Khum Gurung A4.pdf - LUTPub

WASTEW

ATER AND ITS VIABILITY OF IN

TEGRATING ADVAN

CED OXIDATION PROCESSES

Khum Gurung

MEMBRANE BIOREACTOR FOR THE REMOVAL OF EMERGING CONTAMINANTS FROM

MUNICIPAL WASTEWATER AND ITS VIABILITY OF INTEGRATING ADVANCED

OXIDATION PROCESSES

Khum Gurung

ACTA UNIVERSITATIS LAPPEENRANTAENSIS 866

Khum Gurung

MEMBRANE BIOREACTOR FOR THE REMOVALOF EMERGING CONTAMINANTS FROMMUNICIPAL WASTEWATER AND ITSVIABILITY OF INTEGRATING ADVANCEDOXIDATION PROCESSES

Acta Universitatis Lappeenrantaensis 866

Dissertation for the degree of Doctor of Science (Technology) to be presented with due permission for public examination and criticism in the Auditorium of Mikkeli University Consortium (MUC), Mikkeli, Finland on the 1st of October 2019, at noon.

Supervisors Professor Mika Sillanpää

LUT School of Engineering Science

Lappeenranta-Lahti University of Technology LUT

Finland

Assistant Professor Mohamed Chaker Ncibi

LUT School of Engineering Science


Finland

Reviewers Professor Anastasios I. Zouboulis

Department of Chemistry

Aristotle University of Thessaloniki (AUTh)

Greece

Professor How Yong Ng

Department of Civil and Environmental Engineering

National University of Singapore (NUS)

Singapore

Opponent Professor of Practice Anna Mikola

Department of Built Environment

School of Engineering

Aalto University

Finland

ISBN 978-952-335-410-4

ISBN 978-952-335-411-1 (PDF)

ISSN-L 1456-4491

ISSN 1456-4491


LUT University Press 2019

Abstract

Khum Gurung

Membrane bioreactor for the removal of emerging contaminants from

municipal wastewater and its viability of integrating advanced

oxidation processes

Lappeenranta 2019

86 pages

Acta Universitatis Lappeenrantaensis 866

Diss. Lappeenranta-Lahti University of Technology LUT

ISBN 978-952-335-410-4, ISBN 978-952-335-411-1 (PDF), ISSN-L 1456-4491, ISSN

1456-4491

Water scarcity is undoubtedly a global concern due to exacerbated population growth,

economic development, climate change, and rapid urbanization. This has necessitated the

need for establishing more sustainable and well-functioning natural ecosystems by

transforming current water industries based on smart technologies and management

approaches. Municipal wastewater reclamation is one of the promising approaches to

relieve growing pressure on global water resources. However, adequate and efficient

treatment is always essential to achieve high quality reclaimed water, ensuring no

potential risks to human health and aquatic environment due to the presence of emerging

contaminants (ECs), for which legislations are being stricter with time. Since

conventional wastewater treatment plants are not designed to remove ECs, the capabilities

of advanced wastewater treatment technologies, for enhancing the removal of ECs before

releasing the treated effluent into the environment or reclamation, such as membrane

bioreactor (MBR) and its integration with advanced oxidation processes, are emerging

topics of research worldwide.

This dissertation presents a comprehensive practical study on the applicability of the

MBR system for treating municipal wastewater in terms of its performance on the

removal of diverse ECs, such as pharmaceutically active compounds, steroid hormones,

and endocrine disrupting compounds, under different operating conditions, including

Nordic cold environment and varying solid retention times. Additionally, a potential

viability of integrating MBR with two emerging advanced oxidation processes (AOP)

technologies, such as electrochemical oxidation (ECO) and photocatalytic oxidation

(PCO), were studied in batch modes to further enhancing the removal of ECs and to

produce high quality reclaimed water.

A significant membrane fouling, accompanying with about 75% permeability drop, was

observed when process temperature in MBR was below 10 °C, indicating high

deterioration of membrane performance. However, relatively consistent removal

efficiency of human enteric viruses, such as norovirus GI, norovirus GII, and adenovirus,

heavy metals, and selected ECs, was achieved Similarly, the removal and fate of 23

diverse ECs was studied at varying solid retention times (60 and 21 days). It was observed

that at long solid retention time, ECs removal majorly enhanced, while biopolymers

concentrations decreased. Moreover, diverse removal efficiencies of the selected ECs

were observed, which were explained based on their physicochemical properties and

other process operating parameters. The electrochemical oxidation of carbamazepine,

which is one of the highly recalcitrant EC, was studied in batch mode using Ti/Ta2O5-

SnO2 anode. The main operating parameters effecting the removal of carbamazepine,

including current density, initial substrate concentration, pH, and temperature, were

studied. A complete removal (>99.9 %) of carbamazepine, with concentration close to

environmental level (10 µg L-1), in real MBR effluent was observed when electrolyzed

under the optimized conditions (current density: 9 mA cm-2, pH: 6, and temperature: 30

°C) in synthetic solutions. Subsequently, the removal of ECs (carbamazepine and

diclofenac) was studied by heterogeneous photochemical oxidation using 5% Ag2O/P-25

photocatalyst under UV irradiation. The matrix effect, i.e., ECs in deionized water and

MBR effluent, was studied along with the catalyst dosage and initial ECs concentration.

The optimal removal of carbamazepine and diclofenac reached 89.1% and 93.5%,

respectively at catalyst dosage of 0.4 g L-1 in deionized water matrix, and further revealed

that optimal catalyst dosage for ECs removal in MBR effluent matrix increased by 1.5 to

2 fold to achieve the similar removal efficiency as deionized water matrix. High

reusability of 5% Ag2O/P-25 photocatalyst was observed for both ECs. Moreover, various

intermediates of carbamazepine generated during photochemical oxidation were analyzed

and identified, and a possible degradation pathway was proposed.

The current study has expanded the knowledge on the applicability and efficiency of

MBR operation in unique Nordic environment and integration possibilities of MBR with

selected innovative advanced oxidation processes. It has also revealed the need of many

other important related topics for further investigations.

Keywords: Membrane bioreactor, municipal wastewater treatment, emerging

contaminants, advanced oxidation processes, electrochemical oxidation, photochemical

oxidation

Acknowledgements

I have always been fascinated by science and technology. As the life goes on, my interests

and enthusiasm towards science have eventually led me to pursue PhD degree. Achieving

PhD degree has always been my academic dream, although it was not a piece of cake. I

should admit that it was full of challenges, hard work, and endurances. Now, I have

accomplished nearly four years of my PhD study and I have a great opportunity to finally

acknowledge many people who made this journey possible. In this occasion, I am feeling

overwhelmed and would like to convey my sincere gratitude to all of them.

First of all, I would like to express my deepest gratitude to my supervisor, Prof. Mika

Sillanpää for believing on me and providing me this opportunity to start and accomplish

my PhD study under his commendable guidance. I am also grateful to Assistant Prof.

Mohamed Chaker Ncibi for his valuable supports, suggestions and guidance throughout

my PhD journey.

I would like to thank Department of Green Chemistry (DGC), where most of my research

works for the doctoral dissertation was carried out, for facilitating incredible

infrastructures and resources.

I am thankful to Prof. Anastasios I. Zouboulis and Prof. How Yong Ng for reviewing this

dissertation and providing their valuable comments and suggestions that have enriched

the content of the dissertation further.

Many thanks to Professor of Practice Anna Mikola for agreeing to act as the opponent for

the public examination.

I would like to thank Business Finland Oy, which is the most prestigious public funding

agency for promoting high quality research in Finland, for financially supporting this

research work. The research project was granted with the title ‘‘Smart Effluents Project-

New generation wastewater treatment solutions for requirements of 2050’’ and project

decision number of 1043/31/2015. The project was started on October 2015 and ended

by September 2018.

My sincere thanks go to Kenkäveronniemi wastewater treatment plant, Mikkeli, Finland,

owned by Mikkeli water works (Mikkeli Vesilaitos), for being an active company partner

to this project, allow us to install and operate our pilot MBR plant within their premise,

and for providing technical supports during the project. I am grateful to Reijo Turkki

(Director of Mikkeli Vesilaitos), Anne Bergman, Risto Repo, Jani Koski, Päivi and all

other staffs of Kenkäveronniemi wastewater treatment plant, Mikkeli, for their

tremendous support on operating the MBR pilot plant and some laboratory works.

I would like to thank South-Eastern Finland University of Applied Sciences (XAMK) for

its incredible participation to Smart Effluents project. I would like to thank Dr. Heikki

Särkkä and Dr. Hanne Soininen for their continuous support on executing the Smart

Effluents Project steering group meetings. Moreover, especial thanks to all the Smart

Effluents Project consortium group members, Metsäsairila Oy, Suomen Ekolannoite Oy,

BioGTS Oy, and Mipro Oy, for their valuable comments and suggestions throughout the

project.

I would like to thank Dr. Marina Shestakova for helping me on establishing the

electrochemical study setup and guided me during the experimental period. I am thankful

to Dr. Bo Gao for her valuable suggestions on photocatalysis. I like to thank Dr. Deepika

Lakshmi Ramasamy for helping me on some analytical procedures.

I would like to thank all my senior and junior colleagues in the Department of Green

Chemistry for establishing a pleasant working environment with healthy discussions,

mutual support, and good scientific practices, especially Dr. Bhairavi, Dr. Varsha, Dr.

Sidra, Asst. Prof. Yuri, Mirka, Tam, Zhao, and Mahsa. All of you are amazing people.

Also, I like to thank Sanna Tomperi for helping me in many ways during the entire

journey.

Last but not the least, I am indebted to my wife and daughter, Jen and Nohyo, and all my

family members and relatives for being my strength, support and inspiration throughout

this PhD journey and sharing all the ups and downs in my life.

Khum Gurung

September 2019

Mikkeli, Finland

This doctoral dissertation is dedicated in memory of my late mother and

father.

Table of Contents

Abstract

Acknowledgements

Table of contents

List of publications ............................................................................................... 11

List of acronyms and symbols ............................................................................. 13

List of figures ........................................................................................................ 15

List of tables .......................................................................................................... 17

1 Introduction ....................................................................................................... 19

1.1 Problem statement and research motivations ............................................... 19

1.2 Research objectives ...................................................................................... 20

2 Municipal wastewater ....................................................................................... 21

2.1 Occurrence of emerging contaminants in municipal wastewater ................. 23

2.2 Environmental intimidations of ECs ............................................................ 26

2.3 Municipal wastewater treatment ................................................................... 26

2.4 EU legislation and removal of ECs from municipal wastewater.................. 27

3 State-of-the-art of MBRs: advanced municipal wastewater treatment ........ 29

3.1 Introduction to MBRs ................................................................................... 29

3.2 Membrane materials, configuration and operating principles of MBRs ...... 30

3.3 Membrane fouling: a major shortcoming in MBRs...................................... 32

3.4 Fate of ECs in MBRs: application of mass balance ..................................... 34

3.5 Fate of human enteric viruses and heavy metals in MBRs........................... 37

4 Advanced oxidation processes for wastewater treatment .............................. 39

4.1 Electrochemical oxidation (ECO) process ................................................... 39

4.1.1 Active anodes ........................................................................................ 41

4.1.2 Non-active anodes ................................................................................. 42

4.2 Photocatalytic oxidation (PCO) .................................................................... 43

5 Integration potentials of MBR with AOPs ...................................................... 47

6 Materials and methods ...................................................................................... 51

6.2 Assessing the performance of MBR unit in different operating conditions . 51

6.1.1 The Pilot-scale MBR set-up .................................................................. 51

6.2 Assessing integration approaches to MBR ................................................... 52

6.2.1 Lab-scale set-up for batch ECO unit ..................................................... 52

6.2.2 Lab-scale set-up for PCO process ......................................................... 53

6.2.3 Analytical procedures ............................................................................ 54

7 Results and discussion ....................................................................................... 57

7.1 Assessing the performance of MBR at Nordic cold conditions (Paper I) .... 57

7.2 Assessing the performance of MBR under varying solid retention times- fate and

removal of ECs (Paper II) ................................................................................... 61

7.3 Assessing the AOPs as integration alternatives to MBR to enhance the removal of

ECs. .................................................................................................................... 63

7.3.1 Study of ECO process for carbamazepine removal using Ti/Ta2O5-SnO2

electrode (Paper III) ........................................................................................ 64

7.3.2 Study of PCO process for PhACs removal using Ag2O/P-25 photocatalyst

(Paper IV) ....................................................................................................... 66

8 Conclusions and further research .................................................................... 73

References.............................................................................................................. 75

11

List of publications

I. Gurung, K., Ncibi, M.C., Sillanpää, M., 2017. Assessing membrane fouling

and the performance of pilot-scale membrane bioreactor (MBR) to treat real

municipal wastewater during winter season in Nordic regions. Science of the

Total Environment. 579, 1289–1297.

II. Gurung, K., Ncibi, M.C., Sillanpää, M., 2019. Removal and fate of emerging

organic micropollutants (EOMs) in municipal wastewater by a pilot-scale

membrane bioreactor (MBR) treatment under varying solid retention times.

Science of the Total Environment. 667, 671–680.

III. Gurung, K., Ncibi, M.C., Shestakova, M., Sillanpää, M., 2018. Removal of

carbamazepine from MBR effluent by electrochemical oxidation (EO) using

a Ti/Ta2O5-SnO2 electrode. Applied Catalysis B Environment. 221, 329–338.

IV. Gurung, K., Ncibi, M.C., Thangaraj, S.K., Jänis, J., Seyedsalehi, M.,

Sillanpää, M., 2019. Removal of pharmaceutically active compounds

(PhACs) from real membrane bioreactor (MBR) effluents by photocatalytic

degradation using composite Ag2O/P-25 photocatalyst. Separation and

Purification Technology. 215, 317–328.

Author’s contribution

Khum Gurung is the principal investigator, and corresponding author for all the

publications I-IV.

12 List of publications

Related publication

Gurung, K., Ncibi, M.C., Fontmorin, J.M., 2016. Incorporating Submerged MBR in

Conventional Activated Sludge Process for Municipal Wastewater Treatment: A

Feasibility and Performance Assessment. Journal of Membrane Science and

Technology 6, 1-10.

Other publications

Gurung, K., Tang, W.Z., Sillanpää, M., 2018. Unit Energy Consumption as Benchmark

to Select Energy Positive Retrofitting Strategies for Finnish Wastewater Treatment Plants

(WWTPs): a Case Study of Mikkeli WWTP. Environmental Processes. 5, 667–681.

Habib, R., Asif, M.B., Iftekhar, S., Khan, Z., Gurung, K., Srivastava, V., Sillanpää, M.,

2017. Influence of relaxation modes on membrane fouling in submerged membrane

bioreactor for domestic wastewater treatment. Chemosphere 181, 19–25.

Seyedsalehi, M., Paladino, O., Hodaifa, G., Sillanpää, M., Gurung, K., Sahafnia, M.,

Barzanouni, H., 2018. Performance evaluation of several sequencing batch biofilm

reactors with movable bed in treatment of linear alkyl benzene sulfonate in urban

wastewater. International Journal of Environmental Science and Technology.

13

List of acronyms and symbols

Acronyms

AdVs Adenovirus

AFM Atomic force microscope

AOPs Advanced oxidation processes

BDD Boron-doped diamond

BET Brunauer-Emmett-Teller

BOD5 Biological oxygen demand of 5-days

CAS Conventional activated sludge

COD Chemical oxygen demand

CST Capillary suction time

CV Cyclic voltammetry

DOC Dissolved organic carbon

DSA Dimensionally stable anodes

DW Deionized water

ECs Emerging contaminants

EDCs Endocrine disrupting compounds

EDGs Electron donating groups

EDTA Ethylenediaminetetraacetic acid

EDX Energy-dispersive X-ray spectroscope

EPS Extracellular polymeric substance

EWGs Electron withdrawing groups

FOG Fat-oil-grease

FS Flat sheet

FT-ICR Fourier-transform ion cyclotron resonance mass spectrometry

FTIR Fourier-transform infrared spectroscopy

HER Hydrogen evolution reaction

HF Hollow fibre

LOD Limit of detection

LOQ Limit of quantification

MBR Membrane bioreactor

MBBR Moving-bed biofilm reactor

MLSS Mixed liquor suspended solid

MLVSS Mixed liquor volatile suspended solid

MMO Mixed metal oxide

MWCO Molecular weight cut-off

NoV GI Norovirus genome group I

NoV GII Norovirus genome group II

OERs Oxygen-evolution reactions

PCPs Personal care products

14 List of acronyms and symbols

PE Polyethylene

PES Polyethylsulphone

PhACs Pharmaceutically active compounds

PP Polypropylene

PVDF Polyvinylidene difluoride

P-25 Aeroxide® Titania nanoparticles

RME Real MBR effluent

SADm Specific aeration demand per membrane surface area

SCE Saturated Calomel electrode

SEM Scanning electron microscope

SHE Standard Hydrogen electrode

SMP Soluble microbial product

SRT Solid retention time

SVI Sludge volume index

TMP Trans-membrane pressure

TP Total phosphorus

TSS Total suspended solid

WWTP Wastewater treatment plant

Symbols

Ti Titanium

ℎ Planck constant (J.s)

𝜗 Frequency (Hz)

m/z Mass-to-charge ratio

Rt Total membrane filtration resistance (m-1)

𝜇 Dynamic viscosity of water (Pa.s)

𝐽 Membrane flux (Lm-2h-1)

log D Distribution coefficient

log Kd Sludge-water distribution coefficient

E° Standard redox potential (eV)

Rads Adsorbed organic pollutants at anode surface

Pads Oxidized adsorbed organic pollutants at anode surface

∆E Band gap energy (eV)

ecb- Excited electrons to conductance band

hvb+ Photogenerated holes at valence band

15

List of figures

Figure 1: MBR process configurations, (a) Submerged and (b) side-stream ......... 31

Figure 2: Different MBR membrane modules: (a) Tubular, (b) Hollow-fiber, (c) flat-

sheet (Reprinted with permission from (Bérubé, 2010)). ....................................... 31

Figure 3: Temporal and spatial development of membrane fouling in MBRs (Modified

from (Meng et al., 2017)). ...................................................................................... 34

Figure 4: Electrochemical oxidation scheme of organic pollutants on the metal oxide

anode surface (Modified from (Comninellis, 1994)). ............................................ 41

Figure 5: Decision tree for potential integration approaches of MBR and AOP

(Modified from (López et al., 2010)). .................................................................... 48

Figure 6: Schematic diagram of the pilot MBR unit at Kenkäveronniemi WWTP,

Mikkeli, Finland. .................................................................................................... 51

Figure 7: Experimental set-up for electrochemical oxidation. ............................... 53

Figure 8: Experimental set-up for photocatalytic oxidation ................................... 54

Figure 9: Membrane fouling behavior of MBR system under cold weather conditions

and adopted cleaning measures. ............................................................................. 58

Figure 10: Occurrences of viruses in wastewater stream under cold water conditions.

................................................................................................................................ 59

Figure 11: Removal efficiencies of selected ECs in MBR treatment under low-

temperature periods. ............................................................................................... 60

Figure 12: Fate of the selected ECs during MBR treatment at two different SRTs.63

Figure 13: Effect of operating conditions (a) applied current densities; (b) initial

carbamazepine concentrations; (c) initial solution pH; and (d) temperature on the

electrochemical oxidation of carbamazepine. ........................................................ 66

Figure 14: FTIR spectrum (a) and plots of (F (R∞) h𝑣)1/2 vs (h𝑣) for the approximation

of the optical band gap (b) of 5% Ag2O/P-25. ....................................................... 67

Figure 15: Effect of catalyst dosage in different water matrices on photocatalytic

removal of (a & d) carbamazepine in DW and RME; (b & e) diclofenac in DW and

RME; and effect of initial PhACs concentration (c) carbamazepine in RME; (f)

diclofenac in RME. ................................................................................................. 69

Figure 16. Cyclic reusability of 5% Ag2O/P-25 photocatalyst during the photocatalytic

oxidation of (a) carbamazepine and (b) diclofenac. ............................................... 70

Figure 17: Photogenerated intermediates and proposed photodegradation pathway of

carbamazepine by 5% Ag2O/P-25 photocatalyst under UV irradiation. ................ 72

16 List of figures

17

List of tables

Table 1. Characteristics of Municipal wastewater and their sources (Metcalf & Eddy

Inc., 2003; U.S. EPA, 1997) ................................................................................... 22

Table 2. Physicochemical properties of commonly identified ECs in municipal

wastewater. ............................................................................................................. 24

Table 3: Different operating conditions of MBR pilot plant depending on purpose of

studies. .................................................................................................................... 52

18 List of tables

19

1 Introduction

1.1 Problem statement and research motivations

The global water demand has been increasing constantly due to exacerbated population

growth, economic development, climatic change, intensive agricultural practices,

changing consumption patterns and urbanization. Undoubtedly, combating water scarcity

is a global issue (WHO, 2015a; UN GEMS, 2006). As a result, many of the natural water

systems that preserve balanced and thriving ecosystems, have become severely stressed.

The UN report in 2015 revealed that the global clean water scarcity will reach 40% by

2030 (WWAP, 2015). Indeed, water scarcity is not about having too little water, but it is

the scarcity of having good water management practices, which severely endangers

billions of people and the environment (World Water Council, 2019). Water scarcity

afflicted many countries globally, leading poor access to clean potable water and

sanitation (Cosgrove et al., 2014). Globally, about 1.1 billion people lack access to safe

drinking water and 2.7 billion experience water crisis at least a month per year (WHO,

2015a; WWF, 2019). It is obvious that aquatic environments can no longer be perceived

as constant water supply source, and rather these environments are complex matrices that

need careful practices to ensure sustainable and well-functioning ecosystems in the future

(UN GEMS, 2006). This enormous challenge requires a transformation of water

industries based on smart integrated innovative technologies and effective management

outlooks that protects and reuses water for many purposes (Cosgrove et al., 2014).

In these circumstances, municipal wastewater reclamation is one of the promising

approaches to relieve increasing pressure on natural water resources. Nevertheless,

adequate treatment is always required to achieve high quality reclaimed water for reuse

and to ensure no potential threats to human health and aquatic environments (Ma et al.,

2013). Especially, an active evaluation of possible risks of emerging contaminants (ECs)

in the treated effluent of wastewater treatment streams has gained a great attention among

the scientific community (Luo et al., 2014; Nassiri Koopaei and Abdollahi, 2017; Yang

et al., 2017). Moreover, according to EU framework, ECs must be adequately removed

before discharging wastewater or treated effluent to the aquatic environment due to their

persistency, risk of bioaccumulation, and acute or chronic toxicity. Over the decades, EU

water framework legislations are being more stringent day-by-day regarding the removal

of many ECs found in the aquatic environment.

Conventional municipal wastewater treatment processes are mainly aimed for treating

organic matters, suspended solids and nutrients, hence their poor efficiency to treat ECs.

To ensure compliance with the stringent discharge limits for quality water reclamation,

interests in the ability of membrane bioreactors (MBRs) in removing ECs from municipal

wastewater has increased in recent decades (Judd, 2008). MBRs have emerged in the

wastewater treatment sector as one of the most reliable alternatives to the conventional

activated sludge (CAS) processes (Hai et al., 2014). Nevertheless, many ECs are not

efficiently attenuated neither in CAS nor MBRs due to their refractory nature (Xiao et al.,

20 1 Introduction

2019). Therefore, it is essential to develop a treatment system, in which, MBRs can be

integrated with other advanced oxidation processes (AOPs), to enhance the removal of

ECs before discharging the treated effluent is into the environment or reusing it.

Nevertheless, the number of research studies on integrating MBR with AOPs are rather

limited.

1.2 Research objectives

The current thesis investigates the feasibility of advanced wastewater treatment

techniques to treat municipal wastewater containing traces of many ECs that are of

emerging concern. The main research objectives can be divided into two principal

aspects: (1) to assess the functional operation and performance of a conventional MBR

unit in a unique Nordic environment (esp. low water temperature conditions) for treating

real municipal wastewater, in which the pilot MBR unit was operated as an advanced

secondary treatment process and its treatment efficiency, mainly the removal of ECs

under varying operating conditions, such as varying temperatures and solid retention

times (SRT) was assessed; (2) to investigate the possible integration of alternative

advanced water treatment processes for enhancing the ECs removal efficiency of the

conventional MBR with selected AOPs, such as electrochemical oxidation (ECO) and

photocatalytic oxidation (PCO). The main objectives are summarized as follow:

- To study the performance of a conventional MBR to treat the real municipal

wastewater in Nordic winter environment and under varying SRTs (Paper I and II)

- To study the removal of PhACs from MBR treated effluent by utilizing

electrochemical oxidation on newly developed Ti/Ta2O5-SnO2 anode (Paper III)

- To study the removal of PhACs from MBR treated effluent by photocatalytic

oxidation with Ag2O/P-25 photocatalyst (Paper IV)

21

2 Municipal wastewater

Municipal wastewater comprises of either sanitary sewage from households and

commercial buildings or the mixture of sanitary sewage with industrial wastewater,

stormwater and agricultural run-off (Metcalf & Eddy Inc., 2003). Combined sewers are

designed to collect both sanitary sewage and stormwater run-off, while only sanitary

sewage are allowed under separate sewers (US EPA, 2015a). In general, it is reported that

about 70 - 130% of municipal fresh water consumption becomes wastewater (Qasim,

2017).

The characteristics of municipal wastewater varies from site to site depending on land

uses, water consumption patterns, discharges of commercial and industrial wastewater,

extent of separation between stormwater and sanitary sewage, and inclusive of both

diurnal and seasonal fluctuations (Hung et al., 2012). The major constituent of municipal

wastewater is approximately 99.9% of water and the rest fractions includes organics,

inorganics, total suspended solids, together with traces of pharmaceuticals and other

hazardous materials and microorganism (Sperling, 2007). It is difficult to achieve any

effective wastewater management without the detailed understanding of wastewater

characteristics. However, since water-consumption patterns of individuals can rise

haphazardly, the composition and strength of wastewater generated from diverse

standpoint sources fluctuates with time, thus impossible to quantify these fluctuations

accurately. According to US EPA, municipal wastewater can be characterized in terms of

physical, chemical and biological constituents (U.S. EPA, 1997). Table 1 shows the

characteristics of typical municipal wastewater and possible sources.

22 2 Municipal wastewater

Table 1. Characteristics of Municipal wastewater and their sources (Metcalf & Eddy

Inc., 2003; U.S. EPA, 1997)

Characteristics Sources

a) Physical characteristics

Color and Odor

Solids

Turbidity

Temperature

Electrical conductivity

Particle size distribution

Domestic and industrial wastewater, natural

decay of organic matters

Domestic and industrial wastewater, soil

erosion, inflow/infiltration

Domestic and industrial wastewater



Domestic and industrial wastewater, soil erosion

b) Chemical characteristics

Organics:

BOD5 /COD/TOC,

Carbohydrates and proteins,

Fats, oils and grease (FOG)

Refractory organics

(Surfactants, pesticides,

phenols, pharmaceuticals,

PFCs etc.)

Inorganics:

Alkalinity, chlorides

Nitrogen

Phosphorus

pH

Heavy metals

Gases:

Methane, hydrogen sulfides,

oxygen, nitrous oxides

Domestic commercial and industrial wastewater

Domestic, industrial and agricultural

wastewater

Domestic water supply, domestic wastes,

groundwater infiltration, water softener

Domestic and agricultural wastewater

Domestic, commercial and industrial

wastewater, storm water run-off

Domestic, commercial and industrial

wastewater

Industrial wastewater

Domestic water supply, decomposition of

domestic wastewater, surface water infiltration

c) Biological characteristics

Bacteria, protozoa, algae,

helminths

Pathogens (coliform, viruses)

Wastewater treatment plants

Domestic wastewater

23

2.1 Occurrence of emerging contaminants in municipal wastewater

ECs are new chemicals with no associated regulatory standards, and whose effects on

environment and human health are still largely unknown (Deblonde et al., 2011). Over

the past few years, the occurrence of ECs, such as pharmaceutically active compounds

(PhACs), personal care products (PCPs), stimulants, pesticides, steroid hormones,

endocrine disrupting compounds (EDCs) and UV filters, has attained a great concern due

to their ubiquitous presence in various compartments of the aquatic environment (e.g.,

water, sediment and biota) and their potential ecological threats (Barceló and Petrovic,

2008; Deblonde et al., 2011; Tran et al., 2018; Verlicchi and Zambello, 2015). ECs not

only pose threat to aquatic life, but also their constant accumulation in aquatic

environment could evolve antibiotic-resistant microbial strains (Ganiyu et al., 2015).

Currently, more than 700 ECs and their active metabolites are identified in the EU aquatic

environment (UN FAO, 2018). A large number of these ECs have been identified at

source, in treated effluent, and wasted biomass of sewage treatment plants (Clara et al.,

2005b; Trinh et al., 2016; Vieno et al., 2005). The significant spatial and temporal

variations in the concentrations of ECs have been reported by many researchers due to

number of factors, such as rate of production, specific sales and consumption practices,

metabolism, and specific water consumptions (Luo et al., 2014). The physicochemical

properties of some commonly identified ECs in municipal wastewater, most of which

have used as model ECs in this thesis, are summarized in Table 2.

Many ECs are present at extremely low concentrations (µg L-1 to ng L-1) in the

environment, which make their identification and assessment even more challenging

(Rodriguez-Narvaez et al., 2017; US EPA, 2015b). Most of the ECs found in wastewater

and aquatic environment are of anthropogenic origin, which are generally introduced

from various routes, including direct or treated discharge of wastewater from municipal,

hospital, and industrial wastewater treatment plants (WWTPs), surface run-off from

agricultural and veterinary sectors, landfill leachates, and sewer leakage or overflow (Luo

et al., 2014; Tran et al., 2018). Of the many sources, occurrence of ECs at WWTPs are of

main concern since they are frequently detected in both influent and treated effluent of

WWTPs, and subsequently discharge ECs into the environment (Behera et al., 2011). The

category and concentrations of ECs generally found in WWTPs depend on the

socioeconomic composition of the population generating the municipal wastewater (Tran

et al., 2018).


Table 2. Physicochemical properties of some commonly identified ECs in municipal

wastewater.

Category Compound

Molecu

lar

weight

(g.moL-

1)

log D

(at pH

7)

Dissociation

constant

(pKa)

Henry’s Law

constant

(at 25oC)

(atm.m3.moL-1)

1. PhACs

Analge

sics

and

anti-

inflam

matori

es

Paracetamol 151.16 0.47 9.86;1.72 1.92 x 10-11

Diclofenac 296.15 1.77 4.18; -2.26 2.69x10-11

Ibuprofen 206.28 0.94 4.41 5.54 x 10-10

Ketoprofen 254.28 0.19 4.23 1.92 x 10-13

Naproxen 230.26 0.73 4.84 6.08 x 10-12

Antiepilepti

cs

Carbamazepine 236.27 1.89 13.94; -0.49 9.41 x 10-12

Antibiotics Sulfamethoxazole 253.28 -0.22 5.81; 1.39 2.24 x 10-13

Trimethoprim 290.32 0.27 7.04 1.37 x 10-14

Tetracycline 444.44 -2.06 4.50; 11.02 7.35 x 10-29

Ciprofloxacin 331.35 -0.33 6.43; 8.68 2.10 x 10-17

Tylosin 916.11 0.15 13.06; 7.39 n.a.

Ofloxacin 361.37 -0.20 5.19; 7.37 1.76 x 10-16

Norfloxacin 319.33 -0.65 0.16; 8.68 4.13 x 10-16

Oxytetracycline 460.43 -2.25 4.50; 10.80 1.33 x 10-32

Metronidazole 171.15 -0.14 14.44;2.58 2.01 x 10-12

Doxycycline 444.44 -0.92 4.50; 10.84 3.20 x 10-26

Lipid

regulators

Clofibric acid 214.65 -1.06 3.18 2.91 x 10-11

Gemfibrozil 250.34 2.07 4.75 1.92 x 10-11

Bezafibrate 361.82 -0.93 3.29; -2.06 2.12 x 10-19

Simvastatin 418.57 4.72 13.49 4.93 x 10-16

β-Blocker Bisoprolol 325.45 -0.54 9.42 4.54 x 10-15

Atenolol 266.34 -2.09 13.88;9.43 1.34 x 10-17

Metoprolol 267.37 -0.81 13.89; 9.43 1.59 x 10-13

Propranolol 259.35 0.45 13.84; 9.50 6.27 x 10-14

Sotalol 272.36 -2.01 8.28; 9.31 2.63 x 10-14

Diuretics Furosemide 330.74 -0.79 3.04; -2.49 2.57 x 10-19

Hydrochlorothiazide 297.73 -0.03 8.95; -4.08 5.61 × 10-17

Enalapril 376.45 -0.14 3.15; 5.43 1.48 × 10-20

Psychostim

ulants

Caffeine 194.19 -0.63 0.52 1.63 × 10-12

25

Category Compound

Molecu

lar

weight

(g.moL-

1)

log D

(at pH

7)

Dissociation

constant

(pKa)

Henry’s Law

constant

(at 25oC)

(atm.m3.moL-1)

Contrast

media

Iopamidol 777.09 -2.54 10.87; -2.85 6.67 ×10-30

Iopromide 791.12 -2.66 10.62; -2.60 3.29 × 10-35

Antidepress

ants

Fluoxetine 309.33 1.15 10.05 2.25 × 10-11

Paroxetine 329.37 1.16 9.68 3.53 × 10-12

Anticoagula

nts

Warfarin 308.33 0.67 4.50 1.42 × 10-15

Immunosup

pressives

Methylprednisolone 374.48 2.17 12.46 1.88 × 10-17

Methotrexate 454.45 -4.98 3.47; 5.56 n.a.

Beclomethasone 408.92 2.44 12.17 2.22 × 10-18

Cyclophosphamide 261.08 0.73 2.84 9.46 × 10-10

Ifosfamide 261.08 0.76 1.44 9.46 × 10-10

Hydrocortisone 362.47 1.76 12.47 1.74 × 10-17

Bronchodila

tors

Terbutaline 225.28 -1.77 9.11; 9.65 3.59 × 10-14

Clenbuterol 277.19 0.24 13.29; 9.51 9.12 × 10-13

Salbutamol 239.31 -1.77 9.99; 9.62 8.75 × 10-15

Anti-

parasitics

Flubendazole 313.29 3.08 10.66; 4.45 n.a.

Fenbendazole 299.35 2.35 10.80 n.a.

Ketoconazole 531.43 3.80 6.88 2.18 × 10-24

Contracepti

c

Norethindrone 298.43 2.86 13.09 9.5 × 10-12

2. EDCs

Plasticizer Bisphenol A 228.29 3.64 10.29 1.92 × 10-11

3. Steroid

Hormon

es

17β-Estradiol 272.38 4.15 10.27 1.17 × 10-9

17α-Ethynylestradiol 296.41 4.11 10.24 3.74 × 10-10

Estriol 288.39 2.53 10.25 1.75 × 10-11

Estrone 270.37 3.62 10.25 9.61 × 10-10

Progesterone 314.46 3.83 n.a. 1.51 × 10-09

Testosterone 288.43 3.18 15.06 4.89 × 10-11

4. Industri

al

Surfacta

nts

Perfluoro-octanoic

acid (PFOA)

414.07 2.69 0.50 1.13 × 10-05

Perfluorooctane

sulfonic acid (PFOS)

212.28 4.49 <1.0 4.1 × 10-04

n.a.- data not available

(Data are extracted from SciFinder database

https://scifinder.cas.org/scifinder/view/scifinder/scifinderExplore.jsf)


2.2 Environmental intimidations of ECs

The major concern on environmental risk of ECs is evidenced only if they may adversely

pose impact on aquatic life and human health. There are three major risk factors related

to the ECs presence in the aquatic environment viz. persistency, bioaccumulation, and

toxicity (Ebele et al., 2017). Persistency is mainly defined by the physicochemical

properties of ECs, due to which many of them are difficult to remove by conventional

means of treatment. Indeed, their incomplete removal could generate many active

metabolites that could pose potential risks to aquatic life. Even though not all the detected

ECs in environmental matrix are persistent, their continuous replenishment and release to

the environment referred them as ‘‘pseudo-persistent’’ compounds, which pose higher

persistence even when acted by environmental processes, such as photodegradation,

biodegradation, and sorption into sediments (Bu et al., 2016; Ebele et al., 2017). Another

risk is the bioaccumulation of ECs in aquatic biota. Generally, ECs are designed and used

for specific target, but since many of ECs and their metabolites are biologically active,

they can adversely affect non-targeted aquatic organisms (Rodríguez-Mozaz et al., 2016).

Coogan et al reported an abundant bioaccumulation of ECs and their metabolites in algae

samples of WWTPs (Coogan et al., 2007). On the other hand, toxicity is one of the major

concern of ECs, which can emerge due to complex mixtures of ECs that leads to high

synergetic effects. This means that the presence of ECs even at low concentrations could

pose significant toxicity to aquatic microorganisms. Another risk related to the ECs in the

environment is the possible emergence and spread of antibiotic resistant strains in natural

bacterial populations, which is a very emerging topic in recent years. Antibiotic resistance

is aggravated by the misuse and overuse of antibiotic that has continuously undermined

many advances in health and medicine (WHO, 2015b).

2.3 Municipal wastewater treatment

The main objectives of wastewater treatment are to separate, convert, and eliminate

objectionable and hazardous contaminants and infectious pathogenic organisms via a

combination of physical, chemical and biological processes (EU, 1991). Wastewater

produced from municipalities and communities must ultimately be returned to receiving

waters or to land or reuse after the treatment. Wastewater treatment is a multi-step

process, which typically involves primary, secondary and tertiary treatments, according

to the treatment objectives, level of treatment efficiency, and environmental threats on

the receiving water bodies (Sperling, 2007). The municipal wastewater reclamation is an

important concept, in which the collected municipal wastewater is treated and reused for

recharging fresh water sources and recovering value-added resources and energy (Metcalf

& Eddy Inc., 2003).

The primary treatment mostly includes the physical treatment of wastewater in order to

reduce to treatment load to the subsequent steps by selectively removing settleable

organic and inorganic materials and floating (scums) materials via sedimentation. This

27

process can reduce about 25 – 50% of BOD5, 50 – 70% TSS and 65% FOG loads (Aziz

and Mojiri, 2014). However, advanced primary treatment or chemically enhanced

primary treatment could be attained by the addition of coagulants (aluminium sulphate,

ferric chloride or other polymers) to remove mainly phosphorus by precipitation

(Sperling, 2007).

The secondary treatment is intended for the removal of organic pollutants (either

suspended, particulate, or dissolved) and nutrients (nitrogen and phosphorus) (Punmia et

al., 1998). Secondary treatment typically includes biological reactor, where pollutants are

removed by bacteria metabolism, and followed by the secondary sedimentation tank to

separate settleable solids and treated water. The biological treatment process can be

suspended growth, attached growth, and their various combinations. The suspended

growth system includes CAS process (with or without biological nitrogen and phosphorus

removal), extended aeration, and sequencing batch reactors, in which microorganisms

attached to the biomass are maintained in suspension. The biomass and microorganism

assays develop in support medium in attached growth system, including trickling filter,

rotating biological contractor, and submerged aerated biofilters. On the other hand,

combined suspended attached growth system involves processes with internal and up-

flow suspended packing for attached growth such as moving-bed biofilm reactor (MBBR)

and fluidized-bed bioreactors (Metcalf & Eddy Inc., 2003).

2.4 EU legislation and removal of ECs from municipal wastewater

The main objective of EU framework is to protect natural water environment in Europe

and to replenish every river, lake, groundwater, wetland and other surrounding water

bodies (Buttiglieri and Knepper, 2008). Due to persistency, risk of bioaccumulation in

biota, and high synergetic toxicity, ECs must be adequately removed before releasing

with treated effluent to the environment. The EU Water Framework Directive

2000/06/CE announced a list of 33 priority substances or group of hazardous substances,

including heavy metals, pesticides, and polyaromatic hydrocarbon (EU 2000). In

December 2008, directive 2008/105/EC was introduced amending to directive

2000/06/CE for progressively reducing contamination of priority substances with

environmental quality standards (EQS) (EU 2008). In 2013, the directive 2008/105/EC

was further amended to 2013/60/EC by identifying additional 12 substances and a first

watch list of chemicals, including diclofenac, 17- 𝛽 -estradiol (E2) and 17- 𝛼 -

ethinylestradil (EE2) (EU 2013), was established. The watch list of these substances was

set out in Commission Implementing Decision (EU) 2015/495 according to Article 8b of

directive 2013/60/EC that additionally includes 2,6-Ditert-butyl-4-methylphenol, 2-

Ethylhexyl 4-methoxycinnamate, macrolide antibiotics, methiocarb, neonicotinoids,

oxadiazon and tri-allate (EU 2015). In 2018, the Commission Implementing Decision

(EU) 2015/495 was updated to EU 2018/840, in which the substances with adequately

available high-quality monitoring data, such as diclofenac, 2,6-Ditert-butyl-4-

methylphenol, 2-Ethylhexyl 4-methoxycinnamate, oxadiazon and tri-allate, were


removed from the watch list with inclusion of sensitive substances, including amoxicillin,

metaflumizone and ciprofloxacin (EU 2018).

Since municipal WWTPs receive diverse type and concentrations of ECs originated from

synthetic chemicals that have been used for many purposes, these WWTPs are the major

sources of many environmental pollution. More than 100,000 tons of ECs are produced

globally (Radjenović et al., 2008). As confirmed by many researchers, the conventional

wastewater treatment methods, such as CAS processes, are not efficient enough to remove

persistent ECs, and many of these ECs remain poorly degraded and subsequently

discharged into receiving waterbodies with treated effluent (Clara et al., 2005; Vieno et

al., 2007; Vieno and Sillanpää, 2014; Zorita et al., 2009). Polar ECs are mostly

hydrophilic and highly persistent under activated sludge process conditions, which lead

them easily detected even in the treated effluent (Buttiglieri and Knepper, 2008).

Therefore, upgrading of WWTPs and implementation of more sustainable technologies

could be the possible solutions to abate such ECs emissions and to achieve high-quality

treated effluent for the safe reclamation (Radjenović et al., 2009). On this note, MBR is

one of the advanced wastewater technologies that has gained increased interest over three

decades (Xiao et al., 2019).

29

3 State-of-the-art of MBRs: advanced municipal

wastewater treatment

3.1 Introduction to MBRs

MBRs couple biological treatment with membrane separation to produce clarified and

highly disinfected effluent (Judd, 2016). MBR technology is based on suspended growth

process that applies microporous membranes to allow solid/liquid separation (Radjenović

et al., 2008). The combination of a biological reaction to membrane separation

strengthens the biological process and replaces many other subsequent operations, such

as secondary clarifiers (Coutte et al., 2017). MBRs have become a state-of-the-art

technology in water management practices for wastewater treatment and reclamation due

to their remarkable benefits such as high quality treated effluent, relatively small plant

footprint, and less sludge production (Drews, 2010). The MBR enables to maintain a high

MLSS concentration, which is important for slow growing microorganisms (Drews,

2010; Judd, 2008). MBRs can perform biological treatment and disinfection of the

effluent simultaneously with the optimum control of biological degradation and greater

stability and flexibility in operation (Radjenović et al., 2009).

There have been several historical milestones leading to the development of today’s

MBRs. The combination of membrane solid/liquid separators in biological treatment

system has been studied since 1960s (Visvanathan et al., 2000). The era between 1960s

and 1980s is often known as a golden age on membrane science development (Hai et al.,

2014). The first historical coupling of an activated sludge bioreactor with a cross flow

membrane filtration looped side-stream MBRs, was developed by Dorr-Oliveir Inc. in

1969 (Smith et al., 1969). However, the first generation MBRs associated with poor

economics due to higher membrane cost and higher energy consumption. The first

breakthrough of the MBR application emerged in 1989 when Yamamoto et al. introduced

the submerging membranes inside the bioreactor (Yamamoto et al., 1988). Currently,

more than 800 commercial MBRs are implemented in Europe, with the compound annual

global market growth rate up to 12% (Krzeminski et al., 2017). The global MBR market

should grow from 1.9 billion USD in 2018 to reach 3.8 billion USD by 2023 (BCC

Research, 2019). More than 60 super-large scale (capacity ≥ 100, 000 m3 d-1) MBR plants

will be in use by 2026, with Tuas Water Reclamation Plant being largest MBR (1200,

000 m3 d-1) in the world (The MBR site, 2019).

Stricter legislations concerning the effluent discharge, demands for wastewater reuse, and

dramatic reduction of membrane capital costs (~1/10) are the major drivers of rapid MBR

growth worldwide (Park et al., 2015). For instance, EU Urban Wastewater Treatment

Directive (91/271/EEC), amended directive 2006/7/EC for management of bathing water

quality, and EPA Clean Water Protection Act (2009), are the important legal complies

30 3 State-of-the-art of MBRs: advanced municipal wastewater treatment

on treated effluent discharge or reuse limits that expanded the MBR market growth

(Krzeminski et al., 2017).

3.2 Membrane materials, configuration and operating principles of

MBRs

Generally, pressure-driven membrane processes are comprised of microfiltration,

ultrafiltration, nanofiltration, and reverse osmosis. Of these membranes, including

microfiltration (0.1 to 1 µm) and ultrafiltration membranes ( 0.05 to 0.1 µm), which are

widely used to separate particulate materials from water and wastewater, are referred as

low-pressure membranes (Yang et al., 2006). On the other hand, nanofiltration and

reverse osmosis are high-pressure membranes, which are commonly used to separate

more soluble matters (Bérubé, 2010). The commercial MBR membrane pore sizes are

generally between 0.03 and 0.4 µm (Judd, 2016). The most widely used membrane

materials are cellulose, polyamide, polysulphone, polymeric materials, including

polyvinylidene difluoride (PVDF), polyethylsulphone (PES), polyethylene (PE), and

polypropylene (PP). Among the commercial membranes, PVDF membranes are the most

widely used membranes that account almost half of all the products both in flat sheet (FS)

and hollow fibre (HF) configurations (Judd, 2016). Most of the polymeric membranes are

resilient to chemical and physical reactions, but severely prone to fouling due to

hydrophobic nature of these membranes and biopolymers. Therefore, often the surface of

commercially available membrane modules are amended by chemical oxidation, metal

organic framework, or plasma treatment to achieve more hydrophilic surface (Radjenović

et al., 2008).

In general, there are two basic MBR configurations for the membrane module, depending

on the membranes placed either inside or outside the bioreactor, such as immersed or

submerged and side-stream or external MBR system as shown in Figure 1 (Gupta et al.,

2008; Judd, 2008; Melin et al., 2006). Submerged or immersed MBRs are more often

applied in municipal wastewater treatment due to their high compatibility with the

activated sludge process, compactness in design, relatively low energy consumption, and

easy wasting of excel sludge (Gupta et al., 2008). On contrary, external or side-stream

MBRs are stand-alone systems, are generally used to treat industrial high strength

wastewater with poor filterability and require relatively less membrane area than in

submerged MBRs. As pumping and recirculation of activated sludge is the integrated part

in external MBRs, high cost is required due to more energy dissipation (Judd, 2008).

31

Figure 1: MBR process configurations, (a) Submerged and (b) side-stream

Figure 2: Different MBR membrane modules: (a) Tubular, (b) Hollow-fiber,

(c) flat-sheet (Reprinted with permission from (Bérubé, 2010)).

There are three different modules of membranes commercially used in MBRs, such as

tubular membranes, hollow fibre and flat sheet membranes (Figure 2). Tubular

membranes are typically used in external MBRs, where membranes having internal

diameter greater than 5 mm are grouped into modules containing multiple tubes. Hollow

fiber and flat sheet membranes are used typically in submerged MBRs. Hollow fiber

membranes generally have external diameter in the range of 1- 3 mm and confined into

modules of dozens to thousands of fibers (Bérubé, 2010).

a) b) c)


3.3 Membrane fouling: a major shortcoming in MBRs

Despite the promising treatment technology, membrane fouling is inevitable phenomenon

in pressure driven MBRs, which normally increases TMP and reduces permeate flux (Le-

Clech et al., 2006). Membrane fouling is the possible deposition of sludge flocs,

particulates, solutes, microorganisms and cell debris in the membrane pores, and on the

membrane surface (Burman and Sinha, 2018; Meng et al., 2009). However, due to its

complex nature, membrane fouling in MBRs is not yet entirely understood (Meng et al.,

2010).

The major fouling consequences in MBRs involve increase in TMP, rapid deterioration

of permeate flux, and increase in filtration resistance, leading to frequent cleanings

requirement, and high energy consumption for aeration (Drews, 2010). Moreover,

membrane fouling also exacerbates feed pressure, reduces effluent quality, weakens

membrane performance, and reduces membrane effective life (Burman and Sinha, 2018).

The specific energy consumption in MBRs operation generally ranges from 0.6 – 2.3 kWh

m-3 (Krzeminski et al., 2017), which is quite high in comparison to CAS processes (0.3 –

0.6 kWh m-3) (Dai et al., 2015). Even though, the membrane costs have dramatically

reduced these days, which ultimately reduce capital investment cost in MBRs, the overall

operating costs are still high mainly due to membrane aeration for fouling mitigation.

Moreover, the use of aggressive chemicals for repeated membrane cleaning events can

create additional environmental concerns (Drews, 2010).

Membrane fouling can be classified mainly into reversible and irreversible types.

Reversible fouling during filtration mainly occurs due to external deposition of loosely

bound materials on the membrane surface that causes cake layer formation. Reversible

fouling can simply be controlled by physical techniques, which can involve back flushing

or relaxation under crossflow conditions. Contrarily, irreversible fouling mainly occurs

due to strongly attached materials and pore blocking of the membrane, which can only be

removed by chemical cleaning (Burman and Sinha, 2018). Nevertheless, irrecoverable

fouling, which is a subsequently left to be treated by chemical cleaning, is difficult to

remove by any cleaning methods (Drews, 2010).

Furthermore, membrane fouling can be categorized as biofouling, inorganic fouling and

organic fouling depending on the fouling components (Meng et al., 2009). Biofouling

occurs because of biological growth, metabolism and its secretion, which ultimately

accumulate on the membrane and its pores and leading to the formation of biofouling.

Organic fouling is due to the organic substances contained in the biomass and organic by-

products produced by microorganisms, which can easily be deposited on membrane

surface. Colloidal initiate membrane pore clogging and biopolymers, such as protein,

polysaccharides and humics can further propagate organic fouling. Whereas, inorganic

fouling forms due to the chemical precipitation of inorganic ions (anions and cations),

including Ca2+, Mg2+, Al3+, Fe3+, PO43-,SO4

2- and OH-. Inorganic fouling mainly occurs

33

via two mechanisms, such as crystallization and particulate fouling. Of these three

fouling, biofouling and organic fouling are dominant in MBRs (Burman and Sinha, 2018;

Hai et al., 2014). Indeed, membrane fouling is a very complex mechanism that includes

complicated inter-relationships between various factors, such as reactor design and

operational conditions, wastewater composition, sludge rheology, ambient conditions

(Zhang et al., 2014). Membrane fouling can be quantitatively determined according to the

Darcy’s equation as follows (Meng and Yang, 2007):

Rt = TMP

μ J (1)

Where, Rt is the total membrane filtration resistance (m-1); TMP is trans-membrane

pressure (Pa); µ dynamic permeate viscosity (Pa.s); and J is the membrane flux (L m-2 h-

1).

The water temperature greatly influences the efficiency of biological treatment processes

(van den Brink et al., 2011). When temperature drops at about 15°C, methane-producing

bacteria becomes essentially inactive, autotrophic bacteria practically cease functioning

at about 5°C, and even chemoheterotrophic bacteria feeding on carbonaceous material

becomes dormant at 2°C (Metcalf & Eddy Inc., 2003). In MBRs, temperature not only

influences biological process but also effects on membrane performance. However, MBR

process operation in low-temperature zones are comparatively less studied than under

normal ambient conditions (Zhang et al., 2014).

At lower temperatures, MBR efficiency deteriorated mainly due to increased MLSS

viscosity, releasing of more EPS due to increased sludge deflocculation, and reduced back

transport velocity of particles from membrane surface (van den Brink et al., 2011).

Arévalo et al. reported the loss in permeability and increased membrane flux resistance

at temperatures < 15°C (Arévalo et al., 2014) in a full-scale MBR. Brink et al. showed

that at low temperatures, membrane fouling is intensified due to the release of

polysaccharides and submicron particles from sludge flocs (van den Brink et al., 2011).

The bulking of sludge has noticed with significant increase in the sludge volume index

(SVI) at 13°C, which was accompanied by rapid membrane fouling and thereby two-fold

increase in the membrane cleaning frequency (Zhang et al., 2014). The bulking sludge

can cause severe cake fouling due to the accumulation of more irregular shaped sludge

flocs, whereas deflocculated sludge can cause both cake fouling and pore blocking fouling

by the deposition of colloidal particles and dissolved matter onto/into the membrane

surface (Meng and Yang, 2007). Krzeminski et al. observed that the colloidal and soluble

fraction (<1 µm) plays a major role in increasing the membrane resistance during winter

period (Krzeminski et al., 2012). The possible fouling components in MBRs and their

spatial and temporal distributions are schematically shown in Figure 3 (Meng et al.,

2017).


Figure 3: Temporal and spatial development of membrane fouling in MBRs

(Modified from (Meng et al., 2017)).

The most common techniques of cleaning membranes in MBRs are physical and chemical

methods. Physical cleaning includes backflushing or relaxation which is applied for 1-2

min in every 10-15 min, whereas alkaline (pH ~ 12) sodium hypochlorite often followed

by acidic (pH ~3) citric (or oxalic) acid with concentration up to 1% are used as chemical

cleaning (Judd, 2008). However, some innovative membrane cleaning protocols have

been emerged over the traditional methods to emphasize improved efficiency and

minimum energy expenditure. The emerging novel strategies involved the application of

various additives (flux enhancers), biocarriers, sludge granulation, membrane surface

modifications, and quorum quenching techniques (Drews, 2010).

3.4 Fate of ECs in MBRs: application of mass balance

The removal mechanism of ECs in MBRs is a complex phenomenon, which is

characterized by following four major pathways: (i) biotransformation or biodegradation;

(ii) sorption onto the sludge; (iii) volatilization; and (iv) physical retention via membrane

size exclusion (Cirja et al., 2008a). However, the ECs removal via volatilization and

physical retention by membrane is almost insignificant since Henry’s law constant (KH)

values of most of the ECs are absolutely low (< 10-6) ( Table 2) and the molecular weight

cut-off (MWCO) of most of the pressure driven MF and UF membranes are above several

35

thousand Daltons compared to molecular weight of commonly found ECs i.e., 150 to 900

Da. Therefore, biotransformation and sorption onto the sludge are the major removal

mechanisms of ECs in MBR treatments (Li et al., 2015; Luo et al., 2014).

The mass balance approach is useful to assess the fate of ECs considering their percentage

shares of biotransformation/biodegradation, sorption and remained in the treated effluent.

The mass loads (µg d-1) of ECs in influent, effluent and sludge can be calculated by

multiplying their measured concentrations (CInf, CEff, CSludge) with the equivalent flow

data (QInf, QEff, QSludge). Then, the relative distributions of ECs to the influent mass load

can be estimated according to the following equations (2)-(5)(Joss et al., 2006):

Overall removal (%) =( C Inf ∗ QInf − C Eff ∗ QEff)

C Inf ∗ QInf∗ 100 (2)

Removal via sorption onto sludge(%)

= C Sludge ∗ QSludge

C Inf ∗ QInf∗ 100 (3)

Remain in effluent (%)

= C Eff ∗ QEff C Inf ∗ QInf

∗ 100 (4)

Removal via biotransformation(%)

= Overall removal (%)

− Removal via sorption onto sludge (%) (5)

where C Inf ∗ QInf ; C Eff ∗ QEff ; 𝑎𝑛𝑑 C Sludge ∗ Q Sludge are the mass load (µg d-1) of

ECs in influent, treated effluent and sludge, respectively.

The removal mechanism of ECs in MBRs is under the influence of mainly ‘internal

factors’ and ‘external factors’(Luo et al., 2014). The internal factors are related to the

physicochemical properties of ECs, including hydrophobicity, biodegradability, degree

of ionization, and volatility (Cirja et al., 2008a; Li et al., 2015). In general, ECs with High

log D value (log DpH 7 > 3.2) are hydrophobic, poorly soluble and with high sorption

affinity on organic constituents (Cirja et al., 2008a). The higher removal efficiency of

hydrophobic compounds may be ascribed to the sorption of ECs onto sludge that assists

for enhanced biotransformation. Contrarily, the removal of hydrophilic ECs (log DpH 7 <


3.2) can be described based on the qualitative framework reported by Tadkaew et al.: (i)

compounds possessing only EDGs, (ii) compounds having both EDGs and e-withdrawing

(EWDs), (iii) compounds having only strong EWGs (Tadkaew et al., 2011). Likewise,

degree of ionization or polarity nature of ECs can influence their removal in MBR. In

general, polar and non-volatile ECs are the main concerns in wastewater treatment

processes as they can easily escape with treated effluent. The solid-water distribution

coefficient (Kd), which indicates the partition of ECs in water phase and sludge, is a

relative indicator of sorption behavior (Joss et al., 2006). The sorption tendency is

insignificant for compounds having log Kd < 2.48, whereas those having log Kd > 3.2 can

have higher removal (Luo et al., 2014; Tadkaew et al., 2011). Similarly, biodegradability

of ECs depends on their bioavailability, which involves primary uptake of ECs by

microorganism cell, leading to their degradation due to the increased affinity with the

bacterial enzymes (Cirja et al., 2008a). The molecular structure of compound greatly

influences the complexity of compound to biodegradation. In short, ECs with linear short

chains, unsaturated aliphatics, and EDGs, are more biodegradable than ECs with long and

more branched side chains, saturated or polycyclics, and sulfate, halogen or EWGs (Luo

et al., 2014).

The external factors are mainly WWTP-specific that involves process operating

conditions, such as SRT, temperature and pH (Luo et al., 2014). SRT better controls the

size and diversity of microorganism community. In general, the MBRs operated with

extended SRTs can improve building up of slowly growing bacteria (e.g., nitrifying

bacteria) and longer retention of sludge inside the bioreactor, which favors higher removal

of ECs (Suárez et al., 2012). Nitrifying conditions can have positive effect on the removal

of ECs via co-metabolism using ammonium monooxygenase enzyme. The SRTs above

10 days that allow nitrifying conditions can enhance the attenuation of many

biodegradable compounds (Clara et al., 2005a). Nevertheless, there are many studies that

revealed no enhanced elimination of ECs with wide variation of SRTs, indicating that

high SRT does not necessarily mean enhanced removal of ECs (Luo et al., 2014).

Likewise, pH value is another critical parameter influencing the removal of ECs in MBR

treatment. Various protonation states of ECs can be expected depending on their pKa

values. It is expected that at low pH, hydrophobicity of most of the ionizable ECs would

increase that ultimately increases their adsorption onto sludge mass, which indeed favors

increased time for biotransformation (Urase et al., 2005). Modest pH variation can also

have significant effects on the removal of acidic ECs due to the activation of enzymes or

enhanced affinity between sludge and ECs due to protonation phenomenon (Kimura et

al., 2007). On contrary, at varying pH, no changes in the removal efficiency of non-

ionizable compounds is noticed (Taheran et al., 2016).

On the other hand, seasonal variation of wastewater temperature can affect the removal

of ECs in MBRs, which influence mostly biodegradation and partition mechanisms (Luo

et al., 2014). In general, with increasing temperatures, adsorption equilibria are reached

earlier, and microbial activities and biodegradation ECs are enhanced. However, Hai et

al. reported the decreased removal of most hydrophobic ECs when process temperature

reached 45 °C (Hai et al., 2011).

37

3.5 Fate of human enteric viruses and heavy metals in MBRs

The human enteric viruses, including noroviruses and adenoviruses, are most common

pathogenic group associated with several waterborne diseases, such as epidemic

gastroenteritis, conjunctivitis, and respiratory diseases (Kuo et al., 2010; Miura et al.,

2015). Human enteric viruses are largely found in municipal wastewater, which are highly

resistant to inactivation. Adequate reduction of viruses is of a prime regulatory focus from

water reuse due to their acute impact on public health and hygiene (Chaudhry et al.,

2015b). The inadequate WWTPs treatment efficiency often led to the insufficient removal

of viruses from source waters, causing their outbreaks to the environment (Lodder and

Husman, 2005).

MBRs offer absolute barrier to solids and microorganisms, despite MBRs are not

significantly efficient retaining viruses via membrane size exclusion (Herath et al., 1999).

Even though virus removal in MBRs are ascribed to their aggregation and adsorption to

sludge and retention by the gel and cake layer formed on the membrane surface. On the

other hand, formation of fouling layer on the membrane surface plays a vital role on

controlling virus removal, yet current understanding of interactions between viruses and

membrane surface, impact of membrane cleaning as well as membrane imperfections in

MBRs is limited (Hai et al., 2014).

Miura et al. indicated that adsorption to sludge plays a major role in the removal of

norovirus GII in the MBR, in which diverse adsorptive behavior of viruses to sludge

influenced the overall removal efficiency (Miura et al., 2015). Kuo et al. reported the

attachment of adenovirus to sludge and membrane-biofilm as a major mechanism for their

removal in the MBR (Kuo et al., 2010). However, the biomass attachment removal

mechanism is mostly dependent on the type of virus (Chaudhry et al., 2015a). In another

study, Chaudhary et al. suggested the major removal contribution of viruses is majorly by

backwashed membrane, inactivation by biomass due to predation or enzymes, the cake

layer formation and attachment to solids (Chaudhry et al., 2015b).

Heavy metals refers to any metals with relatively high density (>5 g cm-3) and atomic

weights, which are most serious environmental threats due to their recalcitrance and

persistence in the environment (Chipasa, 2003; Fu and Wang, 2011). The risks associated

with the bioaccumulation in the food chains are the major environmental and human

health concerns (Cantinho et al., 2016). The presence of heavy metals at certain

concentrations not only pose environmental concerns but also strongly reduces microbial

activity, leading to adverse effects on biological treatment processes (Chipasa, 2003).

Trace quantities of many heavy metals are identified in municipal wastewater from

various sources, mainly anthropogenic emissions. Of the many heavy metals, cadmium

(Cd), lead (Pb), mercury (Hg) and nickel (Ni) are defined by EU directive 2013 as priority

hazardous elements, which could pose acute to chronic health problems to human and

aquatic life depending on their potency and time of exposure (EU, 2013).


Some studies have confirmed the potential of MBRs for enhanced removal of heavy

metals from wastewater than conventional treatment methods (Di Fabio et al., 2013;

Fatone et al., 2008; Katsou et al., 2011; Santos and Judd, 2010). There are several

mechanisms involved in heavy metal removal by MBRs, such as surface adsorption and

diffusion of metal ions in sludge flocs, adsorption and complexation formation of soluble

metal ions in EPS matrix, and to some extent by chemical precipitation of metals due to

alkaline nature of wastewater (Katsou et al., 2011). Di Fabio et al., who studied the

removal and fate of heavy metals in a pilot-scale MBR revealed that the biosorption of

heavy metals to fouling layer (esp. EPS matrix) is more evident than to sludge (Di Fabio

et al., 2013). However, the pH of the medium greatly influences the number of EPS

binding sites for heavy metals (Comte et al., 2008). Fatone et al. reported that operating

MBRs at high SRTs is not effective strategy to enhance the removal of heavy metals at

low concentrations (Fatone et al., 2008). Addition of vermiculite minerals could enhance

the heavy metal removals due to increased sites for biofilm formation and microorganism

growth, but can increase sludge concentration and membrane fouling (Katsou et al.,

2011).

39

4 Advanced oxidation processes for wastewater treatment

Biological treatment is usually the most economic process to treat ‘readily degradable’

organic pollutants present in wastewater. However, if the wastewater is comprised of

refractory (non-biodegradable) organic pollutants, such as several ECs, then it requires

especial type of treatment to meet the desired discharge limits (Comninellis, 1994). In

this regard, advanced oxidation processes (AOPs), based on aqueous phase oxidation

method on the intermediacy of powerful reactive species, mainly hydroxyl radicals (•OH),

leading to the nonselective degradation of the organic pollutants to their mineralization

(Sirés et al., 2014).

AOP was first proposed by Glaze in 1980s for potable water treatment using ozone

(Glaze, 1987). Since then AOPs are widely applied to treat wastewater as strong oxidants

that can degrade recalcitrant ECs and simultaneously increase the biodegradability (Deng

and Zhao, 2015). AOPs are promising, efficient and environmental-friendly approaches

to treat wastewater contaminated with wide range of ECs, mainly by electrochemical,

photochemical, ozonation, Fenton, or sonochemical reactions (Nidheesh et al., 2018;

Oturan and Aaron, 2014). AOPs can be utilized either as pre- or post-biological treatment

depending on the concentration of ECs in the wastewater influent stream. Pre-treatments

using AOPs are more feasible when the concentrations of ECs are higher than the organic

matters, whereas post-treatments are realistic when the concentrations of ECs are

relatively lower than organic matters (Neoh et al., 2016).

The hydroxyl radical is the second most powerful oxidizing agent after fluorine with a

redox potential of E° (•OH/H2O) = 2.8 (V vs. SHE), and a life time of less than a second

(Brillas and Martínez-Huitle, 2015). These in situ produced •OH radicals then rapidly

reacts with organic pollutants at high reaction rates mainly by the electron transfer or

redox reactions, H atom abstraction (aliphatics), or the addition on an unsaturated bond

(aromatics) to initiate a radical chain oxidation reaction (Nidheesh et al., 2018; Sirés et

al., 2014).

4.1 Electrochemical oxidation (ECO) process

Electrochemical oxidation processes have gained great attention over the last decade for

the decontamination of wastewater containing persistent organic pollutants and for

disinfection purposes (Brillas et al., 2009; Särkkä et al., 2015). ECO processes are

emerging as most appealing environmentally friendly and highly efficient

electrochemical treatments (Martínez-Huitle and Brillas, 2009). The major benefit of

electrochemical treatments is its ability to regulate and generate in situ •OH radicals

without adding chemicals or large amount of catalyst (Oturan and Aaron, 2014). Instead,

ECO process can also benefits on the prevention and remediation of pollution issues since

the electron is a clean reagent (Panizza and Cerisola, 2009). In ECO treatments, the

organic pollutants can be removed by: (i) direct anodic oxidation (or direct electron

40 4 Advanced oxidation processes for wastewater treatment

transfer between anode surface and pollutants), which usually yields relatively poor

oxidation rates, and (ii) chemical reaction with electrogenerated reactive oxygen species

at anode surface such as physisorbed or chemisorbed •OH radicals, which leads to total or

partial oxidation of pollutants, respectively (Martínez-Huitle and Brillas, 2009; Panizza

and Cerisola, 2009).

In direct anodic oxidation, pollutants are oxidized at anode surface after adsorption solely

due to the electron, without the involvement of any other substances. In this process,

oxygen is transferred from water to the organic pollutant via electrical energy, which is

called electrochemical oxygen transfer reaction as described by equation (6) (Panizza and

Cerisola, 2009):

Rads – ne- → Pads (6)

where, Rads and Pads are adsorbed and oxidized adsorbed organic pollutants at anode

surface. Direct electro-oxidation can be possible at low potentials, but usually possesses

low kinetics, depending on the electrocatalytic activity of anode materials. The use of low

cell potentials avoiding oxygen evolution could frequently cause the loss of anodic

activity. This is called poisoning effect and occurs due to the formation of polymer layer

on the anode surface, which has limited their application in wastewater treatment

practices (Martínez-Huitle and Brillas, 2009). The anode surface deactivation depends

on, (i) adsorption properties of the anode surface ,and (ii) nature and concentration of

organic pollutants and their transformation products (Panizza and Cerisola, 2009).

On the other hand, electrochemical oxidation of organic pollutants can be obtained

without the poisoning effect by electrolyzing the aqueous solution at high anodic voltages

with the involvement of oxygen evolution reactions (OERs), which generates in situ •OH

radicals with no oxidation catalysts. A number of anodes favored partial and selective

oxidation of organic pollutants (i.e., conversion method), whereas many others favored

complete mineralization (i.e., combustion/incineration) to CO2 (Brillas and Martínez-

Huitle, 2015; Panizza and Cerisola, 2009). Comninellis found that the nature of anode

material strongly effects both the selectivity and efficiency of the ECO process, and

proposed a comprehensive model of organics oxidation in acidic medium including the

competition with O2 evolution reaction and heterogeneous •OH radicals (Figure 4)

(Comninellis, 1994). The model assumes that in the initial step of oxygen transfer

reaction, water molecules split to form adsorbed •OH radicals at anode oxide surface

(MOx) as equation (7):

MOx + H2O → MOx (•OH) + H++ e- (7)

41

Figure 4: Electrochemical oxidation scheme of organic pollutants on

the metal oxide anode surface (Modified from (Comninellis, 1994)).

Then, the different behaviors of anodes in ECO process was described by considering two

limiting cases of electrodes, defined as ‘’active’’ and ‘’non-active’’ anodes in the

schematic model.

4.1.1 Active anodes

At active anodes, with low O2-overpotentials (such as IrO2, RuO2, and Pt) (Martínez-

Huitle and Brillas, 2009), a strong interaction occurs between the anode surface (MOx)

and •OH radicals (i.e., chemisorbed active oxygen in the lattice of metal oxide). The

chemisorbed •OH radicals may interact with anode, forming so-called higher oxide

(equation (8)). The surface redox couple MOx+1/MOx can act as a mediator in the

conversion of selective oxidation of organics (R with m carbon atom and without

heteroatoms) into short-chain carboxylic acids (equation (9)). Moreover, O2 evolution as

side reaction due to chemical decomposition of higher oxides is also involved in this

reaction (equation (10))(Martínez-Huitle and Ferro, 2006; Panizza and Cerisola, 2009) :

MOx (•OH) → MOx+1 + H++ e- (8)

MOx+1 + R → MOx + RO (9)


MOx → MOx + (1/2) O2 (10)

4.1.2 Non-active anodes

On the other hand, at non-active anodes, with relatively high O2-overpotentials (such as

BDD, PbO2, and SnO2 (Brillas and Martínez-Huitle, 2015; Nidheesh et al., 2018)), weak

interaction exists between physisorbed •OH radicals and the electrode surface. These

anodes do not contribute to the direct anodic oxidation of organic pollutants and do not

promote any catalytic sites for pollutant adsorption (Nidheesh et al., 2018). Therefore, the

formation of higher oxide is excluded and allows the direct oxidation of organics with

adsorbed •OH radicals, which may result in complete combustion to CO2 as follows

(equation (11)):

𝛼MOx (•OH) + R → 𝛼MOx + mCO2 +nH2O + zH++ ze- (11)

where, organic compound R needs 𝛼 = (2m+n) oxygen atoms for complete mineralization

to CO2. The oxidative reaction (9) is much more selective than mineralization reaction

(6). Like in active anode behavior, reaction (11) also undergoes competitive side reaction,

i.e. O2 evolution reaction, resulting in decreased anodic oxidation performance, either

direct oxidation to O2 (equation (12)) or indirect consumption via dimerization to

hydrogen peroxide (equation (13)).

MOx (•OH) → MOx + (1/2) O2 + H++ e- (12)

2 MOx (•OH) → 2 MOx + H2O2 (13)

According to the mechanisms, anodes with low O2-overpotentials (active anodes) (good

catalysts for O2 evolution) allow only partial oxidation of organics, whereas anodes with

high O2-overpotentials (non-active) (poor catalysts for O2 evolution) favor complete

oxidation of organics to CO2 (Panizza and Cerisola, 2009). Therefore, non-active anodes

are ideal for wastewater purification. Nevertheless, since both organic oxidation and O2

evolution reactions compete parallel, many anodes exhibit diverse behavior in practices

(Nidheesh et al., 2018).

It is well confirmed that the type of anode material is the most important factor for

determining the extent of organic pollutants degradation in ECO processes (Comninellis

and Vercesi, 1991). There are wide variety of anode materials including doped and

undoped PbO2, mixed metal oxides of Ti, Ru, Ir, Sn and Sb, Ti/Pt, carbon based anodes ,

and boron-doped diamond (BDD) thin films anodes (Brillas and Martínez-Huitle, 2015).

The PbO2 and BDD anodes, with high O2-overpotential, are the most commonly used

anodes for the electrochemical degradation of organic pollutants. The PbO2 anodes are

inexpensive, easy to prepare, chemically stable, low electrical stability, and a large

surface area (Martínez-Huitle and Brillas, 2009). Likewise, BDD anodes are made by

depositing a thin diamond films on non-diamond materials, usually silicon, tungsten,

43

titanium, tantalum, or glassy carbon, by energy-assisted chemical vapor deposition. BDD

anodes are very promising anode material for the electrochemical treatment of wastewater

contaminated with organic pollutants. They exhibit a great chemical and electrochemical

stability, a wide electrochemical working range ( HER at -1.25 V vs SHE and OER at +

2.3 V vs SHE), high inert surface with low adsorption, and a great mineralization

efficiency compared to other anodes (Oturan and Aaron, 2014; Panizza and Cerisola,

2009). The behavior of PbO2 and BDD anodes during the electrooxidation of organic

pollutant is well confirmed by Panizza and Cerisola (Panizza and Cerisola, 2009).

Nevertheless, despite of the high O2-overpotential, PbO2 anodes show low durability by

surface corrosion and leaching of highly toxic Pb2+ ions into to treated water (Brillas and

Martínez-Huitle, 2015). Similarly, the major drawbacks of BDD anodes includes high

material cost and difficulties in finding an appropriate substrate for thin diamond layer

deposition (Panizza and Cerisola, 2009), which tend to compromise their applicability in

large-scale applications.

Alternatively, MMO electrodes include a broad range of materials able to adsorb oxygen

on their structure and more suitable for anodic oxidation of organic pollutants.

Dimensionally stable anodes (DSA), are the commercial terms for MMO electrodes, due

to their excellent ECO properties and structural integrity (Shestakova, 2016). DSAs

usually constitutes a Ti substrates deposited by a thin semiconducting layer of metal oxide

or mixed metal oxides, such as Ti/TiO2-RuO2, Ti/Ta2O5-IrO2, Ti/IrO2-RuO2, Ti/SnO2-

Sb2O5, and Ti/Ir0.45O2-Ta2O2 (Brillas and Martínez-Huitle, 2015; Shestakova, 2016).

4.2 Photocatalytic oxidation (PCO)

Among other AOPs, heterogeneous photocatalysis is also a well-proven, efficient, and

eco-friendly method for attenuating ambient concentrations of organic contaminants in

aqueous or gas (air) phase (Guillard et al., 1999). Heterogeneous photocatalysis ascribes

to the speeding of photoreaction in the presence of semiconductor catalyst and the light

source. The major advantage of heterogeneous catalysis is the partial photocatalytic

oxidation at beginning or complete mineralization of organic pollutants essentially to

benign substances (Gaya and Abdullah, 2008). In early 70’s, Fujishima and Honda first

unfolded the possibility of water splitting using titania (TiO2) photocatalyst in a photo-

electrochemical solar cell, referred to as the ‘‘Honda-Fujishima effect’’(Fujishima and

Honda, 1972). Subsequently, the heterogeneous photocatalysis by titania photocatalyst

became environmental frontiers (Gaya and Abdullah, 2008; Oturan and Aaron, 2014).

In heterogeneous photocatalysis process, organic pollutants are decontaminated in the

presence of semiconductor photocatalysts, a light energy source, and reactive oxidizing

species. When the irradiated photon energy is greater than the band gap energy (∆E) of

the semiconductor, electrons are excited to the conductance band (ecb-) and positive holes

is generated on the valence band ( hvb+) site vacated by electrons and electron-hole pairs

(ecb-/ hvb

+) are formed (Joo et al., 2019). This chemistry of photo-generated electron-hole


pairs could form reactive oxidants such as •OH radicals, and superoxide (•O2-) radicals

and adsorbed onto the semiconductor surface, and consequently organic pollutants are

then non-selectively oxidized by both photo-generated holes and reactive oxidant species

(Dong et al., 2015). However, the absorption of photons with lower energy than ∆E or

shorter wavelengths causes energy loss in the forms of heat due to rapid recombination

of charge carriers i.e., ecb-/ hvb

+(Ahmed et al., 2011; Cavalcante et al., 2015). On this note,

TiO2 (∆E = 3.2 eV) assisted heterogeneous photocatalysis has been widely applied in

recent years for photocatalytic oxidation of refractory organic pollutants present in

wastewater. TiO2 is easy to produce, inexpensive, chemically and biologically stable, and

has an energy gap comparable to that of solar photons (Oturan and Aaron, 2014). The

activation of TiO2 by UV light irradiation and the photocatalytic degradation mechanism

of organic pollutant can be represented by the following equations (14)-(17) (Ahmed et

al., 2011):

TiO2 ℎ𝑣→ e- + h+ (14)

e- + O2 → •O2- (15)

h+ + H2O → •OH + H+ (16)

h+ / •OH / •O2- + R → Intermediates→ CO2 + H2O+ inorganic ions (17)

In heterogeneous photocatalysis, TiO2 catalyst can be utilized either as aqueous

suspension or in thin film form (Murgolo et al., 2017). The dispersed TiO2 form is easy

to use, provide great surface area, and aerated suspensions prevents the recombining of

ecb-/ hvb

+ to some extent, despite progressive formation of dark catalytic sludge could

diminishes the efficiency of UV intensity. Contrarily, TiO2 thin films do not need to

separate the catalytic particles for reusability, but the catalytic layer should be very stable

and active (Oturan and Aaron, 2014). There are a number of studies performed on

heterogeneous UV/TiO2 photocatalysis of many emerging contaminants (Benotti et al.,

2009; De la Cruz et al., 2013; Doll and Frimmel, 2004; Ghaly et al., 2011; Hapeshi et al.,

2010; Kanakaraju et al., 2014; Rizzo et al., 2009; Xue et al., 2011).

However, like other AOPs, major drawbacks associated with TiO2-based heterogeneous

photocatalysis, includes low quantum efficiency, low adsorption capacity for

hydrophobic pollutants, post- recovery TiO2 particles, and catalysts deactivation (Ahmed

et al., 2011). To overcome these shortcomings, various countermeasures have been

studied in order to enhance photocatalytic efficiency, complete destruction of organic

pollutants, improve light adsorption range, increase stability and reproducibility, and to

improve recycle and reuse capabilities of TiO2 catalyst (Dong et al., 2015). Over the past

few decades, transition metals (e.g. Fe, Ni, Cr, Ag, Au) or their oxides doped TiO2

photocatalysts have been studied for enhancing the removal efficiency of various organic

pollutants (Anpo, 2000; Ren et al., 2015; Khan et al., 2016; Chen et al., 2015; Zhou et al.,

45

2010). TiO2 nanoparticles are modified by simply substitution reaction or intrinsic doping

with different metal cations, which can then form metal oxides or a mixture of oxides.

The transition metals doping improves trapping of excited electrons that inhibits rapid

recombination of ecb-/ hvb

+. Therefore, metal nanoparticles act as an electron traps in

storing and shuttling photogenerated electrons from the TiO2 surface (Dong et al., 2015).

In that context, Ag2O with narrow energy band gap of 1.2 eV (Xu and Schoonen, 2000),

which is a proven photocatalyst for organic pollutant degradation, can be doped on TiO2

nanoparticles for increasing surface area, well-defined crystal structure and homogeneity

in particles size for enhanced photocatalytic effect (Wang et al., 2011). Enhanced

efficiency of dye removal via the heterogeneous photocatalysis using Ag2O doped TiO2

photocatalysts has been confirmed in many previous studies (Ren and Yang, 2017; Zhou

et al., 2010).

47

5 Integration potentials of MBR with AOPs

MBRs are established as environmental-friendly approaches for municipal wastewater

treatment. However, the removal efficiency of many biologically persistent and

hydrophilic ECs with biological treatment alone is not effective (Fatta-Kassinos et al.,

2016). Moreover, although some MBRs can treat wastewater to meet the current

guidelines and producing water for basic industrial reuses, the treated effluent still needs

to be further decontaminated for high grade applications. Generally, single classical

treatment techniques have been inadequate for the removal of persistent ECs, therefore it

is necessary to integrate MBR with other green advanced treatment processes, such as

AOPs, high retention membranes, MBBR, and granulation technology. Since usually

MBR effluent is of high quality (i.e., low level of suspended solids and turbidity), MBRs

offer great technical flexibility to be integrated with other technologies, and significant

synergy can be achieved.

Of the many integration potentials, coupling MBR process with different AOPs have

achieved a great deal of attention over the last decade. The ‘multi barrier approach’

integration of MBR with AOPs aims to enhance the quality of effluent, mitigate the

fouling and to robust the process stability (Neoh et al., 2016). As a multi barrier approach,

it is evidenced that the integrated approach can offer the recycling of oxidation products

back to the biological system for further biodegradation (Laera et al., 2011). The general

integration strategy of wastewater treatment using biological treatment and AOP can be

illustrated by Figure 5. If the wastewater stream contains higher ECs load, then integration

of AOP is beneficial as ‘pre-treatment’ of ECs, which converts recalcitrant ECs to more

biodegradables and subsequently treated in MBR with high biodegradation efficiency.

Whereas, integration of AOP as polishing step to MBR process could further improve the

removal of residual fractions of ECs after MBR treatment. The integration approach not

only benefits the synergy in the treatment performance, but also trade-off the

shortcomings/ challenges of one another (Ganiyu et al., 2015).

48 5 Integration potentials of MBR with AOPs

Figure 5: Decision tree for potential integration approaches of MBR

and AOP (Modified from (López et al., 2010)).

García-Gómez et al. showed an enhanced removal of carbamazepine in synthetic

wastewater from about 6.5 % to 99.99% (complete removal) when MBR system was

coupled with ECO process using Ti/PbO2 anode (García-Gómez et al., 2016). While

MBR was integrated with ozonation in the recirculation stream of the MBR effluent,

residual concentration of anti-viral drug (acyclovir) in the MBR effluent was further

removed by 99%. The major benefit of this integrated system was evident for the removal

of specific ozonation transformation products of acyclovir, which was about 20-fold

higher than MBR alone (Mascolo et al., 2010). Likewise, Pollice et al. reported the

complete degradation of antibacterial (nalidixic acid) when treated with integrated MBR-

ozonation process (Pollice et al., 2012). López et al. demonstrated that the integration of

49

MBR with solar photo-Fenton oxidation process could effectively treat wastewater

polluted with pesticides, by enhancing biodegrability and lessening ecotoxicity of the

treated effluent (López et al., 2010). A laboratory scale MBR system combined with

nanofiltration or reverse osmosis membrane filtration by Alturki et al. reported more than

95% removal of 40 trace organic compounds (Alturki et al., 2010). More than 95% of

carbamazepine removal efficiency was observed by Laera et al. when MBR was

integrated with a lab-scale heterogeneous TiO2-based annular slurry photoreactor (Laera

et al., 2011).

The literature survey indicates that there have been very few studies acquired on the

possibility of integrating MBR with AOPs to enhance the removal of ECs. Other than

AOPs, physical treatment processes (e.g, nanofiltration and reverse osmosis) could

possibly enhance the removal of persistent ECs, but ultimately ends-up in the

concentrates or sludge, which may need to be treated further. Moreover, most of the NF

or RO processes attenuate ECs based on their respective MWCO. On the other hand,

AOPs such as ECO and PCO are eco-friendly processes that non-selectively

degrade/combust highly persistent ECs via hydroxylation or dihydroxylation till their

total mineralization or until they converted into non-toxic and biodegradable short-

chained molecules (Ganiyu et al., 2015).

50 5 Integration potentials of MBR with AOPs

51

Figure 6: Schematic diagram of the pilot MBR unit at Kenkäveronniemi WWTP,

Mikkeli, Finland.

6 Materials and methods

6.2 Assessing the performance of MBR unit in different operating

conditions

6.1.1 The Pilot-scale MBR set-up

A pilot scale submerged MBR unit (ARTAS Ltd. Turkey) was installed at full-scale

Kenkäveronniemi WWTP, Mikkeli, Finland as depicted in Figure 6. The full-scale plant

treats about 5 Million cubic meter of wastewater per year, majorly of municipal sewage

with the influent load of about 63000 population equivalent. The treated wastewater is

discharged into the Saimaa Lake, which is the largest in Finland and fourth largest

freshwater lake in Europe.

The submerged flat-sheet membrane units (0.4 µm pore size), made of chlorinated

polyethylene polymers, with total surface area of 16 m2 (KUBOTA corporation, Japan),

were used for the solid-liquid separation. The in-situ membrane fouling was controlled

physically by providing crossflow scouring air via compressed air and adopting

intermittent suction cycle for filtration. The clean-in-place (CIP) tank was connected with

sodium hypochlorite (NaOCl) and Citric acid dosing tanks for the cleaning of membranes

during fouling occasions. The pilot MBR was operated under different operating

conditions depending on the purpose of studies (Paper I, II, III and IV) as described in

Table 2.

52 6 Materials and methods

Table 3. Different operating conditions of the MBR pilot plant depending on

purpose of studies.

Parameter Units

Values

Paper I Paper II Paper III Paper

IV

HRT Hours 35 42.13 40-45 40-45

SRT Days 25 - 30 60 and 21 55-60 55-60

Avg. Flux L m-2 h-1 7.80 5.84 4-6 4-6

MLSS

concentration mg L-1

5300 -

9800

8550 and

3748 5000-8000

5000-

8000

F/M ratio

kg COD

(kg

MLSS.

d)-1

0.02 –

0.05

0.027 and

0.09

0.02 –

0.09

0.02 –

0.09

Aeration

intensity

(SADm)

m3 m-2 h-1 0.4 – 0.6 0.2 0.2 -0.23 0.2 -

0.23

Process

temperature °C 6.5 - 21 19 ± 2 15-22 15-22

pH Unitless 6.6 – 7.3 6-7.4 6-8 6-8

Suction cycle Minutes 9-On/1-

Off

9-On/1-Off 9-On/1-

Off

9-On/1-

Off

6.2 Assessing integration approaches to MBR

For further polishing the treated effluent from the MBR treatment, two widely used AOPs,

i.e., ECO and PCO processes were assessed for the removal of highly recalcitrant ECs,

including carbamazepine and diclofenac, as model compounds. These AOP units were

performed as ‘post-treatment’ to the residual concentrations of ECs from the MBR treated

effluent.

6.2.1 Lab-scale set-up for batch ECO unit

The lab-scale set-up used for the electrochemical oxidation system is as shown in Figure

7. The experiment was carried out in a glass reactor with cooling jacket arrangement. The

external electrical power was applied by using DC power supply. The inter-electrode gap

of 10 mm was maintained between both anode and cathode electrodes. The aqueous

solutions containing carbamazepine were electrolyzed under different operating

conditions and its attenuation in the concentration was analyzed using UV-vis

53

Figure 7: Experimental set-up for electrochemical oxidation.

spectrophotometer. The newly developed Ti /Ta2O5-SnO2 electrodes were synthesized by

thermal decomposition method using precursor solution of tantalum (Ta) and tin (Sn)

salts on titanium (Ti) plates. The as-prepared electrodes were then characterized for their

microstructure, chemical analysis, and cyclic voltammetry (CV).

6.2.2 Lab-scale set-up for PCO process

The photocatalytic experimental set-up using a self-made photocatalytic reactor is shown

in Figure 8. The photocatalytic degradation of ECs (carbamazepine and diclofenac) in

aqueous solutions (both in deionized water and real MBR effluent) was performed using

5% (w/w) Ag2O/P-25 photocatalysts under UV irradiation. The sample aliquots of ECs

were analyzed by using UV-vis spectrophotometer. The 5% (w/w) Ag2O/P-25

photocatalysts were prepared by a simple pH-mediated chemical precipitation method

and characterized for microstructure, morphology, elemental analysis, surface area, and

energy band gap using different characterization instruments.


Figure 8: Experimental set-up for photocatalytic oxidation

6.2.3 Analytical procedures

The common pollutants in wastewater, such as TSS, COD, NH4-N and TP contained in

influent and effluent of the MBR process, as well as MLSS and MLVSS concentrations

were analyzed according to the Standard Methods (APHA, 1999) (Paper I-IV).

For human enteric viruses, including Adv and NoV (GI and GII) from liquid samples,

pre-treated 500 mL samples (influent and effluent) were concentrated by two-phase

separation method and viral nuclei acids were extracted using the High Pure Viral Nucleic

Acid and High Pure Viral RNA Kit. For sludge samples, 500 mL of sample was

centrifuged to 10 min at 1000g at 4°C. The supernatant was concentrated by two-phase

separation process. For the solid fractions, 50 g of sludge pellet was mixed with 50 mL

of 3% beef extract (pH 9.5) for 30 min at reverse transcriptase (RT) to elute viruses. After

centrifugation for 20 min at 5000g at 4°C, the supernatant was recovered and adjusted to

a pH 7.2. The supernatant samples were then mixed with polyethylene glycol 6000 (PEG

6000) and NaCl, and further rocked overnight at 4°C. After centrifugation for 30 min at

10000g, supernatant was discarded, and the PEG pellet was collected for nucleic acid

extraction (U.S. EPA, 2003). The detection of Advs was carried out by performing the

55

real-time RT-PCR assays in a Rotor-GeneTM 3000 real-time rotary analyzer, whereas

the real-time RT-PCR assays were carried out in one step for the detection of both NoVs

using the QuantiTect Probe RT-PCR Kit. The possible effect of inhibitors in wastewater

samples were assumed to be constant throughout the experimental period (Paper I). On

the other hand, total number of bacteria indicators, such as E.coli and Enterococcus in the

influent, sludge and permeate were enumerated using Colilert ® and Enterolert ®

methods (Paper I).

For heavy metals analysis, liquid influent and effluent samples were acidified (pH<2)

with HNO3 acid, centrifuged at 4000 rpm for 20 min, and then supernatant was filtered

with 0.20 µm cellulose syringe filters. The sludge samples were dried in oven at 105°C.

About 0.5 g of dried sludge in 10 mL HNO3: HCl (1:3) mixture was digested in a

microwave digester. Finally, heavy metals were analyzed using inductively coupled

plasma-optical emission spectrometer (ICP-OES, iCAP 6300, Thermo Electron

Corporation, USA) from the pretreated samples (Fontmorin and Sillanpää, 2015) (Paper

I & III).

The occurrence of diverse ECs from aqueous samples (influent and effluent) and solid

samples (sludge) of the MBR treatment was carried out by standard solid phase extraction

(SPE) followed by LC-MS/MS analysis (U.S. EPA, 2007) (Paper I & II). For SPE,

typically, about 100 mL of effluent or 50 mL of influent or sludge samples were spiked

with the mass-labelled surrogates, for example 13C-carbamazepine, 13C-bisphenol A,

2H-diclofenac, and so on. Additionally, mixing, ultrasonic extraction and centrifugation

were repeatedly employed for sludge samples. The spiked samples were extracted,

conditioned with the mixture of 20 mL methanol and 20 mL deionized water, washed

with 10 mL of 5% methanol and eluted with 15 mL of methanol. The mixture was then

evaporated to dryness under nitrogen flow and 0.5 mL of methanol and 0.5 mL of

deionized water were added. The aliquots were then analyzed using LC-MS/MS (Waters

Acquity UPLC and Xevo TQ, USA). The recovery correction during the calculation was

made with respect to the mass-labelled surrogates. The measurement uncertainty (MU)

was about 40%. The MU % was calculated according to the Nordest TR 537 that covered

every step in the laboratory analysis. Moreover, the residual concentrations of ECs, such

as carbamazepine and diclofenac were analyzed by UV-vis spectrophotometer (Lambda

45, USA) (Paper III & IV). The photocatalytically induced intermediates of

carbamazepine were identified and quantified by using Fourier-transform ion cyclotron

resonance (FT-ICR) mass spectrometry (Bruker Daltonics, Germany) (Paper IV).

The fouling precursors, such as EPS and SMP from sludge were determined by

centrifuging 25 mL of sludge samples at 4000 rpm at 4 °C for 30 min. The supernatant

was filtered through a 0.45 μm glass filter and measured for SMP. The obtained pallets

were resuspended by adding 0.05% NaCl, rapidly mixed with vortex mixture and the

solution was kept in a water bath (~100 °C) for heat treatment for 60 min. Finally, the

supernatant was filtered through a 0.45 μm glass filter and measured for total EPS. Both


SMP and EPS were analyzed using a TOC analyzer (TOC-V Series-CPN, Shimadzu,

Japan) and expressed as mg of DOC g-1 MLVSS (Paper II).

57

7 Results and discussion

7.1 Assessing the performance of MBR at Nordic cold conditions (Paper

I)

In order to investigate the performance of MBR at Nordic cold conditions, the MBR pilot

unit was operated for more than 120 days in winter season of Finland (January to April

2016). This is the first comprehensive study on the performance of pilot MBR on treating

the real municipal wastewater in the unique Nordic environment. The wastewater

temperature inside the bioreactor varied from 7 to 20 °C. On this occasion, the effect of

low sludge temperatures on membrane fouling was evaluated by monitoring TMP and

deterioration of sludge settleability was measured by monitoring SVI. Moreover, the

performance of MBR on removing organic, solids and nutrients loads in the incoming

wastewater, reduction of common enteric viruses, and most importantly the attenuation

of ECs, such as PhACs, EDCs and some toxic heavy metals, were assessed. The

membrane fouling tendency under cold water conditions and adopted cleaning methods

and frequencies are depicted in Figure 9.

While operating MBR in cold-water conditions, many consequences related to the process

operation were observed. During the experimental period, the ambient air temperature

was recorded, which was fluctuated between 0 and -33 °C for more than a week, which

is very typical in winter seasons of Nordic regions. With the decreasing sludge

temperature inside the bioreactor, increasing trend of TMP profile was observed (Figure

9). Two distinct stages of membrane fouling, which is well known tendency in MBRs

(Park et al., 2015), such as slow and steady fouling and rapid (jump) fouling, were clearly

observed. While the sludge temperature was recorded below 10 °C, two consecutive

chemical cleanings for fouled membranes were adapted to reduce very frequent jumping

of TMPs within a week. However, the cleaning cycle was extended to about two weeks

when sludge temperature raised more than 10 °C. It was also noticed that other than cold

water temperature, changing in the specific aeration demand per membrane surface area

(SADm) could effect on membrane fouling behavior to some extent. The permeability of

the membranes was dropped from about 345 L m-2 bar-1 to nearly 50 L m-2 bar-1 while

temperature decreased from 20.5 ± 0.6 °C to below 10°C. About 75% of permeability

drop was observed during the lowest sludge temperature period. After the final manual

cleaning, the permeability was restored gradually and reached nearly to 300 L m-2 bar-1.

The results indicated that the cold-water conditions can deteriorate the hydraulic

performance of MBR due to accelerated membrane fouling propensities. The sludge

settleability was also highly deteriorated, as indicated by the SVI values being increased

from 120 to 168 mL g-1, during the coldest temperature events observed inside the

bioreactor. On the other hand, despite of low temperature operation, relatively high

quality of treated effluent was observed with the corresponding removal efficiencies of

100%, 92.8%, >99.8%, and 90.6% for TSS, COD, NH4-N, and TP respectively.

58 7 Results and discussion

Figure 9: Membrane fouling behavior of MBR system under cold weather

conditions and adopted cleaning measures.

Regarding the fate of human enteric viruses, such as Noroviruses (NoV GI and NoV GII)

and adenovirus (AdVs), all three viruses were constantly detected in the influent samples

during the experimental period as shown in Figure 10. NoV GI was detected in 85% of

the influent samples with an average concentration of 95 GC mL−1, whereas NoV GII and

AdVs were positively detected in 100 % with corresponding average concentrations of

896 and 124 GC mL−1, respectively. The average log removal of 1.82, 3.02 and 1.94 log

units were achieved during MBR treatment for NoV GI, NoV GII and AdVs, respectively.

It was further confirmed that the virus removal efficiency was not correlated with the

degree of membrane fouling. Moreover, none of the concentration of Escherichia coli

and enterococci were detected in the MBR permeate over the experimental period,

showing excellent retaining of these bacteria.

59

Figure 10: Occurrences of viruses in wastewater stream under cold

water conditions.

Similarly, the occurrences and removal efficiencies of four ECs, including PhACs, such

as carbamazepine, bisoprolol, diclofenac, and EDC, i.e., bisphenol A, in the MBR

treatment under cold weather conditions were investigated. All the selected ECs were

frequently identified in the influent wastewater samples, but the relatively higher

concentrations were observed from January up to the end of March, ascribing the high

specific consumption of these ECs during cold season. The mean influent concentrations

of carbamazepine, bisoprolol, diclofenac, and bisphenol A were 0.63, 1.84, 0.43 and 0.95

μg L−1, respectively, where diclofenac was highly detected amongst other refractory

compounds in cold season. In this study, the removal efficiency of individual ECs varied

greatly under MBR treatment. The average removal efficiencies of bisoprolol and

diclofenac were 65 ± 6% and 38 ± 7%, respectively (Figure 11). A decreasing trend in

the removal rates of bisoprolol and diclofenac was observed from Jan-Feb to April, which

might be partially due to the very low sludge temperature inside the bioreactor that might

have affected the biodegradability of the ECs. Despite the low temperature MBR

operation, the bisphenol A removal efficiency was substantially stable (> 97%)

throughout the experimental period, which might be due to its high hydrophobicity that

enhanced its adsorption to biomass and consequently biotransformed. On the other hand,


Figure 11: Removal efficiencies of selected ECs in MBR treatment under

low-temperature periods.

very poor removal efficiency of carbamazepine ranging from -89 to 28% (only for one

occasion) was observed with highest negative removal efficiency during January to

February (Figure 11).

On the other hand, inorganic micropollutants i.e. heavy metals including Zn, Ni, Co, Cu,

As, Pb, Cd, and V, were also consistently detected in influent, permeate and sludge

samples over the experimental period. The average concentration of trace metals in the

incoming wastewater was found in the range of 0.1 to 80 µg L-1, with their relative

abundance in the order of Zn> Ni > Co ≥ Cu >V > Cr > As ≥ Pb > Cd. However, the

removal efficiencies of the trace metals greatly varied from element to element in the

MBR treatment. An average removal efficiency of ≥ 80% was achieved for Cd, Pb, and

V, while ≥ 60% was observed for Zn and Co. Whereas, comparatively lower removal

efficiencies (30 - 50%) were observed for Ni, Cu, As, and Cr under the identical operating

conditions. In this study, chemical precipitation and biosorption mechanisms were

expected to be dominant for the removal of trace metals in the MBR process. Moreover,

metal complexation via metal-ligand co-ordinate bonding and the biofilm layer developed

on the membrane surface could have contributed to the retention/adsorption of insoluble

metal species to some extent other than solubility and stable or unstable forms of trace

metals in the MBR treatment (Bolzonella et al., 2010).

61

Since trace metals are non-biodegradables, their incoming load in MBR process can either

ends up into the sludge or remains in the treated effluent. Therefore, it is crucial to assess

the concentration of trace metals in both treated effluent and produced sludge, especially

when they have some specific utilization. In this study, relatively abundant concentrations

of trace metals were observed in MBR effluent and sludge samples, although their

concentrations were observed below the allowable concentrations limit specified in the

EU and US regulations.

7.2 Assessing the performance of MBR under varying solid retention

times- fate and removal of ECs (Paper II)

Two different SRTs: 60 days and 21 days, were selected to mainly study the removal

efficiency of 23 ECs, including PhACs (antibiotics, β-blockers, analgesics, diuretics,

psychostimulants, antiepileptics, immunosuppressives, anticoagulants) and steroid

hormones (testosterone, progesterone, estrone, and estroil), commonly detected in

municipal wastewater. Overall, a consistently high removal efficiencies of conventional

wastewater pollutants such as TSS, COD, NH4-N, and TP were observed at both SRTs.

Majority of ECs detected in the influent wastewater during the testing SRT periods of 60

and 21 days revealed similar tendencies. Of the 23 ECs, compounds such as iopamidol,

furosemide, caffeine, hydrochlorothiazide, paracetamol, naproxen, ibuprofen and

diclofenac were detected in high levels, thus accounted for more than 50% of the total

ECs. On the other hand, steroid hormones were observed at relatively low levels as

compared to other PhACs.

The removal of selected ECs varied greatly during the MBR treatment process under the

maintained operating conditions. However, the concentrations of many ECs in the treated

effluent were below the limit of quantification (LOQ), which indicates their efficient

removal under MBR treatment regardless of SRT variations. In this study, 17 ECs,

including 13 PhACs (iopamidol, atenolol, tetracycline, caffeine, bisoprolol,

ciprofloxacin, enalapril, ketoprofen, trimethoprim, paracetamol, naproxen, ibuprofen,

and hydrocortisone) and four steroid hormones (estriol, testosterone, estrone,

progesterone) showed the median removal efficiency of more than 90% (90 to ≥ 99.7%).

Three PhACs including metoprolol, furosemide and propranolol showed median removal

efficiencies in the range of 50 – 87%, diclofenac showed poor removal of 39 – 46%, and

two PhACs (carbamazepine and hydrochlorothiazide) were very poorly removed (-8 to

10%). The fate of anti-coagulant (iopamidol) and immunosuppressive (hydrocortisone)

was studied for the first time in the MBR treatment. Out of four β-blockers, the removal

rates of metoprolol and propranolol were slightly higher at longer SRT of 60 days (82 -

84%) than in shorter SRT of 21 days (50 - 60%). Even though the attenuation of

hydrochlorothiazide was very low, its median removal efficiency improved from -1 % at

SRT of 21 days to 10% at SRT of 60 days. Similarly, the median removal efficiency of

carbamazepine enhanced to 2% from - 8% while SRT varied from 21 to 60 days. In most

of the occasions, concentrations of hydrochlorothiazide and carbamazepine in the treated


effluent were higher than in the influent, therefore showing negative removal efficiency.

The major reason for the phenomenon of ‘negative removal’ was ascribed to the

transformation of metabolites, such as glucuronide-conjugates back to the parent

compounds during the biological process, and eventually released with higher

concentrations in the treated effluent (Jelic et al., 2011).

Furthermore, to understand the pathway of ECs removal, mass-balance study was carried

out as described by equations (2)-(5). The shared proportions of ECs relative to the

influent load based on biotransformation, sorption onto sludge and remained in the treated

effluent are elucidated in Figure 12. The dominant mechanism for ECs removal was

presumably via biosorption and biotransformation. The ECs, including atenolol, caffeine,

paracetamol, naproxen, ibuprofen, esteriol and progesterone were removed significantly

with biotransformation rate of 83 to ≥ 99.9%. High biotransformation rate of 58 - 92%

was assigned for iopamidol, furosemide and trimethoprim. Slightly enhanced

biotransformation of metoprolol, hydrochlorothiazide and propranolol was observed at

SRT of 60 days than at SRT of 21 days. On the other hand, some of the ECs were found

to have been removed via sorption onto the biomass (sludge). More than 90% removal

rate was accounted for ciprofloxacin, which could pose serious risk if the excess sludge

has multiple utilization. Some other ECs, such as propranolol, carbamazepine and

testosterone, showed about 10 - 50% sorption efficiency, whereas most of the ECs posed

a very weak sorption affinity to the biomass with the sorption efficiency ranging from 0

- 21%. The physicochemical properties of individual ECs, such as hydrophobicity

(hydrophilic or hydrophobic, characterized by log D values), degree of ionization (pKa),

sorption coefficient (log Kd), could have additionally influenced their removal in MBR

treatment (Cirja et al., 2008b).

The calculated values of log Kd revealed that the selected ECs in MBR treatment were

removed either predominantly via sorption process or via simultaneous biodegradation

followed by partitioning onto sludge (Guerra et al., 2014). However, it was observed that

the calculated sorption coefficients were not correlated with the degree of

biotransformation for many ECs, which indicates that log Kd values of ECs may vary

plant-to-plant depending on different operating conditions (Kim et al., 2014).

Furthermore, noticeable changes in the concentrations of biopolymers such as SMP and

EPS were observed while changing the SRTs. The SMP concentration increased from 5.6

to 15.5 mg (g MLVSS)-1 (about 3 times) and EPS concentration increased from 51 to 103

mg (g MLVSS)-1 while SRT changed from 60 to 21 days, respectively. Similarly, it was

revealed that the dewaterability (measured in terms of CST) of the MBR sludge correlates

well with the SRT values. The sludge with high CST values possesses difficulty in

dewatering due to the presence of more cell bound water.

63

Figure 12: Fate of the selected ECs during MBR treatment at two different SRTs.

7.3 Assessing the AOPs as integration alternatives to MBR to enhance

the removal of ECs.

It was clear from the above studies that the removal of many ECs was not optimal

although they meet most of the recent limit standards. This indicates that the MBR

treatment itself is not the ultimate solution to remediate those highly recalcitrant ECs,

which advocates the need of highly efficient techniques to meet the future discharge

limits. In recent years, developing a more robust treatment processes to ensure high

quality reclaimed effluent from municipal wastewater, especially with the focus on the

removal of ECs has gained increasing interests. In this context, we studied the potential

integration options of two AOP based green technologies: ECO and PCO processes, to

further enhance the removal of ECs from MBR effluent.


7.3.1 Study of ECO process for carbamazepine removal using Ti/Ta2O5-SnO2

electrode (Paper III)

In this study, for the first time, the removal efficiency of one of the highly recalcitrant

PhACs, carbamazepine, in synthetic solutions and the real MBR effluent was assessed

via ECO process using newly developed Ti/Ta2O5-SnO2 electrode. The characterization

of Ti/Ta2O5-SnO2 electrode revealed that the electrocatalytic activity was enhanced due

to doped metallic impurities on the Ti substrate, roughness factor of 12.5 as compared to

smooth surface oxide, and high-overvoltage potentials (appeared between 1.7 and 2 V vs.

SCE) that endorsed the generation of powerful •OH radicals and subsequent oxidation of

carbamazepine. It was also observed that carbamazepine molecules oxidized at the

potential of 1.45 V vs. SCE with distinct anodic current peak. Moreover, the influence of

four major operating parameters, such as current density, initial concentration of

carbamazepine, initial pH of solution and reaction temperature, were also studied.

The current densities of 1, 5, 7, 9, 15 and 17 mA cm-2 were tested to electrolyze

carbamazepine (20 mg L-1) with supporting 0.1 M Na2SO4 electrolyte for 8 h (Figure

13a). With the increasing current density from 1 to 9 mA cm-2, carbamazepine removals

reached about 53 - 71.7%. However, beyond the current density of 9 mA cm-2, no

significant increase in either removal efficiency or kinetics was observed. The reasons

might be attributed to the parasitic reactions of •OH radicals to O2 gas evolution and mass

transfer rate of pollutants towards electrode surface (Song et al., 2010). Furthermore, the

electrical energy consumption to treat one cubic meter of water contaminated with

carbamazepine by electrochemical treatment for 8 h using Ti/Ta2O5-SnO2 was observed

in the range of 3.2 - 172.7 kWh m-3, when current densities varied from 1 - 17 mA cm-2.

The energy consumption was increased from 3.2 - 60.3 kWh m-3 to achieve the

carbamazepine removal efficiency of 53 - 71.7 % with corresponding current density

applied from 1 to 9 mA cm-2. Beyond the current density of 9 mA cm-2, it was noticed

that the energy consumption increased by a factor of more than 2 for achieving the

comparable degree of carbamazepine removal efficiency. Therefore, 9 mA cm-2 was

selected as an optimal current density due to comparably high degree of carbamazepine

removal efficiency and less energy consumption. The reaction kinetics fitted both pseudo-

first order and pseudo-second order model, but latter described better electro-degradation

of carbamazepine with slightly higher regression coefficients (R2). To study the effect of

initial concentration of carbamazepine on the electrochemical treatment, four different

batches of carbamazepine solutions, such as 2, 10, 20 and 30 mg L-1 were electrolyzed

under the optimized current density (9 mA cm-2) (Figure 13b). After the 8 h of

electrolysis, the removal efficiency reached about 80.3%, 73.3%, 71.8% and 70.2% at

initial concentrations of 2, 10, 20 and 30 mg L-1, respectively. The results indicated that

at low initial concentration of substrates, the electrochemical treatment is more efficient.

This might be due to the reason that at low initial concentrations of pollutants,

electrochemical oxidation of organic molecules can be faster than the diffusion of side-

products (Dai et al., 2014).

65

Similarly, to assess the effect of initial solution pH on the electrochemical treatment of

carbamazepine, acidic to alkaline conditions with the pH values of 2, 4, 8 and 10, were

studied (Figure 13c). The results showed that initial pH conditions do not have that

significant effect on the removal efficiency of carbamazepine under electrochemical

treatment. Nevertheless, optimal removal efficiency of 75.2% was achieved at pH 6,

which is very relevant since MBR effluent normally observed at the pH of 6 - 7. On the

other hand, effect of temperature was assessed by electrolyzing carbamazepine solution

at different temperature conditions ranging from 10 - 30°C (Figure 13d). The

carbamazepine removal efficiency of 71.7%, 74.1% and 75.5% were achieved at 10, 20

and 30°C, respectively under the electrochemical treatment for 8 h, indicating the slightly

enhanced trend of treatment efficiency with the increasing solution temperature.

Furthermore, under the optimal current density of 9 mA cm−2 and pH of 6.0, the TOC

removal of carbamazepine were 61.2%, 69.8% and 71.1% at solution temperature of 10

°C, 20 °C and 30 °C, respectively. Since the solution temperature during this work was

about 11 °C, the average degree of mineralization was expected about 61 %. In order to

study the effect of electrode type, the carbamazepine removal efficiency of as-prepared

Ti/Ta2O5-SnO2 and commercially available Ti/PbO2 (Magneto, Netherland) was

compared. After the 8 h of electrolysis, the carbamazepine removal efficiency at

Ti/Ta2O5-SnO2 and Ti/PbO2 anodes reached 71.7% and 77.9%, respectively, showing

relatively high removal efficiency of commercial electrode. However, a significant

leaching of PbO2 was observed when extending electrolysis to 16 h and 24 h, releasing

Pb concentrations of 37.3 and 25.8 μg L−1, respectively, which is higher than the

maximum allowable concentration of Pb (14 μg L−1) in inland and other surface water

(EU, 2013). On the other hand, leaching of Sn, Ti and Ta was observed in the range of

145.7 - 407.3 μg L−1, 53.6 - 92.9 μg L−1 and 2 - 14 μg L−1, respectively while electrolyzing

for 24 h. The high dissolution of Sn to bulk solution was followed respectively by Ti and

Ta, yet there are no specific regulations on these compounds in surface water sources.

Since the removal efficiencies of either electrodes are quite comparable, the as-prepared

Ti/Ta2O5-SnO2 is eco-friendly with no heavy metal leaching into the treated water.


Figure 13: Effect of operating conditions (a) applied current densities; (b) initial

carbamazepine concentrations; (c) initial solution pH; and (d) temperature on

the electrochemical oxidation of carbamazepine.

Furthermore, the optimal conditions (current density= 9 mA cm−2; pH= 6; T= 11 ± 1 °C;

at Ti/Ta2O5-SnO2 electrodes) for the removal of carbamazepine from synthetic solution

were applied to treat the real MBR effluent spiked with carbamazepine (10.75 μg L−1). A

complete carbamazepine removal efficiency of >99.9% was achieved after the 4h of

electrolysis, which required lowest energy consumption of about 57 kWh m-3. The results

revealed that the ECO process based on the use of newly developed Ti/Ta2O5-SnO2

semiconducting electrode is a reliable process to remediate carbamazepine from

contaminated water streams, with promising potential to be integrated with MBR

technology.

7.3.2 Study of PCO process for PhACs removal using Ag2O/P-25 photocatalyst

(Paper IV)

This study advocates the photocatalytic removal of commonly detected PhACs in

municipal wastewater, such as carbamazepine and diclofenac, using newly developed UV

light driven 5% Ag2O/P-25 photocatalyst for the first time ever. Two different water

67

Figure 14: FTIR spectrum (a) and plots of (F (R∞) h𝑣)1/2 vs (h𝑣) for the

approximation of the optical band gap (b) of 5% Ag2O/P-25.

matrices, including deionized water (DW) and real MBR effluent (RME) contaminated

with selected PhACs, were tested. For the preliminary screening experiments 0.4 g L-1 of

four different photocatalysts including Ag2O/P-25, Sn3O4/P-25, PdO/P-25 and NiO/P-25

were dispersed into 50 mg L-1 of carbamazepine and diclofenac and the extent of

adsorption under dark conditions, UV-photolysis and UV irradiated photocatalysis were

studied. In any case, the removal of carbamazepine and diclofenac via adsorption alone

was only about 2 – 9%. Similarly, UV-photolysis for 180 min showed the carbamazepine

and diclofenac degradation rates of nearly 13% and 25%, respectively. Whereas,

photocatalytic degradation under UV irradiation for 180 min using four as prepared

photocatalysts showed significant attenuation of carbamazepine and diclofenac. Since 5%

Ag2O/P-25 revealed a comparatively higher removal efficiency of the selected PhACs

than other photocatalysts, hence it was selected for the subsequent experiments.

The characterization of 5% Ag2O/P-25 by SEM and EDX showed the agglomeration and

homogenous dispersion of Ag+ nanoparticles on the surface of P-25. The FTIR analysis

showed the main absorption peaks at 3000, 1533.67, 1420, 1050 and 433.33 cm-1,

supporting the presence of chemical bonds (O-H stretching vibration, stretching of C=O

group or O-H bending and M-O stretching vibration, respectively) in the bulk

compositions (Figure 14a). The BET analysis indicated that the surface area of P-25

increased slightly by the deposition of Ag2O nanoparticles from 47.27 to 54.1 m2 g-1. The

energy band gap estimated based on the reflectance spectra of 5% Ag2O/P-25 was about

2.96 eV, which revealed the high possibility of harnessing absorption spectra in UV-vis

range (Figure 14b).

a) b)


In this study, operating parameters involved in the photocatalytic removal of selected

PhACs using 5% Ag2O/P-25, such as solution matrix, catalyst dosage, and initial PhACs

concentration were experimented as shown in Figure 15. The removal of carbamazepine

and diclofenac in two solution matrices (DW and RME) was assessed at the catalyst

dosage ranging from 0.2 to 1.0 g L-1. In DW matrix (Figure 15 a-d), carbamazepine and

diclofenac removal efficiencies reached 80.4 - 89.1% and 87.5 – 93.5% at catalyst dosage

ranging from 0.2 - 0.4 g L-1, respectively. Since the removal efficiency of both

carbamazepine and diclofenac did not improve significantly beyond the catalyst dose of

0.4 g L-1, so this value was considered as optimal dose for the photocatalytic removal of

both PhACs in DW matrix. For RME matrix, the removal efficiency of carbamazepine

reached 89.74% from 76.6% while increasing catalyst dosage from 0.4 - 0.8 g L-1, but no

significant improvement was observed beyond 0.8 g L-1. Whereas, optimal diclofenac

removal efficiency of 90.7% was achieved at catalyst dose of 0.6 g L-1. Therefore, the

optimal dosage of 5% Ag2O/P-25 in RME matrix to achieve comparable extent of

carbamazepine and diclofenac removal efficiency as in DW matrix, was increased nearly

by 2 and 1.5-fold, respectively. Furthermore, the effect of initial concentration on the

photocatalytic removal of PhACs by 5% Ag2O/P-25 was studied in RME matrix (Figure

15 e-f). The decreasing tendency in the removal efficiency of carbamazepine and

diclofenac was observed while gradually increasing their initial concentrations. The

removal efficiency of carbamazepine and diclofenac decreased about 38% and 23% while

initial concentrations were increased from 20 to 50 mg L-1. Moreover, the photocatalytic

degradation kinetics of PhACs at 20 min of initial reaction fitted the Langmuir-

Hinshelwood kinetics model very well with regression coefficients of the linear fitting ≥

98% for both carbamazepine and diclofenac. The mineralization efficiencies of about

68% and 65% were observed during the photocatalytic degradation of carbamazepine in

DW and RME matrices, respectively, whereas slightly lower mineralization efficiencies

of diclofenac i.e., 60% and 55% were accounted in DW and RME matrices, respectively.

69

Figure 15: Effect of catalyst dosage in different water matrices on photocatalytic

removal of (a & b) carbamazepine in DW and RME; (c & d) diclofenac in DW and

RME; and effect of initial PhACs concentration (e) carbamazepine in RME; (f)

diclofenac in RME.


Figure 16. Cyclic reusability of 5% Ag2O/P-25 photocatalyst during

the photocatalytic oxidation of (a) carbamazepine and (b) diclofenac.

The quenching experiments indicated that the addition of ethylenediaminetetraacetic acid

(EDTA) and methanol (for suppressing h+ and •OH), the photocatalytic degradation of

carbamazepine was substantially reduced nearly by 70% and 90% in 60 min, respectively,

whereas only about 18% reduction was observed with the O2-purging (suppressing •O2-).

On the other hand, diclofenac removal was substantially decreased to 62% in the presence

of EDTA, but no significant effect in removal efficiency (~ 7%) was observed due to

methanol and O2-purging. The results indicated that the induced h+ and •OH are the

dominant reactive oxidation species during the photocatalytic degradation of

carbamazepine, whereas h+ mainly attributed to diclofenac degradation.

Similarly, to assess the reusability of 5% Ag2O/P-25, cyclic photocatalytic experiments

were conducted for both the PhACs (Figure 16). The 5% Ag2O/P-25 did not show any

significant loss in the photocatalytic degradation efficiency of carbamazepine until five

repetitive cycles, exhibiting a remarkable reusability. Whereas, gradually decreased

photocatalytic removal efficiency (dropped by about 35% until 5th cycle) of diclofenac

was observed during cyclic reuse of 5% Ag2O/P-25. Nonetheless, since the inhibition rate

of the photocatalytic activity was almost stable after 2nd cycle, it can be concluded that

the as-prepared 5% Ag2O/P-25 is moderately stable for the attenuation of diclofenac and

could be useful for the long-term industrial application. Moreover, in either cases, the

recovery of photocatalyst was achieved by centrifuging the solution just for 15 min, which

is probably the more cost efficient than high filtration costs associated with TiO2

nanoparticles.

71

The intermediate identification study on the photocatalytic degradation of carbamazepine

(one of the selected PhACs) using high-resolution FT-ICR mass spectrometry detected

the common transformation products of carbamazepine, such as 10,11- dihydro-

carbamazepine -10,11-epoxide (m/z 253), acridine-9-carboxaldehyde (m/z 208), and

hydroxyacridine-9-carboxaldehyde (m/z 224). The identification of many intermediate

products indicated the complexities involved during the photocatalysis of carbamazepine.

Moreover, a possible multi-step carbamazepine photodegradation path was proposed

based on the unambiguously identified intermediate products in this study and suggested

in the other literatures as shown in Figure 17.


Figure 17: Photogenerated intermediates and proposed photodegradation pathway

of carbamazepine by 5% Ag2O/P-25 photocatalyst under UV irradiation.

73

8 Conclusions and further research

MBRs are very potential technology for municipal wastewater treatment and reclamation.

However, the depth of understandings on their performance and operational

consequences in the regions where low temperature conditions remain for longer period

of the year are very limited. For this aim, the pilot scale MBR plant was operated under

varying operating conditions to study the membrane fouling and efficiency of MBR on

removing ECs. It is noticed that severe membrane fouling is associated with MBR

operation during cold water conditions. However, MBR showed excellent removal of

ECs, heavy metals and human enteric viruses even under extreme cold water conditions

including other common pollutants. Diverse removal efficiencies of ECs were observed

during the MBR treatment and several influencing factors were evaluated.

Biotransformation and sorption to sludge were found to be the predominating reasons of

the removal of ECs, although metal complexation, retention by membrane fouling layer

and membrane material have also attributed on the attenuation of heavy metals and

viruses. Moreover, physicochemical properties of ECs and operational conditions had

also contributed on removal by MBR treatment. Overall, a very good quality of treated

effluent was achieved by MBR treatment despite of varying operating conditions, even

though the removal efficiencies of many ECs were not optimal.

This indicated that the MBR treatment is not the ultimate solution to remediate many of

highly recalcitrant ECs. Therefore, integration possibilities of MBR with some AOPs,

viz. ECO and PCO to ensure high quality reclaimed effluent with emphasizing the

removal efficiency of ECs from municipal wastewater treatment, were explored. In ECO,

a newly developed Ti/Ta2O5-SnO2 electrode was used for the electrochemical degradation

of carbamazepine. The optimal removal of carbamazepine in aqueous solution was

reached 75.5% when electrolyzed under the current density of 9 mA cm-2, pH of 6, 30 °C

for 8 hrs. A complete removal of carbamazepine in real MBR effluent was observed when

electrolyzed under the optimized conditions in aqueous solutions. On the other hand, In

PCO, heterogenous 5% Ag2O/P-25 photocatalyst was used for the photocatalytic

degradation of carbamazepine and diclofenac from aqueous solutions and real MBR

effluent matrices. About 89.1% and 93,5% removal of carbamazepine and diclofenac was

achieved using 5% Ag2O/P-25 at catalyst dosage of 0.4 g L-1 under UV irradiation for

180 min. However, to achieve the same degree of attenuation, the optimal catalyst

dosages for both carbamazepine and diclofenac were increased by the factor of 2 and 1.5,

respectively. The photo-induced holes and ●OH were the primary oxidation species

involved in the photocatalytic reactions. Moreover, some important degradation by-

products of carbamazepine were identified and quantified, and multi-step degradation

pathway was proposed. Therefore, the batch experiments on ECO and PCO systems

revealed the enhanced removal of selected ECs than in the MBR treatment alone, which

indicated their potentials of integrating with MBR technology to improve effluent quality

and the better stability of the treatment process.

74 8 Conclusions and further research

This dissertation has expanded the knowledge on the applicability and efficiency of MBR

operation at rather low water temperature conditions in Nordic environment and

integration possibilities of MBR with selected AOPs. However, it has also unfolded the

need of many other important topics for further investigations. Of the many research

topics related to the current work, some of them are briefly described below.

As substantial membrane fouling was observed in MBR during winter periods, more

research could be focused on finding fouling mitigation strategies. On this note, feasibility

of integrating MBR with electrocoagulation can be studied, which is expected to

simultaneously destroy major organic fouling fractions (e.g. EPS and SMP), enhance ECs

removal efficiency, and generate less sludge. Similarly, the ability of low-cost adsorbents,

such as biochars derived from wasted sludge, could be studied as subsequent fouling

mitigation approach. On the other side, scaling-up of integrated MBR-AOPs options

could be studied in a pilot-scale. Further, optimizing the operation of pilot-scale

integrated system, regeneration of electrodes and photocatalysts, and overall cost

evaluation of the system should be investigated. Similarly, toxicity associated with the

treated effluent after AOPs could be a challenge due to the generation of more toxic

transformation by-products. Therefore, further studies should be focused on detail

identification of the generated by-products of parent ECs and their relative toxicities in

the treated effluent.

75

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Publication I

Khum Gurung, Mohamed Chaker Ncibi, Mika Sillanpää

Assessing membrane fouling and the performance of pilot-scale membrane

bioreactor (MBR) to treat real municipal wastewater during winter season

in Nordic regions

Reprinted with permission from

Science of the Total Environment

Vol. 579, pp. 1289-1297, 2017

© 2017 Elsevier

Assessing membrane fouling and the performance of pilot-scalemembrane bioreactor (MBR) to treat real municipal wastewater duringwinter season in Nordic regions

Khum Gurung a,⁎, Mohamed Chaker Ncibi a, Mika Sillanpää a,b

a Laboratory of Green Chemistry, School of Engineering Science, Lappeenranta University of Technology, Sammonkatu 12, FI-50130 Mikkeli, Finlandb Department of Civil and Environmental Engineering, Florida International University, Miami FL-33174, USA

H I G H L I G H T S

• Performance of a pilot-scale MBR wasstudied at low temperatures in Nordicregion.

• Substantial membrane fouling was ob-served at low temperatures (below10 °C).

• About 75% of the permeability drop wasobserved due to fouling.

• High removal rates were achieved forpathogens.

• High removal rates for the emergingmicropollutants were observed.

G R A P H I C A L A B S T R A C T

a b s t r a c ta r t i c l e i n f o

Article history:Received 23 October 2016Received in revised form 17 November 2016Accepted 17 November 2016Available online 29 November 2016

Editor: Jay Gan

In this study, the performance of a pilot-scale membrane bioreactor (MBR) to treat real municipal wastewaterwas assessed at low temperatures (7 to 20 °C) in Nordic regions. First, the effect of low temperatures on mem-brane fouling was evaluated by monitoring trans-membrane pressure. A significant membrane fouling was ob-served when the sludge temperature inside the MBR unit was below 10 °C with a 75% permeability drop, thusindicating high deterioration of the membrane performance at low temperatures. Moreover, increasing valuesof sludge volume index (SVI) during low temperatures showed high deterioration of sludge settleability. As forthe pollution removal, MBR achieved high performances primarily for pathogens and emergingmicropollutants.The average log reductions of 1.82, 3.02, and 1.94 log units were achieved for norovirus GI, norovirus GII, and ad-enoviruses, respectively. Among the four trace organic compounds (TrOCs), the average removal efficiencies ofbisoprolol, diclofenac and bisphenol A were 65%, 38%, and N97%, respectively. However, carbamazepine wasnot efficiently removed (‐89% to 28%). Regarding trace metals, an average removal of N80% was achieved forCd, Pb, and V. For the rest of the metals, the removal capacities were between 30 and 60%.

© 2016 Elsevier B.V. All rights reserved.

Keywords:Membrane bioreactorMembrane foulingLow temperatureMunicipal wastewaterPathogensEmerging micropollutants

Science of the Total Environment 579 (2017) 1289–1297

⁎ Corresponding author.E-mail addresses: [email protected] (K. Gurung), [email protected] (M. Sillanpää).

http://dx.doi.org/10.1016/j.scitotenv.2016.11.1220048-9697/© 2016 Elsevier B.V. All rights reserved.

Contents lists available at ScienceDirect


j ourna l homepage: www.e lsev ie r .com/ locate /sc i totenv

1. Introduction

Temperature influences theperformanceof biological treatment in con-ventional activated sludgeprocesses (Arévalo et al., 2014;Metcalf andEddyInc., 2003).Moreover, inMBRs, temperaturenot only affects the bioconver-sion process but also influences microbial community, fouling rate andsludge morphology (Zhang et al., 2014). The biological activity of the acti-vated sludge tend todecrease inwinter periods. Indeed,when temperaturedrops to about 15 °C, methane-producing bacteria become essentially dor-mant, and at about 5 °C, the autotrophic bacteria practically cease function-ing. Even the chemoheterotrophic bacteria performing on carbonaceousmaterial becomes inactive at 2 °C (Metcalf and Eddy Inc., 2003).

Membrane fouling, which is the potential deposition of the feedstream on the membranes, is one of the major causes behind the dete-rioration of MBR performance (permeability). In this context, it was re-ported that the low temperatures can cause rapid membrane fouling ascompared to mesophilic temperature under which proteins are thedominant foulants (Gao et al., 2014). There has been many previousstudies on membrane fouling (Drews, 2010; van den Brink et al.,2011; Zhang et al., 2014). Membrane fouling at low temperatures arealso found to induce changes in the cake layer thickness and/or porosityof the membranes (van den Brink et al., 2011). Moreover, other factorsattributed to the membrane fouling at low temperature include in-creased viscosity, reduced sludge stabilization, reduced particle size ofsludge and reduced mass transfer efficiency (Zhang et al., 2014).

The fate of human enteric viruses such as noroviruses and adenovi-ruses are of major interest because they are common causes of water-borne outbreaks. In MBR system, the removal mechanisms ofpathogens can be either via attachment to biofilm developed on themembrane surface or adsorption in/onto the biomass flocs and tosome extent via retention by micro-membranes (Miura et al., 2015).

The active evaluation of possible risk of emerging organicmicropollutants or TrOCs such as pharmaceutically active compounds(PhACs) and endocrine disrupting chemicals (EDCs) in treated effluents ofwastewater treatment systems has become a great concern (Trinh et al.,2012). In winter seasons in cold regions, higher concentrations of emergingmicropollutants aredetecteddue to thedecrease inbacterial biodegradationactivities under low temperature (Kruglova et al., 2016), in addition to theincreasing consumption of pharmaceutical drugs, especially antibiotics, inwinters (Suda et al., 2014). In Finland, carbamazepine (CBZ), diclofenac(DCF), and bisoprolol (BIS) are three commonly used PhACs ((FIMEA),2015). The toxicity of these PhACs on fish, algae, and bacteria in concentra-tions b1 mg L−1 have been recently reported (Kruglova et al., 2016).

On theother hand, tracemetals are recalcitrance andpersistence inor-ganic micropollutants (Fu and Wang, 2011) that could cause acute andchronic toxicity (Karvelas et al., 2003). Moreover, their presence at cer-tain concentrations is not only of environmental concern but also strong-ly reduces microbial activity, thus resulting adverse effects on biologicalwastewater treatment processes (Chipasa, 2003). In MBRs, the removalof trace metals are highly enhanced primarily due to the reduction insuspended solids from effluents (Di Fabio et al., 2013).

In this context, the present research aims at studying the perfor-mance of a submerged pilot MBR to treat real municipal wastewaterunder the low temperature conditions during winter season in Finland.The investigation mainly focused on assessing the membrane fouling,pathogens removal and fate of emerging micropollutants (organic andinorganic). To the best of our knowledge, this is the first detailedstudy on MBR treating real municipal wastewater under low tempera-ture periods in Nordic environment.

2. Materials and methods

2.1. The pilot scale MBR plant and operating conditions

In this research, a pilot-scale MBR was employed to treat a real mu-nicipal wastewater inMikkeli Finland. The detailed experimental set-up

arrangement of the MBR plant was explained elsewhere (Gurung et al.,2016). The capacity of theMBR pilot plant was designed to treat 3m3 ofwastewater per day. The MBR unit used flat sheet microfiltration mem-branes having pore size of 0.4 μm and effective surface area of 16 m2.The MBR unit was fed with primary clarified wastewater of the full-scale Kenkäveronniemi WWTP. The stabilized sludge was wasted frommembrane compartment daily to maintain designated sludge retentiontime (SRT). The treated permeate water and wasted sludge were col-lected in the permeate and sludge tanks, respectively and thendischarged to the WWTP sewer.

TheMBR pilot unit was operated for N120 days from January to April2016, coinciding with the winter season in Finland (temperature of in-fluent wastewater around 7–10 °C). The water level in the membranetank was maintained constant (b5% change in water volume duringthe study). No backwashing was performed, however, the chemicalwashings of membranes were carried out whenever the TMP reacheda level of ≤20 kPa by using sodium hypochlorite followed by citricacid. The pH of the biological reactors was adjusted using NaOH solu-tion. During the manual cleaning, membrane cartridges were takenout of the bioreactor and then accumulated cake layer was cleaned byusing brush and water jet. The operating conditions of pilot operationis summarized in Table 1.

2.2. Analytical methods

The major pollutants in wastewaters such as total suspended solids(TSS), chemical oxygen demand (COD), ammonium nitrogen (NH4-N),and total phosphorus (TP) in influent and permeate were determinedaccording to the Standard Methods (American Public HealthAssociation et al., 1999). Liquid samples were collected for 24 h usingautomatic composite samplers in weekly basis. The SVI was determinedweekly frommixed liquor grab samples according to the previously de-scribed method (Fontmorin and Sillanpää, 2015). The mixed liquorsuspended solids (MLSS) concentrations of mixed liquor from both aer-obic and membrane bioreactors were determined once a week accord-ing to Standard Methods (American Public Health Association et al.,1999).

Membrane fouling was quantified as a progressing phenomenon bymonitoring trans-membrane pressure (TMP), while the permeate fluxwas maintained almost constant over the experimental period. Mem-brane performance was assessed according to the following equation(Park et al., 2015):

Lp ¼ JTMP

¼ 1μw � Rt

ð1Þ

Lp ismembrane permeability (Lm−2 h−1 bar−1); J ismembraneflux(L m−2 h−1); TMP is trans-membrane pressure (Pa); μw is permeateviscosity (Pa·s); Rt is total filtration resistance (m−1).

The analysis of noroviruses (GI and GII) and adenoviruses were per-formed from influent, permeate and mixed liquor according to themethods described elsewhere (Miura et al., 2015; Simmons et al.,

Table 1Operational conditions of the MBR pilot plant.

Parameter Units Values

Operational period Days 120Hydraulic retention time (HRT) Hours 35Sludge retention time (SRT) Days 25–30Flux (continuous) L m−2 h−1 7.80MLSS concentration mg L−1 5300–9800F/M ratio kg COD/kg MLSS 0.02–0.05Aeration intensity m−3 m−2 h−1 0.4–0.6Sludge temperature °C 6.5–21/7.0-20pH Unitless 6.6–7.3Suction cycle Minutes 9-ON/1-OFF

1290 K. Gurung et al. / Science of the Total Environment 579 (2017) 1289–1297

2011).Moreover, the concentrations of E. coli and Enterococcus in the in-fluent, sludge and permeate were enumerated by using Colilert® andEnterolert® methods.

The occurrence of TrOCs both in liquid and solid phases were ana-lyzed from influent, permeate and sludge, respectively using standardsolid phase extraction (SPE) and LC-MS/MS methods (EPA, 2007).

The analysis of heavy metals in the influent, permeate and sludgesamples was conducted with an inductively coupled plasma-opticalemission spectrometer (ICP-OES, iCAP 6300, Thermo Electron Corpora-tion, USA) from the pretreated samples (Fontmorin and Sillanpää,2015).

3. Results and discussion

3.1. MBR operation and membrane fouling at low temperatures

Fig. 1 shows the variations in sludge temperature, TMP profile andaeration intensity duringMBR operation. At the beginning of this exper-iment, the average sludge temperaturewas 11±1.5 °C, the specific aer-ation demand per membrane surface (SADm) was fixed at about0.6 m3 h−1 m−2 and the TMPwas about 2 kPa until 9th February. How-ever, two weeks later, the atmospheric temperature began to dropsharply. The outside air temperature was varying between 0 °C and−33 °C for more than a week, which is the typical Nordic weather phe-nomenon during winter seasons. Due to this sudden temperature drop,some pumps and connecting pipes were frozen and severely clogged.Thus, the system malfunctioned and stopped for more than a week(18th to 25th January). Heating cables were installed to the stretch ofpipes from feed point to the pilot plant inlet. The broken pumps werechanged with new ones and the clogged pipes were cleaned with hotwater. Additionally, three heaters (25 W capacity each): two insidethe pilot plant and one at the influent pumping station were installed

and the process resumed again. The average temperature of mixed li-quor reached 20.5 ± 0.6 °C immediately after the process restarteduntil 8th February, which was expected due to the heating effects ofair mass flow along with heaters and heating cables. Thus, the SADm

was reduced and maintained at 0.4 m3 h−1 m−2 from 9th February to6th April in order to ensure the lowest possible temperature inside thebioreactors. As a consequence, the sludge temperature started decreas-ing gradually from 10th February and reached nearly 10 °C by 18th

March. In addition, the decreasing temperature of the sludge was alsoattributed to the incoming flow of melting ice to the municipal waste-water. Occasionally, heaters and heating cables were intermittentlyused depending on outside air temperature in order to prevent seriousdamages related to freezing.

The first TMP jump occurred between 23rd February and 18thMarch.The temperature shock was evident when temperature dropped toabout 10 °C, which is consistent with the results of Gao et al. (Gao etal., 2014). As the membrane fouling accelerated rapidly and the TMPreached nearly to 17 kPa (1st peak), first chemical washing of mem-branes was performed. After the washing TMP dropped to about6 kPa, but increased rapidly and finally reached nearly to 12 kPa (2nd

peak) within a week from the 1st peak. Yet, the sludge temperaturewasmeasured to be 7.7± 1.2 °C during the 2nd peak and the propensityof the membrane fouling was very rapid. Therefore, it was concludedthat the rapid increase in the TMPwas highly attributed to the low tem-perature of sludge. Besides, due to the Easter holidays, the second in-situ washing of membranes was performed during the 2nd foulingpeak (a bit early when compared to the 1st peak). Again, the TMP in-creased gradually and reached nearly to 13 kPa within 12 days fromthe 2nd peak. The sludge temperature was increased to about 8–12 °Cand stayed almost stable after second chemicalwashing. Themembranefouling occurrence between the 1st peak and 2nd peak was within aweek. However, it extended to nearly two weeks between 2nd and 3rd

Fig. 1.Membrane fouling behavior and cleaning measures. (Note: SADm is not in scale).

1291K. Gurung et al. / Science of the Total Environment 579 (2017) 1289–1297

peak. The extended time between the fouling peaks can be attributed tothe increased sludge temperature. The results showed that membranefouling was important during low temperature periods, which is inagreement with other studies (Gao et al., 2014; van den Brink et al.,2011; Zhang et al., 2014). As the sludge temperature stayed at about12 °C (after 2nd fouling peak) for two consecutive weeks, maintaininglow temperatures in the pilot was no more possible as the weatherstarted to gradually warm up. Therefore, after 3rd fouling peak, theSADm was increased back to 0.6 m3 h−1 m−2. This increase helped re-ducing the TMP by approximately 2.5 kPawithin a week. Thus, the foul-ing phenomena occurring during the MBR operation were not onlyattributed to the low temperature conditions but also, to some extent,related to the reduction in aeration intensity. During the chemicalwash-ings the TMP dropped nearly to 6 kPa and then raised rapidly, illustrat-ing that this cleaning procedure is not so effective. As for the manualwashing of membranes, it reduced the TMP to nearly 3 kPa. Moreover,SADm was maintained at of 0.6 m3 h−1 m−2 and the TMP was almoststable (stayed around 3 kPa until the end of the experiment). In thiscontext, it was reported that when wastewater temperature decreases,the mixed liquor viscosity increases, resulting in the rapid building offouling resistance, leading to the rise of TMP (van den Brink et al.,2011; Zhang et al., 2014). Membrane fouling at low temperatures canalso be accelerated due to the reduced shear stress proclaimed by courseair bubbling, low particle back transport velocity andhigh hydrophobic-ity of themixed liquor (Zhang et al., 2014). In addition, the pore size canshrink in the low temperatures due to intrinsic characteristics of mem-branes. At low temperatures, a higher amount of submicron particlesare found in the mixed liquor composition, which enhances the rapidfouling by pore blocking or narrowing of themembrane pores, however,cake layer formation becomes dominant after the initial fouling (vanden Brink et al., 2011).

At the beginning, the permeability was about 345 L m−2 h−1 bar−1,and the correlation with temperature was not so clear. During the 1st

peak of fouling, the permeability dropped to 50 L m−2 h−1 bar−1.After the 1st membrane cleaning, the permeability increased back to130 L m−2 h−1 bar−1, but rapidly declined to 75 L m−2 h−1 bar−1

until the 2nd peak of fouling. About 75% of permeability drop was ob-served during the period with the lowest temperatures (from 9th

March to 11th April), as compared to the initial permeability. About50% of permeability differences were reported between summer andwinter periods in the full scale MBRs (van den Brink et al., 2011). Afterthe manual cleaning of membranes, the permeability was restoredgradually and reached nearly to 300 L m−2 h−1 bar−1, though unableto regain the initial value. The results showed that the acceleratedmem-brane fouling due to the low temperatures deteriorated the hydraulicperformance (permeability) of membranes. Therefore, extra

membranes are recommended to be installed for the colder periods inorder to ensure constant membrane permeability, especially in coun-tries where long low temperature periods exists throughout a year.Membrane fluxes can be lowered using extra membranes, resulting ina slower propensity of membrane fouling due to the distribution oftrans-membrane pressures. This leads to maintain relatively high andconstant permeability.

Fig. 2 shows the variations of the MLSS concentrations and SVI ofmixed liquor. The average SVI was found to be increasing from120mLg−1 to 168mL g−1when the average sludge temperature variedbetween 15 and 10 °C. The higher SVI values at low operating tempera-ture indicates that the compressibility and settleability of sludge arehighly deteriorated (Zhang et al., 2014). More fine particles, colloids orpolymers are released when operating at low temperatures due toloose morphology of mixed liquor, leading to high tendency of mem-brane fouling (Drews, 2010). More irregular shaped sludge flocs mightbe produced due to bulky sludge which facilitates the development ofhighly dense cake layer on the membrane surface, thus resulting in amore significant fouling phenomenon (Zhang et al., 2014). On theother hand, the MLSS concentration was found to be decreasing fromabout 9800 mg L−1 at the beginning to 5600 mg L−1 when the sludgetemperature dropped by approximately 10 °C (from 20 °C to 10 °C).However, the MLSS concentration remained almost stable when thesludge temperature was about 8 °C to 12 °C. The results are consistentwith previous study where reduced biomass concentrations is reportedduring low temperature operations (Zhang et al., 2014). Under the lowtemperatures less energy is released from the oxidation of organic mat-ter, which lowers the rate of cell synthesis and thus deteriorate thesludge yield (Zhang et al., 2014). Similar observations reported byKrzeminski et al. (2012) revealed that sludge concentrations are highlyreduced in winter than in summer in a full-scale MBR plant.

The treatedwater quality regarding the removal of TSS, organicmat-ter and inorganic nutrients (NH4-N and TP) is shown in Table 2. Despiteof low temperature operation, high quality treated water was achieved.100% removal of TSS was accomplished, thus revealing an excellent re-tention capability from inert suspended solids by the micro-filtration(MF)membranes. About 93% of COD removal was achievedwith the ef-fluent concentration of 18.7 ± 2.9 mg L−1. The present work is in goodagreement with a previous study (Arévalo et al., 2014), in which no sig-nificant effect was observed on the removal of COD under a wide tem-perature variation (9–33 °C), when full-scale MBR was operated atSRT and HRT of 35 days and 35 h, respectively. High removal rate ofNH4-N (nearly 100%) during the experimental period indicated an effi-cient nitrification process in theMBR system, regardless of low temper-ature conditions. Higher nitrification under low temperature operationwas expected due to the relatively longer SRTmaintained during the ex-perimental period. Maximum growth and accumulation of nitrifyingbacteria (nitrifiers) are achieved when MBR is operated at relativelylonger SRT of 35 days which offsets slow growth of nitrifiers due tolow temperature, thus leading to excellent nitrification (Arévalo et al.,2014). Furthermore, the average removal rate of TP reached nearly91% with the concentration of 0.43 mg L−1 in permeate, even thoughno chemical coagulants were added to the MBR system. Therefore, en-hanced biological phosphorus removal process is expected at longerSRT during MBR operation due to the enrichment of some phosphate-accumulating organisms (PAOs).

3.2. Removal of human viruses: Noroviruses (NoV GI and NoV GII), and Ad-enoviruses (AdVs)

NoVsGI andGII, and AdVswere consistently detected in all the influ-ent samples, and their occurrences are elucidated in Fig. 3. The detectionrates varied depending on varieties of samples and behaviors of differ-ent viruses types (Miura et al., 2015). In the present work, the detectedbut not quantified virus concentrationswere assigned as positive resultsFig. 2. Evolution of MLSS concentrations and SVI in MBR process.


for viruses, except those under the limit of detection (i.e. ND: notdetected).

NoV GI was detected in 85% of the influent samples with an averageconcentration of 95 GC mL−1 and high positive results during FebruaryandMarch. NoV GI in themixed liquor samples were below the limit ofquantification with an average concentration of 48± 28 GCmL−1. NoVGI was detected only in two permeate samples, during February andApril. The log removal of NoV GI was ranging from N0.67 to N2.73(1.82 ± 0.6) log units. Similarly, NoV GII was constantly detected in100% of the influent samples with an average concentration of896 GC mL−1 having gradually increasing trend during January toMarch, and then declining tendency during April. NoV GII in themixed liquor samples were mostly below limit of quantification withan average concentration of 25 ± 44 GC mL−1. In permeate, NoV GIIwas detected in three samples with an average concentration of 28 ±37 GC mL−1. The log removal of NoV GII was achieved in the range ofN2.1 to N3.62 ( 3.02± 0.5) log units. On the other hand, AdVs were de-tected in 100% of the influent samples with an average concentration124 GC mL−1. The concentrations of AdVs varied consistently through-out the low temperature period. AdVs were highly detected in themixed liquor with an average concentration of 196 GC mL−1. Interest-ingly, only one positive sample was detected for AdVs (on 23rd March)from the permeate sample. The log removal of AdVs ranged fromN0.76 to N2.76 (1.94 ± 0.7) log units.

The tendency of variations monitored for NoVs and AdVs in the in-fluent wastewater during winter period is consistent to the previousstudies (Kuo et al., 2010; Miura et al., 2015; Simmons et al., 2011). Inthis study, the overall removal efficiencies of NoV GI (N 0.67 to N2.73log unit), NoV GII (N2.1 to N3.62 log units), and AdVs (N0.76 to N2.76log units) 'were achieved'. Miura et al. (2015) reported the log reduc-tions of N0.2 to N3.4 log units for NoV GII. Ottoson et al. (2006) reportedmean log removal of 1.0 to 1.1 log units for NoVs in pilot-scaleMBRwitha 0.45 μm pore size membranes. Sima et al. (2011) reported the log re-moval of 0.9 to ≥6.8 log units for NoVs in full-scaleMBR treatment. Sim-ilarly, in the full-scale MBR with nominal pore size (0.04 μm)membranes, Simmons et al. (Simmons et al., 2011) reported the averagelog removals of 3.0 and 4.7 log units for AdVs and NoV GII, respectively.

Virus adsorption to mixed liquor plays an important role in the re-moval of viruses by MBR process (Miura et al., 2015; Ottoson et al.,

2006; Simmons et al., 2011). In this study, NoV GI was detected underlimit of detection in the mixed liquor, whereas NoV GII was observedat very low concentrations. The log removal of NoV GI and NoV GIIwas about 0.5 log units and 1.7 log units, respectively in the mixed li-quor. One reason might be attributed to the presence of inhibitors inthe samples, which may affect the sensitivity of the analytical methodor even cause false-negative results (Schrader et al., 2012). In thisstudy, the influent to the MBR was fed from primary settling tank ofthe main process (where returned sludge was recycled periodically),which might have traces of calcium ions (inorganic inhibitors). Thepresenceof organic inhibitors such as polysaccharides, humic acids, pro-teins (Schrader et al., 2012),which are highly available during cold tem-perature (Zhang et al., 2014), might have effects on PCR. On the otherhand, AdVs were significantly higher in the mixed liquor samples thanin the influent, showing their high adsorption affinity towardsmixed li-quor (Kuo et al., 2010). However, as the surface characteristics of virusesvaries greatly, therefore their adsorption to the complex biomass mayvary too (Miura et al., 2015). It is likely that viruses which are adsorbedto sludge more efficiently removed as waste sludge in the MBR pilotprocess (Miura et al., 2015). This indicates that even under the low tem-perature conditions, appropriate control of MLSS concentration in MBRprocess can enhance efficient virus removal. Furthermore, differences inthe log removal efficiencies related to the morphology of viruses wasnot clearly observed. The removal of NoV GII (27–40 nm) was foundto be higher than the AdVs (90–100 nm). This result is contradictorywith the findings of Simmons et al. (2011), where higher removal ofhuman adenoviruses was achieved compared to NoV GII in MBR pro-cess. This tendency could be explained by the variations in influent con-centrations in the one hand, and that of the limits of detection on theother hand. The combination of those variations might lead to higheror lower removal efficiency of viruses during the calculation(Simmons et al., 2011).

In the present study, NoV GI, NoV GII and AdVs were detected in theMBR permeate, though the occurrence was not consistent, which con-firms that MBR is not an absolute barrier for retaining virus passage(Simmons et al., 2011). During this research work, variations in themembrane fouling degree were observed with the TMP ranging from2 to 18 kPa. Despite the cleaning of membranes, the overall removal ef-ficiency of viruses in the MBR process did not change. Moreover, pas-sage of all three viruses were observed in the permeate samplesbefore or after the membrane fouling peaks. These results suggestedthat the virus removal efficiency was not related to the degree of mem-brane fouling, which is in good agreement with the previous study(Miura et al., 2015). Further research would be needed for hourly ordaily samples before and after the cleaning of membranes to determinethe precise correlation between fouling and virus removal.

Furthermore, the average concentrations of the Escherichia coli andenterococci were observed in the influent wastewater in the range ofN4.8 × 105 MPN 100 mL−1 and 1.9 × 104 to N4.8 × 105 MPN

Table 2Basic performance of MBR process regarding the removal of main pollution load.

Parameter Influent concentration[mg L−1]

Effluent concentration[mg L−1]

Removal rate[%]

TSS 126.9 ± 19.6 b1 100COD 265.6 ± 53.1 18.7 ± 2.9 92.8 ± 1.4NH4-N 45.9 ± 9.2 b0.02 N99.96TP 4.7 ± 0.8 0.43 ± 0.1 90.6 ± 2.4

Fig. 3. Concentrations of NoV GI, NoV GII, and AdVs. Black triangles - influent; grey diamonds –mixed liquor; white squares - permeate; dashed line - Limit of Quantification; ND - belowLimit of Detection.


100 mL−1, respectively. However, none of them were detected in theMBR permeate over the experimental period. Nonetheless, the highconcentrations of Escherichia coli (1.6 × 103 to 8.05 × 105 MPN100 mL−1) and enterococci (1.07 × 103 to 4.3 × 104 MPN 100 mL−1)were detected in the sludge supernatant. The average bacteria removalefficiency of the MBR process during low temperature period was N5.6log units, which is in the good agreement with other study (Ottosonet al., 2006).

3.3. Fate of organic micropollutants: TrOCs

The occurrence and removal efficiencies of CBZ, BIS, DCF, and BPA intheMBR process are shown in Fig. 4a–b. All the selected TrOCs were de-tected in the influents throughout the experimental period, althoughthe relatively higher concentrations were measured from January upto the end of March as shown in Fig. 4a. The mean influent concentra-tions of CBZ, BIS and BPA were 0.63, 0.43, and 0.95 μg L−1, respectively.Similarly, for DCF, the concentration was 1.84 ± 0.4 μg L−1, which wasthe highest among the other refractory compounds during cold season.

The mean concentrations of CBZ, BIS, DCF and BPA in the MBR per-meate were 0.77, 0.14, 1.12, and b0.05 μg L−1, respectively. Most ofthe TrOCswere found comparatively low in the permeate than in the in-fluent (excluding CBZ), indicating the enhanced efficiency of the MBRprocess to remove many emerging TrOCs from the real wastewaters.

TrOCs removal efficiency differ from plant to plant and even countryto country as it highly depends on wastewater characteristics and spe-cific operating procedures implemented by particular treatment pro-cesses (Tambosi et al., 2010). The overall removal rates in this studyvaried largely among the individual TrOCs. The mean removal efficien-cies of CBZ, BIS, DCF and BPA achieved during low temperatures(interpreted asmonthly average) are shown in Fig. 4b. Negative remov-al efficiencies of CBZ were mostly observed, except a relatively low effi-ciency of about 9% and 28% during last two weeks of March. CBZremoval rates varied from −89 to 28% (average of −25%) with thehighest negative efficiency during January–February. The average re-moval efficiencies of BIS andDCFwere 65±6% and 38±7%, respective-ly. Decreasing tendencies of the removal rates were observed both forBIS and DCF from Jan–Feb to April. On the other hand, BPA concentra-tions in the permeate were significantly below b0.05 μg L−1 (bLOQ) re-gardless of low temperature operation, thus substantially stable BPAremoval efficiency of N97% (N95 to N98%) was achieved.

A mass balance of TrOCs in MBR was performed (Table 3) accordingto the method described by Kimura et al. (2007), where quantity of theTrOCs adsorbed on mixed liquor solids are measured and comparedwith the eliminated quantity estimated by considering the influentminus the permeate concentrations. The presented mass balances

must be interpreted as rough estimates as they were calculated fromthe averaged values considering a completely mixed and steady state.Table 3 clearly elucidates great variations between the overall elimina-tion and the elimination due to sorption regardless of the compounds.For all the selected TrOCs, a very weak removal via biomass sorptionwas observed. Thus, the removal mechanism for BIS, DCF and BPA wasattributed primarily due to the biotransformation of TrOCs in the MBRprocess. Similar studies on MBR demonstrated that sorption is not thedominant mechanism, but biotransformation is crucial in the elimina-tion of TrOCs regardless of the treatment conditions (Clara et al., 2005;Kimura et al., 2007; Trinh et al., 2016). The biological elimination ofTrOCs could be due to the directmetabolization or by co-metabolizationwhere bacteria use TrOCs as direct carbon source, or just break themdown via specific enzymes which having potential to degrade or trans-form trace organic molecules (Vieno and Sillanpää, 2014). It is observedthat membrane filtration alone does not allow any further retention ofTrOCs due to the size exclusion (Clara et al., 2005).

3.3.1. CBZ removalIn the present study, the MBR unit showed insufficient elimination

of CBZ with high concentrations in permeate. In most of the permeatesamples, the concentrations of CBZ were observed even higher than inthe influent samples. Moreover, many researchers have reported CBZas ‘non-removal’ compounds in MBRs by documenting even negativeremoval cases (Clara et al., 2005). The increased amount of CBZ in per-meate samples were most likely due to the conversion of CBZ glucuro-nides and other conjugated metabolites to the parent compound via

Fig. 4. Concentrations of CBZ, BIS, DCF and BPA in influents during low-temperature periods [a]; removal efficiencies achieved for selected TrOCs (monthly average) [b].

Table 3Mean concentrations (±SD) and Mass balances of selected TrOCs in MBR process.

Selected TrOCs

CBZ BIS DCF BPA

Treated water flow (L d−1) 2520Wasted sludge (kg d−1) 0.80LOQ aqueous (μg L−1) 0.05 0.02 0.1 0.05LOQ sludge (μg kg−1) 0.1 0.05 0.2 0.1Influent concentration (μg L−1) 0.63

± 0.20.43± 0.2

1.84± 0.4

0.95± 0.2

Permeate concentration (μg L−1) 0.77± 0.1

0.14± 0.1

1.12± 0.4

b0.05

Mixed liquor concentration(μg kg−1)

0.88± 0.2

0.17± 0.1

2.11± 0.6

0.17± 0.1

Overall Elimination in MBR(μg d−1)

−353 730 1814 2268

Elimination due to sorption(μg d−1)

0.71 0.14 1.70 0.14


enzymatic cleavage in the activated sludge process in MBRs (Vieno etal., 2007). The small percentage of CBZ removal during theMBR processis associatedwith retention by the sludge and elimination during sludgewasting (González-Pérez et al., 2016). In addition, large concentrationsof CBZ found in mixed liquors and permeate suggested the recalcitrantbehavior of CBZ to the biological transformation or biodegradation(González-Pérez et al., 2016). The poor removal of CBZ is partly due toits hydrophilic characteristics (Log Kow b 2.5) and chemical stability(González-Pérez et al., 2016).

3.3.2. DCF removalThe average influent concentration of DCF (1.12 μg L−1) was higher

than the values (0.23 to 1 μg L−1) reported by other researchers in mu-nicipal sewage treatment plants. In this study, the partial removal effi-ciency of diclofenac varied from 13% to 89% (average 38%), which issimilar to the results in the previous studies performed during the win-ter season (Clara et al., 2005; Trinh et al., 2016). In lab-scale MBR underSRTs of 30–90 days at varying temperatures (8 to 12 °C), Kruglova et al.(2016) reported substantially slow biological degradation of DCF. Like-wise, the elimination of DCF was higher in Jan–Feb than March–April.One reasonmight be themixed liquor temperature being higher around15 °C during Jan–Feb. The other reason could be the high MLSS concen-trations (approximately 8300 mg L−1) in the MBR process. In March,the process temperature was even below 10 °C for more than a week.Longer SRTs allow high microbial population and biomass concentra-tions and thuswould enhance the biological interactionswith the struc-turally-complex pharmaceuticals (Kimura et al., 2007). Majority of thestudies reported in the literature revealed higher removal performanceof DCFwith longer SRTs in MBR processes. However, it is worth noticedthat some contradictory results with no correlation between the elimi-nation of DCF and SRT were also reported (Vieno and Sillanpää, 2014).

3.3.3. BIS removalThe present work is probably the only study performed so far on BIS

removal in MBR operated at low temperature conditions. The removalefficiency of BIS varied over the investigation periodwith higher remov-al rates observed during Jan–Feb (73%) than in March (64%) and April(55%), during which sludge temperature varied approximately from20 °C to 12 °C. The explanation for the higher removal of BIS inMBRpro-cess might be attributed to the higher biodegradability at higher tem-perature and vice-versa.

3.3.4. BPA removalRelatively higher and stable removal of BPA (N97%) was achieved as

compared to other compounds regardless of low temperature condi-tions. The average BPA removal efficiency of N90% was documentedby (Clara et al., 2005; Trinh et al., 2016) in the MBR processes duringcold periods of the year, which is consistent with our result. Due to itshigh hydrophobicity (Log D N 3), BPA might have a tendency to sorbon suspended solids of mixed liquor matrix and subsequently be re-movedwith thewithdrawal of excess sludge (Cirja et al., 2008). Howev-er, about 95% of BPA was found to be eliminated via biotransformation,relative to influent load in this study. Thus, biotransformation is thedominant pathway of BPA removal. One reason for the poor eliminationof BPA via sorption process in MBR process might be due to its verypolar nature (Cirja et al., 2008). Biotransformation was also reportedas the main pathway for BPA removal by other researchers (Trinh etal., 2016).

3.4. Fate of inorganic micropollutants: trace metals

3.4.1. Concentrations of trace metals and removal efficienciesConcentrations of the selected trace metals in the influent and per-

meate, corresponding removal efficiencies, and themaximumallowableconcentrations (MAC) according to Water Quality Guidelines or Stan-dards set by international organizations are shown in Table 4. All the

trace metals were detected in the collected samples with an occurrencefrequency of 100%. Concentrations of trace metals in the influent werein the range of 0.1 to 80 μg L−1 (only dissolved fractions), which areconsistent with the range of trace metal concentrations documentedby other researchers (Di Fabio et al., 2013). With an average concentra-tion of 75 μg L−1, Znwas themost dominant,whereas Cdwas the lowest(0.1 μg L−1) detected among other tracemetals. On the other hand, lowconcentrations of Cr, Pb and Cd in influent wastewater explained thatno industrial or mining wastewaters are channeled with themunicipal wastewater. The relative abundance of trace metals ininfluent and permeate samples followed the general order ofZn N Ni N Co ≥ Cu N V N Cr N As ≥ Pb N Cd. Similar order of abundancesof trace metals have been reported by other researchers in commonmunicipal wastewaters (Chipasa, 2003; Karvelas et al., 2003). The con-centrations of Zn, Ni, Cu, and Cowere highly fluctuating in the permeatesamples, whereas the concentrations of V, Cr, As, Pb, and Cdwere almoststable.

The removal efficiencies of trace metals in MBR process varied withrespect to each element. An average removal of N80% was achieved forCd, Pb, and V, whereas about 60% removal was registered for Zn and Co.On the other hand, between 30% and 50% of Ni, Cu, As, and Cr were re-moved under the same operating conditions. Similar results are report-ed by other researchers (Bolzonella et al., 2010). In fact, the removalefficiencies are largely dependent on the initial concentrations ofmetals. In the present work, since the feeding source of wastewaterwas the effluent from primary clariflocculators of WWTP, the removalof some metals is expected to start before reaching MBR system.When a raw wastewater undergoes clariflocculation, a portion of tracemetals is abated by adsorption onto solid particles (Chipasa, 2003; DiFabio et al., 2013). Moreover, the concentrations of trace metals in per-meatewere found to be far lower than themaximumallowable concen-trations reported in the water quality guidelines.

Trace metals removal in MBRs can occur via various processes in-cluding precipitation of metals inside bioreactor, biosorption of metalions on sludge or bacterial flocs, and retention of the insoluble metalspecies by membranes (Bolzonella et al., 2010). Trace metals in waste-water are present in different phases (i.e. dissolved or particulate)with different water solubility at pH of 7–9 (Karvelas et al., 2003). ThepH critically affects the metal solubility, the surface functional groupsof bacteria, and the competition of cations and protons (Hammaini etal., 2007; Malamis et al., 2011). In the present study, the pH inside thebiological reactors was maintained at 6.6–7, therefore, both precipita-tion and biosorption mechanisms were expected for the removal oftrace metals. Katsou et al. (2011) reported an average removal of80%–98% for Cu and Pb and to a lower extent 50%–77% for Ni and Znby precipitation at pH 7–9. Malamis et al. (2011) has reported that be-yond pH 6 the sludge biosorption rate is decreasing and chemical pre-cipitation becomes dominant mechanism for metal removal. Thegradual increase in pH generates more OH− ions in the mixture whichform complex precipitates with metals (Fu and Wang, 2011). A signifi-cant metals removal via biosorption can be suggested for the presentwork, since the precipitation mechanism should be dominant (andmuch faster than biosorption) in the pH of range 8–11 (Fu and Wang,2011). Comte et al. (Comte et al., 2008) reported more favorablebiosorption of metals at pH range of 6–9, in which activated sludge isnegatively charged and provides potential sorption sites for metals.However, dominant metals removal mechanism via biosorption is ob-served at pH 4 for Pb and 6 for Cd, Cu and Zn when sludge is employedas sorbent (Hammaini et al., 2007; Malamis et al., 2011).

Higher concentrations of Zn andNi spotted in permeatemight be at-tributed to the metal ions penetrated through the membranes due totheir high solubility at aforementioned pH, which is consistent withthe observation reported previously (Katsou et al., 2011). However, inaddition to the solubilization, these observations can be partially attrib-uted to the unstable forms of Zn and Ni. On the other hand, removal ofCr, Pb, As, Cd, As, and Co remained relatively high which was possibly


due to the low pH required for solubilizing and their availability in sta-ble form within the given pH. Enhanced removal of trace metals is re-ported with the increasing MLSS concentration as more sites areavailable for the sorption of soluble metal ions (Katsou et al., 2011;Malamis et al., 2011). On the contrary, Hammaini et al. (2007) reportedrelatively steady or reducedmetal sorption capacitywith the increase inMLSS concentration due to the screen effect between cells.

Promising removal efficiencies of trace metals by different types ofmembranes such as ultra-filtration, reverse osmosis, nanofiltrationand electrodialysis have been reported (Fu and Wang, 2011; Katsou etal., 2011). However, the higher removal efficiencies of most of themetals in the present study was attributed to the complete retentionof TSS via MF membranes, which is in agreement with a previousstudy (Bolzonella et al., 2010). Furthermore, a substantial removal oftrace metals is reported by a research study (Malamis et al., 2011), inwhich MF membrane was combined with activated sludge. On theother hand, an enhanced biosorption of tracemetals on the problematicbiofilm developed on membrane surface (fouling layer) than to thesludge flocs is reported in pilot-scale MBR (Di Fabio et al., 2013). Eventhough fouling has negative influence on membrane permeability, afraction ofmetals removal could be originating frommembrane fouling.

3.4.2. Trace metals in sludge (mixed liquor)Average trace metal concentrations in activated sludge of MBR pro-

cess are summarized in Table 5. To assess the toxicity of excess sludgefromMBR process, it is very important to monitor trace metal accumu-lation in the bioreactor for the long term functioning (Bolzonella et al.,2010). Trace metals in sludge are present in the several forms includingmetal precipitates in the sludge flocs, soluble metal ions, complexes ofsoluble metals and accumulated soluble metal in the microbial cells(Chipasa, 2003). Trace metals are non-biodegradable compounds, thusthe total trace metal loading of feed wastewater will either end up inthe sludge or remain in the treated effluent. In fact, tracemetals releasedvia excess sludge is more crucial than through treated effluent to the

point-source of discharges (Karvelas et al., 2003). The frequency of oc-currence of all trace metals in the sludge samples was 100% and samefor the liquid samples. From the analytical results, Zn was the mostabundant metal in sludge, whereas Cd exhibited the lowest abundance.However, the concentrations of Zn, Ni, Cu, Pb, and Cd were found to beunder the allowable concentrations specified in the EU standard. As nolimits are set for Co, Cr, As by EU so far, their concentrations were com-pared with US regulations and also found under the MAC (the concen-tration threshold for V was not mentioned).

4. Conclusion

In this study, the performance of pilot-scale MBR unit, operated atlow temperature conditions in Nordic regions, was investigated forthe treatment of real municipal wastewater. An accelerated membranefouling was observed during low temperature periods and while main-taining aeration intensity below the normal value. This resulted in asubstantial decrease in membrane permeability (about 75%), leadingto an increased frequency of membrane cleaning. The average reduc-tions of 1.82, 3.02, and 1.94 log units were achieved for norovirus GI,norovirus GII, and adenoviruses, respectively. Removal of Escherichiacoli and enterococci was N5.6 log units. The average removal efficienciesof BIS, DCF and BPA were 65%, 38%, and N97%, respectively. However,the MBR pilot did not show efficient removal of carbamazepine.

Overall, and based on European and American regulations, highquality permeatewas achieved, demonstrating the aptitude ofMBR sys-tem to perform satisfactory wastewater treatment during the winterseason in Nordic regions and reclaim treated water and sludge for theagricultural and industrial sectors.

Acknowledgements

This workwas supported by TEKES-Finnish (1043/31/2015) fundingagency for innovation, Finland, the research fund for SMART EFFLUENTS

Table 4Occurrence and removal of 9 trace metals in the MBR pilot plant (n = 42); comparison of permeate concentrations with Water Quality Guidelines or Standards set by internationalorganization.

Metal Influent concentration (average± Sd)[μg L−1]

Permeate concentration (average± Sd)[μg L−1]

Removal efficiency (average± Sd)[%]

Maximum allowableconcentration[μg L−1]

Reference

Zn 75.86 ± 17.5 31.43 ± 14.3 59.6 ± 11.6 3000 (WHO, 2011)a

Ni 14.93 ± 3.9 9.75 ± 1.8 33.7 ± 6.1 34 (EU, 2013)b

Co 6.66 ± 2.3 2.6 ± 1.4 60.9 ± 17.7 Not mentionedCu 6.19 ± 3.5 3.87 ± 2.0 31.3 ± 32.6 2000 (WHO, 2011)a

V 2.73 ± 1.0 0.35 ± 0.1 85.7 ± 5.8 Not mentionedCr 1.64 ± 0.7 0.81 ± 0.2 47.5 ± 9.6 50 (WHO, 2011)a

As 0.68 ± 0.2 0.34 ± 0.1 45.7 ± 18.9 10 (WHO, 2011)a

Pb 0.65 ± 0.3 0.11 ± 0.1 80 ± 10.9 14 (EU, 2013)b

Cd 0.1 ± 0.1 0.01 ± 0.0 86.3 ± 7.0 1.5 (EU, 2013)b

a Guidelines for drinking-water quality.b 2013/39/EU directive of the European parliament and of the council.

Table 5Average trace metal concentrations in sludge with maximum allowable concentrations as per International Standards.

Metals Average ± Sd[mg kg−1 dry weight sludge]

Maximum allowable concentration[mg kg−1 dry weight sludge]

Reference

Zn 600.7 ± 89.3 2500–4000 (EU, 1986)Ni 29.5 ± 4.5 300–400 (EU, 1986)Co 24.9 ± 3.7 77 (Harrison et al., 1999)Cu 130.7 ± 13.1 1000–1750 (EU, 1986)V 30.9 ± 4.8 Not mentionedCr 45.4 ± 6.9 75 (Harrison et al., 1999)As 3.8 ± 0.8 41 (Harrison et al., 1999)Pb 8.8 ± 1.3 750–1200 (EU, 1986)Cd 0.6 ± 0.1 20–40 (EU, 1986)


project. The authors like to thank National Institute for health and wel-fare (THL), Finnish environment services (SYKE) andMikkeliWWTP fortheir support on laboratory works. Especial thanks go to all the staff ofMikkeli WWTP, Finland.

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Publication II

Khum Gurung, Mohamed Chaker Ncibi, Mika Sillanpää

Removal and fate of emerging organic micropollutants (EOMs) in

municipal wastewater by a pilot-scale membrane bioreactor (MBR)

treatment under varying solid retention times



Vol. 667, pp. 671-680, 2019

© 2019 Elsevier

Removal and fate of emerging organic micropollutants (EOMs) inmunicipal wastewater by a pilot-scale membrane bioreactor (MBR)treatment under varying solid retention times

Khum Gurung ⁎, Mohamed Chaker Ncibi, Mika SillanpääDepartment of Green Chemistry, School of Engineering Science, Lappeenranta-Lahti University of Technology, Sammonkatu 12, FI-50130 Mikkeli, Finland

H I G H L I G H T S

• Fate of 23 EOMs in pilot MBR treatmentat varying SRT was studied.

• At long SRT, EOMs removal majorly en-hanced while SMP and EPS concentra-tion lessened.

• Hormones and antibiotics were highlyremoved than diuretics and anti-epileptics.

• Iopamidol andhydrocortisone (N90% re-moval) were studied for the first time inMBR.

• MBR Effluent fulfilled the water qualityrequirements for reuse.

G R A P H I C A L A B S T R A C T

a b s t r a c ta r t i c l e i n f o

Article history:Received 29 December 2018Received in revised form 13 February 2019Accepted 19 February 2019Available online 20 February 2019

Editor: Paola Verlicchi

This study investigates the removal and fate of 23 emerging organic micropollutants (EOMs) including a widerange of pharmaceuticals (antibiotics, β-blockers, analgesics, diuretics, psychostimulants, antiepileptics, immu-nosuppressives, anticoagulants), and steroid hormones detected inmunicipal wastewater by a pilot-scale mem-brane bioreactor (MBR) plant at two different solid retention times (SRTs) of 60 and 21 days. Different removalefficiencies of the selected EOMs were observed and explained based on their physicochemical properties (suchas distribution coefficient, log D; dissociation constant, pKa; solid-water distribution coefficients, and Kd) alongwith process operating parameters. The dominant removal mechanisms of EOMs were biotransformation andsorption onto the sludge, which were confirmed by the mass balance study. Moreover, changes in the sludgeproperties, as a consequence of different SRTs, were evaluated based on variations in soluble microbial products(SMP), extracellular polymeric substances (EPS), and capillary suction time (CST). Finally, the quality of theMBReffluent was compared with some established guidelines, which confirmed the fulfilment of water quality re-quirements for reuse purposes.

© 2019 Elsevier B.V. All rights reserved.

Keywords:Municipal wastewater treatmentMembrane bioreactorEmerging organic micropollutantsBiotransformation/biodegradationSolid retention times

1. Introduction

In recent decades, the occurrence of emerging organicmicropollutants(EOMs) in the aquatic environment has received a global attention be-cause of its possible environmental intimidations. Globally, a large num-ber of EOMs have been frequently detected at source, and in treated

Science of the Total Environment 667 (2019) 671–680

⁎ Corresponding author.E-mail addresses: [email protected], [email protected] (K. Gurung),

[email protected] (M. Sillanpää).

https://doi.org/10.1016/j.scitotenv.2019.02.3080048-9697/© 2019 Elsevier B.V. All rights reserved.



j ourna l homepage: www.e lsev ie r .com/ locate /sc i totenv

effluent and wasted sludge at wastewater treatment plants (WWTPs)(Clara et al., 2005b; Trinh et al., 2012, 2016). Moreover, the occurrenceof EOMs could endanger the reuse of reclaimed wastewater in achievingsustainable water management within the WWTPs (Behera et al.,2011).Most of the EOMs found inwastewater are of anthropogenic originfrom various sources such as households, hospitals, veterinary surgeries,farms and pharmaceutical manufacturing facilities, which eventuallyend up at WWTPs (Clara et al., 2005b). Since conventional WWTPs arenot specifically designed to eliminate EOMs, many of these EOMs passthrough WWTPs and are thus eventually discharged into the environ-ment (Behera et al., 2011; Jelic et al., 2011; Miège et al., 2009; Sim et al.,2010; Vieno et al., 2005; Yang et al., 2011, 2017; Ying et al., 2009). Somecountries such as Australia and Switzerland have adopted dischargeguidelines for some EOMs (NHMRC/EPHC/NRMMC, 2008;Oekotoxzentrum Centre Ecotox, 2016). In the EU context, environmentalquality standards (EQS) for 45 priority EOMs including first watch listsubstances like diclofenac, 17-β-estradiol, and 17-α-ethinylestradiol,were stipulated in Directive 2013/39/EU (EU, 2013), but the majority ofother EOMs have not yet been addressed. However, recent amendmentin directive 2013/39/EU has introduced more watch list substances, inwhich diclofenac has removed while macrolide antibiotics, methiocarb,neonicotinoids, metaflumizone, amoxicillin and ciprofloxacin have beenadded (EU, 2018).

To ensure compliance with future discharge requirements, interestin the ability of MBRs in the removal of EOMs frommunicipal wastewa-ter has increased in recent decades. MBRs are becoming a mature tech-nology in water management practices for wastewater reclamation andreuse due to their ability to produce superior-quality effluent, a rela-tively small footprint and little sludge (Judd, 2008; Meng et al., 2009).The removal mechanisms of EOMs involved inMBR treatment are com-plex and include biotransformation/biodegradation, adsorption tosludge, volatilisation, and physical retention by the membrane (Cirjaet al., 2008). However, biotransformation and adsorption to biosolidsare often reported to be the dominant mechanisms for the removal ofEOMs in the MBR process (Clara et al., 2005b; Kim et al., 2014;Radjenović et al., 2009). On that note, physicochemical properties(such as distribution coefficient, log D; dissociation constant, pKa;solid-water distribution coefficients, and Kd) and functional groups as-sociated with EOMs showed notable effects on their removal (Li et al.,2015; Tadkaew et al., 2011). At the pH above pKa, phenolic hydroxylgroup of some hormones dissociates that leads to the charge repulsionbetween the negatively charged hormone and negatively chargedmembrane (Schäfer et al., 2011). Naproxen and diclofenac (pKa valuesof 4.2 and 4.15, respectively) deprotonated a at pH 7–8 and rejectedby negatively charged membrane surface, whereas carbamazepine(pKa = 13.9) showed no ionisation at same pH level, hence attained apoor removal (Röhricht et al., 2009). At pH 6–7, tetracyclines (pKa =4.5) are not charged, hence removed only via adsorption to sludge(Kim et al., 2005). The sorption onto sludge can considered to be insig-nificant for EOMs with logKd b 2.48 (Joss et al., 2006). The overall re-moval efficiencies of most of hydrophilic chemicals (log D b 3.2) wereabove 92% in the MBR (Trinh et al., 2016).

Often previous studies on the removal of EOMs by MBRs focused onthe aqueous phase (Bo et al., 2009; Radjenovic et al., 2007; Schröderet al., 2012; Trinh et al., 2012) and their mass load adsorbed in thesolid phase was often neglected, which restricts the differentiation be-tween the removal mechanisms via adsorption onto sludge and bio-transformation. There is limited literature investigating the fate ofEOMs in MBRs when treating real municipal wastewater (Kim et al.,2014; Radjenović et al., 2009; Trinh et al., 2016). Some researchershave estimated the contributions of the adsorption and biotransforma-tion of EOMs based on values of Kd from literature or by direct batch ex-periments (Clara et al., 2005b; Kimura et al., 2007). Additionally,previous studies on the fate of EOMs in MBR either mainly focused onpharmaceuticals and personal care products (PPCPs) (Radjenovićet al., 2009; Suárez et al., 2012; Tambosi et al., 2010; Trinh et al.,

2016) or are based on lab-scale batch experiments (Hai et al., 2018;Phan et al., 2014; Tadkaew et al., 2011;Wijekoon et al., 2013). The influ-ence of solid retention times (SRTs) on the removal mechanisms ofPPCPs by MBRs has been studied by a few researchers (Clara et al.,2005b; Schröder et al., 2012; Tambosi et al., 2010), but these studiescounted either estimated values of Kd or negligible amounts of adsorbedfractions of PPCPs onto sludge, as well as a spiking of PPCPs in the MBRsystem.

Based on the above identified research gaps, the current study inves-tigates the removal dynamics of 23 frequently detected EOMs in thetreatment of real municipal wastewater at an MBR pilot plant operatedat twodifferent SRTs. To thebest of our knowledge, the removal and fateof EOMs such as iopamidol, hydrocortisone, furosemide, ciprofloxacin,tetracycline, and enalapril have rarely (or not) been studied in thepast by MBR processes. A mass balance study was assessed to furtherconfirm the removal mechanisms associated with different EOMs. Be-sides, the influence of physicochemical properties on the removal ofEOMswas discussed. Additionally, the influence of SRTs on sludge char-acteristics including dewaterability, major foulants such as soluble mi-crobial products (SMPs), and extracellular polymeric substances (EPS),was discussed. Finally, the quality of MBR-treated effluent in terms ofEOMs was compared with some established guidelines.


2.1.MBR system, operational protocol, selected EOMs and sample collection

A pilot-scaleMBR systemwas employed in this study. The schematicdiagram indicating major components of the MBR pilot plant is shownin Fig. 1. The key features of the MBR pilot plant is described in our pre-vious work (Gurung et al., 2016). Typically, the submerged flat-sheetmembrane units (0.4 μm pore size) with membrane surface area of16m2 (KUBOTA Co., Japan) were used for solid-liquid separation. A per-meate pump (Thomas, Germany) was used to draw water in an inter-mittent suction mode. For in-situ cleaning of membranes, the clean-in-place (CIP) tankwas also installed and connectedwith Sodiumhypo-chlorite and Citric acid dosing tanks. Scouring air was supplied to main-tain biological activities and to mitigate membrane fouling. The feed tothe MBR pilot plant was taken continuously from the outlet of the pri-mary clarifier at the Kenkäveronniemi WWTP, Mikkeli, Finland. Thisplant treats about 15,000 m3 of wastewater per day.

In this study, the pilot plant was operated in two different SRTs of60 days and 21 days in order to investigateMBRperformance in treatingthe pollution load of nutrients, organics, solids and, preferably, the re-moval of EOMs. Table 1 summarises the in-depth operational conditionsof theMBR pilot plant for this study. The pilot plant was operated a pro-tocol similar to that reported by other researchers (Van den Broecket al., 2012). The SRT of 60 days was adapted N10 weeks before the ac-tual sampling periods in order to stabilize the activated sludge. How-ever, relatively less adaption period of sludge (about a week) wasmaintained for SRT of 21 days by wasting required amount of sludgeout of the system. Each of the two different SRTs were maintained forstable sludge period of 4 weeks. During the investigation, the perme-ation flow rate and the transmembrane pressure (TMP) were nearlyconstant (4–5 kPa), so no in-situ cleaning of membranes was necessary.

A set of 55 EOMs representing a wide range of pharmaceuticals, pes-ticide and steroid hormones with diverse physicochemical propertieswere studied. The selection of these EOMs was made based on their in-creasing annual consumption, potential adverse impacts on health ofhuman and aquatic animals, and stringent future regulations in the EUWater Framework Directive (EU, 2013) on pharmaceuticals andhormones. The physicochemical properties of studied EOMs aresummarised in Table S1 (Supplementary Information). Weekly 24-hcomposite aqueous samples of influent (1 L), MBR permeate (1 L) andgrab samples of sludge (0.5 L) were collected for 8 weeks (4 samplesof each SRT). The sum total of samples included 8 influent, 8 effluent


and 8 sludge samples. All the collected samples including liquid andsludge phases were stored in dark and cold conditions (~ −18 °C) be-fore the analysis.

2.2. Mass balance calculation

To assess quantitatively the fate of the EOMs, the percentage sharesof biotransformation/ degradation, sorption and remaining in effluentwere estimated using a mass balance approach (Jelic et al., 2011). Themass flow rates of EOMs in the influent, effluent and sludge were calcu-lated by multiplying measured concentrations in given streams by thecorresponding flow data. Then, mass load lost per unit of time (μgd−1) due to all the processes that possibly occur during theMBRprocessand their relative distributions to the influentmass loadwere calculatedaccording to the following Eqs. (1)–(4).

Overall removal %ð Þ ¼ CInf � QInf−CEff � QEff ÞðCInf � QInf

� 100 ð1Þ

Removal via sorption onto sludge %ð Þ ¼ CSludge � QExcess sludge

CInf � QInf� 100 ð2Þ

Remain in effluent %ð Þ ¼ CEff � QEff

CInf � QInf� 100 ð3Þ

Removal via biotransformation %ð Þ¼ Overall removal %ð Þ−Removal via sorption onto sludge %ð Þ ð4Þ

where CInf ∗ QInf; CEff ∗ QEff; and CSludge ∗ QExcess sludge are the massflow rate (μg/d) of influent, treated effluent and sludge, respectively.QExcess sludge is excess sludge production (g d−1).


Chemical oxygen demand (COD), total suspended solids (TSS), am-monium nitrogen (NH4-N), total nitrogen (TN), total phosphorus (TP),mixed liquor suspended solids (MLSS) and mixed liquor volatilesuspended solids (MLVSS) were determined according to StandardMethods for the Examination of Water and Wastewater (APHA, 1999).The CST of sludge was measured using a capillary suction timer (Type304 M, Triton, UK). The concentrations of EPS and SMP of the sludgewere determined by previously described methods (Ramesh et al.,2006). At first, 25 mL of sludge sample was centrifuged at 4000 rpm at4 °C for 30min. The supernatantwasfiltered through a0.45 μmglassfilterand measured for SMP. The obtained pallet was resuspended by adding0.05% NaCl, mixed with vortex mixture and the solution was kept in awater bath (~100 °C) for heat treatment for 60 min. Finally, the superna-tantwasfiltered through a 0.45 μmglassfilter andmeasured for total EPS.Both SMP and EPS were analysed in term of dissolved organic carbon(DOC) using a TOC analyser (TOC-V Series-CPN, Shimadzu, Japan).

The concentrations of EOMs in influent, permeate and sludge sam-ples were analysed using standard solid phase extraction (SPE)followed by UPLC/MS/MS (modified EPA 1694 and modified EPA 539).The measurement uncertainty (MU) was 40%. The MU %was calculatedaccording to the Nordest TR 537 that covered every step in the labora-tory analysis.


3.1. Basic performance of MBR

As real municipal wastewater was fed into the MBR, the influentcomposition varies over time, and does the pollution removal efficiency.

Fig. 1. Schematic diagram of the pilot MBR unit with indicated sampling points.

Table 1Operational conditions of the MBR pilot plant.

Parameter Units Values min – max (avg.)

SRT: 60 days SRT: 21 days

Hydraulic retention time(HRT)

Hours 40.1–45.5 (42.13)

Flux (Continuous) L m−2 h−1 5.4–6.13 (5.84)MLSS concentration mg L−1 7808–9688

(8550)2271–4865(3748)

MLSS/MLVSS _ 0.57–0.83(0.67)

0.3–0.68(0.56)

F/M ratio kg COD. (kg MLVSS.d)−1

0.022–0.034(0.027)

0.06–0.11(0.09)

Specific aeration demand(SADm)

m h−1 0.15–0.23 (0.20)

Dissolved oxygen (DO) mg O2 L−1 1.2–4.6 (2.9)Reactor temperature °C 14–22 (19)pH Unitless 6.0–7.4 (6.6)Intermittent filtration Minutes 9-ON/1-OFF


Table 2 summarises the composition of influentwastewater during eachperiod of SRT and the treatment efficiency of the MBR on conventionalpollutions such as COD, NH4-N, TN, and TP. The fluctuations on influentcomposition were not significant (Table 2). The pH of the effluent wasaround 7. The removal efficiencies of TSS, COD and NH4-N were consis-tently high (N90%,N99.96%, respectively) at both SRTs. For two SRTs, thesuspended solids in the effluent samples were below the limit of quan-tification (LOQ), showing the excellent physical barrier of the MBRsystem. The pilot MBR plant was not designed to remove TN, butabout 15% of removal on both SRTs was achieved due to a simultaneousnitrification-denitrification environment created by intermittent aera-tion inside the nitrification tank. A relatively consistent removal of TPwas also observed (N87%) on both SRTs due to enhanced biologicalphosphorus removal (EBPR) in theMBRprocess. Overall, a good effluentquality was observed at both the SRTs with comparable treatmentefficiencies.

3.2. Occurrence of EOMs in the MBR

The median concentrations of the EOMs detected in the influent ofthe MBR pilot plant are presented in Fig. 2. The concentrations ofEOMs in WWTPs depend on several factors such as the consumptionpattern of pharmaceuticals, the wastewater influent load etc., whichmight vary within a region or country and from country to country(Radjenović et al., 2009). Of the 55 EOMs, only 23 compounds were an-alytically detected in the median range from 23 to 145,000 ng L−1

(Fig. 2) and the remaining 32 compounds were not detected abovetheir LOQs during the testing of both SRTs (Table S2). Fig. 2 indicatesthat the occurrences of themajority of EOMs including pharmaceuticalsand steroid hormones during the SRT of 60 days showed similar tenden-cies to those observed during the SRT of 21 days, which is explained bythe fact that all the sampling sessionswere performed during the springseason. The median concentrations of the EOMs belonging to anti-coagulants, diuretics, psychostimulants and analgesics categories suchas iopamidol (2625 ± 2367 ng L−1), furosemide (1864 ± 91 ng L−1),caffeine (140,437 ± 6453 ng L−1), hydrochlorothiazide (2035 ±757 ng L−1), paracetamol (93,600 ± 3394 ng L−1), naproxen (3395± 276 ng L−1), ibuprofen (14,400 ± 565 ng L−1) and diclofenac(1620 ± 394 ng L−1), respectively were observed at the highest levelsduring the testing periods. N50% of the total EOMs in the influent wasaccounted for these eight compounds. The high concentrations of caf-feine, naproxen and ibuprofen are consistent with previous papersthat studied the occurrence and treatment of EOMs in real municipalwastewater (Clara et al., 2005b; Kim et al., 2014; Radjenović et al.,2009; Trinh et al., 2016; Vieno et al., 2005). The observed caffeine con-centration is in agreement with previous literature (Kim et al., 2014),and is also about four times higher than the values reported in anotherstudy (Trinh et al., 2012). Antibiotics such as tetracycline, ciprofloxacin,and trimethoprim were detected at b1000 ng L−1, which is very com-mon in municipal wastewater (Guerra et al., 2014). The concentrationsof EOMswere arranged from lower to higher log D pH 7 (distribution co-efficient) values of compounds (Fig. 2). The log D pH 7 is the correctedvalue of log Kow (partition co-efficient) for ionisation at pH 7 (Trinhet al., 2016). The concentrations of steroid hormones including testos-terone, progesterone, estrone, and estroil having log D pH 7 values N3,

were observed at a consistently low level (23 to 510 ng L−1) duringboth testing periods.

3.3. Removal of EOMs in MBR treatment

The removal efficiencies of selected EOMs observed during the MBRtreatment at two different SRTs (60 days and 21 days) are shown inFig. 3a–b. SRT is a very important operational parameter in MBR pro-cesses influencing the removal of EOMs (Clara et al., 2005a; Schröderet al., 2012; Taheran et al., 2016; Tambosi et al., 2010). The concentra-tions of the majority of the EOMs in the treated effluent were observedto be lower (even b LOQs to many EOMs) than in the influent at bothSRTs, indicating the efficient attenuation of EOMs by the MBR, regard-less of the changing of the SRTs. In general, increasing SRT can improvethe biodiversity of slowly growing bacteria, sludge growth and the lon-ger retention of sludge that favours the removal of EOMs (Kimura et al.,2007; Maeng et al., 2013; Taheran et al., 2016; Tambosi et al., 2010).However, some previous researchers confirmed that the removal of ibu-profen, diclofenac, naproxen, ketoprofen and carbamazepine is not sig-nificantly affected by changing SRT during MBR treatment (Li et al.,2015).

Of the selected EOMs in this study, 17 EOMs, including steroid hor-mones, showed consistent removal of N90%. Furthermore, three com-pounds showed moderate removal between 50% and 87% (metoprolol,furosemide, and propranolol), one showed poor removal of 39–46%(diclofenac), and two (carbamazepine and hydrochlorothiazide) werevery poorly removed (almost not at all) (−8 to 10%). Psychostimulant(caffeine) was detected at highest concentration among other EOMs atboth SRTs, but efficiently attenuated with a removal efficiency of up to99.7%. These results are in agreement with many previous works (Kimet al., 2014; Trinh et al., 2012, 2016). An effective attenuation of caffeineis achievable even at a relatively short SRT (Maeng et al., 2013).

Moreover, EOMs including anti-coagulant (iopamidol) andimmunosupressive (hydrocortisone), whose fate was studied for thefirst time in the MBR process, were efficiently removed in the range of90–96%. Regarding β-blockers, atenolol and bisoprolol were attenuatedat consistently high removal efficiency (92–97%) at both SRTs, whereastwo other β-blockers (metoprolol and propranolol) were attenuatedslightly higher at a longer SRT of 60 days (82–84%) than in shorter SRTdays of 21 (50–60%). Very high removal efficiencies of analgesics (para-cetamol, ibuprofen, ketoprofen, naproxen)were observed ranging from96 to N99.9%. The higher removal of paracetamol is due to its structurethat allows unrestricted access of bacteria and enzymes to the stericallyunprotected molecule that is then subsequently modified (Tambosiet al., 2010). Moreover, the removal of ibuprofen, ketoprofen andnaproxen is partially accompanied by the heterotrophic bacteria com-munity, which is highly favourable in extended SRTs, and less so byslow-growing ammonia-oxidising bacteria (Falås et al., 2012). Thehigher removal results of analgesics are in agreement with those re-ported by many researchers (Kim et al., 2014; Radjenović et al., 2009;Schröder et al., 2012; Tambosi et al., 2010). On the other hand,diclofenac (analgesic) was attenuated at relatively low efficiency com-pared to other analgesics with median removal efficiencies of 38% and45% at SRTs of 60 days and 21 days, respectively. A great discrepancyin the removal efficiency of diclofenac via MBR was reported in some

Table 2Basic performance of MBR system (mean ± standard deviation) at two different SRTs.

Parameter SRT:60 days SRT: 21 days

Influent (mg L−1) Effluent (mg L−1) Removal efficiency (%) Influent (mg L−1) Effluent (mg L−1) Removal efficiency (%)

TSS 128.0 ± 37.9 b1 100 133.33 ± 14.1 b1 100COD 299.3 ± 55.2 22.7 ± 5.2 92.43 330.0 ± 19.6 23.0 ± 5.3 93.03NH4-N 29.7 ± 6.3 b0.01 N99.96 38.85 ± 3.2 b0.01 N99.96TP 3.94 ± 0.7 0.47 ± 0.2 88.12 3.86 ± 1.4 0.49 ± 0.1 87.40TN 45.0 ± 7.5 38.33 ± 7.1 14.81 50.0 ± 0.74 42.5 ± 4.4 15.00


Fig. 2.Occurrences of EOMs inMBR process at two different SRTs (60 days and 21 days) arranged according to their Log D (at pH7) values. EOMswith concentration lower than LOQ are notincluded.

Fig. 3.Median removal efficiencies of EOMs inMBR process at different SRTs. The EOMswith their detection in influent b LOQ are not presented. The LOQvalues of EOMs in the effluent areused to calculate % removal whenever measured concentrations are below LOQ. Error bars represents the standard deviation calculated from the duplicate samples taken once in a week.


literature (0 to 87%) (Clara et al., 2005a; Trinh et al., 2016). A slow ornon-biotransformation rate of diclofenac has been reported by other re-searchers (biotransformation rate constant, Kbiol b 0.1 g−1 ss d−1 inMBR) (Trinh et al., 2016; Vieno and Sillanpää, 2014).

The diuretics (furosemide and enalapril), which were rarely studiedin MBRs, were highly attenuated at both the SRTs with median removalefficiencies of 87% and 96%, respectively. These results are consistentwith the previous study (Kim et al., 2014). On the other hand, the atten-uation of diuretic (hydrochlorothiazide) was very low or hardly re-moved at all in this study. During SRT of 21 days, the removalefficiency ranged from−52% to 13% (median− 1%), whichwas slightlyenhanced at SRT of 60 days achieving removal efficiency between−19%and 41% (median 10%). The comparably poor removal of hydrochloro-thiazide (~5%) is reported by other researcherswhen treatingmunicipalwastewater with full-scale MBR (Radjenović et al., 2009). The increasedconcentrations of hydrochlorothiazide in treated effluent, more thanthe influent levels (i.e. negative removal efficiency), might be due tothe transformation of metabolites such as unmeasured products of thehuman metabolism and/or metabolites that could be converted backto their parent compounds during the biological treatment process.This phenomenon of “negative removal” was also reported in otherstudies (Jelic et al., 2011).

The attenuation of an anti-epileptic drug (carbamazepine) was verylow and frequently not removed, regardless of varying SRTs. At an SRTof 60 days, the median removal efficiency of carbamazepine was closeto 2% (not removed in two samples out of four), whereas non-removalwas observed (median – 8%) at an SRT of 21 days. This means that, inmost of the events, effluent concentrations ofMBRwere greater than in-fluent levels. The major metabolite of carbamazepine in humans is10,11 epoxy-carbamazepine, which is excreted principally as glucuro-nide and further to glucuronide-conjugates via the hydroxylation ofthe aromatic ring. These glucuronide-conjugates can presumably becleaved in a biological process and so increase in the effluents (Ternes,1998). The recalcitrant behaviour and non-removal of carbamazepinein MBR treatment due to poor degradability have been reported inmany other studies (Clara et al., 2005b; Radjenovic et al., 2007;Wijekoon et al., 2013).

Concerning the removal of antibiotics, consistently high removal ef-ficiencieswere achieved at 87 to 98%. Tetracycline, ciprofloxacin and tri-methoprim were attenuated at a comparable level at both SRTs withremoval efficiencies of 98%, 93%, and 94%, respectively. Comparable re-sults in the removal of tetracycline and ciprofloxacin were reported byKim et al., which was the only research studied to-date in theMBR pro-cess (Kim et al., 2014), where the SRT was kept relatively slow at6–8 days. This indicates that the removals of tetracycline and ciproflox-acin were not significantly affected by changes in SRTs. Nevertheless, awide range of removal efficiencies were reported in literature fromnon-removal to nearly 67% (Kim et al., 2014; Radjenović et al., 2009). Al-though trimethoprim is often presumed to be recalcitrant to activatedsludge bacteria, its poor biotransformation was reported to be achiev-able by slow-growing bacteria (Radjenović et al., 2009).

The median removal efficiency of steroidal hormones (testosterone,progesterone, estriol, estrone) in the MBR process at both SRTs is pre-sented in Fig. 3. Results indicated that steroid hormones were attenu-ated effectively by the MBR, achieving steady removal efficiency ofover 98%. These results are in agreement with other studies performedon MBRs (Maeng et al., 2013; Trinh et al., 2012, 2016).

3.4. Behaviour of EOMs in MBR treatment: application of mass balances

The removal mechanisms of EOMs inMBR treatment form a complexprocess characterised by four major pathways: (i) biotransformation ordegradation; (ii) sorption onto the sludge (via excess sludge removal);(iii) volatilisation or stripping by aeration; and (iv) physical retentiondue to size exclusion by membranes (Cirja et al., 2008; Li et al., 2015).Since the Henry's law constant (KH) values of the selected are low (b

10−6) (Table S1), the removal of these compounds by volatilisation is in-significant. Moreover, the typical molecular weight cut-off (MWCO) ofMF and UF membranes are well above several thousand daltons (Da)(Taheran et al., 2016), so the retention of EOMs (MWCO between 200and 800Da) in theMBRprocess cannot be expected due to size exclusion.Therefore, the dominant removalmechanisms of EOMswere presumablyvia biotransformation and sorption onto the sludge, which was con-firmed by the mass balance calculations as shown in Fig. 4. The dailymass loads of EOMs via biotransformation, sorption onto sludge and re-maining in effluent are interpreted as shared proportions relative to theinfluent load (Eqs. (1)–(4)).

Among the selected EOMs, the highest significant biotransformationrates (83–99.9%, median 99%) were observed for atenolol, caffeine,paracetamol, naproxen, ibuprofen, estriol and progesterone. Similarly,high biotransformation rate was accounted to iopamidol, furosemideand trimethoprim (58–92%, median 86%). Moderate biotransformationwas observed for diclofenac (median 33%). Enhanced biotransforma-tions of metoprolol, hydrochlorothiazide, and propranolol wereachieved at SRT of 60 days (78%, 27% and 51%) compared to SRT of21 days (37%, 0% and 0%, respectively). On the other hand, carbamaze-pine showed high resistance to biotransformation, so was highly de-tected in the effluent. The biotransformation proportions observed fornaproxen, ibuprofen, trimethoprim, enalapril, atenolol, metoprolol andfurosemide are consistent with previous studies (Jelic et al., 2011; Josset al., 2006; Kim et al., 2014).

On the other hand, some of the EOMs were observed to be removedvia sorption to sludge as indicated in Fig. 4. EOMs can be sorbed tosludge and subsequently removed while wasting the excess sludgefrom the bioreactor. The mechanism of sorption occurs via absorptionand adsorption (Li et al., 2015). Absorption occurs due to the hydropho-bic interactions of the aliphatic and aromatic groups of EOMs with theliphophilic cell membrane of somemicroorganisms and the fat fractionsof sludge, whereas adsorption involves the electrostatic interactions ofthe positively charged (e.g., amino groups) to negatively charged sur-face of microorganisms (Cirja et al., 2008). Most of the EOMs were fre-quently detected in the sludge under LOQ (10 to 500 ng g−1, median100 ng g−1) (Fig. S1). Other EOMs were detected in the range of38–3750 ng g−1 (median 275 ng g−1) at which ciprofloxacin followedby hydrochloro-thiazide were the highest detected in the sludge. Thisresult is in agreement with a previous study, where ciprofloxacin wasmost frequently detected in MBR sludge, although the concentrationlevel was considerably higher (N 103 ng g−1) (Kim et al., 2014). Fromthe results, ciprofloxacin posed a serious challenge due to very strongaffinity to the sludge with a sorption fraction of N90%. Relatively highersorption of ciprofloxacin is reported by many researchers (Kim et al.,2014). Biosorption as a dominant mechanism for the removal of cipro-floxacin is reported by many authors (Githinji et al., 2011; Guerraet al., 2014). Similarly, propranolol, carbamazepine and testosteroneshowed moderate sorption efficiency to the sludge (10–50%, median32%), whilst weak sorption affinities were observed for rest of theEOMs (0–21%, median 6%).

In order to better understand the fate of EOMs due to sorption to thesludge, sludge-water distribution coefficients or sorption coefficients(log Kd) were experimentally assessed by dividing the concentrationof EOMs in sludge (ng kg−1) by their concentrations in effluent water(ng L−1) (Loke et al., 2002; Radjenović et al., 2009). The correspondinglog Kd for high sorption fractions of N90% and moderate fraction(10–50%) were 3.0 to 4.9 (Fig. 4), demonstrating their removal bystrong sorption to sludge. These results are comparable with the find-ings of other researchers (Kim et al., 2014). Nevertheless, many com-pounds which showed a high degree of removal via biotransformation(e.g. tetracycline, bisoprolol, enalapril, ketoprofen, ibuprofen, hydrocor-tisone, estriol, estrone and progresterone) accounted for the high log Kd

values (3.9 to 4.3), indicating that these EOMs are attenuated simulta-neously by biodegradation and partitioning onto biosolids (Guerraet al., 2014). On the other hand, the relatively low value of log Kd (2.6


to 3.0) was calculated for diclofenac, carbamazepine and hydrochloro-thiazide, indicating their poor removal efficiencies via sorption. In gen-eral, the higher biotransformation of EOMs was followed by lowersorption potential, which was not in accordance with the calculatedlog Kd. These results suggest that log Kd values may vary plant-to-plant depending on operating conditions and the extent of removal ef-ficiencies (Kim et al., 2014). Indeed, the biotransformation of EOMs isenhanced by the sorption process due to the longer retention time ofsolids in the MBR process (Tadkaew et al., 2011).

3.5. Other factors influencing the removal of EOMs in MBR

Hyrophphobic interaction is also one of the key mechanisms con-trolling biosorption of EOMs, hence the removal efficiencies inMBRpro-cesses (Cirja et al., 2008). Most of the EOMs quantified in the influentload at both testing periods were hydrophilic (log D pH 7 b 3.2) (Haiet al., 2018; Tadkaew et al., 2011; Wijekoon et al., 2013) ranging frompharmaceuticals to steroid hormones. Of the four hormones, only tes-tosterone, estrone and progresterone are hydrophobic (log D pH 7

N 3.2) (Fig. 3). Hydrophobicity can be defined by log Kow for non-ionicand log D for ionic compounds (Taheran et al., 2016). All three hydro-phobic EOMs showed consistently high removal efficiency of 96–99%(median 97%). Similar results have been reported in other studies(Suárez et al., 2012). The higher removal efficiency of hydrophobic com-pounds might be due to: (i) the dominating mechanism of sorption tothe sludge that facilitates enhanced biological degradation, and (ii)possessing only e-donating groups (EDGs) that enhance better oxida-tion (Tadkaew et al., 2011). High log D is characteristic of hydrophobicEOMs, poor hydrosolubility and high sorption tendency on organic con-stituents of sludge matrix (Cirja et al., 2008). Moreover, the calculatedlog Kd values of these EOMs were above 4, indicating substantiallyhigh sorption affinity to the sludge.

On the other hand, the removal of hydrophilic EOMs (log D pH 7 b 3.2)varied from ‘limited removal’ (carbamazepine and hydrochlorothiazide)to almost complete removal (paracetamol, caffeine and ibuprofen). It isvery obvious that the removal of hydrophilic EOMs varies significantlysince they have diverse molecular structures and functional groups(Table S1). Varying removal efficiencies accounted for in hydrophilicEOMs can be explained based on the qualitative framework by Tadkaewet al. (Tadkaew et al., 2011) as: (i) compounds possessing only EDGs(slightly hydrophobic such as estriol) could achieve high removal(~99%); (ii) compounds possessing both EDGs and e-withdrawing(EWGs) such as amide and chloride in their molecular structure(Tadkaew et al., 2011;Wijekoon et al., 2013) including paracetamol, ibu-profen, diclofenac, hydrocortisone, naproxen, propranolol, trimethoprim,ciprofloxacin, bisoprolol, caffeine, furosemide, metoprolol, tetracycline,atenolol and iopamidol with varying removal efficiencies ranging from37 to 99.9%; (iii) compounds possessing only strong EWGs such as hydro-chlorothiazide and carbamazepine (0–10%). Therefore, the dominatingremoval mechanism of hydrophobic compounds is strongly influencedby their intrinsic biodegradability as sorption to sludge is less significant(Tadkaew et al., 2011).

Another aspectmay be the degree of ionisation (pKa) of the selectedEOMs, which controls their ionisation state as a consequence of pH var-iation in the solution. For EOMs including ibuprofen, diclofenac,naproxen, ketoprofen, enalapril, caffeine, furosemide, tetracycline andciprofloxacin possessing a pKa of between 0.52 and 6.43 (Table S1), re-moval efficiencies varied from38 to 99%. The results are partially consis-tent with the previous work, where acidic EOMs showed low removalefficiency in neutral pH condition. This might be related to the factthat these EOMs are negatively charged at neutral pH due to the pres-ence of carboxyl functional group in their structures (Urase et al.,2005). Likewise, EOMswith pKa values of 7 to 15 (Table S1) showedvar-ied removal efficiencies ranging from ‘non-removal’ to N97%, with the

Fig. 4. Fate of the selected EOMs during MBR process at two different SRTs. The log Kd values to the right showed the distribution of calculated sludge-water distribution coefficients/sorption coefficients of EOMs.


lowest being carbamazepine and hydrochlorothiazide and the highestbeing paracetamol. Carbamazepine and hydrochlorothiazide bothhave a basic amine functional group, which may be protonated to gaina positive charge under neutral pH (Nghiem and Khan, 2007; Uraseet al., 2005). This phenomenonmay have increased their hydrophilicitywhich was confirmed by their noticeably higher concentrations in theMBR effluents (Fig. S1).

The molecular weights (MWs) of the selected EOMs varied from151 g moL−1 (paracetamol) to 777 g moL−1 (iopamidol) in this study(Table S1). However, the removal tendencies of EOMs in relation toMWs were not clear. For example, paracetamol and caffeine were al-most completely removed although their MWs were the lowest onthe list (151 and 194 g moL−1), whereas the removal efficiencies ofother EOMs with MWs ranging from 206 to 777 g moL−1 varied fromnon-removal to N99.6%. EOMs with higher MW may provide morebranches that allow microbes to selectively cleave certain target sitesand subsequently initiate degradation (Tadkaew et al., 2011).

Furthermore, the average MLSS concentrations inside the reactorwere 8.5 ± 0.8 g L−1 and 3.7 ± 1.1 g L−1 at SRTs of 60 days and21 days, respectively. Results showed that the removal of metoprolol,hydrochlorothiazide, propranolol and carbamazepine was enhancedwhile operating the MBR at high SRT (i.e. high MLSS concentration)(Fig. 3). At increasing SRTs, the amount of sludge wasted from the bio-reactor decreases, which allows longer retention and subsequent degra-dation of less polar EOMs in the system. The higher specific areaprovided by increasingMLSS concentration enablesmore enzymatic ac-tivities (Cirja et al., 2008).

3.6. Influence of SRTs on sludge characteristics

Fig. 5 shows the variations in SMP and EPS during the MBR treat-ment at two different SRTs. Although the TMP profile of theMBR systemwas quite stable during the testing periods of both the SRTs, noticeablechanges in the concentrations of SMP and EPSwere observed. SMP con-centration increased from 5.6 to 15.5 mg gMLVSS−1 (about 3 times)while SRT was lowered from 60 days to 21 days. Similarly, more EPSwere released at an SRT of 21 days (103mg gMLVSS−1) than the longerSRT of 60 days (51 mg gMLVSS−1). A comparable trend of EPS variationwas reported by other researchers (Massé et al., 2006). As SRT in-creased, concentrations of SMP and EPS decreased as most of the sub-strates were consumed for the metabolism of microorganisms (Parket al., 2015). Nevertheless, extremely high SRTs can accumulate highsoluble and particulate matters due to the increase in MLSS concentra-tion, leading to an increase in viscosity that demands high scouring airfor membranes (Krzeminski et al., 2012). However, since effluent flowwas relatively stable, which substantiates the robustness of the MBR

process, no clear correlation of SMP and EPS to membrane fouling wasobserved in this study. The consistent rate of fouling might have beenadequately mitigated by continuous physical membrane-cleaning pro-cesses such as intermittent membrane relaxation and the use of scour-ing air involved in the pilot operation. Many researchers reported thatthe release of SMP and EPS concentration decreases as the sludgestays longer in the bioreactor, and increases when lowering the SRT(Grelier et al., 2006; Meng et al., 2009). The presence of EOMs in MBRmay affectmicrobial activities and the structure community of microor-ganisms in activated sludge, which increases the endogenous respira-tion rates and subsequently releases EPS (Avella et al., 2010).However, studying the extent of EOM influence on releasing SMP andEPS was out of the scope of this study.

Furthermore, the dewaterability of MBR sludge was also assessed atboth SRTs in terms of their corresponding CST values (Fig. 5). The aver-age CST of the sludge operated at an SRT of 60 days (31 s) was abouttwice as high as at an SRT of 21 days (15 s). An increasing trendof sludgeCST with increasing SRT is reported by other researchers, which indi-cates that the sludge with high CST is difficult to dewater due to thehigh content of bound water (Ng et al., 2006). Also, increasing SMPand EPS concentrations indicated thedecreasingCST of sludge due to in-creased fractions of fine and colloidal particles, and vice versa (Grelieret al., 2006).

3.7. MBR effluent quality in compliance with Australian and Swissguidelines

To evaluate the quality of MBR-treated effluent in terms of EOMsconcentrations, the observed concentrations of EOMs were studied incompliance with the Australian (NHMRC/EPHC/NRMMC, 2008) andSwiss guideline values (Oekotoxzentrum Centre Ecotox, 2016)(Table 3). Nevertheless, no guidelines or standards have yet been pro-posed for EOMs selected by the EU. A wider range of pharmaceuticalsand steroid hormones was established in the Australian guidelinesthan in the Swiss guidelines. Most of the studied EOMs were observedwell below the guideline values of both the standards (Table 3). No ste-roid hormoneswere detected above their LOQs, whichwere lower thanboth the guidelines. Among diuretics, enalapril was detected below theLOQ (10 ng L−1), whereas hydrochlorothiazide and furosemide werefrequently detected in the range of 185–1700 ng L−1. Similarly, immu-nosuppressive (hydrocortisone) was not detected above its LOQ, butno guidelines for diuretics and immunosuppressives are available orhave been proposed so far. Caffeine was detected at a concentration of60–328 ng L−1, which was lower than the guideline value of350 ng L−1, but it showed excellent removal efficiency of N99% in MBRtreatment since the influent concentration was very high (140,437 ±6453 ng L−1). Regarding poorly removed carbamazepine, the observedconcentration of 480–665 ng L−1 was lower than both guidelines. Over-all, the results indicated that the quality of effluent by MBR treatmentfulfilled the water quality requirements for reuse purposes, regardlessof varying SRTs.

4. Conclusion

The removal and fate of selected EOMs by the MBR pilot plant atvarying SRTs were studied. Only 23 EOMs were detected in the influentwastewater during the testing periods above their LOQs. Diverse re-moval dynamics for selected EOMs were observed. The application ofmass balance assessment proved that biotransformation and sorptiononto sludge were the dominant mechanisms associated with the re-moval of EOMs. Even though the EOMs removal efficiencies were notthat significant at SRTs of 21 and 60 days, enhanced removal of the ma-jority of EOMs were observed at longer SRT. Therefore, SRT of 60 dayswould be preferred to be implemented, which would additionally helpin releasing less concentrations of SMP and EPS. This study has ex-panded the understanding of the removal and fate of the EOMs (someFig. 5. Variation of SMP, EPS fractions, and CST as a function of varying SRTs.


of which were never or rarely studied before in MBR) via MBR treat-ments under different operating conditions (i.e. SRTs), which certainlyhelps in achieving more sustainable water management.

Acknowledgements

The authors would like to thank Business Finland Oy(Innovaatiorahoituskeskus) (formerly Tekes- the Finnish FundingAgency for Technology and Innovation) for funding this work as partof the Smart Effluents Project. Special thanks to Reijo Turkki, Risto Repo,Anne Bergman and Päivi Luukkonen for their continuous support in op-erating and maintaining the MBR pilot plant during the sampling pe-riods. We also thanks Eurofins Environment Testing Finland Oy for itslaboratory analysis support.

Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2019.02.308.

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Table 3Comparison of median concentrations of EOMs inMBR effluentwith Australian and Swissguidelines.

Compound Concentration (ng L−1)

MBREffluent -SRT 60(SRT 21)

Australian guidelinevalues(NHMRC/EPHC/NRMMC,2008)

Swiss guidelinevalues(OekotoxzentrumCentre Ecotox,2016)

Iopamidol 200 (350) 400 × 103 n.p.Atenolol 4 (25) n.a. 330 × 103

Tetracycline b10 (b10) 105 × 103 n.p.Metoprolol 100 (350) 25 × 103 750 × 103

Furosemide 185 (230) n.a. n.p.Caffeine 328 (60.5) 350 n.p.Bisoprolol 11 (19.5) 630 × 103 n.p.Ciprofloxacin 50 (51.5) 250 × 103 360Enalapril b10 (b10) n.a. n.p.

Hydrochlorothiazide1700(1350)

n.a. n.p.

Ketoprofen 5 (5) 3.5 × 103 n.p.Trimethoprim 28.5 (42) 70 × 103 210 × 103

Propranolol 33 (51.5) 40 × 103 12 × 103

Paracetamol 325 (50) 175 × 103 n.p.Naproxen 19.5 (84) 220 × 103 860 × 103

Ibuprofen 59 (50) 400 × 103 1700 × 103

Hydrocortisone b10 (b10) n.a. n.p.Diclofenac 760 (1050) 180 × 101 n.p.Carbamazepine 480 (665) 100 × 103 2000 × 103

Estriol b5 (b5) 50 n.p.Testosterone b1 (b1) 7 × 103 n.p.Estrone b5 (b5) 30 n.p.Progesterone b1 (b1) 105 × 103 n.p.

n.a. values not available; n.p. not proposed; ‘b’ denotes below LOQ.


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Yang, Y., Ok, Y.S., Kim, K.-H., Kwon, E.E., Tsang, Y.F., 2017. Occurrences and removal ofpharmaceuticals and personal care products (PPCPs) in drinking water and water/sewage treatment plants: a review. Sci. Total Environ. 596–597, 303–320. https://doi.org/10.1016/j.scitotenv.2017.04.102.

Ying, G.-G., Kookana, R.S., Kolpin, D.W., 2009. Occurrence and removal of pharmaceuti-cally active compounds in sewage treatment plants with different technologies.J. Environ. Monit. 11, 1498–1505. https://doi.org/10.1039/B904548A.


Publication III

Khum Gurung, Mohamed Chaker Ncibi, Marina Shestakova, Mika Sillanpää

Removal of carbamazepine from MBR effluent by electrochemical

oxidation (EO) using a Ti/Ta2O5-SnO2 electrode


Applied Catalysis B: Environmental

Vol. 221, pp. 329-338, 2018

© 2018 Elsevier


Applied Catalysis B: Environmental

journal homepage: www.elsevier.com/locate/apcatb

Research Paper

Removal of carbamazepine from MBR effluent by electrochemical oxidation(EO) using a Ti/Ta2O5-SnO2 electrode

Khum Gurunga,⁎, Mohamed Chaker Ncibia, Marina Shestakovaa, Mika Sillanpääa,b

a Laboratory of Green Chemistry, School of Engineering Science, Lappeenranta University of Technology, Sammonkatu 12, FI-50130, Mikkeli, Finlandb Department of Civil and Environmental Engineering, Florida International University, Miami, FL, 33174, USA

A R T I C L E I N F O

Keywords:Electrochemical oxidationTi/Ta2O5-SnO2 electrodesCarbamazepineOperating parametersEnergy consumptionMBR effluent

A B S T R A C T

This study aims at investigating the electrochemical oxidation (EO) of carbamazepine (CBZ) synthetic solutionsand real membrane bioreactor (MBR) effluent using newly developed Ti/Ta2O5-SnO2 electrodes to enhance CBZremoval. The characterization of the prepared Ti/Ta2O5-SnO2 electrodes was performed by using scanningelectron microscope, energy dispersive X-ray spectroscope, atomic force microscope, and cyclic voltammetryanalyses. The main operating parameters influencing the CBZ removal efficiency in synthetic solutions using Ti/Ta2O5-SnO2 electrodes were evaluated including the applied current density, initial CBZ concentration, pH, andtemperature. The optimum removal of CBZ (20 mg L−1) and TOC reached 75.5%, and 71.1%, respectively, after8 h of electrolysis, under current density of 9 mA cm−2, pH 6, temperature of 30 °C, and using 0.1 M Na2SO4 assupporting electrolyte. Increasing current density and temperature influenced the CBZ removal, unlike pH whichdid not have significant influence on CBZ removal. The performance of Ti/Ta2O5-SnO2 electrode was comparedwith conventional Ti/PbO2 electrode in terms of CBZ removal efficiencies and stability of the electrodes. Theresults showed that under the same operating conditions, the CBZ removal efficiency of Ti/PbO2 electrode wasslightly higher than of Ti/Ta2O5-SnO2 (77.9 and 71.7%, respectively). Nonetheless, the use of this newly de-veloped electrode is more energy-efficient as it required the lowest energy consumption of 60.3 kWh m−3 toachieve optimum CBZ removal, in addition to fact that no heavy metals were leached (unlike the PbO2 elec-trode). Furthermore, a complete degradative removal of real MBR effluents spiked with CBZ was achieved whenelectrolyzed under the optimized conditions of CBZ synthetic solutions. Overall, the EO based on the use of Ti/Ta2O5-SnO2 electrode was found to be a reliable approach to remove CBZ from contaminated waters, withpromising potential for integration with MBR technology to remediate CBZ.

1. Introduction

Pharmaceutically active compounds (PhACs) have gained a greatdeal of attention among researchers in recent years as emerging mi-cropollutants with long-term potential threats to aquatic environmentand human health [1–4]. These PhACs are introduced to aquatic eco-systems primarily from the effluents of wastewater treatment plants(WWTPs). In addition, direct and unsafe discharge of untreated was-tewaters from agriculture, industries and hospitals are also reported tocontribute to the release of PhACs into the aquatic ecosystems[2,3,5–7].

Carbamazepine, which is used for the treatment of epilepsy, de-pression, trigeminal neuralgia and wide variety of mental disorders, isone of the frequently identified PhAC in various aquatic environments[2,3,7]. The world-wide annual consumption of CBZ is more than 1000tons and about 28% of it ends up in WWTPs as unmetabolized forms.

[3,5]. In Finland, the annual consumption of CBZ is around 3.3 tons,with comparatively more concentrations in winter seasons than insummer at WWTP effluents [8]. CBZ is highly recalcitrant to biode-gradation in conventional activated sludge processes and photo-de-gradation (more than 100 days) [2]. In fact, many researchers havedocumented difficulties in removing CBZ, with biodegradation removalof only less than 10% [3,4,6,8]. Thus, it is crucial to apply highly ef-ficient degradation technologies for the removal of CBZ in order toachieve the highest mineralization or generate non (or less) toxic in-termediates.

In this context, MBRs have been studied for the removal of variousPhACs including CBZ by many researchers [9–11]. However, a rela-tively low removal (0–28%) of CBZ was reported in MBR processeswhen treating synthetic and real wastewaters [2,9,11]. Thus, biode-gradation and membrane filtration alone are not efficient enough forthe removal of CBZ. Therefore, in order to meet the future legislations

http://dx.doi.org/10.1016/j.apcatb.2017.09.017Received 1 July 2017; Received in revised form 29 August 2017; Accepted 5 September 2017

⁎ Corresponding author.E-mail addresses: [email protected] (K. Gurung), [email protected] (M. Sillanpää).

Applied Catalysis B: Environmental 221 (2018) 329–338

Available online 06 September 20170926-3373/ © 2017 Elsevier B.V. All rights reserved.

MARK

regarding the stringent limits of PhACs from WWTP effluents, MBRprocess integrated with various advanced oxidation processes (AOPs)has gained increasing interest.

AOPs are oxidation-based technologies applied in aqueous phase inorder to destruct organic pollutants via the in-situ generated highlyreactive oxygen species, mainly hydroxyl radicals (%OH) [12]. They arepromising and environmentally friendly methods for the treatment ofwastewaters contaminated with a wide range of recalcitrant PhACs.Among the various AOPs, EO processes are of growing interest for waterand wastewater decontamination due to their low cost and high effi-ciency [3]. In EO processes, dissolved organic pollutants are mainlyoxidized by (i) direct anodic oxidation on the anode surface by chargetransfer, and (ii) reaction with physio and/or chemisorbed hydroxylradical (%OH) generated from water oxidation (Eq. (1)) [13,14]. Hy-droxyl radicals are extremely reactive oxidants (2.8 V (vs. SHE)) withthe life time of less than a second, which makes their detection verychallenging [15]. The subsequent oxidation mechanism of organiccompounds depends on the behavior of anodes, which are divided as“non-active” and “active” [13–16]. In both cases, firstly, water mole-cules splitted and leading to the formation of OH at the anode oxidesurface (MOx) (Eq. (1))[17].

H2O + MOx =MOx (%OH) + H+ + e− (1)

At ‘active’ anodes, OH radicals are chemisorbed (strongly adsorbed),thus allow only selective oxidation of organics (R), which involves in-termediate step of higher oxide formation (Eqs. (2) and (3))[14,16].

MOx (%OH) → MOx+1 + H+ + e− (2)

MOx+ 1 + R→ MOx + RO (3)

On the other hand, at ‘non-active’ anodes, the formation of higheroxide is excluded and OH radicals are physisorbed without interactionon the anode surface [14,16]. Moreover, ‘non-active’ anodes with highO2-overpotential towards water oxidation allow good electrochemicalstability and can generate many strong (OH) radicals predominantlyleading to the complete combustion of organics (Eq. (4)) [15–17].

MOx (%OH) + R = MOx + CO2 + H2O + zH+ + ze− (4)

The major problems of EO process are the polarization and passi-vation of electrodes, related either to the poor mass transfer, corrosiveenvironment, formation of polymer films or the inevitable formation ofoxide interlayers due to the considerable porosity in supported catalystfilms, [14,15,18]. In any cases, those phenomena result in the at-tenuation of the efficiency and service life of electrodes.

Various anodes such as PbO2, and more recently, boron-dopeddiamond (BDD) have shown an interesting electro-degradation effi-ciency for CBZ [2,3,6]. Nevertheless, the higher cost of BDD electrodesand the possible leaching of lead (Pb) from PbO2 coated electrodes tendto compromise their applicability in large-scale applications [19,20]. Infact, PbO2 is one of the widely studied anode on the electrochemicaldegradation of PhACs [2,6,21,22]. PbO2 anodes are inexpensive, rela-tively easy to prepare, has low electrical resistivity, and has a largesurface area [23]. In this context, Ti/Ta2O5-SnO2 (SnO2 semi-conductordoped with Ta2O5) are eco-friendly electrodes, which were proven to beefficient in the oxidation of some organic compounds [18,24].

The main aim of the current research is to evaluate the efficiency ofthis novel electrode to remove persistent CBZ drug from synthetic so-lutions and the real MBR effluent. First, the EO of synthetic solutionloaded with CBZ was investigated by monitoring and optimizing themajor influencing factors including current density, pH, and tempera-ture, under various initial concentrations of CBZ. In addition, the de-gradation efficiency and stability of Ti/Ta2O5-SnO2 were comparedwith Ti/PbO2 electrodes. Furthermore, optimized operating conditionsin EO process were applied for the treatment of real MBR effluentspiked with CBZ. This study is the first research work on the EO of CBZusing novel Ti/Ta2O5-SnO2 electrodes in synthetic solutions and real

wastewaters.


2.1. Reagents and materials

All the reagents and materials used were of analytical grade. CBZ(meets USP testing specifications), Titanium foil (99.7% trace metalbasis, 2 mm thick), SnCl2·2H2O (≥99.99% trace metal basis), TaCl5(≥99.99% trace metal basis), NaOH (≥98% anhydrous), H3PO4 (85%),Na2SO4 (≥99% anhydrous) were purchased from Sigma-Aldrich, USA.Commercial Ti/PbO2 electrodes (5 cm × 10 cm× 2 mm) were sup-plied by Magneto, Netherlands.

A synthetic stock solution containing CBZ (20 mg L−1) was pre-pared using Milli-Q water (18.2 MΩ.cm), stirred at 500 rpm for 24 h atroom temperature [5]. Then, 0.1 M (14.2 mg L−1) of Na2SO4 (bettersupporting electrolyte than NaCl regarding economical, effectivenessand environmental aspects) was added to increase conductivity of thesolution. In this regard, the possible formation of organic chlorine by-products is also limiting the use of NaCl in electrochemical processes[5]. Finally, the synthetic stock solution was stored in the refrigerator at4 °C for further use.

2.2. Preparation of Ti/Ta2O5-SnO2 electrodes

Ti/Ta2O5-SnO2 electrodes were prepared by thermal decompositionof precursor solutions on Ti substrates [25]. For the preparation of theelectrodes, the methodology is detailed in a previous study conductedin our laboratory [26]. The optimized nominal composition of Ta (7.5at.%) and Sn (92.5 at.%) were used to prepare precursor solutions bydissolving SnCl2·2H2O and TaCl5 salts in absolute alcohol (pro analysis,Panreac), where total concentration of metal ions was kept constant(0.04 M). At first, Ti substrate plates (5 cm × 10 cm× 2 mm) werepre-treated by mechanical polishing and followed by degreasing in10 wt.% NaOH for 10 min and etching in boiled 18 wt.% HCl (proanalysis, Fluka) for 30 min, and finally rinsed carefully using Milli-Qwater. Secondly, the precursor solution was applied over only one sideof the pre-treated Ti substrates by drop casting method. The precursorsolution of 1 mL volume was dropped on to the Ti substrate, uniformlydistributed using the tip of a Pasteur glass pipette, dried in oven at 80 °Cfor 5 min, and then annealed in muffle furnace at 550 °C for 5 min. Theprocess was repeated until 8 layers of ultrathin film deposition of theprecursor solution, which attributed highest electro-catalytic propertiesof the electrodes [26]. After the final layer, electrodes were additionallyannealed at 550 °C for 10 h to enhance the formation of compositeoxides of Ta and Sn on the electrode surface via thermal decomposition.The final composite oxides on Ti substrate gave the wide band gap n-type SnO2 semi-conductor extrinsically doped with Ta2O5. After 10 h,electrodes were cooled down and kept in an ultrasonic bath for 5 min inorder to remove the surface impurities and deformations, and then ovendried at 105 °C. Thus, the precursor material deposited only one side ofTi substrates had surface area of 50 cm2 and the deposited material wasequal to 1.10 mg cm−2 ± 5%.

2.3. Physicochemical and electrochemical characterization of the electrodes

Microstructure of the Ti/Ta2O5-SnO2 electrodes was analyzed byusing scanning electron microscope (SEM) (Hitachi, SU3500, Japan)using 5 kV of acceleration voltage. The chemical analysis of selectedareas of the electrode surface was performed with energy-dispersive X-ray spectroscope (UltraDry™ Compact EDS Detector, Thermo Fisher,USA). The AFM images were taken by using atomic force microscope(AFM) (NX10, Park, Korea) using the non-contact mode.

The electrochemical characterization of the Ti/Ta2O5-SnO2 elec-trode was performed by conducting cyclic voltammetry (CV) mea-surements using an Autolab (PGSTAT 12 Potentiostat/Galvanostat,

K. Gurung et al. Applied Catalysis B: Environmental 221 (2018) 329–338

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Metrohm, Switzerland) connected with GPES software (Eco Chemie,Netherlands). All the CV measurements were performed in a threeelectrode cell (100 mL). A coiled platinum wire was used as counterelectrode and saturated calomel electrode (SCE) as reference electrode,whereas Ti/Ta2O5-SnO2 electrode (1 cm× 1 cm) was used as workingelectrode. First, the voltammograms were recorded at different scanrates from 20 to 100 mV s−1 in 0.1 M Na2SO4 aqueous solution between0.6 and 0.8 V (vs. SCE), where only double-layer currents (non-faradic)were observed for estimating the relative roughness factor. Second, thevoltammograms were plotted between the anodic currents againstelectrical voltage between 0.2–2.5 V (vs. SCE) at scan rate of100 mV s−1 with or without CBZ in 0.1 M Na2SO4 aqueous solution. Ineach of the assays, aqueous solution inside the CV cell was deox-ygenated by argon bubbling prior to the measurements.

2.4. Electrochemical degradation assays

The experimental set-up for the electrochemical degradation systemis shown in Fig. S1. The CBZ-containing solutions were electrolyzedusing an open glass reactor of 500 mL capacity with cooling jacket forthe circulation of thermostated water, under a 500 rpm stirring. Thereaction temperature was maintained at 11 ± 1 °C. The electro-chemical unit consisted of Ti/Ta2O5-SnO2 electrode as anode and pureTi plate (same area) as cathode, with an inter-electrode gap of 10 mm.The electrical power was applied via DC power supply (EX752M-Multimode PSU, TTI, UK) with maximum current and potential ratingof 4A and 150 V, respectively. All the electrochemical degradation as-says were performed under galvanostatic conditions. Each experimentwas run for 8 h with sampling every hour.


2.5.1. CBZ measurementsThe removal of CBZ was analyzed by UV–vis spectrophotometer

(Lambda 45, Perkin Elmer, USA) by measuring the absorbance of UVlight by CBZ at 210 nm. A calibration curve was created by scanningdifferent concentrations of CBZ from 1 to 30 mg L−1 (Fig. S1,

Supplementary data) in order to calculate the residual CBZ concentra-tion and thus the removal efficiency. As for the real MBR effluent, sincethe spiked samples have low concentration of CBZ (10 μg L−1), theywere analyzed with LC–MS/MS (Acquity UPLC/Xevo TQMS, USA) using5 cm C18 column (Acquity UPLC BEH, USA).

2.5.2. Other measurementsNon-Purgeable Organic Carbon (NPOC) was measured with TOC-

analyzer (TOC-V Series-CPN, Shimadzu, Japan) to evaluate the miner-alization efficiency. The pH was measured using pH meter (pH 110,VWR). The electrical conductivity was monitored by using conductivitymeter (Euroscan Con 6, EUTECH, USA).

The energy consumption (EC) needed for the electrochemicaltreatment per volume of working solution was calculated based on Eq.(5) [15].

=−EC (kWh m ) U It

V3 cell

s (5)

where Ucell is the average potential difference of the cell (in V), I is theapplied current (in A), t is the electrolysis time (in h), and Vs is theworking solution volume (in dm3).


3.1. Microstructural characterization

SEM and AFM micrographs of Ti/Ta2O5-SnO2 electrode before andafter the 8 h of electrolysis using the current density of 9 mA cm−2 areshown in Fig. 1a-f. In SEM images, the electrode surface before elec-trolysis (Fig. 1a-b) showed foam-like porous microstructure with largenumbers of relatively homogenous agglomerates mainly related to thedensity of Ta2O5 crystals, which can partially increase the surfaceroughness. In fact, the porosity and relative surface roughness dependon the active oxide layers. When the electrode was subjected to elec-trolysis for 8 h, the surface exhibited structural changes in the surfaceshowing heterogeneously-distributed porous network and more roughstructure as shown in Fig. 1d–e.

Fig. 1. SEM and AFM images of Ti/Ta2O5-SnO2 electrodes before electrolysis (a–c) and after 8 h of electrolysis (d–f) under current density of 9 mA cm−2.


331

Likewise, the AFM analysis before electrolysis shows a more eventopography of the electrode surface as shown in Fig. 1c. Whereas, thesurface structure of the electrode after the 8 h of electrolysis depictsmore deformed surface with visible deep and extended valley (darkspots in Fig. 1f). The depth of undulations was noted to be±1.5 μmand ± 4 μm on the electrode surface before and after the electrolysis,respectively.

3.2. Elemental characterization

The EDX elemental microanalysis was conducted to determine thebulk composition of metal oxide thin films on Ti/Ta2O5-SnO2 electrode.The EDX elemental mappings on the surface of Ti/Ta2O5-SnO2 elec-trode under the current density of 9 mA cm−2, before and after the 8 hof electrolysis, are as shown in Fig. 2a–c and d–f, respectively. Theatomic molar percentages of the precursor metals from EDX were foundto be Sn- 91.7 at.% and Ta-8.3 at.% before and Sn-91.5 at.% and Ta-8.5at.% after the electrolysis. The elemental composition of metals showed

reasonable agreement with the nominal mixture used [26]. Moreover,the atomic ratio of Sn/Ta from EDX (10.9) was more closed to thenominal ratio (12.33). This suggests that the Ti substrate is coveredmore uniformly with Sn, whereas Ta deposition is more dispersed.Overall, the metal fractions did not change noticeably before and afterthe electrolysis, though morphological changes existed.

3.3. Electrochemical characterization

The electrocatalytic activity of any electrode depends mostly onfactors such as relative roughness factor of the electrode surface and thevoltammetric currents associated with the degree of electroactive sites[27]. For the roughness factor, the current values for different sweeprates were measured at E = 0.7 V (vs. SCE), where an approximatesymmetry for anodic and cathodic currents was recorded as shown inFig. 3a. By plotting the anodic currents with respect to the scanningrates, a regression equation is deduced (shown in Fig. 3a, inset) as,

= + × =− −vI (A) 0.0001 (Vs ) 2.6 10 , R 0.99771 6 2 (6)

Fig. 2. EDX mappings of Ti/Ta2O5-SnO2 electrodes before (a–c) and after 8 h of electrolysis (d–f) under current density of 9 mA cm−2.

Fig. 3. Cyclic voltammograms of Ti/Ta2O5-SnO2 in(a) double-layer region (0.6–0.8 V) at scan rates from20 to 100 mV s−1 in 0.1 M Na2SO4 solution. Inset:Linear regression of I versus v measured atE = 0.7 V; (b) CBZ (30 mg L−1, pH 5.7) with sup-porting electrolyte 0.1 M Na2SO4, scanned between0.2–2.5 V at scan rate of 100 mV s−1.


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The linear relationship between electric current and sweep rateshows that the Faradic current in this region is negligible, thus only theelectric double-layer current is measured [27]. The double-layer char-ging current is much related to the double-layer capacitance of theelectrode/solution interface as conductive layer was made of metaloxides [28]. So, the normalized capacitance of 0.0001 F cm−2 wascalculated by dividing the slope value of Eq. (6) with the geometricsurface area of the electrode. Further, the capacitance of the electrodesurface was compared with that of a smooth surface oxide layer, withthe capacitance of 8 μF cm−2 [27], which gave a relative roughnessfactor of 12.5. Therefore, the roughness of electrode surface has slightlyimproved by Ta doping over SnO2 layer, which in turn have attributed agood electrocatalytic activity to the Ti/Ta2O5-SnO2 electrodes. How-ever, roughness factors of metal oxide films might vary by factor ofseveral thousands with high discrepancy in magnitude between them[28].

Furthermore, CV tests were again performed to investigate theelectrochemical property of CBZ molecule at Ti/Ta2O5-SnO2 electrode,as shown in Fig. 3b. The oxygen evolution reactions (OERs) appearedbetween 1.7 and 2 V (vs. SCE) in both the solutions. In the presence ofCBZ, a small anodic current peak was noticed at potential of 1.45 V (vs.SCE) in the voltammogram (Fig. 3b). The absent of more distinct peaksfor CBZ oxidation might probably be due to the CBZ oxidation at apotential close to OER region. Furthermore, it can be noticed that therewas no reduction peak recorded in the voltammogram during reversescanning, which might be due to the irreversible characteristics of CBZ[29].

3.4. Effects of operating parameters on the removal of CBZ

3.4.1. Effect of current densityIn electrochemistry, applied current density is the most important

parameter which controls electron transfer and the generation of re-active oxidants [30], thus directly influences the removal rates of pol-lutants. Such major influencing role of current densities during theelectrochemical removal of CBZ has been reported in many studies[5,6,20,23,31]. In this study, different current densities ranging from 1to 17 mA cm−2 were applied in order to select an optimum currentdensity with higher rate of CBZ removal as shown in Fig. 4.

It was observed that after the 8 h of electrolysis, the CBZ removalreached 53.0%, 67.0%, 68.5%, and 71.7% with corresponding currentdensities of 1, 5, 7, and 9 mA cm−2, respectively. This trend is in ac-cordance with the fact that increasing current density increases theorganic pollution removal efficiency by the generation of more reactiveoxygen species (esp. %OH radicals) [3,7,23]. The generated powerful %OH radicals then attack CBZ molecules and destruct it. Likewise, theCBZ removal rates were 73.8% and 72% at current densities of 15 and

17 mA cm−2, respectively. Thus, beyond current density of 9 mA cm−2,neither removal rate nor the kinetics increased noticeably. Two possiblereasons might have associated with this phenomenon. First, the un-desirable parasitic reactions of %OH radicals to O2 gas or dimerizationto weak oxidants such as %OH2 radicals, which could compete the mainoxidation reaction (via %OH radicals), and can reduce the oxidationefficiency of the targeted pollutants [6,14,23,32]. Second, as anodicoxidation takes place heterogeneously with physio-adsorbed %OH ra-dicals at electrode surface, electrochemical degradation rate afterreaching the limiting current density might be directly proportional tothe mass transfer rate of pollutants towards the electrode surface [23].The current density of 9 mA cm−2 was selected as optimum currentdensity for further study to ensure both an improved removal efficiencyand a low energy consumption.

Furthermore, in order to evaluate the degradation kinetics of CBZ,the kinetic data were analyzed by using different kinetic models:pseudo-first order kinetic model and pseudo-second order kineticmodel. The following Eqs. (7) and (8) were used for the pseudo-firstorder and pseudo-second order kinetic models,

=Ln CC

k tt

01 (7)

− =1C

1C

k tt 0

2 (8)

where C 0 is the initial concentration of CBZ; C t is the concentration ofCBZ at time t; t is the reaction time; k1 is the pseudo-first order rateconstant (s−1); and k2 is the pseudo-second order rate constant(L mol−1 s−1). The regression coefficients (R2) for both pseudo-firstorder and pseudo-second order kinetics were compared correspondingto the different current densities as indicated in Table 1. The calculatedvalues of R2 from the pseudo-first order kinetic model were slightlylower than the value obtained from the pseudo-second order kineticmodel. These results indicated that the pseudo-second order modelbetter describes the degradation of CBZ by EO process. This is in

Fig. 4. Effect of applied current densities on the removal of CBZduring electrolysis (CBZ (0): 20 mg L−1 + 0.1 M Na2SO4; pH-5.7;T = 11 ± 1 °C) on Ti/Ta2O5-SnO2 anodes. Inset: fitting of the dif-ferent current densities to a pseudo-second order kinetic model.

Table 1CBZ removal efficiency (%), regression coefficient (R2), pseudo-rate constant (kpseudo),and half-life, as a function of current densities.

Current densities (mA cm−2) 1 5 7 9 15 17

% CBZ removal 53.0 67.0 68.5 71.7 73.9 72.1Pseudo-First order model, R2 0.998 0.971 0.980 0.971 0.970 0.950Pseudo-Second order model, R2 0.997 0.996 0.993 0.995 0.991 0.983k2 (L mol−1 s−1) 0.47 0.91 1.12 1.42 1.57 1.46Half-life (h) 6.9 3.6 2.9 2.3 2.1 2.3


333

agreement with the previous work by J. D. García-Espinoza et al. [6],who also reported that the second-order was the best model to fit thekinetics in the electrochemical removal of CBZ with Ti/PbO2 anodes.However, the pseudo-first order kinetic model for CBZ removal using anEO process have also been reported in other works [5,34,35]. In thepresent work, since R2 values for both the models are comparable, thepseudo-first order model is also valid to describe the reaction kinetics ofCBZ degradation. In this regard, it is worth noting that within 8 h of thereaction time, the formation of intermediate compounds could slightlydistort the kinetic analysis.

The tendency of the enhanced CBZ removal efficiencies corre-sponding to the increased current densities are further supported withthe results of pseudo-second order rate constant values, k2 (Table 1).The increasing trend in the values of k2 as a function of applied currentdensities indicates that amplified OERs are associated with increasedcurrent densities. With the increasing O2 gas evolution, the masstransport phenomena increases indicatively leading to the enhanceddiffusion flux of CBZ molecules towards anode surface, which subse-quently increases k values [20]. From Table 1, at optimum currentdensity of 9 mA cm−2, the value of k2 was 1.42 L mol−1 s−1 and thehalf-life was found to be 2.3 h. It indicates that 50% of the CBZ removalwas reached within 2.3 h of electrolysis under the current density of9 mA cm−2.

3.4.2. Effect of initial CBZ concentrationIt is obvious that the concentrations of CBZ fluctuates periodically

or seasonally at wastewater treatment plants [89,36]. Therefore, it ismore significant to investigate the effect of the initial concentration ofCBZ on the removal efficiency of electrochemical treatment process. Inthis context, different concentrations of CBZ ranging from 2 to30 mg L−1 were electrolyzed under the constant current density of9 mA cm−2 for 8 h, in order to evaluate the influence of initial con-centrations of CBZ on the removal efficiency, as shown in Fig. 5.

The removal rates of about 80.3%, 73.3%, 71.8%, and 70.2% wereachieved after the 8 h of electrolysis when initial concentrations of CBZwere 2, 10, 20, and 30 mg L−1

, respectively (Fig. 5, inset). The removalrate was much faster during the first 2 h of operation in all of the fourdifferent assays, however, decreased from 80.3 to 70.2%, when theinitial CBZ concentrations increased from 2 to 30 mg L−1. About 70%of CBZ removal was achieved within 3.5 h of electrolysis for 2 mg L−1

of initial CBZ concentration, whereas 8 h were needed to achieve thesame CBZ removal for initial CBZ concentration of 30 mg L−1. Theseresults are consistent with other works in which electrochemical de-gradation gave better results at low initial concentration [21,23,31].

At higher initial concentrations, the mass transfer of organic sub-stances increases, which ultimately inhibits the interaction between thetargeted pollutant and anode active sites [21,23]. This can generate

more intermediate products, which in turn competes with CBZ in thereaction with %OH radicals and can reduce the CBZ removal efficiency[21]. So, while increasing the initial concentration of CBZ underidentical current densities, the oxidation process can be limited by theformation of %OH radicals [20]. Further, the electrode surface might bemediated with passivation (diffusion layer) due to the deposition ofmore reaction products. On the other hand, at lower concentrations, thehigher CBZ removal rate was achieved, which indicates that the greateramount of CBZ molecules are attacked by the %OH radicals. This resultcan be explained by the phenomena that at low initial concentrations,electrochemical degradation rates of organic molecules can be fasterthan the diffusion of side-products [21].

3.4.3. Effect of pHpH values can have a significant influence on the electrochemical

treatment processes [19]. To assess the effect of pH on the removalefficiency of CBZ at Ti/Ta2O5-SnO2 anode, different initial pH condi-tions ranging from acidic to basic media (2–10) were investigated.Fig. 6 depicts the degradation kinetics of CBZ (20 mg L−1) when elec-trolyzed for 8 h under the current density of 9 mA cm−2 with sup-porting electrolyte of 0.1 M Na2SO4. In Fig. 6 (inset), the residualconcentrations of CBZ after the 8 h electrolysis are found to be de-creasing from about 6.3, 5.7, and 5.0 mg L−1 at initial pH values of 2, 4,and 6, respectively. However, the residual concentrations are in in-creasing trend in basic media ranging from 5.7 to 6.0 mg L−1 at cor-responding pH values of 8 and 10, respectively. Under the identicalconditions, the optimum removal of 75.2% was achieved at pH value of6. Whereas, the removal efficiencies were 68.8%, 71.2%, 71.5%, and70.1% at the pH values of 2, 4, 8, and 10, respectively. Thus, theseresults indicate that the influence of pH was not significant for elec-trochemical removal of CBZ, which is in agreement with other studies[3,6,23]. The reason might be the high acid-dissociation constant (pKa)of CBZ (13.94), for which no significant dissociation is generally re-ported, either in acidic or basic conditions [6]. In practice, this ten-dency will be useful in real treatment cases since MBR processes arenormally operated at pH of 6–7 [9]. On the other hand, the decreasingtrend of the CBZ removal efficiency with increasing pH values might beattributed to the significant shift of OER onset potential to less positivevalues (low O2-overvoltage), which favors more oxygen formation thanpollutant oxidation [37]. It was also noticed that the solution pH didnot change significantly in acidic assays over the 8 h of time. Never-theless, the solution pH was found to be changing into a more acidiccharacter (5–6) when performed in alkaline media (pH 8–10 initially),which might be due to the formation of weak acid products such ascarboxylic acids [32,38].

Fig. 5. Effect of the initial CBZ concentrations on the removal ten-dency during electrolysis (CBZ (0): 1–30 mg L−1; 0.1 M Na2SO4; pH:5.7; T = 11 ± 1 °C; current density: 9 mA cm−2) on Ti/Ta2O5-SnO2

anodes. Inset: CBZ removal efficiencies at different initial CBZ con-centrations (2, 10, 20, and 30 mg L−1).


334

3.4.4. Effect of temperatureThe reaction rate constants are exponentially dependent on the

temperature of the medium following the Arrhenius law. The similarprinciple can be expected in the electrochemical processes. Therefore,to investigate the effect of temperature on the electro-oxidation,20 mg L−1 of CBZ solution was electrolyzed on Ti/Ta2O5-SnO2 anodeunder the current density of 9 mA cm−2 at three different temperaturesranging from 10 to 30 °C. Fig. 7 shows the effect of different tempera-tures on the CBZ degradation kinetics during 8 h of electrolysis. Theremoval efficiencies of 71.7%, 74.1%, and 75.5% were achieved whencorresponding operating temperatures were maintained at 10 °C, 20 °C,and 30 °C (Fig. 7, inset). Thus, increasing the temperature had a slightpositive influence on the kinetic rates. One possible reason might be abetter mass transfer of pollutants towards the anode surface at highertemperature, due to high molecular agitation which could enhance theoxidation efficiency [39]. Another reason might be the indirect oxida-tion reaction of organics in the bulk solution by electrogenerated oxi-dants peroxodisulphates (S2O8

2−), which is one of the powerful oxi-dizing species, from the oxidation of electrolyte (Na2SO4). Indeed, itwas reported that the rate of oxidation of organic pollutants via per-oxodisulphates increases with positive increments in the solution tem-perature [31]. In terms of hydrodynamics of the bulk solution, viscositydecreases with increasing solution temperature, which enhances masstransfer rate of organic species towards electrode surface thus con-tributes higher oxidation efficiency [20].

Moreover, the degree of mineralization was also analyzed at thisstage under the optimum operating conditions. Using current density of9 mA cm−2, pH of 6.0 and initial CBZ concentration of 20 mg L−1,optimum removal of TOC were 61.2%, 69.8% and 71.1% under thesolution temperatures of 10 °C, 20 °C, and 30 °C, respectively (Fig. 7,inset). Therefore, the average rate of mineralization during this workwas about 61%, as the operating temperature was maintained at about11 ± 1 °C. It can be concluded that increase in the temperature haspositive effect on the degree of mineralization too. Nevertheless, themineralization efficiency was always lower than that of CBZ (71.1%and 75.5%, respectively). The most possible reason might be attributedto the fact that carboxylic acids are highly recalcitrant, thus they de-grade slower than parent compounds [32].

3.5. Comparison between Ti/Ta2O5-SnO2 and Ti/PbO2 electrodes

In order to evaluate the effect of different electrode materials on theCBZ removal efficiencies, Ti/Ta2O5-SnO2 and Ti/PbO2 electrodes werecompared. The stability of both electrodes were also compared byperforming leaching tests. Fig. 8 shows the removal kinetics of20 mg L−1 CBZ in aqueous 0.1 M Na2SO4 solution, pH 6.0, T = 10 °C,and using current density of 9 mA cm−2. The removal trends of both Ti/Ta2O5-SnO2 and Ti/PbO2 electrodes were almost identical until 2 h ofelectrolysis. However, after 8 h of electrolysis, the CBZ removal effi-ciency reached 71.7% and 77.9% on Ti/Ta2O5-SnO2 and Ti/PbO2

Fig. 6. Effect of initial pH on the removal tendency of CBZ duringelectrolysis (CBZ (0): 20 mg L−1 + 0.1 M Na2SO4; current density:9 mA cm−2; T = 11±1 °C) on Ti/Ta2O5-SnO2 anodes. Inset: Residualconcentrations and corresponding CBZ removal% after 8 h of elec-trolysis at different pH.

Fig. 7. Effect of temperature on the removal tendency of CBZ duringelectrolysis (CBZ (0): 20 mg L−1 + 0.1 M Na2SO4; current density:9 mA cm−2; pH: 6) on Ti/Ta2O5-SnO2 anodes. Inset shows% removalof CBZ and TOC at different temperatures.


335

anodes, respectively (Fig. 8). From the results, it can be concluded thatunder the identical operating conditions, the Ti/PbO2 electrode is moreeffective on CBZ removal. This tendency could be explained by the factthat the Ti/PbO2 anodes have greater electrocatalytic activity than Ti/Ta2O5-SnO2 anodes due to their higher relative roughness factor [40].The increase in surface roughness can produce larger amount of phy-sisorbed %OH radicals under the identical current density [21], whichcould react rapidly with all organic species at the vicinity of anodesurface [31].

Furthermore, stability of both the electrodes was examined in termsof the leaching of Ta, Sn, Ti and Pb, by monitoring their ion con-centrations in the bulk solution. Each electrodes were electrolyzed for24 h by reusing the same electrodes over three days (8 h per day withfresh solutions each run), as mentioned in Fig. 9a,b.

In Fig. 9a, the dissolution profile of Ti/PbO2 anode exhibits noleaching of Pb during the first 8 h of electrolysis. However, the dis-solution of PbO2 was significant to bulk solution during 16 h and 24 hof the electrolysis, releasing higher amounts of Pb (concentrations of37.3 and 25.8 μg L−1, respectively). Likewise, traces of Ti ions werealso measured from the solution with maximum of 14.2 μg L−1 duringthe first 8 h of electrolysis. These results show that current density of9 mA cm−2 is capable of dissolving PbO2 from the electrode surface. Asignificant concentration of Pb leached from PbO2 oxide coated elec-trodes has also been reported in another work [20]. It is worth notingthat the maximum allowable concentration of Pb in surface water re-sources, according to EU directive is 14 μg L−1 [41].

On the other hand, leaching profile of Ti/Ta2O5-SnO2 shows theincreasing trend for Sn and Ti, as compared to Ta (Fig. 9b). Theleaching of Sn was the highest with concentrations of 145.7, 318.2, and407.3 μg L−1, respectively after 8, 16 and 24 h of electrolysis. Similarly,Ti concentration increased from 53.6 to 92.9 μg L−1 after 8 and 24 h of

electrolysis, respectively. In contrast, the leaching of Ta was measuredonly in the range of 2–14 μg L−1, indicating a low dissolution in to thebulk solution. In this regard, currently there are no specific regulationson the concentrations of Ti, Sn, or Ta in surface water sources. More-over, the CBZ removal efficiency of Ti/PbO2 and Ti/Ta2O5-SnO2 wascomparable, nonetheless the latter is more environmentally friendlysince the leaching phenomenon (frequently associated with EO) will notgenerate heavy metals into the treated waters.

3.6. The energy consumption

The electrical energy consumed to treat specific volumes pollutedsolutions is the prime concern to assess the viability of any electro-chemical degradation processes. The EC evolved during the electro-chemical degradation of CBZ at varying current densities (Cf. Section3.4.1) are shown in Fig. 10. It indicates that the minimum electricalenergy consumption of 3.2 kWh m−3 was required to reach 53% of CBZremoval while using the current density of 1 mA cm−2. The EC in-creased from 26.3 to 60.3 kWh m−3 when current densities increasedfrom 5 to 9 mA cm−2. Beyond the current density of 9 mA cm−2 until17 mA cm−2, the EC increased from 60.3 kWh m−3 and reached to172.7 kWh m−3, even though the removal efficiencies were not im-proved significantly. For the optimum current density of 9 mA cm−2,the energy required to degrade 71.7% of CBZ was 60.3 kWh m−3. It hasbeen observed that the current densities higher than the limiting one(9 mA cm−2 in the present study) can be associated with high energyconsumption. The reason can be that a large mass of OH radicals mightbe evolved to O2 due to parasitic reactions, then a large cell voltage isoften used mostly for O2 generation, which does not help further inoxidizing the organic pollutant [31].

Fig. 8. The comparison on the removal tendencies of CBZ duringelectro-oxidation (CBZ (0): 20 mg L−1 + 0.1 M Na2SO4; current den-sity: 9 mA cm−2; pH: 6, T = 10 °C) on Ti/Ta2O5-SnO2 and Ti/PbO2

anodes.

Fig. 9. Leaching profile of Pb, Ta, Sn, and Ti from (a)Ti/PbO2 and (b) Ti/Ta2O5-SnO2 anodes when elec-trolyzed for 24 h by reusing the same electrodes butusing fresh working solutions (CBZ(0):20 mg L−1 + 0.1 M Na2SO4) for each assays; undercurrent density of 9 mA cm−2.


336

3.7. Applicability of EO process in real MBR effluent for CBZ removal

In practice, the optimum operating conditions during the removal ofCBZ from synthetic solution were applied for the treatment of CBZ inthe real MBR effluents, collected from a local MBR pilot. The char-acteristics of the MBR effluent, detailed specifications and operatingconditions of the MBR pilot plant are described in previous study [9].The MBR effluent samples were then spiked with a precise amount ofCBZ to have an initial concentration of 10 μg L−1. Moreover, two dif-ferent working solutions, with or without supporting electrolyte, wereprepared and experimented. The details of experimental conditions andresults are presented in Table 2.

From Table 2, it can be noticed that the EO method, which wasoperated under optimal conditions, has shown an excellent CBZ re-moval efficiencies in the real MBR permeate, independent of the in-fluence of supporting electrolyte. About 90% removal of CBZ was at-tained after 2 h of electrolysis, which further reached 99.99% within4 h of electrolysis, for both assays (data not shown). These results areconsistent with the findings of García-Gómez et al. [2], where lowconcentrations of CBZ spiked into the synthetic wastewater was elec-trolyzed by using EO method. Nevertheless, the energy consumptionwas higher when no supporting electrolyte was used as compared to theuse of it (Table 2). High salinity of effluent wastewater could increasethe conductivity and thus can lower the energy consumption [1].Nonetheless, the conductivity of the MBR effluent in this work wasreported to be 601 μS cm−1, which was not enough to maintain aconstant electrical pressure during the redox reactions. On the otheraspect, the cost of using supporting electrolyte (approx. US$ 70 permetric ton) is much cheaper than the consumption of electricity (ap-prox. US$ 0.05 per kWh).

4. Conclusion

The electrochemical oxidation of CBZ from synthetic solutions and

real MBR effluents using a newly developed Ti/Ta2O5-SnO2 electrodeswas investigated under galvanostatic conditions. The characterizationof the Ti/Ta2O5-SnO2 electrodes were performed by using SEM, EDX,AFM, and cyclic voltammetry analyses. The main operating parametersinfluencing the CBZ removal efficiency in synthetic solutions on thetested anodes were evaluated as a function of applied current density,initial CBZ concentration, pH, and temperature. The results showed thatthe maximum removal of CBZ (20 mg L−1) and TOC reached 75.5%,and 71.1%, respectively, after 8 h of electrolysis, current density of9 mA cm−2, solution pH 6, temperature of 30 °C, and using 0.1 MNa2SO4 as supporting electrolyte. The performance of Ti/Ta2O5-SnO2

electrode was compared with Ti/PbO2 electrode in terms of CBZ re-moval efficiencies and stability. The results showed that, under thesame conditions, the CBZ removal efficiency of Ti/PbO2 electrode wasslightly higher when compared with the investigated electrode, but thelatter was more environmentally friendly. Besides, achieving an optimalCBZ removal efficiency of 71.7% required the lowest energy con-sumption of 60.3 kWh m−3.

Furthermore, the complete degradative removal of CBZ wasachieved in spiked real MBR effluents, under the optimal operatingconditions. Overall, the EO based on the use of novel Ti/Ta2O5-SnO2

electrode was found to be a reliable process to remove CBZ from con-taminated waters, with promising potential for integration with MBRtechnology to remediate CBZ.

Acknowledgements

The authors are very grateful for the financial support provided byTEKES-Finnish Funding Agency for Innovation 1043/31/2015(Helsinki, Finland), a research fund for the Smart Effluents Project. Mr.Toni Väkiparta, for SEM/EDX analysis of electrodes, and Mr. Risto Repo(Mikkeli Water works), for helping in operating the MBR unit andproviding the effluent samples.

Fig. 10. The electrical energy consumption trend against the CBZremoval efficiency during the electrochemical oxidation (CBZ (0):

20 mg L−1; 0.1 M Na2SO4; pH: 5.7; T = 11 ± 1 °C) on Ti/Ta2O5-SnO2 anode under different current densities.

Table 2Electrochemical oxidation experiments on real MBR effluent for the validation of CBZ removal (Current density: 9 mA cm−2; pH: 6; T: 11 ± 1 °C; at Ti/Ta2O5-SnO2 anodes).

Parameter Initial concentration in MBR effluent Final concentration after 8 h of electrolysis Electrolyte Removal (%) EC (kWh m−3)

CBZ (μg L−1) 10.75 ± 0.35 < 0.07 (LOD)a No electrolyte 99.99 109.4< 0.07 (LOD)a 0.1 M Na2SO4 99.99 57.2

a (LOD) Limit of detection.


337

Appendix A. Supplementary data

Supplementary data associated with this article can be found, in theonline version, at http://dx.doi.org/10.1016/j.apcatb.2017.09.017.

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338

Publication IV

Khum Gurung, Mohamed Chaker Ncibi, Senthil K. Thangaraj, Janne Jänis, Mahdi

Seyedsalehi, Mika Sillanpää

Removal of pharmaceutically active compounds (PhACs) from real

membrane bioreactor (MBR) effluents by photocatalytic degradation

using composite Ag2O/P-25 photocatalyst


Separation and Purification Technology

Vol. 215, pp. 317-328, 2019

© 2019 Elsevier


Separation and Purification Technology

journal homepage: www.elsevier.com/locate/seppur

Removal of pharmaceutically active compounds (PhACs) from realmembrane bioreactor (MBR) effluents by photocatalytic degradation usingcomposite Ag2O/P-25 photocatalyst

Khum Gurunga,⁎, Mohamed Chaker Ncibia, Senthil K. Thangarajb, Janne Jänisb,Mahdi Seyedsalehic, Mika Sillanpääa

a Department of Green Chemistry, School of Engineering Science, Lappeenranta University of Technology, Sammonkatu 12, FI-50130 Mikkeli, FinlandbDepartment of Chemistry, University of Eastern Finland, Yliopistonkatu 7, FI-80101 Joensuu, Finlandc School of Environment, Tsinghua University, Beijing 100084, China

A R T I C L E I N F O

Keywords:PhACsPhotocatalytic degradationAg2O/P-25 photocatalystsMatrix effectTransformation products

A B S T R A C T

Pharmaceutically active compounds (PhACs) are emerging pollutants causing serious challenges to wastewatertreatment plants due to poor biodegradability. In this study, the enhanced removal of highly recalcitrant andcommonly monitored PhACs, carbamazepine (CBZ) and diclofenac (DCF) by heterogeneous photocatalysis wasinvestigated using 5% Ag2O/P-25 photocatalyst. The photocatalyst was characterized by scanning electronmicroscope (SEM-EDX), Brunauer-Emmett-Teller (BET), Fourier transfer infrared spectroscopy (FTIR) andUV–vis diffuse reflectance spectra (UV-DRS). The effects of catalyst dose, initial pollutants concentration, andmineralization during the photocatalytic degradation of PhACs were investigated. The matrix effect was assessedin deionized water (DW) and real membrane bioreactor effluent (RME). Optimal CBZ and DCF removals of89.10% and 93.5%, respectively for 180min of UV irradiation were achieved at catalyst dosage of 0.4 g L−1 inDW matrix. However, the optimal catalyst dosages for CBZ and DCF in RME matrix were increased by factor 2and 1.5, respectively, to achieve the same degree of removal. Declining trends of removal rate were observedwhen initial concentrations of both the PhACs were increased under optimal catalyst dosages, and kinetics seemto fit the Langmuir-Hinshelwood model. Photo-induced holes and %OH were the dominant oxidation speciesinvolved in the photocatalytic degradation of the PhACs. A plausible reusability of 5% Ag2O/P-25 photocatalystwas observed for both the PhACs. Moreover, various aromatic/aliphatic intermediates generated during thephotodegradation CBZ were identified using fourier-transform ion cyclotron resonance (FT-ICR) mass spectro-metry, and a possible multi-step degradation pathway was proposed. Overall, the removal of PhACs using 5%Ag2O/P-25 photocatalyst showed promising results in real wastewater.

1. Introduction

Pharmaceutically active compounds are used for the prevention,diagnosis, and treatment of several diseases in humans and animals.The global annual average per capita consumption of pharmaceuticalsis estimated at about 15 g, whereas industrialized countries accountedbetween 50 and 150 g [1]. The annual per capita expenditures onpharmaceutical ranged from US$ 7.61 in low-income countries to US$431.6 in high-income countries [2]. As most of the PhACs are non-vo-latile and often charged molecules composed of diverse moieties, theseemerging pollutants are highly recalcitrant and not entirely degradedafter the medical use, hence the excretion of unchanged (parent)compounds and metabolites and their channeling to the aquatic

environment [3]. In recent years, the fate of the PhACs and their me-tabolites has received a great deal of interest from the scientific com-munity. Though PhACs are normally detected in the concentrationrange of ng L−1 or µg L−1, they can pose serious issues due to theirpotential adverse effects on human health and ecosystem [4,5]. How-ever, other factors such as increased persistence, time of exposure,biotransformation, and degradation mechanisms can also influence theimpacts of PhACs, other than their concentration alone [5]. Among thelarge number of PhACs, CBZ and DCF are the most frequently identifiedin aquatic environments [1].

CBZ is an antiepileptic drug often used to treat epilepsy and sei-zures, whereas DCF is a non-steroidal anti-inflammatory drug (NSAID)commonly used as analgesic, anti-rheumatic, and antiarthritic [6,7].

https://doi.org/10.1016/j.seppur.2018.12.069Received 22 October 2018; Received in revised form 24 December 2018; Accepted 25 December 2018

⁎ Corresponding author.E-mail addresses: [email protected], [email protected] (K. Gurung), [email protected] (M. Sillanpää).

Separation and Purification Technology 215 (2019) 317–328

Available online 03 January 20191383-5866/ © 2018 Elsevier B.V. All rights reserved.

T

CBZ has been proposed as an anthropogenic indicator in water bodies[8]. Moreover, DCF is included in the watch list of priority substancesas emerging pollutants in EU directive that needs careful environmentalmonitoring in its member states [9].

The wastewater treatment plants (WWTPs) are the primary receptorof PhACs via direct and unsafe discharge of untreated wastewaters fromdomestic households, industries, agriculture lands, and hospitals[10,11]. PhACs have been found in the effluents of municipal WWTPsbecause of their incomplete removal [5]. It is well known that con-ventional WWTPs are not fully efficient in treating most of the PhACs[1,12,13]. Even Membrane bioreactors (MBRs), which are regarded ashaving superior capabilities in terms of performance and flexibility[14], are not effective in the removal of certain PhACs [8,15]. In thisregard, many researchers have recommended the development oftreatment systems where biological degradation is integrated with ad-vanced oxidation processes (AOPs) as a pre-treatment of effluents, inorder to improve biodegradability and overall removal efficiency ofPhACs, before their discharge into water bodies [4,16]. Despite theircost, the use of AOPs is still justified as complete degradation of per-sistent pollutants is not always needed [4]. The prospective applicationof various methods such as Fenton oxidation [17,18], ozonation [19],chemical oxidation [20], and TiO2 photocatalysis [7] processes havebeen proposed to transform PhACs into non-toxic and biologically in-active compounds.

TiO2 is the most commonly used photocatalyst (EG=3.2 eV) be-cause of its cost, and remarkably active and stable features [7,21,22].Heterogeneous UV/TiO2 photocatalysis has become the most attractiveAOP to remove varieties of organic pollutants from water and waste-water [16]. In heterogeneous UV/TiO2 photocatalysis, upon irradiationof TiO2 semiconductor by UV–vis light, an electron (e−) is promoted tothe conduction band and a hole (h+) is generated on the valence bandsite vacated by the electron, thereby forming highly reactive oxygenspecies (ROS) such as hydroxyl (%OH) and superoxide (%O2

–) radicals[21,23–25]. Therefore, heterogeneous photocatalysis is based on thechemistry of photo-generated h+ and/or the %OH, which promotes non-selective oxidation of pollutants [3,26]. Yet, there are some drawbacksassociated with conventional TiO2 such as rapid recombination ofcharge carriers (ecb−/hvb+), difficulty in separating the reacted TiO2

particles, inefficient quantum yield, and wide energy band gap, hin-dering its industrial applications [4,7,27]. Therefore, several R&D stu-dies focused on developing highly efficient UV–vis light driven photo-catalysts by doping TiO2 with narrow band gap semiconductors, whichis crucial having less energy use and highly efficient photocatalyticoxidation of pollutants [27,28].

Many studies have suggested that fine particles of transition metalsor their oxides dispersed on the surface of photocatalyst matrix couldact as electron traps, which effectively reduces the combination ofcharge carriers and improves photocatalytic activity [22,29,30]. Ag2Onanoparticles doped on TiO2 particles can enhance the optical absorp-tion spectrum and lattice structural stability of TiO2, improving pho-tocatalytic degradation of hazardous pollutants under UV–vis light ir-radiation [21,23,31]. Indeed, Ag2O is a proven photocatalyst for thedegradation of organic pollutants under visible light irradiation havinga narrow band gap of 1.2 eV [32]. The physicochemical changes in theTiO2 particles after doping, such as increasing surface area, decrease ofTiO2 crystals, and homogeneity in particles surface size, are responsiblefor the enhancement of photocatalytic efficacy [4,33]. Additionally,recovery and reusability of TiO2 can be enhanced due to improvedsegregative specialty with impurity doping [30]. The photocatalyticefficiency of Ag2O doped TiO2 has been well confirmed by nitrate re-duction [34] and degradation of organic dyes [28,30,35]. Nevertheless,to the best of our knowledge, the application of Ag2O doped TiO2 (P-25)nano-particles in photocatalytic degradation of CBZ and DCF from realMBR effluent has not been studied so far.

The main objective of this study was to investigate the removal ofcommonly used PhACs including CBZ and DCF from two aqueous

matrices, i.e. DW and RME, by heterogeneous photocatalytic degrada-tion using Ag2O/P-25 photocatalyst. The composite heterojunction 5%Ag2O/P-25 photocatalyst was synthesized by a simple pH-mediatedchemical precipitation reaction method and its photocatalytic effi-ciency for the removal of PhACs from both matrices were examined.The emphasis was given on the influence of catalyst dose and miner-alization in both DW and RME, the effect of initial pollutants con-centration, stability of catalyst, and determination of reactive oxygenspecies and reaction mechanism. Moreover, various intermediate pro-ducts formed during the photodegradation of CBZ were identified, anda possible multi-step degradation pathway was proposed.


2.1. Chemicals

All the chemicals used were analytical grade and purchased fromSigma-Aldrich Chemical Co. (St. Louis, USA), unless specified. Thechemicals used for this experiment include CBZ (≥99%, meets USPtesting specifications), DCF (2-[(2,6-dichlorophenyl) amino]benzenea-cetic acid sodium salt), Aeroxide ® P-25 as commercial TiO2 (21 nmprimary size (TEM), ≥99.5% trace metal basis), AgNO3 (99.999% tracemetal basis), NaOH (≥98%, anhydrous), Acetic acid (natural,≥99.5%), N,N Dimethylformamide (anhydrous, 99.8%), Lithiumacetate dehydrate (99.999% trace metal basis), Titanium(IV) butoxide(reagent grade, 97%), Palladium(II) chloride (≥99.9%), Hydrochloricacid (reagent grade, 37%), Tin (II) chloride dehydrate (≥99.99% tracemetal basis), Sodium citrate tribasic dehydrate (ACS reagent, ≥99.0%),Nickel nitrate (99.999% trace metal basis), Methanol (≥99.8%, anhy-drous), and Ethylenediaminetetraacetic acid sodium salt dehydrate (EPreference standard).

2.2. Synthesis of photocatalysts

5% (w/w) Ag2O/P-25 photocatalysts were synthesized via a simplepH-mediated chemical precipitation reaction method [34]. The proce-dure typically includes the dispersion of 1.0 g of P-25 in 117mL ofdeionized water and the solution was stirred for 4 h. Subsequently,540 µL of 0.2M AgNO3 solution was added dropwise into the suspen-sion. The mixture was then stirred in dark for 30min in order to reachAg+ sorption equilibrium. Then, a measured amount of 0.1 M NaOHsolution was slowly added into the Ag+ pre-equilibrated mixture andstirred for 1 h at room temperature. The mixture was filtered through0.45 µm nylon membranes (VWR, USA) and washed repeatedly withdeionized water and ethanol. The obtained pallets were then kept indeep-freezer for 24 h and lyophilized. After grinding and sievingthrough 500 µ sieve, the dried powder was used for subsequent ex-periments.

Furthermore, three other photocatalysts were also synthesized byadopting the methods from literatures including Sn3O4/P-25 [33],PdO/P-25 [34], and NiO/P-25 [35] by doping different metallic sub-strates over pure P-25.

2.3. Model solutions and matrix

Model solutions containing target PhACs (CBZ and DCF) in twodifferent matrices were adopted for this study. The properties of targetpollutants in this study are shown in Table 1. The aqueous solutionmatrices were prepared by spiking CBZ or DCF in deionized (ultrapureMilli-Q) water and in real effluent of the submerged MBR pilot plantlocated in Kenkäveronniemi WWTP (Mikkeli, Finland). The detaileddescriptions of MBR pilot plant and operating conditions are describedin our previous studies [10,15]. The main physicochemical character-istics of the MBR effluent are listed in Table S1 (Supplementary in-formation)

K. Gurung et al. Separation and Purification Technology 215 (2019) 317–328

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2.4. Characterization of photocatalysts

Microstructure and morphology of the synthesized photocatalystswere examined by scanning electron microscope (SEM) (Hitachi,SU3500, Japan) using acceleration voltage of 5 kV. The elementalanalysis of photocatlyst was performed using energy-dispersive X-rayspectroscope (Thermo Fisher,USA). The Brunauer-Emmett-Teller (BET)surface area (SBET) was determined by using Tristar II Plus. The UV–visdiffuse reflectance spectra were measured in the range of 250–800 nmusing Agilent Cary 5000 spectrophotometer with DRA2500 integratingsphere. The Fourier transfer infrared spectroscopy (FTIR) was per-formed using Bruker Platinum ATR VERTEX 70.

2.5. Photodegradation experiments

The photocatalytic degradation experiments were performed in aself-made continuous stirred photochemical reactor as shown in Fig. S1(Supplementary information). The photocatalytic reaction chamberconsists of a UV lamp (rated power is 14W, rated voltage is 216–253 Vat 50 Hz, λmax of 254 nm, Trojan, Canada) placed in a quartz sockettube with one end closed, magnetic stirrer, rotor, and a reaction vesselwith double-jacketed cooling system. The model solution (300mL) wasadded inside the photochemical reactor and the temperature inside thejacketed reactor was maintained at 21 ± 2 °C by continuous circula-tion of water. Prior to irradiation, measured quantity of photocatalystwas dispersed into the solution and magnetically stirred in dark for30min to achieve adsorption–desorption equilibrium. Upon the irra-diation, an aliquot of 3mL reaction solution was sampled at given timeintervals. The reaction solution aliquots were then filtered through0.45 µm membranes prior to analysis. No buffer was used and the pHvalues ranged from 5 to 7 (before and after) during all set of the ex-periments.

2.6. Trap experiments

The radicals and holes trapping experiments were designed to elu-cidate the types and roles of reactive species, and mechanisms involvedduring the photocatalytic degradation of CBZ and DCF using 5% Ag2O/P-25. In practice, several chemical compounds were tested during thephotocatalytic processes, including ethylenediaminetetaacetic acid(EDTA) as a hole (h+) scavenger [28], methanol (MeOH) as hydroxylradical (%OH) scavenger [36], and O2-purging (%O2

– scavenger) via N2

[37].

2.7. Reusability experiments

Cyclic experiments were performed to check the reusability of theinvestigated photocatalyst over the degradation of CBZ and DCF.Typically, 0.24 g of 5% Ag2O/P-25 was dispersed into 0.3 L CBZ

aqueous solution (20mg L−1) and reaction was started as in photo-degradation experiment for 180min. Samples were collected at theinterval of 60min, 120min, and 180min. After the first run, suspensionwas centrifuged at 3500 rpm for 30min, washed thoroughly withdeionized water, and again dispersed into the another fresh CBZ solu-tion to start the second run. Similarly, the experiment was repeated fivetimes.

2.8. Analytical procedures

The changes in the concentration of CBZ and DCF during photo-catalytic reaction were analyzed using a UV–Vis spectrophotometer(Lambda 45, Perkin Elmer, USA). The absorbance was measured at285 nm and 276 nm for CBZ and DCF, respectively, since these wave-lengths correspond to the characteristic peaks of these drugs in the UVspectrum. Moreover, it should be noted that the absorbance measure-ments of the treated solutions might not indicate the absolute con-centration of parent CBZ or DCF as some of the treated intermediateproducts may also absorb at similar wavelengths, thus contributing toincreased absorbance. Therefore, the reported concentrations used tocalculate percentage removal may account both residual parent pollu-tants and some of the related reaction by-products, which could un-derestimate to actual degradation efficiency of the parent compounds.Dissolved organic carbon (DOC) was also measured using TOC analyzer(Shimazdu, Japan) to assess the extent of mineralization.

The degradation products of CBZ were analyzed by 12-T BrukersolariXTM XR hybrid Qh-Fourier Transform Ion cyclotron resonance(FT-ICR) mass spectrometer (Bruker Daltonics, Bremen, Germany)coupled with an electrospray ionization (ESI) source (Apollo-II), oper-ated in the positive-ion mode. The photocatalytic degradation productsof CBZ (10 µL) were prepared in methanol (diluted 10:200 v/v). Aceticacid (1 v-%) was added to the samples to aid protonation. The sampleswere directly infused into the ion source at a flow rate of 2 µLmin−1.The ion accumulation time in the hexapole was 0.8 s and the time offlight from the collision cell to the ICR cell was set at 0.260ms. Drynitrogen was used as drying (4.0 Lmin−1) and nebulizing gas (80 °C,1.0 bar). For each spectrum, 200 time-domain transients (4 MWordeach) were co-added and full-sine apodized. The mass spectra wereexternally calibrated using an ES tuning mix (Agilent Technology).Bruker FTMS Control™ 2.1 software was used for the instrument controland data acquisition. The mass spectra were further processed andanalyzed by using Bruker DataAnalysis 4.4 software.


3.1. Screening of photocatalysts: dark adsorption, UV- photolysis, and UVphotocatalysis

In a preliminary stage, photocatalytic degradation efficiencies of

Table 1The physico-chemical properties of target PhACs in this study.

Compound/formula Therapeutic Use Molecular weight(gmol−1)

Water solubility (mg L−1,20 °C)

Chemical structure

Carbamazepine (C15H12N2O) Antiepileptic, anticonvulsant drug 236.27 125 ± 2a

Diclofenac sodium (C14H10Cl2NNaO2) Non-steroidal anti-inflammatory and analgesicsdrug

318.13 50,000

a Ref. [52].


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four different photocatalysts including Ag2O/P-25, Sn3O4/P-25, PdO/P-25, and NiO/P-25 were tested on the selected pollutants. First, in orderto evaluate the extent of adsorption and UV-photolysis of target pol-lutants, adsorption under dark conditions and UV-photolysis tests werecarried out (Fig. 1). Even though experiments were run for severalhours, adsorption-desorption equilibrium was reached in 30min for allthe cases. At the catalyst dose of 0.4 g L−1, the percentage adsorptionremoval of 50mg L−1 of CBZ and DCF, was around 2–9%. Similarly, theconversion rates of CBZ and DCF during 180min of UV-photolysis werenearly 13% and 25%, respectively. The results concluded that both ofthe target pollutants were slightly converted under UV-photolysis.Comparable extents of CBZ or DCF removal by adsorption on TiO2

based photocatalysts and UV photolysis have been reported in manystudies [38–40]. On the other hand, the photocatalytic degradationkinetics of as prepared catalysts were tested by irradiating the modelsolutions under UV lamp for 180min. Of the selected five photocatalyst,the highest photodegradation efficiency of CBZ (i.e. 55%) was observedwith Ag2O/P-25 (Fig. 1). Similarly, the degradation of DCF by PdO/P-25 and Ag2O/P-25 was more than 80%. Based on the screening ex-periments, since 5% Ag2O/P-25 photocatalyst accounted the highestphotocatalytic activity either in case of CBZ or DCF, therefore selectedfor the subsequent experiments involving both target micropollutants.Moreover, the selection of Ag2O/P-25 over PdO/P-25 and NiO/P-25was made because of the comparable efficiency and compatibility toboth the model pollutants.

3.2. Characteristics of selected 5% Ag2O/P-25 photocatalyst

The major characteristics of as-prepared 5% Ag2O/P-25

photocatalyst such as surface chemical compositions and morphologywere characterized by X-ray diffraction (XRD), transmission electronmicroscopy (TEM), and X-ray photoelectron spectroscopy (XPS) ana-lysis in the previous work [34,35]. However, we further characterizedthe as-prepared photocatalyst by other analytical techniques such asSEM, EDX, BET, FTIR, and UV-reflectance spectra.

The morphology and elemental microanalysis of the selected 5%Ag2O/P-25 photocatalyst were investigated by SEM and EDX. The SEMmicrographs with different magnifications are as shown in Fig. 2. Thesurface of P-25 nanoparticles acts as underlying substrate to absorbAg2O nanoparticles and an average diameter of 4.4–20 nm could evenlydisperse on the surface of P-25 during pH-mediated precipitation pro-cess [30,34]. Presence of Ag nanoparticles on the surface of P-25photocatalyst was confirmed by the EDX elemental analysis, whichreveals the bulk compositions of Ti, O, C, and Ag elements (Fig. 2i). TheTiO2 nanoparticles provide numerous nucleation sites, which could beresponsible for the homogeneous dispersion of Ag2O nanoparticles onthe surface of TiO2 nanoparticles [30]. Furthermore, elemental map-pings of 5% Ag2O/P-25 confirms the homogenous dispersion of Ag+

particles on P-25 as depicted by Fig. 2(d–h).To further confirm the chemical composition of 5% Ag2O/P-25

photocatalyst, FTIR spectroscopy was performed as depicted in Fig. S2-a. The main absorption peaks were observed at 3000, 1533.67, 1420,1050 and 433.33 cm−1. The peak located at 433.33 cm−1 could beassigned to MeO (i.e. either TieO or AgeO) stretching vibration[41,42]. The broad absorption band at about 2976.75 cm−1 is char-acteristic of OeH stretching vibration of the TieOH. The vibrationalpeaks at 1558.30, 1419.43 and 1052 cm−1 are due to stretching of C]O group, OeH bending and stretching of CeO bond, respectively[42,43].

The BET surface area, pore size, and pore volume of 5% Ag2O/P-25were estimated as reported in Table S2 (Supplementary information).The SBET of pure TiO2 (P-25) increased slightly from 47.27m2 g−1 to54.1 m2 g−1 due to the deposition of Ag2O nanoparticles over P-25surface. Likewise, the average pore size and pore volume of P-25 wereincreased from 8.25 to 9.27 nm and 0.10 to 0.13 cm3 g−1, respectively.However, this observation indicated that the contribution of the Ag2Oparticles to the specific surface area of P-25 was relatively low. Theincreased surface area, pore size and pore volume showed that Ag2Oparticles were well dispersed over P-25 surface with limited inter-ference with the existing pores. Similar trends in the characteristicmodifications of P-25 after doping with Ag2O have been reported inother studies [44].

For the energy band gap (Eg) determination of semiconductors, thereflectance spectra of photocatalysts were converted to the absorbancespectra using the Kubelka-Munk functions [45] and then Eg evaluatedby using Tau plot [44–46]. The undoped P-25 showed a band gap of3.2 eV with the absorption peak at∼ 395 nm. After doping with Ag2O,the wavelength threshold of the absorption peak of P-25 was between400 and 500 nm, with the corresponding energy band gap at 2.96 eV asshown in Fig. S2-b (supplementary information). The decreased bandgap energy inferred the chemical interactions between P-25 and Ag2Onanoparticles. Therefore, lower energy transitions are possible in Ag2O/P-25 semiconductors as Ag2O could create localized energy levels in theP-25 bandgap, which evidently increases the absorption spectra inUV–vis range [44].

3.3. Effect of catalyst dose, solution matrix, and initial substrateconcentration on the removal of selected PhACs

In order to ensure the highest absorption of photons, without gen-erating an substantial amount of waste, determining the optimal dose ofcatalyst in an adopted photocatalytic system is of high importance[38,47]. In fact, several researchers have investigated the reaction ki-netics as a function of catalyst dosing in photocatalytic processes [48].In this study, the optimal dose of 5% Ag2O/P-25 for the removal of CBZ

Fig. 1. Dark adsorption, UV-photolysis, and photocatalytic degradation trendsof different photocatalyts for (a) CBZ, and (b) DCF. (Operating conditions:[CBZ]0= [DCF]0= 50mg L−1, [(5% w/w) Ag2O/P-25]= 0.4 g L−1).


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and DCF was determined for two different solution matrices (i.e. DWand RME) as shown in Fig. 3. The catalyst dose of 0.2–1.0 g L−1 wasassessed for DW matrix, whereas 0.4–1.2 g L−1 was used in the case ofRME matrix. Since organic matter in RME could inhibit the degradationof organic pollutants, higher amount of catalyst dose is essential toachieve enhanced removal efficiencies [4]. Therefore, a higher range ofcatalyst dosing was selected to assess the influence of organic mattercontained in the RME matrix. Table 2 shows the collective removal oftarget pollutants and the extent of mineralization in two solution ma-trices under different operating conditions.

The photocatalytic degradation kinetics of CBZ in DW and RMEmatrices are shown in Fig. 3a and b, respectively. In DW matrix, andafter an irradiation time of 180min, the removal of CBZ reached from80.4% to 89.10% while increasing the catalyst dosing from 0.2 to0.4 g L−1. Likewise, for the case of DCF, the removal rate increasedfrom 87.5% to 93.5% in DWmatrix for the similar range of catalyst. Thehigher removal rate of DCF compared to CBZ in identical operatingconditions exhibited greater affinity of DCF molecules towards photo-catalytic degradation by 5% Ag2O/P-25. From Fig. 3 and Table 2, it isclear that the photocatalytic degradation of both CBZ and DCF beyondthe 0.4 g L−1 of catalyst dose remained constant. Therefore, 0.4 g L−1 of5% Ag2O/P-25 was concluded to be the optimum catalyst dose for theremoval of both PhACs in DW matrix. Similar tendencies have beenreported in earlier studies [4,47]. The limiting value 0.4 g L−1 might beattributed to fact that photocatalytic rate increases with catalyst dosingat the beginning and then decreases (or saturates) at higher doses due

to unfavorable phenomenon, such as light scattering due to enhancedreflectance and increased opacity of the suspension [48,49]. Also, theagglomeration of solid particles due to excess dose may result in a re-duction of active surface sites available for photon absorption, whicheventually reduces photocatalytic efficiency [33]. Even though thenumber of active sites increases with more catalyst dosing, the pene-tration of light is to be compromised because of the shielding effect[33,48]. The trade-off between these two divergent phenomena resultsin an optimal catalyst dosing for photocatalytic processes [33]. Theshielding by TiO2 particles may occur due to deactivation phenomenonof activated TiO2 molecules by collision with ground state molecules[48,50]. Nevertheless, the optimal catalyst dosing is highly dependenton the concentration and source/type of the pollutant and the operatingconditions as well [33]. Therefore, the optimal catalyst dose of0.4 g L−1 achieved for both CBZ and DCF in DW matrix might be dif-ferent for other conditions.

On the other hand, relatively slow kinetics was observed in theexperiments performed with RME matrices of both CBZ and DCF(Fig. 3b and d). In Table 2, an increasing trend of CBZ removal rate wasclearly observed from 76.6 to 93.9% when the catalyst dose was in-creased from 0.4 to 0.8 g L−1. However, the removal rate decreased bynearly 3% when catalyst dose was subsequently increased beyond0.8 g L−1. On the other hand, DCF removal rate was improved greatlywhen the catalyst dose was increased from 0.4 g L−1 to just 0.6 g L−1,though no further remarkable enhancement was observed beyond0.6 g L−1 of catalyst dose. Therefore, different optimal catalyst dose for

Fig. 2. SEM micrographs of different magnifications (a–c); Elemental mappings (d–h): (d) total mapping, (e–h) C, O, Ti, and Ag mappings; and (i) EDX microanalysisof (5% w/w) Ag2O/P-25 selected and as-prepared photocatalyst.


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CBZ (i.e. 0.8 g L−1) and DCF (i.e. 0.6 g L−1) were established in RMEmatrix.

As clearly seen in Table 2, there is an obvious deviation in the op-timum catalyst dose of DW and RME matrices. The need for optimalcatalyst dose in RME matrix, in order to achieve approximately thesame level of removal efficiency in DW matrix, is increased by factor of2 and 1.5 for CBZ and DCF, respectively. Therefore, the competition

between organic matter and the PhACs on the surface of catalyst wasevidently observed in the RME matrix during the photocatalytic pro-cess. This may be attributed to either the organic carbon content ofRME samples, or the presence of various species that may act as naturalscavengers and other reactive moieties [12]. The organic matter in realeffluent not only competes for the active species generated in the pro-cess but could also be adsorbed on the surface of photocatalysts, whichinhibits the photocatalytic activity [4]. Additionally, a possible inter-ference of inorganic ions on catalyst surface deactivation, and a com-petition for UV absorption should also be taken into consideration [51].

Subsequently, the optimized catalyst doses for CBZ and DCF re-moval in RME matrix were used to study the influence of initial sub-strate concentration in photocatalytic degradation kinetics, as depictedin Fig. 4. The initial substrate concentration was varied in the range of20–50mg L−1 at optimal catalyst doses of 0.8 g L−1 and 0.6 g L−1 forCBZ and DCF, respectively. In Fig. 4a and b, decreasing trends in theremoval of CBZ and DCF were observed when initial substrate con-centrations were gradually increased. The removal efficiency of CBZwas decreased by 38%, when the concentration was increased from 20to 50mg L−1 (Fig. 4a, inset). Whereas, only 23% reduction was ob-served in the DCF removal efficiency with the same range of con-centration increments (Fig. 4b, inset). These trends may be attributed tothe increasing occupied sites of the catalyst due to more pollutantmolecules adsorbed on the surface of catalyst, which inhibits the gen-eration of reactive species [12]. On the other hand, high pollutantsconcentrations can lead to a decreasing penetration of photons into thebulk solution and hence the solutions become more impermeable to UVirradiation due to an induced inner filtering effect. Thus, the photo-excitation of catalyst particles by photon energy will be reduced leadingto a diminished photodegradation efficacy [52].

The photocatalytic oxidation of organic contaminants via hetero-geneous catalytic process usually showed that the corresponding data

Fig. 3. Effect of catalyst dosing and water matrix on the photocatalytic degradation of (a) CBZ in DW, (b) CBZ in RME; and (c) DCF in DW, and (d) DCF in RME.(Operating conditions: [CBZ]0= [DCF]0= 20mg L−1; pH=5–7; [5% Ag2O/P-25]=0.2–1.0 g L−1 for DW matrix and 0.4–1.2 g L−1 for RME matrix;Temperature=21 ± 2 °C).

Table 2Removal rates (%±SD) of CBZ and DCF under the varying catalyst dosages andthe extent of mineralization (%) in two different water matrices.

Solutionmatrices

Targetpollutants

Catalystconcentration(g L−1)

Removalefficiency (%)

Mineralization (%)

DW CBZ 0.2 80.40 ± 0.5 67.900.4 89.10 ± 1.50.6 88.60 ± 1.40.8 91.70 ± 1.51.0 89.02 ± 4.7

DCF 0.2 87.60 ± 2.2 64.800.4 93.50 ± 0.10.6 93.60 ± 0.10.8 93.04 ± 0.11.0 93.40 ± 0.3

RME CBZ 0.4 76.60 ± 5.2 60.300.6 85.40 ± 6.50.8 89.74 ± 0.41.0 90.50 ± 0.41.2 90.95 ± 2.3

DCF 0.4 86.60 ± 0.3 55.200.6 90.70 ± 4.50.8 88.30 ± 3.41.0 90.40 ± 0.401.2 92.10 ± 1.10


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fit the Languir-Hinselwood (L-H) kinetic model as shown in Eq. (1)[12,53]:

= +

Cr

Ck

1Kk

eq

0

eq

r r (1)

where r0 is the initial reaction rate, Ceq is the equilibrium concentrationof substrate in solution after the completion of dark experiments, kr isthe reaction rate constant at maximum coverage, and K is equilibriumconstant for adsorption of the substrate onto the catalyst surface.

In order to check if the investigated photocatalytic degradationfollows the L-H kinetic model, the experimental data with initial con-centrations of target pollutants (20–50mg L−1) in RME matrix (Fig. 5)were assessed and the initial rates were calculated at 20min of initialreaction. A graphical plot of Eq. (1) for runs at various initial substrateconcentrations is shown in Fig. 5, which indicates that the photo-catalytic degradation well fitted the Langmuir-Hinshelwood kineticmodel. Moreover, the degradation kinetics beyond 20min of irradiationwere observed to follow a curve rather than a straight line whenmodelled with pseudo-first order (PFO) kinetic for both CBZ and DCFcompounds. The possible formation of aromatic organic intermediates,which can be detected at the similar wavelengths of parent compounds,could affect the corresponding final absorbance values and conse-quently the reaction kinetics [54].

3.4. Mineralization of PhACs

In order to characterize the mineralization extent of CBZ and DCF indifferent solution matrices, the DOC measurements were carried out

during photocatalytic treatment. The optimal catalyst doses wereadopted for different solution matrices and target pollutants. Fig. 6aand b shows the normalized DOC removals during the photocatalytictreatment of CBZ and DCF, respectively. About 68% of mineralizationrate was observed when the CBZ removal was 90% in DW matrix,

Fig 4. Effect of initial substrate concentration in photocatalytic degradation of(a) CBZ, and (b) DCF. (Operating conditions: All are RME matrix; [5% Ag2O/P-25]= 0.8 g L−1 for CBZ; and 0.6 g L−1 for DCF). Removal percentages of CBZand DCF versus initial concentrations are shown in insets of (a), and (b), re-spectively.

Fig. 5. Fitting of L-H model for photocatalytic degradation kinetics of CBZ andDCF. (Operating conditions: RME solution matrix, [5% Ag2O/P-25)]0= 0.8 g L−1 for CBZ, and 0.6 g L−1 for DCF). Plot of Eq. (1).

Fig. 6. Mineralization during photocatalytic degradation of (a) CBZ, and (b)DCF in two solution matrices. (Operating conditions: [5% Ag2O/P-25]= 0.4 g L−1 in DW matrix for both CBZ and DCF; and 0.8 g L−1 and0.6 g L−1 in RME matrix for CBZ and DCF, respectively).


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whereas the mineralization rate was reduced to 65% when CBZ removalwas 94% in RME matrix (Table 2). Similarly, the mineralization rates ofDCF for DW and RME matrices were 60% and 55%, whilst their re-movals were 93% and 90%, respectively. The removal percentages ofboth CBZ and DCF molecules by photocatalysis were noticeably higherthan the corresponding removal percentage of DOC. This might beexplained by the formation of aromatic organic intermediates fromparent molecules of target pollutants, which leads to the increase oforganic molecules in the solution [12,40]. Moreover, results showedthat the mineralization rates of both the target compounds were in-hibited to some extent in RME matrix as compared to DW matrix. Thisfact evidently confirms the influence of organic matter (cf. Table S1)present in the RME matrix.

3.5. Role of reactive oxygen species and mechanism involved in thephotocatalytic degradation

The influence of reactive species (h+, %OH, %O2–) involved during

the photocatalytic degradation of CBZ and DCF using 5% Ag2O/P-25 isshown in Fig. 7. In Fig. 7a, the photocatalytic degradation of CBZ underUV irradiation for 60min was substantially suppressed with the addi-tion of EDTA and MeOH by nearly 70% and 90%, respectively, ascompared to the control test. However, the degradation percentagedecreased slightly (18%) when O2-purging test was performed. Theresults indicated that trapped holes and hydroxyl radicals adsorbed onthe surface of the 5% Ag2O/P-25 were the dominant oxidative speciesduring the photocatalytic degradation of CBZ molecules. Similarly, thedegradation rate of DCF in the presence of EDTA was substantiallydecreased to 62% as compared to the control test (Fig. 7b). However,the effect of either MeOH or O2-purging was not significant (∼7%),

which suggests that the trapped holes were the main oxidative speciesprompted due to UV irradiation of 5% Ag2O/P-25, and thus accoun-table for the photocatalytic degradation of DCF. Interestingly, theconversion of DCF was not influenced due to the addition of MeOH. Toconfirm this result, other scavengers such as isopropyl alcohol and tert-Butanol were also tested, however similar effects were observed (datanot shown). This might be due to the higher affinity of photo-generatedROS with DCF molecules, compared to the added scavengers. Moreover,the major role of photogenerated holes during the photocatalytic oxi-dation of TiO2 based photocatalyts has been reported [55]. In addition,the organic pollutants could also absorb actinic photons from UV irra-diation and transfer energy or electrons to produce reactive oxidationspecies, which in turn degrade the organic pollutants. This phenomenonis referred to as self-sensitized photooxidation of pollutants [56].

On the basis of above results, a possible photocatalytic mechanismscheme of the heterojunction 5% Ag2O/P-25 photocatalyst was pro-posed as shown in Fig. S3 (Supplementary information). When com-posite Ag2O/P-25 was irradiated by UV, both e− and h+ were excited inthe conductance band (CB) and valence band (VB), respectively ac-cording to Eqs. (2) and (3). In normal case, most of e− – h+ pairs re-combine rapidly to produce heat. As ECB (TiO2, −0.4 eV vs. SHE) islower than E (O2/%O2

–) (−0.046 eV vs. SHE), %O2

– could produce fromdissolved O2 [57,58]. However, the ECB (Ag2O, +0.2 eV vs. SHE) washigher than E (O2/%O2

–) , thus unable to produce %O2

– from dissolved O2

[58,59]. The electron produced on CB of TiO2 can transferred easily toAg2O as ECB of TiO2 is positioned above that of Ag2O [57]. Subse-quently, the partial lattice Ag+ in Ag2O nanoparticles on the surface ofTiO2 reacts with photoinduced electrons and reduced to metallic Ag(Eq. (4)). Further, the metallic Ag acted as an electron pool andtransferred the produced electrons to combine with O2. The Ag2O aselectron absorbent inhibits the recombination of the photogenerated e−

– h+ pairs, which allows the holes to effectively oxidize organic pol-lutants, and thus the photocatalytic activity is highly improved [30].Meanwhile, the generated electrons are trapped by dissolved O2 toproduce %O2

–, and transformed to reactive oxygen species %OH bymultistep reductions (Eq. (5)–(8)). Moreover, the EVB (TiO2, +2.8 eV vs.SHE) was higher than the E (%OH/H2O) (+2.68 eV vs. SHE) [57,59],indicating that the photogenerated holes at VB of TiO2 could oxidizeH2O to %OH according to Eq. (9). Subsequently, photoinduced holes inthe VB of TiO2 can easily move and accumulate at the VB of Ag2O sinceEVB (Ag2O, +1.0 eV vs. SHE) is less positive than that of TiO2. As aresult, the recombination potential of e− - h+ pairs was largely re-duced, thus greatly enhancing quantum efficiency [28]. Moreover,photoinduced holes transferred and accumulated in the VB of Ag2Ocould directly oxidize the target pollutants to a great extent. The pho-togenerated reactive species (h+, %OH, %O2

–) could degrade CBZ or DCFmolecules into intermediates and finally to mineralization (Eq. (10)).

→ +− +hTiO e

h2

ϑ(2)

→ +− +hAg O e

h2

ϑ(3)

Ag2O+H2O+2e−→ 2Ag+2OH– (4)

e−+O2→%O2

– (5)

%O2–+H+→ %OOH (6)

%OOH+H++e−→H2O2 (7)

H2O2+ e−→ %OH+OH– (8)

h++H2O→ %OH+H+ (9)

h+, %O2–, %OH+CBZ or DCF→ Intermediate products→ CO2+H2O

(10)

Fig. 7. Trap experiments to assess the role of reactive species during the pho-tocatalytic degradation of (a) CBZ, and (b) DCF. (Operating conditions: [CBZ]0= [DCF] 0=20mg L−1; [5% Ag2O/P-25]= 0.4 g L−1).


324

3.6. Reusability of 5% Ag2O/P-25 catalyst on PhACs removal

The reusability of any catalyst during photocatalytic process is animportant factor to evaluate its performance in practical applications.For this aim, cyclic photocatalytic experiments were performed for bothtarget pollutants on 5% Ag2O/P-25, as shown in Fig. 8. After, fiveconsecutive cycles of photocatalytic degradation of CBZ, the catalystdid not exhibit any significant loss of its activity (Fig. 8a). The resultsare indicative of the potential practical application of Ag2O/P-25 cat-alyst based on its remarkable reusability for CBZ degradation. Theplausible photocatalytic activity of the selected catalyst might be at-tributed to the self-stability exhibited by Ag2O nanoparticles once thecomplex structure of Ag-Ag2O is formed during the photocatalyticprocess [60]. Moreover, the separation of catalyst from aqueous solu-tion was achieved in all the cycles with just 15min of centrifugation,which leads to the conclusion that the complex catalyst synthesized caneasily be recovered without high filtration costs, as normally associatedwith TiO2 nanoparticles [22].

On the other hand, the photocatalytic removal efficiency of DCFinhibited gradually during cyclic reuse of composite catalyst (Fig. 8b).The removal percentage of DCF dropped by 35% up to 5th cycle ascompared to 1st cycle. Nevertheless, as the inhibition rate of the pho-tocatalytic activity was acceptable and remained almost stable after2nd cycle, the synthesized 5% Ag2O/P-25 demonstrated a moderatelystable activity towards DCF degradation, and thus could be re-commended for the long-term application. However, the inhibitorymechanism observed during the photocatalytic process of DCF underthe identical working conditions of CBZ was not clear. This might bedue to the possible de-activation of catalyst by surface blockage, irre-versible adsorption on active sites, and aggregation of catalyst particles[47]. A clear transformation in a physical appearance of 5% Ag2O/P-25was observed before and after the reaction (Fig. S4). The interferingcompounds formed during rapid initial photocatalytic process (organicby-products etc.) may block the active surface sites, allowing neithertarget pollutants nor reactive species to adsorb and react with photo-generated charges. Interfering compounds may act as a bridge or ad-hesive between the particles, and also can scavenge the generated ROSsuch as %OH, or %O2

–, either discarding them or transforming into lessactive ones [61]. The above hypothesis might be partially supported bythe results from trapping experiment (cf. Fig. 7b), in which the influ-ence of %OH, and %O2

– was negligible on the photocatalytic degradationof DCF.

3.7. CBZ intermediates identification and possible degradation pathway

Of the selected PhACs, transformation products formed during thephotocatalytic degradation of CBZ (as an illustrative) using Ag2O/P-25

under UV irradiation were analyzed using high-resolution FT-ICR massspectrometry and detected based on their accurate masses and deducedelemental formulae (Table S3, supplementary information). Thecommon transformation products of CBZ such as 10,11- dihydro-CBZ-10,11-epoxide (m/z 253), acridine-9-carboxaldehyde (m/z 208), hy-droxyacridine-9-carboxaldehyde (m/z 224) [38,39,50,62] were de-tected. The photocatalysis transforms CBZ to degradation productsshowing other ecological properties [50]. The large number of com-pounds formed during the photocatalytic degradation evidences thecomplexity involved in the photocatalysis of CBZ, thus indicating thepossible occurrence of multi-step degradation routes and inter-connected pathways [16].

A photocatalytic degradation mechanism of CBZ shown in Fig. 9was proposed based on the degradation products unambiguouslyidentified photoproducts in this study and proposals of the previousliteratures [38,39,62,50]. The short-lived proposed intermediates are inthe brackets, while stable intermediates are inside the parentheses.Since substituents could appear in different positions in the aromaticrings, they have been represented as crossing the parenthesis marks. Asverified by the quenching experiments, the major reactive species in-volved in photocatalytic degradation of CBZ in aqueous Ag2O/P-25suspensions were photogenerated h+ and %OH radicals.

In photocatalytic process, hydroxylation of aromatic rings by %OHradicals promotes the sequential ring cleavage reactions [7]. Theremight be two different routes (a, and b) when photogenerated %OHradicals react with CBZ to form hydroxy-CBZ. First, %OH radical attacksthe aromatic ring of CBZ, which forms intermediate 1 via route b. Theintermediate 1 can further be transformed into intermediate 6 (oxcar-bamazepine). Alternatively, intermediate 2 (epoxide-CBZ) and inter-mediate 5 ((2-carboxyphenyl)amino-5-hydroxybenzoic acid) form viaroute a due to short-lived reaction intermediate, hydroxy-CBZ. Thereaction of intermediate 2 with %OH radicals can form other degrada-tion products, such as intermediate 3 (acridine-9-carboxaldeyde). Thering cleavage in the intermediate 3 is followed by the loss of amidegroup, which leads to an undetected intermediate, which further reactswith %OH to form intermediate 4 (hydroxyacridine-9-carboxaldehyde).Due to ring cleavage, intermediate 5 can further be transformed intoother intermediates 6 (anthranilic/salicylic acid) and 7 (cathecol). Fi-nally, the complex hydroxylated intermediates and low molecular masscompounds can be oxidized through ring-rupturing reactions to ali-phatic compounds and further oxidation to mineralization (CO2 andH2O).

4. Conclusion

A heterojunction 5% Ag2O/P-25 was selected as an efficient pho-tocatalyst by a screening assessment of four different photocatalysts.

Fig. 8. Cyclic experiments in the photocatalytic degradation of (a) CBZ, and (b) DCF over 5% Ag2O/P-25.


325

Dark adsorption, UV photolysis and photocatalytic kinetics of targetpollutants were investigated as preliminary tests. The selected 5%Ag2O/P-25 catalyst was prepared and characterized by SEM, EDX, BET,and UV–vis diffuse reflectance spectra. The effects of catalyst dose,

initial pollutants concentration, and mineralization during the photo-catalytic degradation of CBZ and DCF were investigated in both DW andRME solution matrices. The optimal CBZ and DCF removals of 89.10%and 93.5%, respectively for 180min of UV irradiation were observed at

Fig. 9. Proposed photodegradation pathway of CBZ by Ag2O/P-25 under UV light irradiation. [CBZ]0=20mg L−1, Catalyst dosage=0.4 g L−1.


326

catalyst dose of 0.4 g L−1 in DW matrix. However, the optimal catalystdoses for CBZ and DCF in RME matrix were respectively increased byfactor 2 and 1.5, to achieve the same level of removal, which may bedue the influence of organic matters and natural scavengers in the realeffluents. A noticeable decreasing trends in removal rates were ob-served with the increased initial concentration of both PhAC, and ki-netics seem to fit the Langmuir-Hinshelwood model. A substantial im-provement in mineralization was also achieved. The results showed thatphotoinduced holes and %OH were the dominant oxidation species inthe photocatalytic degradation of PhACs. A possible reaction me-chanism and energy band matching of complex 5% Ag2O/P-25 was alsodemonstrated. Good reusability was exhibited by the selected photo-catalyst during photocatalytic degradation of target PhACs, ensuring itsaptitude for practical applications. Moreover, multiple routes of theCBZ transformation and various intermediates formed during photo-catalytic degradation were studied and a possible multi-step degrada-tion path was proposed.

Acknowledgement

This work was financially supported by Business Finland Oy(Innovaatiorahoituskeskus) (formerly Tekes – the Finnish Funding Agencyfor Technology and Innovation) – A research funding for the SmartEffluents Project. Authors heartily thank Reijo Turkki (Director ofMikkeli Vesilaitos), Risto Repo, Anne Bergman, and Päivi Luukkonenfor their consistent support to this work. The FT-ICR MS facility issupported by Biocenter Finland/Biocenter Kuopio and EuropeanRegional Development Fund (Grant A70135).

Appendix A. Supplementary material

Supplementary data to this article can be found online at https://doi.org/10.1016/j.seppur.2018.12.069.

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866M

EMBRAN

E BIOREACTOR FOR THE REMOVAL OF EM

ERGING CON

TAMIN

ANTS FROM

MUN

ICIPAL Khum

Gurung

ISBN 978-952-335-410-4ISBN 978-952-335-411-1 (PDF)

ISSN-L 1456-4491ISSN 1456-4491

Lappeenranta 2019

Khum Gurung A4.pdf - LUTPub

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