for emerging UV absorbers degradation from
aqueous and membrane fouling elimination, J. Hazardous Materials 344 (2018) 1229-1239.
[193] R.M. Huertas, M.C. Fraga, J.G. Crespo, V.J. Pereira, Sol-gel membrane modification
for enhanced photocatalytic activity, Separation and Purification Technology 180 (2017) 69-
81.
[194] J. Xu, T. Wu, J. Shi, K. Teng, W. Wang, M. Ma, J. Li, X. Qian, C. Li, J. Fan,
Photocatalytic antifouling PVDF ultrafiltration membranes based on synergy of graphene
oxide and TiO2 for water treatment, J. Membrane Sci. 520 (2016) 281-293.
[195] S. Filice, D. D'Angelo, S. Scarangella, D. Iannazzo, G. Compagnini, S. Scalese, Highly
effective and reusable sulfonated pentablock copolymer nanocomposites for water
purification applications, RSC Adv. 7 (2017), 45521-45534.
[196] F. Li, Z. Yu, H. Shi, Q. Yang, Q. Chen, Y. Pan, G. Zeng, L. Yan, Chem. Eng. J. 322
(2017) 33-45.
[197] G. Liu, K. Han, H. Ye, C. Zhu, Y. Gao, Y. Liu, Y. Zhou, A Mussel-inspired method to
fabricate reduced graphene oxide/g-C3N4 composites membranes for catalytic decomposition
and oil-in-water emulsion separation, Chem. Eng. J. 320 (2017) 74-80.
[198] H. Wu, D. Chen, N. Li, Q. Xu, H. Li, J. He, J. Lu, Hollow porous carbon nitride
immobilized on carbonized nanofibers for highly efficient visible light photocatalytic
removal of NO, Nanoscale 8 (2016) 12066-12072.
[199] J. Hu, D. Chen, N. Li, Q. Xu, H. Li, J. He, J. Lu, In situ fabrication of Bi2O2CO3/MoS2
on carbon nanofibers for efficient photocatalytic removal of NO under visible-light
irradiation, Appl. Catal. B: Environmental 217 (2017) 224-231.
[200] X. Wang, G. Wang, S. Chen, X. Fan, X. Quan, H. Yu, Integration of membrane
filtration and photoelectrocatalysis on g-C3N4/CNTs/Al2O3 membrane with visible-light
response for enhanced water treatment, J. Memb. Sci. 541 (2017) 153-161.
54
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
List of Figures
Figure 1.Typical sources of PEPs. Reprinted with permission from Reference [8], copyright
(2016) Elsevier.
Figure 2. Yearly production of PEPs with environmental impact (million tons) indicators [re-
drawn from http://ec.europa.eu].
Figure 3. Schematics of MBSP spectrum including process name, size range and potential
solute rejected over prescribed range of pores. Reprinted with permission from Reference
[28], copyright (2016) Royal Society of Chemistry.
Figure 4: Summary of macrocontaminant removal during (a) UF and (b) RO treatment.
Reprinted with permission from Reference [72], copyright (2013) Elsevier.
Figure 5: Hybrid NF combined with adsorption and AOP used for wastewater treatment.
Reprinted with permission from Reference [91], copyright (2014) Elsevier.
Figure 6: Schematic representation of FO-RO hybrid system.
Figure 7: Schematic representation of FOwEO process showing enhanced rejection and
elimination of antibiotics simultaneously. Reprinted with permission from Reference [134],
copyright (2015) Elsevier.
Figure 8: Submerged membrane photocatalytic reactor. Reprinted with permission from
Reference [8], copyright (2008) Royal Society of Chemistry.
55
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
List of Tables
Table 1: Various pharmaceutically active compounds (PhACs) detected and identified from different sources along with their continental distributions. Reprinted with permission from Reference [2], copyright (2011) Elsevier.
Table 2. Characteristic PEP levels in water produced at different stages in textile processing Industries. Reprinted with permission from Reference [25], copyright (2010) Elsevier.
Table 3: Qualitative rejection prediction based on octane-water partition coefficient (log Kow) and experimental retention values for the prioritized PEPs. Reprinted with permission from Reference [100], copyright (2007) Elsevier.
Table 4. The list of recent TiO2-based hybrid photocatalytic membranes used for removal of various environmental pollutants.
Table 5. The list of recent different metal oxides and nanocarbons-based hybrid photocatalytic membranes used for removal of various environmental pollutants.
56
1
2
3
456
78
910
111213
1415
1617
18
19
20
21
22
23