removal of pharmaceuticals and personal care ...

In: Wastewater Treatment Editor: Lydia M. Barrett

ISBN: 978-1-63482-467-5 © 2015 Nova Science Publishers, Inc.

The exclusive license for this PDF is limited to personal website use only. No part of this digital document may be reproduced, stored in a retrieval system or transmitted commercially in any form or by any means. The publisher has taken reasonable care in the preparation of this digital document, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained herein. This digital document is sold with the clear understanding that the publisher is not engaged in rendering legal, medical or any other professional services.

Chapter 4

REMOVAL OF PHARMACEUTICALS AND PERSONAL CARE PRODUCTS FROM WASTEWATER Ma Ying1, Jian Wu1, Ken D. Oakes3 and Anming Hu1,3* 1

Institute of Laser Engineering, Beijing University of Technology, Beijing, P. R. China Verschuren Centre, Department of Biology, Cape Breton University, Sydney, Canada 3 Department of Mechanical, Aerospace and Biomedical Engineering, University of Tennessee, Knoxville, TN, US

2

ABSTRACT Pharmaceuticals and personal care products (PPCPs) have attracted much recent attention as widespread emerging environmental contaminants, both due to their near ubiquitous detection in surface waters adjacent urban areas, but also their potential to generate endocrine modulating responses at low concentrations in exposed organisms. Although usually detected in environmental matrices only at low ng/L or µg/L ranges, adverse effects in exposed aquatic organisms, especially with respect to estrogenic compounds, have been widely reported and prompted concerns over the potential for human health effects. Consequently, there has been increasing research attention paid to cost-effectively removing PPCPs from wastewater prior to its‘ discharge to the environment. Numerous studies have examined the occurrence and fate of PPCPs in wastewater and adjacent receiving environments, focusing on their removal by conventional and advanced treatment processes at varying scales ranging from lab to bench experiments to full treatment plant manipulations. This review will discuss various removal mechanisms such as surface absorption, membrane filtration, advanced oxidization treatments (including photocatalytic degradation and electrolysis) with advantages and limitations of each treatment approach (or combinations thereof) in removing PPCPs explored. Elucidating the fate and effects of PPCPs in the environment, and the potential of recent technologies to limit their environmental contamination, are key elements in protecting future ecosystem and human health.

*

Corresponding author: [email protected].

82

Ma Ying, Jian Wu, Ken D. Oakes et al.

1. INTRODUCTION In the past few decades, aging demographics and consumer demand has lead to sharp increases in the production and consumption of pharmaceuticals and personal care products (PPCPs), along with the detection of parent compounds, metabolites, and degradation products in soil, sediment, and water matrices. It was not until Daughton and Ternes [1] published a review on environmental pharmaceuticals and personal care products (PPCPs) in 1999 that widespread public and researcher attention was focused on the distribution, particularly in surface waters, of these chemicals. The United States Environmental Protection Agency (USEPA) formally defines PPCPs as including not only prescription drugs and biologics, but also diagnostic agents, ―nutraceuticals,‖ fragrances, sun-screen agents, and numerous other consumer ingested or applied compounds which ultimately pass through municipal waste water treatments plants (MWWTPs) where they are detected, almost invariably only in trace concentrations, in the environment. In recent years, PPCPs in aquatic environments have been recognized as emerging contaminants due to their high polarity, low volatility, and near continuous infusion within MWWTP effluents. [2] With the advent of new analytical techniques, these compounds can be detected at extremely low concentrations in very complex matrixes. [3,4] As pharmaceuticals are designed to exert specific physiological effects on humans and animals at trace concentrations, their inadvertent release to the environment, and subsequent uptake by exposed aquatic organisms or human populations via drinking water intakes downstream of MWWTPs may produce unintended consequences in biota or medically compromised patients. As some PPCPs resist microbial degradation in MWWTPs, and in light of their consequent continuous input to the environment, PPCPs were considered to be a potential threat to human and animal health, depending on the properties of each compound, in both high and low concentrations. [5,6] Consequently, many questions regarding the environmental fate of PPCPs, their ecotoxicological and human health risks, and the ability of current water and wastewater treatment infrastructure to effectively remove these compounds remain unanswered. [4] Within this context, this chapter will summarize the current knowledge and uncertainties regarding PPCPs in aquatic environments, as well as, the effectiveness of varying water treatment processes (from conventional water treatment to advanced oxidation processes, with special attention to emerging nanotechnologies4) to remove PPCPs from water.

2. OCCURRENCE OF PPCPS IN AQUATIC ENVIRONMENTS 2.1. PPCPs in Water Supply Sources The occurrence of PPCPs in aquatic environments has been investigated and described in many countries globally; [7] although animal feeding operations [8] can be the dominant source of PPCPs in rural areas, in more urban areas, most PPCPs in surface waters originate from MWWTP effluents, although other sources and pathways are possible (Fig 1). These PPCPs are introduced to MWWTPs mainly through ingested medicines excreted as unmetabolized parent compounds (or reactive metabolites) via urine, effluents from pharmaceutical production facilities, hospitals and direct disposal of medication in toilets,

Removal of Pharmaceuticals and Personal Care Products from Wastewater

83

where many will eventually be introduced to surface water. MWWTP sludges and manure applied to land can also be significant non-point sources of PPCPs, especially during severe rain-events, where residues can leach into surface water and groundwater. Further, PPCPs may also be introduced to groundwater via aquaculture facilities and septic systems. [9] As worldwide environmental contaminants, the occurrence and concentration of PPCPs in water sources can vary significantly, ranging from a few ng/L to high μg/L values (Table 1), and can differ significantly even in the same river, typically increasing with additional downstream MWWTP effluent additions. [10]

Medicinal products

Pharmaceutical production facilities

hospitals

Wastewater treatment plants(WWTPs)

Surface water

Veterinary products

Waste disposal

manure

soil

WWTP sludge

Ground water

Septic systems

Aquaculture failities Drinking water

Figure 1. Scheme showing possible sources and pathways for the occurrence of pharmaceutical residues in the aquatic environment. Adapted from Ref. [3].

Table 1. PPCP concentration ranges in source water PPCP compound Clofibric acid Phenazon Salicylic acid Ibuprofen Paracetamol Acetaminophen Caffeine Sulfamethoxazole Trimethoprim Naproxen

Concentrations in source water (μg/L) 3.2–26.7 0.06-0.155 1.601-89.133 0.002–34.0 0.069–26.09 0.027–65.2 0–38.0 0.0017-2.17 0.002-0.18 0–135.2

References [11] [12] [12,13] [14-16] [17-18] [11,19] [11-12,20-21] [12,16,22] [12,19,21] [11,23-25]

84

Ma Ying, Jian Wu, Ken D. Oakes et al. Table 1. (Continued) PPCP compound Ciprofloxacin Bisphenol A Diclofenac Carbamazepine Gemfibrozil Estrone 17β-estradiol Triclosan Nonylphenol Octylphenol

Concentrations in source water (μg/L) 0.017-0.036 0–147.2 0.003–5 0.014–0.5 0.3-1.652 0.0006–4.7 0.0004–4.5 0.39–37.8 0.006–37 0.0008–0.7

References [29-30] [11,26-27 ] [12,15.28] [12,18,29-31] [14,20] [11,16,32] [11,16.33] [12,31,34-35] [26-27,36-37] [26-27,36-37]

2.2. PPCPs in Drinking Water Drinking water treatment plants (DWTPs) remove microorganisms and organic contaminants from raw water sources to meet drinking water standards, with the treatment process employed mainly a function of the best technology affordable that is adequate for the local raw water characteristics. Conventional treatment processes usually include coagulation (flocculation and sedimentation), filtration and disinfection steps, occasionally also including ion exchange (or water softening) processes. [38] Despite these treatment processes, studies demonstrate some recalcitrant PPCPs are not completely removed, [39,40] although relative to MWWTPs, there is less known of the behavior of PPCPs in drinking water. One reason may be the lack of systematic monitoring programs at many municipalities, or insufficient analytical sensitivity to detect PPCPs in drinking water, which are usually present at concentrations of sub-ng/L levels. [41] However, at least 25 PPCPs are reported more than once in drinking water in the reviewed literature, with the β-blocker atenolol being the most frequently detected compound in drinking water with concentrations above the limit of quantification (LOQ), followed by salicylic acid and carbamazepine, one of which were characterized at detectable levels in more than 30% of the drinking water samples. Table 2 reports the maximum measured concentrations and frequency of PPCP detection in drinking water samples from France [15], Spain [42], China [43] and the United States [44] in recent years. Table 2. Maximal measured concentrations and frequency of PPCPs detection in drinking water samples from France, Spain, China, and the United States PPCP compounds Atenolol Atorvastatin Bisphenol A Butylparaben Carbamazepine Carbazochrome Chlofibric acid Chloramphenicol

Maximum Concentration (ng/L) 23.00 18.00 6.00 28.00 32.00 0.89 19.00 2.00

Frequency 25 8 2 5 15 2 5 1

Removal of Pharmaceuticals and Personal Care Products from Wastewater PPCP compounds Diclofenac Fenofibric acid Fluoxetine Gemfibrozil Hydrochlorthiazide Ibuprofen Ketoprofen Meprobamate Metoprolol Methylparaben Naproxen Oxazepam Paracetamol Phenytoin Pravastatin Propylparaben Salicylic acid Sotalol Sulfamethoxazole Sulpiride Thiamphenicol Triclocarban Triclosan Trimethoprim

Maximum Concentration (ng/L) 18.00 16.00 0.82 2.10 7.00 39.00 7.00 42.00 1.00 12.00 11.00 2.50 45.00 19.00 0.20 9.00 31.00 3.00 3.00 0.17 7.00 13.00 1.20 1.00

85

Frequency 4 3 2 10 10 8 3 14 2 6 10 2 4 15 1 5 21 4 6 3 6 2 1 1

3. REMOVAL OF PPCPS DURING TREATMENT PROCESSES Both MWWTPs and DWTPs are designed to remove organic and inorganic suspended materials, flocculated matter, and pathogens from wastewater or drinking water sources; however, neither drinking nor wastewater treatment processes are specifically designed to remove PPCPs from water (although both do to some degree, varying with the physicochemical properties of individual PPCPs). Due to their high chemical stability and low biodegradability, most PPCPs are unlikely to be completely removed by conventional treatment processes (Figure 2). In recent decades, several technologies were developed to meet more stringent water treatment demands, including advanced oxidation processes (AOPs) and membrane technologies. [45]

3.1. Advances Oxidation Processes (AOPs) Advanced oxidation processes can be defined as ―near ambient temperature and pressure water treatment processes utilizing the generation of hydroxyl radicals in sufficient quantity to effect water purification‖ [46], and are generally used to remove organic and sometimes inorganic materials via oxidation by hydroxyl radicals (OH•), which are usually produced with the help of oxidants (e.g. O3, H2O2) and/or energy sources (e.g. UV light) or catalysts (e.g. TiO2, LaFeO3). AOPs can reduce contaminant concentrations from several hundred ppm

86


to less than 5 ppb, earning the moniker ―the water treatment processes of the 21st century‖ [46]. “Filters”

Drinking water

Wastewater treatment

Surface water Bank filtration Drinking water treatment

Figure 2. ―Filters‖ able to remove pharmaceutical compounds on the path from wastewater to drinking water. The different filters mainly reflect engineered solutions, but also pathways and environmental compartments where mechanisms of biodegradation, sorption, photolysis, and oxidation can occur. Adapted from Ref. [40].

3.1.1. Photocatalysis Photocatalysis is a change in the rate of a chemical reaction when the photocatalyst absorbs ultraviolet, visible, or infrared radiation. Titanium dioxide (TiO2) is the most widely utilized semiconductor photocatalyst in water treatment applications due to its low cost and commercial availability in various crystalline forms, particle characteristics, non-toxic nature, and high photochemical stability. [47] From a mechanistic viewpoint, photocatalysis occurs through the absorption of photo energy greater than the TiO2 band gap, to produce valence band holes and conduction band electrons. The holes and electrons can either: (1) recombine, with the resultant absorbed energy dissipated as heat, or (2) make their separate ways to the TiO2 surface where they can react with and degrade waterborne contaminant species adsorbed on the catalyst surface. [48] Valence band holes can also react with water and produce hydroxyl radicals, which are among the most powerful reactive oxygen species, capable of oxidizing a wide range of substrates. In addition to TiO2, other semiconductor catalysts such as ZnO and LaFeO3 have also been employed in water treatment applications. [49, 50] 3.1.1.1. TiO2 Photocatalysts under UV Irradiation Typically, UV illumination is the primary energy source driving photocatalytic activity, with pure TiO2 in the anatase phase possessing a band gap energy of 3.2eV, and numerous studies have investigated the potential of using TiO2 as a photocatalyst for removal of waterborne PPCPs. In most of these studies, the focus has been on determining optimal operating conditions, such as photocatalyst loading, initial concentration of the PPCP under investigation, and solution pH and light intensity to maximize degradation and mineralization. [51]


87

Hu et al. [52,53] synthesized highly entangled TiO2 nanowires (10-20 nm diameters, 100 μm lengths) and their photocatalytic degradation kinetics towards PPCPs under UV irradiation, demonstrating both that selected PPCPs can be degraded in a process described by pseudo-first order kinetics, and that TiO2 nanowire membranes have potential for water treatment applications. As with TiO2 nanowires, TiO2 anatase phase nanobelts (30-100 nm widths and 10 μm lengths) exposed to 365 nm UV illumination can also photocatalyze the oxidation of PPCPs. [54] These TiO2 nanobelts under high reaction temperatures, alkaline conditions, and in a concentration-dependent manner, facilitated faster photodegradation of theophylline, while amoxicillin conversely had mineralization retarded as solution pH was increased from 5 to 7.5. [55]

O₂

CB e¯

hν₃

e¯

e¯

O₂¯

OH¯

hν₁ h⁺

hν₂

h⁺

VB

H₂O₂

h⁺

OH¯ OH*

Figure 3. Photocatalystic mechanisms: hυ1 pure TiO2, hυ2 metallic-doped TiO2, hυ3 non-metallic-doped TiO2. Valance Band (VB) and Conduction Band (CB) illustrated by horizontal lines.

The anti-inflammatory Naproxen is one of the most detected pharmaceuticals in water, and studies on photocatalytic and photolytic degradation of naproxen demonstrated that under Xe-lamp (290-400 nm), naproxen degradation by photolysis was twice as efficient as by photocatalysis. [56] Low adsorption of naproxen on the catalytic surface, as well as possible recombination and deactivation of the OH• radical may have have contributed to the lower photocatalytic efficiency. In contrast with naproxen, diclofenac showed excellent removal by UV/TiO2 treatment. [57] The highest photocatalytic efficiency (residual diclofenac concentration = 0.4%) was obtained with a TiO2 loading/initial concentration of 624 mgL-1/8.17 mg L-1 with complete mineralization achieved within 2 h of irradiation. Degradation efficiencies of other PPCPs such as iopromide, acetaminophen, sulfamethoxazole, and carbamazepine under UV/TiO2 treatment were also investigated, [58] with more efficient degradation observed when the reactor influent was in buffered electrolyte than when in wastewater effluent (Figure 4) demonstrating the influence of wastewater effluent matrix in inhibiting photocatalytic deactivation.

88


Figure 4. Degradation efficiencies in buffered electrolyte and wastewater effluent. Adapted from Ref. [57].

3.1.1.2. Modified TiO2 Photocatalysts under Visible Irradiation As UV accounts for only 5% of visible spectrum radiation, incident solar energy is not efficient for the photoexcitation of unmodified TiO2; consequently, there has been a great deal of interest in preparing a TiO2 photocatalyst capable of efficiently utilizing sunlight. One means of reducing the wide band gap of TiO2, which requires an energy threshold only associated with the UV portion of the visible spectrum, is by modifying or ―doping‖ TiO2 with other constituents. Both metallic and non-metallic doping can reduce the energetic band gap by either raising the position of the valence band, or by lowering the position of the conduction band. As shown in Figure 3, electrons are excited from the defect state, imparted by metal doping, to the conduction band (CB) with excitation by a photon of energy hv2. Transition metal doping can also improve the trapping of electrons and prevent electron–hole recombination during irradiation, resulting in enhanced photoactivity. With non-metallic doping, electrons are excited in the valance band (VB) by impurity energy level hv3 (Fig 3) under UV irradiation, experiencing excitation from impurity energy levels under visible light irradiation. [59] Platinum ion-doped TiO2 (Ption-TiO2), one of several metallic-doping methods, demonstrated good photocatalytic activities with respect to the degradation of dichloroacetate under visible light irradiation. [60] The visible activity of Ption-TiO2 can be ascribed to the role of doped Pt ions as a charge generating center, producing free and trapped charges (Pt3+). (

) (

)

In addition to Ption-TiO2, other metallic-doped TiO2 compounds, such as Zr-doped TiO2, have also had their photocatalytic performance evaluated using the degradation of ibuprofen under UV-Vis light as a model. [61] Relative to P25, Zr-TiO2 demonstrated improved performance due to its increased specific surface area and improved adsorption properties. Iodine doping is a commonly employed non-metallic doping method using iodic acid as a


89

dopant [62,63] Studies suggest rutile iodine-doped TiO2 nanowires produce the highest methylene blue degradation relative to other synthesized nanomaterials. Their enhanced photocatalytic activity was attributed to existing oxygen vacancies, iodine multi-valences in IO-Ti bonds, and 3d state Ti3+ sites in the TiO2 lattice. Another further means of achieving visible light-driven photocatalytic activity is coupling TiO2 with narrow band gap semiconductors, which should have more negative and less positive valence band potentials. [64] As shown in Fig. 5, the mechanism underlying this technique enhances photocatalytic efficiency by promoting the separation of photo-induced electron–hole pairs through changing the carrier transfer pathway. For example titanium dioxide, when coupled with tungsten trioxide, as TiO2–WO3 composite nanotubes, completely removed a mixture of PPCPs (caffeine, metoprolol and ibuprofen) through photocatalytic ozonation using simulated solar light and photocatalytic ozonation in less than 40 min (with up to 64% TOC removal after 2 h). [65] However, the obvious drawbacks of semiconductor coupling are the reduction in capacity of conduction band electrons of the coupled semiconductor, while the high oxidation capacity of valence band holes of TiO2 cannot be utilized, according to the Z-Scheme electron transfer model. [66]

3.1.1.3. LaFeO3 Perovskite Photocatalyst under Visible Irradiation Perovskite-type oxides of the general formula ABO3 are exciting materials for water treatment owing to their stability and high photocatalytic activity in water. LaFeO3, a typical ABO3-type perovskite oxide, has attracted considerable attention among researchers due to its broad potential for diverse applications such as catalytic oxidation, surface electronic states, and gas-sensitive characters. Additionally, LaFeO3 has been found to be photocatalytically active in the presence of visible light due to its unique optoelectronic properties and narrow band gap. [67] To date, there is a paucity of data on PPCP removal by LaFeO 3, despite many studies focussing on the visible light photocatalytic activity of LaFeO3. [68,69] Hu et al., [49] synthesized LaFeO3 and La2FeTiO6 by glucose sol–gel methods and evaluated their potential as photocatalysts by assessing p-chlorophenol degradation under visible light. La2FeTiO6 demonstrated 62.1% degradation of p-chlorophenol with 56.6% chemical oxygen demand (COD) removal after 5 h of visible-light irradiation, while LaFeO3 degraded 49.0% of pchlorophenol with 45.1% COD removal under identical conditions. The bigger surface area and smaller grain diameter contribute to the excellent photocatalytic activities of La2FeTiO6. Supported perovskite oxide, LaFeO3/SBA-15, was also investigated for photocatalytic efficacy in the oxidation of organic dyes using hydrogen peroxide as the oxidant. [70] Comparative tests on various organic dyes demonstrated after 60 min of reaction time that LaFeO3/SBA-15 is most efficient in degrading Rhodamine B (80% removal) relative to Methylene Blue (66% removal), Reactive Brilliant Red X-3B (42% removal),, and Direct Scarlet 4BS (42% removal). 3.1.2. Photolysis Processes Photolysis is the interaction of light (natural or artificial) with a molecule to induce photochemical reactions leading to the molecules degradation to intermediate products, and ultimately, to complete mineralization. [71] Physical disinfection of drinking water is largely dependent on ultraviolet radiation-driven photolysis, which has the advantage of minimizing the formation of disinfection by-products relative to chlorination. [72] However, a typical dose of 40 mJ/cm2 applied for drinking water treatment, while sufficient for microbial

90


disinfection, is largely ineffective for removal of most targeted organic compounds. For example, 17α-ethinylestradiol (EE2), diclofenac, sulfamethoxazole, and iopromide experienced only modest removals (0.4–27%) in dilute solutions of buffered water undergoing photolysis at pH 7.0. [73] Some macrolide antibiotics (i.e. such as clarithromycin, erythromycin and azithromycin) exposed to UV alone exhibited very low (4-7%) removal efficiencies, even with the considerable UV dose of 2768 mJ/cm2 for 15 min, which is a much higher dose than the 40-140 mJ/cm2 required for typical disinfection. [74]

O₂¯ O₂

e¯

CB

e¯ CB VB

OH¯ h⁺

VB

Coupled semiconductor

OH˙

TiO₂ Figure 5. Schematic illustrating the visible light photocatalytic mechanisms of TiO2 coupled with a semiconductor.

The efficiency of UV photolysis can be enhanced when irradiation is combined with hydrogen peroxide (H2O2), to form high reactive hydroxyl radicals (OH•), thus facilitating non-specific degradation of a broad suite of organic compounds.

The efficiency of UV/H2O2 co-treatment depends upon the absorbance spectrum of the PPCP or other organic compounds, along with the quantum yield of photolysis, the concentration of hydrogen peroxide employed, and the other dissolved and suspended constituents comprising the water matrix. Relative to UV treatment alone, the combination of UV/H2O2 yielded over a 90% removal efficiency for 39 of 41 evaluated pharmaceuticals at a UV dose of 923 mJ/cm2, the exceptions being norfloxacin and caffeine (69% and 67% removal efficiencies, respectively). As illustrated in Fig. 6, the addition of H2O2 substantially enhanced removal of pharmaceuticals during UV irradiation processes, to the extent that it may be possible to reduce UV energy inputs required for effective water treatment. [74] However, the presence of natural organic matter (NOM) in water will quench OH• radicals and decrease degradation efficiency [75]; however, Doll and Frimmel [76] suggested that


91

while NOM can act as a radical scavenger, it can also serve as a precursor of reactive species, which can enhance degradation of recalcitrant PPCPs such as carbamazepine by the production of photochemically-induced reactive species.

Figure 6. Removal efficiency of 41 pharmaceuticals during UV and UV/H2O2 water treatment processes. Reproduce permission from Ref. [74].

3.1.3. Ozonation Processes Ozone is a strong oxidant able to decompose into hydroxyl radicals, which are stronger oxidizing agents than ozone itself. Therefore ozonation incorporates the so-called indirect oxidation afforded by hydroxyl radicals, and their attack on certain functional groups of organic molecules through an electrophilic mechanism. [77] Unlike ozone, hydroxyl radicals are non-specific oxidants able to degrade a broader range of PPCPs and other organic micropollutants via radical addition, hydrogen abstraction or electron transfer mechanisms. [78] Hydroxyl radical production by ozone can be described as:

Ozonation is commonly employed in drinking water treatment systems as an alternative disinfection to free chlorine, mitigating odor and taste aesthetic complaints. Considerable research to date demonstrates the interest of researchers and water managers alike in the application of O3, O3/H2O2, and O3/UV for removal of PPCPs from water. Snyder et al. [79] conducted bench-scale investigations of PPCP ozonation in surface water and wastewater matrices demonstrating that the majority of investigated PPCPs were more than 90% removed at O3 exposures commonly used for disinfection. Notably, almost complete (>99%) removal of some fairly ubiquitous and recalcitrant PPCPs including carbamazepine, diclofenac, naproxen, sulfamethoxazole and trimethoprim demonstrated the effectiveness of ozone as shown in Fig. 7. However, other PPCPs and common micropollutants were relatively resistant

92


to ozonation, with iopromide, meprobamate, and tris-chloroethylphosphate (TCEP) exhibiting, less than 60% removal at the highest ozone concentrations investigated (2.5 mg/L). The addition of H2O2 to O3 to achieve advanced oxidation was of little benefit for those PPCPs resistant to ozonation, relative to O3 alone. O3/H2O2 marginally enhanced removal of dilantin, diazepam, DEET, iopromide and meprobamate, while advanced oxidation actually decreased removal efficacies of pentoxifylline, caffeine, and androstenedione relative to ozonation alone. [79] 100

Precent Removal (%)

80

60

40

20

Ac e An tam dr in os op te h ne en di Be o A nz tr ne o( az a) in py e C ar C ren ba af e m fei az ne ep in e D D T D DEE ia ze T D p ic am lo fe n D ac ila Ib nt up in Io rof M pr en ep om r o id ba e N ma O apr te xy ox b Su Pro en en lfa ge zon m ste e et ro ho n xa e z Te T ole st C Tr ost EP im er et on ho e pr im

0

H2O2 Dose 0mg L-1 H2O2 Dose 0.5mg L-1

Figure 7. PPCPs removal efficiencies using O3 and O3/H2O2 processes. Adapted from Ref. [79].

In O3/H2O2 system, the removal efficiency of COD and dissolved organic carbon (DOC) depends on the degree of OH• radical generation, which can be described as the following reaction: [80]

Degradation efficiency differs among different H2O2 to O3 molar ratios, with O3/H2O2 applied at a 0.5 molar ratio demonstrating the highest efficiency for removal of organics from wastewater effluent (Fig. 8), which suggests that the ratio of H2O2 injected with respect to ozone dosage is critical. [81]

93


Figure 8. Effect of O3 and O3/H2O2 treatment on COD and DOC removal from wastewater effluent. Adapted from Ref. [81].

UV-irradiation of O3 promotes the formation of OH• radicals via the rapid decomposition of O3. The observed reaction pathways for the UV/O3 process were similar to those of the ozonation process, but with faster kinetics, presumably induced by non-selective OH• radical reactions. [81]

Table 3. Electrical Energy and Operating Cost Required for O3/UV Processes. Adapted from Ref. [82] Applied process O3 dose (mg L-1) Total electrical energy(kWh m−3) No. of PPCPs removed ≥90%/No. of detected PPCPs Operating cost (Yen m−3)

O3 alone 2 4

6

O3/UV 21.5W 2 4

6

O3/UV 65W 2 4

6

0.03

0.06

0.09

0.37

0.4

0.43

1.06

1.09

1.12

24/37

32/37

35/37

15/35

23/35

31/35

24/38

34/38

35/38

0.5

0.9

1.4

5.6

6.0

6.5

15.9

16.3

16.8

Advanced oxidation (UV/O3) removal of PPCPs is superior to O3 alone due the additional degradation afforded by direct UV photodegradation, but most importantly, OH• radical oxidation. While advanced oxidation is more effective, when considering cost-effectiveness, ozonation alone removed more PPCPs per unit cost (Table 3). [82]

94


3.1.4. Electrocatalysis Electrocatalysis is catalytic action generated on an electrode surface achieved through changing the rate and selectivity of electrochemical reactions, and is closely connected with the adsorption of reacting, intermediate and end products. [83] The main difference between electrocatalytic and heterogeneous catalytic reactions is the direct anodic oxidation of electrocatalysis where the pollutants are adsorbed on the anode surface and destroyed by the anodic electron transfer reaction. 3.1.4.1. Photoelectrocatalytic (PEC) with TiO2 Nanofilm Modified Electrode To prevent the fast recombination of the photo-generated electrons/holes (e−/h+) and improve the photocatalytic efficiency of TiO2 films, the photoelectrocatalytic (PEC) technique was devised by applying an external bias potential onto TiO2 film-coated electrodes, combining both electrochemical and photocatalytic technologies. [84] The potential of a PEC oxidation process for the treatment of carbamazepine was evaluated using TiO2 photo-anodes prepared by pulsed laser deposition (PLD). [85] The PLD TiO2 coatings were found to be of anatase structure consisting of nanocrystallites of approximately 15 nm diameter. The contribution of treatment time and pollutant concentration accounts for 70.6% and 23.3% of carbamazepine PEC degradation, respectively, whereas the contributions of cathode material and current intensity were 2.4% and 0.75%, respectively. The PEC process applied under optimal conditions is able to oxidize 73.5% ± 2.8% of carbamazepine and ensure partial mineralization of 21.2% ± 7.7%. The TiO2 thin film can also be deposited onto surfaces via the liquid phase deposition (LPD) process. [86] In evaluating degradation of the model compound 4-aminoantipyrine, PEC processes exhibited a much higher efficiency of 62.1% after 120 min treatment, while electrolysis, direct photolysis, photocatalysis were 3%, 38% and 50.8% efficient, respectively. Recently, magnetic loading has been demonstrated as a novel means of preparing TiO2 electrodes, using the magnetic force between an external magnet and TiO2/SiO2/Fe3O4 (TSF) nanoparticles to facilitate the desired adherence of catalyst to the electrode, which has been successfully applied in the degradation of diclofenac. [87] 3.1.4.2. PEC with TiO2 Nanotube Arrays (TNAs) Electrode In a typical TiO2-based PEC process, there are two key factors affecting overall treatment efficiency; the TiO2 photocatalyst and the reactor configuration. Recently, highly ordered titania nanotube arrays (TNAs) grown by the electrochemical anodisation of titanium foils have attracted considerable attention due to their unique microstructures. [88,89] Bai et al., [90] developed a novel thin-layer PEC reactor with a double-faced TNA electrode and successfully applied it for the degradation of tetracycline. The titania nanotubes are highly ordered and well aligned (Figure 9), with diameters of ∼100 nm, wall thicknesses of ∼10 nm and lengths of ∼420 nm. The efficiency of organic compound degradation by the new PEC with TNA reactor was proven much higher than conventional PEC processes under similar treatment conditions. With UV illumination of 5 mW cm-2, the PEC with TNA tetracycline removal rate reached 54.8-96.4% when using 20–120 mg L−1 tetracycline in wastewater, higher than that achieved by conventional PEC reactors (about 14.6-80.4%) within a 1 h timeframe. Similar to pristine crystalline TNAs, cathodised dark TNAs also have proven excellent photocatalytic performance. [91] The cathodised dark TNAs were obtained by rapid


95

cathodisation for 30 s, which resulted in significantly enhanced optical absorbance, covering the full spectrum of visible light. The degradation of four PPCPs (phenol, ibuprofen, carbamazepine and caffeine) showed that the dark crystalline TNAs have a 107-131% enhancement in degradation kinetics for the targeted organic contaminants relative to the pristine crystalline TNAs (annealed in air).

3.2. Filtration Processes Filtration in water treatment usually describes some mechanical or biological processes through which undesirable constituents are removed by physical hinderance or absorption into a biological film grown on (or in) the filter medium.

3.2.1. Sand Filtration Sand filtration is a typical treatment of secondary waste water effluent used to remove suspended solids and persistent turbidity after clarification. It is possible to remove a portion of dissolved PPCPs associated with the retained solids, but this contribution is relatively small. Considering the removal efficiencies of 24 different PPCPs during sand filtration, the hydrophobicity of these organic compounds is likely to be a dominant factor governing their adherence to colloidal particles, and hence filterability. [92] This is consistent with the survey conducted by Oulton et al., [93] where the highest removal efficiencies were obtained for hydrophobic compounds which are hydrophobic enough to partition from water to sludge phases, making these higher log Kow compounds more bioavailable to sludge microorganisms, and hence more susceptible to degradation during conventional activated sludge treatments with adequate hydraulic and solids retention times.

3.2.2. Membrane Filtration Membrane filtration has been used for removal of pathogens, micropollutants and salts in drinking water treatment over the past several decades, but in recent years, novel membrane technologies have also attracted interest for removing PPCPs from wastewater. Membranes can be divided into four classes according to their molecular weight cut-offs (MWCO): microfiltration (MF), ultrafiltration (UF), nanofiltration (NF) and reverse osmosis (RO). Micro- and ultrafiltration are usually installed in low hydraulic pressure contexts as pretreatment devices, while nanofiltration and reverse osmosis are often used with high pressures to facilitate micropollutant removal. Based on physical dimensions of most PPCPs, only NF and RO membranes would be considered suitable for PPCP removal based purely on size exclusion mechanisms, and negating other factors. Yoon et al., [94] investigated removal efficiencies for 52 PPCPs varying in physicochemical properties such as size, hydrophobicity, and polarity using NF and UF membranes in a dead-end stirred-cell filtration system. Experiments were performed at environmentally relevant initial PPCP concentrations typically ranging from 2 to 0.1 μm 2-100 nm

Molecular mass >5000 KDa 5-5000 kDa

Microfiltration Ultrafiltration