Environment International 69 (2014) 28–39

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Environment International journal homepage: www.elsevier.com/locate/envint

Review

Fate of diclofenac in municipal wastewater treatment plant — A review Niina Vieno a,⁎, Mika Sillanpää b,1 a b

Envieno, Logomo Byrå, Köydenpunojankatu 14, FI-20100 Turku, Finland Lappeenranta University of Technology, Laboratory of Green Chemistry, Innovation Centre for Safety and Material Technology, Sammonkatu 12, FI-50130 Mikkeli, Finland

a r t i c l e

i n f o

Article history: Received 12 November 2013 Accepted 20 March 2014 Available online xxxx Keywords: Pharmaceutical Biotransformation Sorption Conventional activated sludge Membrane bioreactor Attached-growth bioreactor

a b s t r a c t Diclofenac (DCF) is a common anti-inflammatory pharmaceutical that is often detected in waste wasters, effluents and surface waters. Recently, DCF was included in the watch list of substances in EU that requires its environmental monitoring in the member states. DCF is also known to harmfully affect several environmental species already at concentrations of ≤ 1 μg/l. This review focuses on the occurrence and fate of DCF in conventional wastewater treatment processes. Research done in this area was gathered and analyzed in order to find out the possibilities to enhance DCF elimination during biological wastewater treatment. More precisely, human metabolism, concentrations in wastewater influents and effluents, elimination rates in the treatment train, roles of sorption and biotransformation mechanisms during the treatment as well as formation of transformation products are reported. Additionally, the effect of process configuration, i.e. conventional activated sludge (CAS), biological nutrient removal (BNR), membrane bioreactor (MBR) and attached-growth bioreactor, and process parameters, i.e. solids retention time (SRT) and hydraulic retention time (HRT) are presented. Generally, DCF is poorly biodegradable which often translates into low elimination rates during biological wastewater treatment. Only a minor portion is sorbed to sludge. MBR and attached-growth bioreactors may result in higher elimination of DCF over CAS or BNR. Long SRTs (N 150 d) favor the DCF elimination due to sludge adaptation. Longer HRTs (N 2–3 d) could significantly increase the elimination of DCF during biological wastewater treatment. Bioaugmentation could be used to enhance DCF elimination, however, this requires more research on microbial communities that are able to degrade DCF. Also, further research is needed to gain more information about the deconjugation processes and biotic and abiotic transformation and the nature of transformation products. © 2014 Elsevier Ltd. All rights reserved.

Contents 1.

2. 3.

4.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Human metabolism of DCF . . . . . . . . . . . . . . . . . . . . . . . . 1.2. Occurrence of DCF in wastewaters, effluents, environment and drinking waters 1.3. Ecotoxicological effects of DCF . . . . . . . . . . . . . . . . . . . . . . Sorption of DCF to sewage sludge . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Removal of DCF in sludge treatment . . . . . . . . . . . . . . . . . . . . Biological elimination and transformation of DCF . . . . . . . . . . . . . . . . . 3.1. Biotransformation mechanism of DCF . . . . . . . . . . . . . . . . . . . 3.2. Biotic and abiotic transformation products . . . . . . . . . . . . . . . . . Elimination of DCF during biological wastewater treatment . . . . . . . . . . . . . 4.1. Effect of process configuration . . . . . . . . . . . . . . . . . . . . . . 4.2. Effect of process parameters . . . . . . . . . . . . . . . . . . . . . . . 4.2.1. Solids retention time (SRT) . . . . . . . . . . . . . . . . . . . . 4.2.2. Hydraulic retention time (HRT) . . . . . . . . . . . . . . . . . .

⁎ Corresponding author. Tel.: +358 50 5448431. E-mail addresses: [email protected] (N. Vieno), mika.sillanpaa@lut.fi (M. Sillanpää). 1 Tel.: +358 400 205215.

http://dx.doi.org/10.1016/j.envint.2014.03.021 0160-4120/© 2014 Elsevier Ltd. All rights reserved.

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N. Vieno, M. Sillanpää / Environment International 69 (2014) 28–39

5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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37 37

1. Introduction

1.1. Human metabolism of DCF

Diclofenac (2-(2-(2,6-dichlorophenylamino)phenyl)acetic acid) (DCF) is a common non-steroidal anti-inflammatory drug (NSAID) that is used as oral tablets or as a topical gel. It is sold under the commercial names of Acoflam, Algosenac, Almiral, Ana-Flex, Anthraxiton, Antiflam, Arcanafenac, Arthrex, Arthrifen, Arthtotec, Diclabeta, Diclac, Dicloabac, Diclodoc, Diclofenac–Ratiopharm, Diclofenbeta, Diclomex, Diclowal, Dicuno, Difen, Diklotab, Dolgit–Diclo, Eese, Effekton, Jutafenac, Monoflam, Motifene Dual, Rewodina, Sigafenac and Voltaren. Its physicchemical properties are presented in Table 1 and molecular structure in Fig. 1. In the point of view of environmental regulations in EU, pharmaceuticals and hormones are a highly topical group of compounds. A proposal was made during the revision of the Water Framework Directive (2000/60/EC) in European Union that would have classified DCF along with two estrogenic hormones as priority substances. In the end, in Directive 2013/39/EU, DCF and the hormones were included in the watch list of substances that will be established alongside the list of priority substances (EU, 2013). Substances in the watch list shall be monitored by the EU member states in their surface waters for a maximum of four years. However, no environmental quality standards (EQS) were assigned for the watch list substances. However, during the revision process, EQS value of 100 ng/l for inland waters and 10 ng/l for coastal water were proposed for DCF. Due to the wide interest of regulators and the public to DCF, this review focuses solely on this common antiinflammatory drug. Its fate in the human body and during the municipal wastewater treatment is reviewed. Additionally, mechanisms of sorption and biotransformation as well as formation of transformation products are discussed. The effect of process configuration, i.e. conventional activated sludge (CAS), biological nutrient removal (BNR), membrane bioreactor (MBR) and attached-growth bioreactor as well as process parameters, i.e. solids retention time (SRT) and hydraulic retention time (HRT) are reviewed. No review has been published that would concentrate on DCF, its fate and transformation processes in conventional wastewater treatment and would have aimed to identify the possibilities to increase its elimination by enhancing the existing treatment processes. Previously published reviews have focused in reporting the concentrations of wide range of micropollutants (Ratola et al., 2012) or their removals in various biological systems (Li et al., 2014; Onesios et al., 2009; Verlicchi et al., 2012).

DCF is administered topically or orally and undergoes almost complete biotransformation in the human body. Topical gel adsorption was found to be 6–7% (Davies and Anderson, 1997). The remaining part is either washed off the skin or is attached to clothing. Of the orally administered dose, between 65 and 70% is excreted in urine and 20–30% in feces as the parent drug or as metabolites (Davies and Anderson, 1997; Stierlin and Faigle, 1979). The majority of DCF is metabolized in the human body and only b 1% of the orally administered dose is excreted as un-metabolized DCF. As a result of phase II metabolism involving glucuronic acid and taurine, glucuronide and sulfate conjugates of DCF are formed. These conjugates make up to 11% of the administered dose (Davies and Anderson, 1997; Stierlin and Faigle, 1979). The World Health Organization has defined a daily dose for diclofenac of 100 mg. Of this dose, less than 1 mg is eliminated from the human body as DCF and about 11 mg as DCF conjugates. The rest of the administered dose is excreted as metabolites of DCF or their conjugates. Metabolic pathway of DCF in human body is presented in Fig. 1. Six phase I metabolites of DCF have been detected in human plasma, urine and/or feces. According to Davies and Anderson (1997), the pattern of DCF metabolites in human urine is the same after topical and oral administration. In total, the six metabolites of DCF and the conjugates of these account for 90% and 65% of the total administered DCF in urine and in feces, respectively (Blum et al., 1996; Davies and Anderson, 1997; Faigle et al., 1988; Stierlin and Faigle, 1979). The main human metabolites of DCF are 4′-OH-DCF and 5-OH-DCF. Both of them are excreted mainly in conjugated form and only less than 1% is excreted unchanged (Stierlin and Faigle, 1979). Important metabolites are also 3′-hydroxy-DCF and 4′,5-dihydroxy-DCF. The two remaining metabolites, 3′-OH-4′-OCH3-DCF and 4′-OH-3′-OCH3-DCF are excreted in urine only in trace amounts (Blum et al., 1996; Faigle et al., 1988). In some animal models, 4′-OH-DCF has been shown to have 30% of the anti-inflammatory and antipyretic activity of DCF (Davies and Anderson, 1997). However, according to Wiesenberg-Boettcher et al., the antiinflammatory activity of the main metabolites of DCF is at least 10 times lower compared to DCF activity (Wiesenberg-Boettcher et al., 1991).

Table 1 Physico-chemical properties of diclofenac. Parameter

Value

Reference

Chemical formula CAS no

C14H10Cl2NO2 15307-86-5 15307-79-6 (disodium salt) 2.37 mg/L 4.15 4.51 2.7 2.3 1.2 2.1 2.3–2.5 1.3–2.2

– – – SRC (2013) SRC (2013) SRC (2013) Ternes et al. (2004a) Radjenovic et al. (2009) Ternes et al. (2004a) Radjenovic et al. (2009) Radjenovic et al. (2009) Carballa et al. (2008)

Water solubility pKa logKow logKd,primary sludge logKd,secondary sludge logKd, MBR logKd,digested sludge

1.2. Occurrence of DCF in wastewaters, effluents, environment and drinking waters In reviewed studies, the measured maximum concentrations of DCF in municipal wastewaters vary between 0.44 and 7.1 μg/l and the mean concentrations are between 0.11 and 2.3 μg/l (Table 2). According to Sim et al. (2011), the maximum concentrations in hospital wastewater reached 6.88 μg/l and in pharmaceutical manufacturer's wastewater 203 μg/l in South Korea. The values are significantly higher than normally detected in municipal wastewater. On the other hand, Zorita et al. (2009) measured similar concentrations (around 0.2 μg/l) in both hospital and municipal wastewater in Sweden. Municipal wastewater concentrations reflect the consumption of DCF by the residents in the particular sewer system. The consumption rates vary greatly between countries and also within countries. This makes it difficult to determine typical wastewater concentrations. The yearly consumption of DCF has been reported to vary between 195 and 940 mg per inhabitant in different countries (Carballa, 2005; Clara et al., 2005b; Finnish Medicines Agency, 2013; Khan and Ongerth, 2004; Sakshaug, 2012; Ternes, 1998). In the effluents of municipal wastewater treatment plants, DCF is among the most frequently detected pharmaceuticals (Verlicchi et al.,

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N. Vieno, M. Sillanpää / Environment International 69 (2014) 28–39

3’-hydroxy-4’-methoxy-diclofenac (3’-OH-4’-OCH3-DCF)

4’-hydroxy-3’-methoxy-diclofenac (3’-OH-4’-OCH3-DCF)

Diclofenac (DCF)

Diclofenac-1-β-o-acyl glucuronide (DCF-G)

4´-hydroxy-diclofenac (4´-OH-DCF)

5-hydroxy-diclofenac (5-OH-DCF)

3´-hydroxy-diclofenac (3´-OH-DCF)

4´,5-dihydroxy-diclofenac (4´,5-OH-DCF)

R, R´, R´´= not definitely identified ligands, presumably one per molecule Fig. 1. Metabolic pathway of diclofenac (DCF) in human body. Adapted from Blum et al. (1996), Davies and Anderson (1997), Faigle et al. (1988) and Stierlin and Faigle (1979).

2012). Due to its incomplete elimination during the treatment, effluent concentrations rarely fall below the detection limits of few nanograms per liter when analyzed using LC–MS/MS or GC/MS. According to Table 2, the effluent maximum concentrations vary between 0.12 and 4.7 μg/l and mean concentrations between b 0.002 and 2.5 μg/l. According to Verlicchi et al. (2012), out of 73 pharmaceuticals reviewed, DCF had the eighth highest average mass load (240 mg/1000 inh) in the secondary effluent of municipal wastewater treatment plants. In addition, the metabolites of DCF can also enter the environment via WWTP effluents. Langford and Thomas (2011) reported that in Norway, 5-OH-DCF was measured up to a concentration of 3.7 μg/l in WWTP effluents. Reported surface water concentrations of DCF generally fall below 100 ng/l (Hernando et al., 2006; Hilton and Thomas, 2003; Kim et al., 2007; Lin et al., 2005; Rabiet et al., 2006; Vieno, 2007). Some studies report higher, but still generally lower than 500 ng/l, surface water concentrations (Bendz et al., 2005; Buser et al., 1998; Kosjek et al., 2005; Metcalfe et al., 2003; Öllers et al., 2001). Ternes (1998) reported a maximum concentration of 1200 ng/l in German rivers receiving sewage effluents. In Karachi, Pakistan, a concentration of 8500 ng/l of DCF was measured in Korangi drain that receives untreated residential and industrial effluents (Scheurell et al., 2009). Also, the hydroxylated human metabolites, 3′-OH-DCF and 4′-OH-DCF, were detected in the sample at concentrations of 300 and 1800 ng/l, respectively. Generally, groundwater concentrations have been low or under the detection limits (Lin et al., 2005; Loos et al., 2010; López-Serna et al., 2013; Rabiet et al., 2006). Few studies report higher groundwater concentrations. In the groundwater of Barcelona, diclofenac was detected up to a concentration of 380 ng/l (López-Serna et al., 2013). Also, 4-OH-DCF was found up to a concentration of 147 ng/l. Similar DCF concentrations (maximum of 380 ng/l) have also been measured in groundwaters in Germany (Heberer, 2002). In drinking waters, DCF concentrations have been reported as below or just above the detection limits (1–7 ng/l) (Benotti et al., 2009; Hernando et al., 2006; Kosjek et al., 2005; Lin et al., 2005; Vieno, 2007; Vulliet et al., 2011).

1.3. Ecotoxicological effects of DCF Haap et al. (2008) reviewed literature studies that have evaluated the ecotoxicology of DCF. They concluded that depending on the species, exposure duration and endpoint used, effect concentrations varied between 1 μg/l and 80 mg/l. Generally, median effective concentrations, i.e. EC50 values, using acute toxicity tests for Daphnia magna varied between 22 and 80 mg/l. Thus, a risk assessment based on routine tests using for example D. magna mortality, would likely underestimate the environmental toxicity of DCF. EC50 values are far higher than the environment concentrations reported for DCF. Also chronic toxicity tests (exposure time of 48 h–10 days) using mortality, reproduction and embryo and larvae mortality was reported to result in the lowest observed effect concentrations (LOECs) of N2000 μg/l for DCF (Ferrari et al., 2003). Based on these studies, Ferrari et al. (2003) calculated a predicted no effect concentration (PNEC) for DCF as 116 μg/l, i.e. 1000-fold higher than are normally measured in the environment. Table 3 summarizes studies that have shown ecotoxicological effects of DCF at environmentally relevant concentrations using different biomarkers as endpoints for exposure. Triebskorn et al. (2004) reported a LOEC of 1 μg/l for rainbow trout (Oncorhynchus mykiss) using changes in liver ultrastructure, liver glycogen and kidney protein as the endpoint. Triebskorn et al. (2007) concluded that DCF showed more effects on fish liver, kidney and gills than carbamazepine, clofibric acid and metoprolol. LOECs for damages in liver, kidney and gills were all at 1 μg/l. Already extremely low DCF concentrations, i.e. 10 ng/l, was reported to impair osmoregulatory ability of a green shore crab Carcinus maenas (Eades and Waring, 2010). The authors suggested that this caused the increase in the haemolymph osmolality recorded for the crab. Lipid peroxidation was used as a biomarker for oxidative stress in the studies of Quinn et al. (2011) and Feito et al. (2012). Quinn et al. (2011) recorded significantly increased levels of lipid peroxidation, when zebra mussels (Dreissena polymorpha) were exposed to a concentration of 1 μg/l of DCF. Feito et al. (2012), on the other hand, reported a reduction of lipid peroxidation in zebrafish embryos already at

N. Vieno, M. Sillanpää / Environment International 69 (2014) 28–39

31

Table 2 Concentrations of pharmaceuticals in municipal sewage treatment plant influents and effluents in different countries. Country

Austria Canada China China China Europeb Europec Europed Finland France Germany Germany Germany Greece Greece Pakistan South Korea South Korea South Korea Spain Spain Spain Spain Spain Spain/Croatia Sweden Sweden Switzerland Switzerland Switzerlanda Taiwan UK USA USA

Influent (μg/l)

Effluent (μg/l)

Min

Max

Mean

0.91

4.1

0.11

0.44

0.29

0.28

0.34

0.32

Median

0.28 0.03a

Max

0.78 0.005 0.035

3.5 0.36 0.46

0.12

0.17 5.5 1.4 1.5 0.62 0.49 4.7

0.15 0.23

0.64

0.42

0.46

7.1

Reference

Min

0.14 0.21

3.02 2.10 2.3e

0.86 n.d.

2.17 5.2

0.09

0.52

0.20

3.60

b0.14 b0.03 n.d.

0.74 0.72 0.56 0.50

0.15 n.d.

0.28

0.24 0.08

0.193

0.05 0.14

0.53 0.23

0.21 b0.015 0.006

0.25

1.1 0.38 2.3e 0.13 1.8 2.2 1.2 0.62 0.74 0.43 0.50

e

1.02

0.10

0.53

0.19

0.32

Median

0.18 0.03a 0.15 0.68 0.29 0.35

2.5 0.81 1.6e 0.41 0.11 0.04 0.19 n.d.

0.09

0.34 0.22 0.32 0.12e 0.49

0.16 0.23 4.70

Mean

0.31 0.10 0.05

0.93 0.70 0.56

0.41

0.46

n.d.

0.12

b0.002 e

0.09

0.58

0.11 0.28

0.25

0.09e 0.012

n.d.

Clara et al. (2005a) Metcalfe et al. (2003) Sui et al. (2011) Yan et al. (2014) Duan et al. (2013) Andreozzi et al. (2003) Hernando et al. (2006) Paxéus (2004) Vieno (2007) Rabiet et al. (2006) Heberer, 2002 Ternes (1998) Quintana and Reemtsma (2004) Samaras et al. (2013) Kosma et al. (2014) Scheurell et al., 2009 Kim et al. (2007) Sim et al. (2011) Nam et al. (2014) Gómez et al. (2007) Kuster et al. (2008) Gracia-Lor et al. (2012) Martín et al. (2012) Rosal et al. (2010) Petrovic et al. (2006) Bendz et al. (2005) Zorita et al. (2009) Buser et al. (1998) Öllers et al. (2001) Soulet et al. (2002) Lin et al. (2005) Hilton and Thomas (2003) Yu et al. (2006) Yu et al. (2013)

n.d. = not detected, ‘b’ = value is below the detection limit. a Estimated from a figure. b Data from France, Greece, Italy and Sweden. c Data from Spain, Belgium, Germany and Slovenia. d Data from France, Greece, Italy, Sweden and Denmark. e n = 1.

a concentration of 0.03 μg/l and an exposure time of 90 min. They also reported phytotoxicological effects of DCF. An increase in mitochondrial activity was reported for fern Polystichum setiferum after a 48-h exposure to a DCF concentration of 0.3 μg/l. Chronic effects for DNA of fern were noted already at a concentration of 0.03 μg/l. A study by Ericson et al. (2010) showed that already at a concentration of 1 μg/l, DCF affected negatively the byssus strength of the Baltic Sea blue mussel (Mytilus edulis trossulus). The effect was more pronounced at higher DCF concentrations. This implies that the mussels impacted by DCF had reduced ability to attach to the underlying substrate. DCF may cause damages in fish eggs or embryos. Hallare et al. (2004) reported that when zebrafish (Danio rerio) embryos were exposed to concentrations of b1000 μg/l of DCF, no significant embryotoxic or proteotoxic effects were noted. However, according to Nassef et al.

(2010), DCF can cause damages in fish eggs or embryos if maternally transferred into the eggs. They injected DCF into the eggs of Japanese medaka fish (Oryzias latipes). At an exposure concentration of 12 ng/ egg, survival rate of the embryos was decreased to 67%, abnormal eye development was reported in 9%, hemorrhage in 16%, yolk-sac shrinkage in 11% and delayed development in 24% of the embryos. Significantly prolonged hatching times and the number of larvae failed to swim upward was already noted at a concentration of 5 ng/egg compared to control. In addition to harmful effects to environmental organisms, DCF has been shown to bioconcentrate in mussels and fish. Ericson et al. (2010) measured significantly higher concentrations of DCF in mussels compared to water. At an environmental concentration of 1 μg/l and exposure duration of 8 d the bioconcentration factor was recorded to be in

Table 3 Studies showing ecotoxicological effects (LOEC = lowest observed effect concentration) of DCF on several species at environmentally relevant concentrations. Species

Exposure time

Endpoint

LOEC

Reference

Oncorhynchus mykiss Carcinus maenas Dreissena polymorpha Mytilus edulis trossulus Polystichum setiferum

28 d nr 96 h 8d 48 h 90 min

1 μg/l 0.01 μg/l 1 μg/l 1 μg/l 0.3 μg/l 0.03 μg/l 0.03 μg/l

Triebskorn et al. (2004) Eades and Waring (2010) Quinn et al. (2011) Ericson et al. (2010) Feito et al. (2012)

Danio rerio embryos

Changes in liver ultrastructure, liver glycogen and kidney protein Haemolymph osmolality Lipid peroxidation Byssus strength Mitochondrial activity DNA quantification Lipid peroxidation

nr = not reported.

Feito et al. (2012)

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average of 175 μg/g wet weight. Bioconcentration factors reported in fish (rainbow trout) liver, kidney, gills and muscle tissue at aqueous DCF concentration of 1 μg/l have been reported as: 2732, 971, 763 and 69, respectively (Schwaiger et al., 2004). Kallio et al., 2010 studied the occurrence of DCF and its metabolites in the bile of rainbow trout. In aquarium conditions using spiked DCF of 1.9 μg/l, DCF and its metabolites (4′-OH-DCF, 5-OH-DCF and glucuronides of DCF and the metabolites) were detected in the bile samples (Kallio et al., 2010). Concentrations varied between 40 and 620 μg/l. The total bioconcentration factor (BFC) for DCF and its metabolites were estimated to be 950 (the ratio between the total measured concentrations in the bile and the measured exposure concentration in water). In the later study of the authors (Brozinski et al., 2013), DCF could be detected in the biles of bream and roach caught from a lake that receives municipal wastewater effluents. DCF concentrations in the lake ranged from 22 to 302 ng/l whereas the concentrations in the bile of bream and roach were up to 95 and 148 μg/l, respectively, i.e. roughly 1000 times higher than the aqueous concentrations. In the environmental waters, DCF has been shown to rapidly phototransform and to form several transformation products (Buser et al., 1998; Svanfelt, 2013). Not only DCF itself but also these environmental transformation products may pose a risk to aquatic organisms. Schmitt-Jansen et al. (2007) reported a six-fold increase in phytotoxicity evaluated using chlorophyte Scenedesmus vacuolatus when DCF was exposed to sunlight for 53 h. To summarize, humans excrete free and conjugated DCF and its metabolites are excreted to municipal wastewater. DCF is shown to be ubiquitous in municipal wastewater treatment plant effluents as well as in the aquatic environment. DCF has been shown to pose harmful effects on organisms used in various test systems and to bioconcentrate in fish and mussel at environmentally relevant concentrations. Therefore, the load of DCF into the aquatic environment should be reduced. 2. Sorption of DCF to sewage sludge In general, sorption of pharmaceuticals on sludge depends on the lipophilicity and acidity of the compound as well as the ambient conditions such as pH, ionic strength, temperature and the presence of complexing agents, and the properties of the sludge (Carballa et al., 2005; Carballa et al., 2008). Sludge properties in a treatment plant vary according to the type of plant and its method of operation. Primary sludge is collected from the bottom of a primary sedimentation basin. Secondary sludge is produced in the biological secondary treatment. Its character and quantity depend on the biological treatment method that can be for example activated sludge, MBR or moving bed biofilm reactor (MBBR). Sludge removed from the treatment is often treated by digestion (digested sludge) and/or composting (composted sludge). Characteristics of sludge types differ greatly. For example, pH of the primary sludge is in general lower and fat and grease content higher than that of activated or digested sludge. Compared to activated sludge, the MBR sludge has a smaller particle size and thus a higher surface area for adsorption (Kimura et al., 2007). Also, the content of inert matter is often higher compared to activated sludge due to longer sludge age applied in MBR process (Joss et al., 2006). All in all, sorption of compounds is expected to be different to different sludge types. Sorption of compounds to sludge occurs via absorption and adsorption. In absorption, the aliphatic and aromatic groups of the compound interact hydrophobically with the lipophilic cell membrane of the micro-organisms and the lipid fraction of the sludge (Ternes et al., 2004a,b). In adsorption, the positively charged groups of the chemical interact electrostatically with the negatively charged surfaces of the micro-organisms (Ternes et al., 2004a,b). Sorption accounts both absorption and adsorption and it can be estimated by Kd value. It is the ratio of compound's concentration in the solid and in the aqueous phase at equilibrium conditions (Carballa et al., 2005). According to Ternes et al. (2004a), sorption of a compound to sludge can be

considered negligible when Kd value is b 500 l/kgSS (i.e. logKd b 2.7). According to Joss et al. (2005), Kd should be over 300 l/kg (i.e. logKd N 2.5) for an efficient sorption. Diclofenac is slightly soluble in water and has a moderately low octanol–water coefficient (Table 1). Measured and calculated logKd values of DCF to different sludge types in wastewater treatment vary between 1.2 and 2.7 (Table 1) (Carballa et al., 2008; Radjenovic et al., 2009; Ternes et al., 2004a,b). Highest Kd values have been reported for primary sludge. In Ternes et al., 2004a, sorption of diclofenac was b5% to secondary sludge whereas the sorption was 5–15% to primary sludge. Radjenovic et al. (2009) also measured higher sorption potential of DCF to primary than to secondary sludge. In a sewage treatment plant in Sweden, DCF concentration was noted to decrease by 50% during the pre-treatment of sewage by grit removal and primary sedimentation (Zorita et al., 2009). This behavior suggests that DCF primarily interacts with the sludge via adsorption where pH plays a significant role (Ternes et al., 2004a). The carboxylic acid moiety of DCF is negatively ionized at neutral pH and hence the compound repels the negatively charged sludge. At acidic pH, DCF becomes electronically neutral thus allowing it to adsorb to the sludge. In primary treatment, process pH is in general lower than in biological treatment unit. For example, in Ternes et al. (2004a), pH values were 6.6 and 7.5 at the primary and secondary treatment, respectively. Urase et al. (2005) noted that in MBR reactor the lower pH enhanced the DCF removal. Removal was only 10% at a pH range of 6.8–8.0 but was increased to 80% at the MBR reactor with pH in the range of 4.3–5. At neutral pH the carboxyl group of DCF is dissociated and consequently the compound is negatively charged. The authors suggested that the removal was due to adsorption to sludge and the adsorption was promoted by the lower pH. Additionally, Urase and Kikuta (2005) noted in their laboratory batch experiments that partitioning of DCF to activated sludge increased by more than 20 times when pH was decreased from 6.7 to 4.4. Sorption of DCF to MBR sludge has been found to be slightly higher than to activated sludge (Table 1). However, there are differences between MBR sludges. Radjenovic et al. (2009) measured lowest concentration of DCF in the sludge that was collected from an MBR operating at prolonged SRT and high concentration of suspended solids. The authors suggested that this was due to the higher biodegradation potential of the sludge resulting in lower amount of soluble DCF for sorption. At wastewater treatment plants, sorption of DCF to sludge have been reported but to a low extent (Gracia-Lor et al., 2012; Kimura et al., 2007; Martín et al., 2012; Suárez et al., 2012; Verlicchi et al., 2012). Samaras et al. (2013) reported that 8–19% of DCF was bound to particles in raw wastewater. In dewatered sludge, DCF concentration was 30 μg/kg and in raw and digested sludge below the detection limit. Yu et al. (2013) reported a mean concentration of DCF in sewage sludge as 48.4 μg/kg. Martín et al. (2012) found that DCF concentrations were always below the detection limits (1.22–33.1 μg/kg) in different sludges collected from wastewater treatment plant. In activated sludge, concentrations of 150–450 μg/kg have been reported (Radjenovic et al., 2009; Ternes et al., 2005). In primary sludge, Radjenovic et al. (2009) measured DCF at a concentration of 200 μg/kg and at MBR sludge at 100–175 μg/kg. In digested sludge, concentrations of 200 μg/kg have been reported (Carballa et al., 2007; Ternes et al., 2005). Radjenovic et al. (2009) calculated that DCF load in the aqueous effluent and in the treated sludge from a Spanish wastewater treatment plant was in average of 46.5 g/d and 6.4 g/d, respectively. Thus, only about 10% of the environmental load from wastewater treatment plant can be estimated to occur via solids. 2.1. Removal of DCF in sludge treatment Excess sludge from the treatment process is often anaerobically digested to stabilize sludge and to produce methane for energy production. Depending on the temperature, digestion can be mesophilic (optimum T = 25–40 °C) or thermophilic (optimum T = 55–65 °C). Higher

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temperature accelerates biochemical reactions and has been noted to slightly enhance the removal of solids and organic material in comparison to mesophilic process (Carballa et al., 2007). According to Carballa et al. (2007), DCF removal rates in both mesophilic and thermophilic processes are similar and can reach 80%. According to Samaras et al. (2013), other NSAID ibuprofen and naproxen were removed by N80% during mesophilic sludge digestion. However, DCF concentration was already below the detection limit in the raw sludge and thus process performance could not be evaluated. However, in Carballa et al. (2007) high removal rates were only detected at long operating times suggesting that sludge acclimation followed by broader biodiversity is needed for DCF removal. Sorption of DCF to digested sludge has been found to be fairly low (Table 1) even though Carballa et al. (2008) had the following hypothesis. The digested sludge has a different composition, structure and morphology and it can be modified by different pretreatment techniques. Thus, the adsorption of DCF to digested sludge could be higher than to primary and secondary sludge. Therefore, primary elimination method in sludge treatment is most likely biotransformation rather than sorption. 3. Biological elimination and transformation of DCF The objectives of the biological treatment of municipal wastewater are to coagulate and remove the non-settleable colloidal solids, to reduce the organic content and the nutrients nitrogen and phosphorous. These are accomplished using a variety of micro-organisms, principally bacteria, of which there may be 300 species present. The predominant group of bacteria is heterotrophs that mainly feed on organic carbon molecules. Inorganic matter is taken in by autotrophs, such as ammonia oxidizing bacteria that oxidize ammonia into nitrite. Heterotrophs often out-compete autotrophs which have lower growth rate and are often more sensitive to process conditions and variations. Nitrifiers are the most important group of autotrophic bacteria in biological wastewater treatment. Several bacteria genera are responsible for nitrification, for example Nitrosomonas, Nitrobacter and Nitrospira. Nitrosomonas oxidizes ammonia to nitrite (NO− 2 ), which is further converted to nitrate (NO− 3 ) by Nitrobacter and Nitrospira. Nitrifying bacteria are sensitive organisms whose optimum pH and temperature ranges are narrow. They are also sensitive to inhibitors. In fact, some pharmaceuticals have been noted to inhibit the performance of ammonia oxidizing bacteria (Wang and Gunsch, 2011). For example, gemfibrozil and naproxen were noted to decrease ammonia removal by 45% in a pilot scale sequencing batch reactors mimicking wastewater treatment plants operations (Wang and Gunsch, 2011). COD removal was not affected by the addition of pharmaceuticals, suggesting that heterotrophic bacteria are more robust to these compounds. Kraigher et al. (2008) noticed that addition of pharmaceutical mix of ibuprofen, naproxen, ketoprofen, diclofenac and clofibric acid into a pilot scale wastewater treatment system reduced the diversity in bacterial communities already at an individual pharmaceutical concentration of 5 μg/l. Most significantly, genus Nitrospira was found only in the reactor without pharmaceuticals representing 8% of the total community. Genus Nitrospira plays a key role in nitrite oxidation during biological wastewater treatment. Reduction in species diversity is often reported in response to stress factors, e.g. phenol shocks and elevated temperatures. A reduction in bacterial diversity may affect the essential functions of the activated sludge wastewater treatment systems. However, in the later study of the authors (Kraigher and Mandic-Mulec, 2011) addition of pharmaceuticals in the reactors was not noted to decrease the nitrification activity of activated sludge bacteria and the removal of ammonia was constantly N 90%. However, in reactors with pharmaceutical con− centration of 50 μg/l, concentrations of N–(NO− 2 + NO3 ) was significantly higher than in reactors with no pharmaceuticals and in reactors with higher pharmaceutical concentration (i.e. 200 and 500 μg/l). The − authors could relate the high concentration of N–(NO− 2 + NO3 ) into the abundance of Nitrospira sublineage II in reactors with 50 μg/l of

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pharmaceuticals. Other reactors only contained Nitrospira sublineage I. They suggested that, at lower concentrations, Nitrospira sublineage II could use pharmaceuticals as carbon source or could co-metabolize them. As the concentrations increase, sublineage I or some other competitor could co-metabolize pharmaceuticals and the advantage of Nitrospira II sublineage is lost. Other explanation that was given by the authors was the suggestion that pharmaceuticals could negatively influence the activity and/or structure of denitrification bacteria at low pharmaceutical concentration. This could have caused the increase in nitrate concentration and changes in Nitrospira community structure. However, no nitrite accumulation and thus inhibition of nitrite oxidation by pharmaceuticals was noted. The authors did not consider the possibility of abiotic transformation of pharmaceuticals and, more specifically, the potential of some pharmaceuticals, such as diclofenac, to undergo abiotic nitration processes in the presence of nitrite. This phenomenon is further discussed in Section 3.2. At higher pharmaceutical concentration the consumed nitrite into abiotic processes is also increased. The biological elimination of pharmaceuticals in municipal wastewater treatment plants could occur by direct metabolization or by cometabolization. In the former, bacteria use the compound as their primary carbon source whereas in the latter, bacteria break down or partially convert the compound but do not use it as the primary carbon source (Jones et al., 2007; Ternes et al., 2004a,b). Co-metabolism occurs via enzymes that bacteria secrete to break down large organic molecules into monomers that are small enough to be ingested. Many of these enzymes have the potential to degrade or to transform small organic pollutant molecules present in wastewater. For example, E. coli secrete β-glucuronidase enzyme that is capable to deconjugate the βglucuronated pharmaceuticals excreted by the human body. Thus, this could result in releasing the active pharmaceutical into the wastewater. Ammonia monooxygenase (AMO) is secreted by autotrophic ammonia oxidizing bacteria during the oxidation of ammonium to hydroxylamine in aerobic nitrification. Hydroxylamine is further converted to nitrite. AMO is capable of degrading pollutants such as pharmaceuticals via hydroxylation. It has been suggested to be one of the main enzymes in cometabolic reactions for removal of micropollutants from wastewater (Fernandez-Fontaina et al., 2012; Roh et al., 2009; Tran et al., 2009). The ability of bacteria to secrete a particular enzyme may be latent, i.e. the bacterium requires the presence of the particular compound in the water to switch on the genes for the synthesis of the enzyme required for its digestion. In case of process changes of biological wastewater treatment such as increase in sludge age or sudden increase in toxic compounds, those species of bacteria that have the ability to secrete the enzymes to break down a novel food source will grow more rapidly. This process is known as adaptation or acclimation. Current assumption is that due to their trace level concentrations in municipal wastewater, pharmaceuticals are degraded by microbial enzymes through co-metabolism (Fernandez-Fontaina et al., 2012; Onesios et al., 2009; Quintana et al., 2005; Roh et al., 2009; Tran et al., 2009). For example, Quintana et al. (2005) carried out metabolic and co-metabolic biodegradation tests for pharmaceuticals according to ISO 7827 using sludge, which was drawn from a reactor treating wastewater in which all the studied pharmaceuticals were found. They concluded that when pharmaceuticals were used as the sole substrate no degradation occurred during the 28-d test period. On the other hand, in the co-metabolism test, bezafibrate, naproxen and ibuprofen were degraded by 30–96% in 28 days. 3.1. Biotransformation mechanism of DCF Generally, biodegradation of DCF has been found to be slow or nonexisting in biodegradability studies (Buser et al., 1998; Joss et al., 2005; Lee et al., 2012; Perez and Barcelo, 2008; Quintana et al., 2005). Quintana et al. (2005) concluded that DCF could not be degraded neither through metabolic nor co-metabolic route. Also, low

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Table 4 Values of biological degradation constant kbiol of diclofenac at different biological process configurations. kbiol (l g−1SS d−1)

Biological process configuration

Reference

b0.1 b0.1 b0.1 0.04 1.2 0.31–0.52

Conventional activated sludge Membrane bioreactor Anoxic + aerobic Anoxic conditions Nitrifying conditions Nitrification culture

Joss et al. (2006) Joss et al. (2006) Suárez et al. (2012) Suárez et al. (2010) Suárez et al. (2010) Tran et al. (2009)

biodegradation of the main human metabolite 4′-OH-DCF has been reported (Lee et al., 2012). Its pseudo-first order rate constant under biotic conditions has been measured as 0.018 d−1 which translates into a halflife of nearly 40 d (Lee et al., 2012). Biological degradation constant, kbiol, can be used to estimate the biodegradation potential of a compound. Measured kbiol values for DCF are presented in Table 4. Joss et al. (2006) suggested a classification scheme for micropollutants based on the kbiol (as l g−1SS d−1): kbiol b 0.1: no substantial biodegradation (b20%) 0.1 b kbiol b 10: partial biodegradation (20–90%) kbiol N 10: ready biodegradation (N90%). For DCF, Joss et al. (2006) as well as Suárez et al. (2012) measured kbiol values of b 0.1 meaning that the compound does not undergo biodegradation. Suárez et al., 2010 applied following classification scheme to estimate biological degradation of pharmaceuticals: kbiol b 0.5: hardly biodegradable 0.5 b kbiol b 1: moderately biodegradable 1 b kbiol b 5: highly biodegradable kbiol N 5: very highly biodegradable. According to Suárez et al. (2010) kbiol of DCF was 0.04 l g−1SS d−1, i.e. no biodegradation, in anoxic conditions. However, kbiol value of 1.2 l g −1 −1 was measured for DCF in aerobic nitrifying conditions suggestSS d ing high biodegradation potential of DCF. Also, Tran et al. (2009) noted that DCF is moderately biodegradable in experiments that used enriched nitrifying activated sludge (kbiol of 0.31–0.52 l g− 1SS d− 1). Suárez et al. (2010) concluded that enrichment of specific nitrifying bacteria in the reactor were probably responsible of the high biodegradation potential of DCF. Good nitrifying activities have been noted to increase the biodegradation rates of several micropollutants (Fernandez-Fontaina et al., 2012). It has been suggested that the increased removal of several pharmaceuticals under nitrifying conditions could be due to their degradation by ammonia monooxygenase secreted by ammonia oxidizing bacteria (Fernandez-Fontaina et al., 2012; Tran et al., 2009). Although, ibuprofen which is a highly biodegradable anti-inflammatory pharmaceutical, seems to be degraded primarily by heterotrophic bacteria (Roh et al., 2009, Tran et al., 2009). In the case of DCF, some studies argue that nitrifiers are not the sole organisms responsible for DCF biodegradation (De Graaff et al., 2011; Falås et al., 2012). Falås et al. (2012) noted low or negligible removal of pharmaceuticals in the reactor, which contained carriers from a partial nitritation/anammox sludge liquor treatment. According to the authors, this indicates that the ammonia oxidizing bacteria have very limited ability to degrade/transform DCF even at high ammonia concentration and realistic nitritation rates. They concluded that nitrifying bacteria contribute very little or not at all to the removal of DCF in nitrifying biological treatment systems. Consistent with this study are the results reported by De Graaff et al. (2011) who studied the fate of hormones and pharmaceuticals in anaerobic treatment, partial nitritation reactor and anammox reactor. No significant removal of DCF was detected during the treatment. In

Tran et al. (2009), increase in initial ammonia concentration was noticed to significantly increase the DCF removal. Since increase in ammonia concentration increases the AMO concentration the increase could be suggested to be due to degradation by AMO. In addition, inhibition of AMO by allylthiourea addition was noted to decrease the DCF removal from about 75% to 25%. On the other hand, the authors performed a test where they added sodium acetate into the reactor to induce the growth of heterotrophic bacteria. DCF removal was significantly increased after the acetate addition suggesting the significant role of heterotrophs in DCF biodegradation. It therefore seems, that both autotrophic and heterotrophic bacteria contribute to the biological degradation/transformation of DCF. Further research is needed to identify the micro-organism communities that are able to degrade DCF. This information could be used in bioaugmentation of wastewater with these communities to enhance the DCF biotransformation. RodríguezRodríguez et al. (2012) used bioaugmentation of sewage sludge with fungi Trametes versicolor to degrade several pharmaceuticals. During the 42 d bioremediation experiment, DCF was eliminated by 22% without the addition of T. versicolor. Elimination was increased to up 61% by using the fungal treatment. 3.2. Biotic and abiotic transformation products It seems that the degradation pathway of DCF in biological wastewater treatment is a complex process. Unfortunately only few of the biodegradation studies have reported on the structural identity of DCF metabolites and other transformation products (Table 5). Two of the Table 5 Metabolites of biological transformation of diclofenac and their suggested molecular structures. Figure after the letter ‘M’ refers to the molecular weight of the metabolite. Molecular structure or chemical formula

Information about the metabolite Metabolite M190 – Biological wastewater treatment Perez and Barcelo (2008)

Metabolite M250 – Laboratory-scale biological treatment with activated sludge – Isomeric structure, the position and nature of –CH2-group not defined. Kosjek et al. (2009)

C14H11NO2Cl2

Metabolite M275 – Laboratory-scale biological treatment with activated sludge Kosjek et al. (2009) Metabolite M278 – Laboratory-scale biological treatment with activated sludge Kosjek et al. (2009)

Metabolite M324 – Biological wastewater treatment Perez and Barcelo (2008)

Metabolite M340 – Biological wastewater treatment Perez and Barcelo (2008)

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metabolites, i.e. M340 and M324 can be related to nitrifying wastewater bacteria. Metabolite M324 is formed by O-nitritation of the hydroxyl group of the carboxylic acid moiety in DCF. According to the authors (Perez and Barcelo, 2008), this is an untypical metabolic pathway brought about by micro-organisms. Metabolite M340, on the other hand, was concluded to be a product of microbial nitration at one of the aromatic rings. In general, metabolites undergoing ring hydroxylation has not yet been reported for DCF which would have been an evidence on monooxygenase-mediated reaction. There may be another explanation for the formation of nitroproducts during wastewater treatment. Namely, these transformation products of DCF during the nitrification–denitrification process could form via abiotic nitration in the presence of nitrite rather than via biological degradation. This is supported by the study of Gaulke et al. (2008) who conducted batch tests on the biodegradation of 17αethinylestradiol (EE2) with pure cultures of ammonia oxidizing bacteria from Nitrosomonas and Nitrosospira. They reported that co-metabolic degradation of EE2 occurred neither with the addition of ammonia nitrogen at 10 mg/l nor without the addition. On the other hand, ammonia was completely oxidized to nitrite. EE2 could be degraded to some extent when ammonia concentration was increased to 200–500 mg/l. The authors concluded that EE2 transformation was not due to the enzymatic attack by the bacteria but rather by abiotic nitration of EE2 with nitrite produced by the bacteria. Similar could happen to DCF in the wastewater nitrification unit. Also, it has been suggested that abiotic nitration products could return to parent compound when the concentration of nitrite reduces (Barbiere et al., 2012). Barbiere et al. (2012) studied the effect of subsurface artificial groundwater infiltration under nitrate reducing conditions on pharmaceuticals. A sudden drop after the 1.5 d of incubation to about 50% of initial concentration of DCF was noted. However, by day 10, the concentration of DCF had returned to the initial concentration. Thus, after 10-day incubation, no overall removal of DCF was noted. The authors also observed that this drop in concentration occurred currently and oppositely to the reduction of nitrite. Similar trend was noted to sulfamethoxazole, which is also an aromatic amine substance. Authors could identify a degradation by-product to be a nitro analog of DCF, which was suggested to occur via nitration at one of the aromatic rings and thus have the structure of the metabolite M340 (see Table 5). The authors concluded, that a phenomenon that has been previously noted to other aromatic amines, namely formation of nitroproducts in the presence of nitrite, occurs to DCF. Nitro-derivatives return to parent compounds when the concentration of nitrite drops. This phenomenon could also occur in wastewater treatment during nitrification and denitrification process. This can lead to wrong estimation of the actual treatment efficiency of DCF as nitro

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derivatives can transform back to the parent compounds when released into the environment.

4. Elimination of DCF during biological wastewater treatment Generally, high variation in elimination rates has been noted in municipal wastewater treatment plants (Table S1 and Fig. 2). Considering the complex picture of DCF biodegradation/biotransformation discussed in the previous section it seems only natural that high variation occurs when only the influent and effluent concentrations are considered in the calculations of elimination rates. Elimination rates of up to about 80% can be reached at the treatment; however, values in the range of 20–50% are more common. Also, increase in DCF concentration and thus negative elimination has been reported (Clara et al., 2005a; Zorita et al., 2009). Zorita et al. (2009) measured an increase in DCF concentration from 100 to 485 ng/l in activated sludge treatment consisting of subsequent anoxic and aerobic process units for biological nitrogen removal. The authors suggested that this was due to deconjugation of glucuronide or sulfate conjugates of DCF and/or desorption of DCF from the particles. According to Lee et al. (2012) a rapid deconjugation of the glucuronide conjugate to release DCF can occur at wastewater treatment. In microbiological degradation tests, they noted that the glucuronide-DCF deconjugated to form equimolar DCF within 7 d of incubation. Reaction rate constant were 0.28 d−1 (t1/2 ≈ 2.5 d) in the inoculum and 0.12 d−1 (t1/2 ≈ 5.8 d) under sterile conditions. This suggests that abiotic hydrolysis occurs but the reaction is enhanced by micro-organisms. According to Stierlin and Faigle (1979), b 1% of the administered DCF dose is excreted as free DCF and about 11% as conjugated DCF (glucuronide and sulfate). If deconjugation occurs, a 10-fold increase in DCF concentration is possible in the sewer and/or in the biological wastewater treatment. One possible source of DCF could also be biological transformation of another analgesics, namely aceclofenac. Perez and Barcelo (2008) noted that this compound underwent a rapid ester cleavage to DCF during biological wastewater treatment. However, aceclofenac is not as common a pharmaceutical as DCF. For example, its consumption has not been reported in Finland or in Norway (Finnish Medicines Agency, 2013; Sakshaug, 2012). Aceclofenac is sold under the commercial names of Biofenac, Airtal, Barcan, Cartrex, Falcol, Gerbin, Gladio, Kafenac, Preservex, Sanein and Sovipan. Concentrations measured in the wastewater have also been significantly lower than DCF concentrations. In a Spanish wastewater treatment plant, aceclofenac concentrations in average of 33 ng/l were measured whereas DCF concentrations were in average of 349 ng/l (Perez and Barcelo, 2008).

100% 90% 80% 70% 60% 50% 40%

48% 36%

36%

30% 20% 10% 0% Conventional activated sludge

Activated sludge with biological nutrient removal

Membrane bioreactor

Fig. 2. Elimination rates in wastewater treatment plants with different biological treatment processes. Filled symbols and the percentage value refer to the average elimination rates for the treatment method in question. Note that this percentage does not contain the negative elimination rates reported in Clara et al. (2005a), Kosma et al. (2014) and Zorita et al. (2009). References are listed in Table S1.

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Therefore, aceclofenac may not be the main source for increased DCF concentrations reported in wastewater treatment. 4.1. Effect of process configuration Nowadays, conventional activated sludge (CAS) and activated sludge with biological nutrient removal (BNR) are the most commonly applied biological process in full-scale wastewater treatment plants. CAS is an aerobic suspended-growth treatment process. BNR is s suspended-growth treatment process that combines aerobic, anaerobic and anoxic treatment units. Recently, interest in membrane bioreactors (MBRs) has increased. MBR is a membrane process using micro- or ultrafiltration together with suspended-growth biomass. The most important advantage of MBR over CAS or BNR is the complete retention of suspended solids and thus low effluent turbidity. In addition, MBR has less sludge production, 50–70% smaller footprint size and better effluent quality in terms of bacteria, viruses and sometimes also DOC and COD (Bernhard et al., 2006; De Wever et al., 2007). Also, MBR can be used together with powdered activated carbon (PAC), which reduces the membrane cleaning or replacement frequency and could enhance the removal of micropollutants as well (Nguyen et al., 2013). For example, in Nguyen et al. (2012), MBR alone was noted to eliminate DCF by 4–26%. The addition of PAC increased the removal to 96%. Disadvantages of MBR over CAS and BNR are higher cost, higher requirements in operation and maintenance as well as power consumption compared to conventional system (Clara et al., 2005b). A moving bed biofilm reactor (MBBR) is an example of attached-growth biomass process where biomass grows on specially designed carriers that move freely with the reactor water volume and biofilm is grown on the carrier surfaces. Advantages of MBBR are simplicity, compactness, growth of aerobic and anaerobic organisms in the same system, and negligible hydraulic headloss (Zupanc et al., 2013). Table S1 and Fig. 2 show that no single biological treatment configuration outperforms the other. For MBBR, only one study was found that reported the removal of DCF (Zupanc et al., 2013). In the study, MBBR resulted in 74 to 85% removal of DCF compared to 36% in activated sludge system. The authors suggested that the micro-organisms in the biofilm that was developed on the surfaces on carriers were able to exploit DCF as organic substrate. According to Fig. 2, similar average removal rates calculated from literature studies were reported for CAS and BNR processes. However, many individual studies report a higher removal of DCF under nitrifying conditions (Maeng et al., 2013; Suárez et al., 2010; Tran et al., 2009). On the other hand, DCF is not degraded under anoxic conditions, i.e. under denitrifying conditions (Suárez et al., 2010). Since only influent and effluent concentrations have been used to evaluate the elimination rates, the formation of by-products as well as the deconjugation and subsequent release of DCF have not been considered. As was discussed in the previous section, the abiotic nitration of DCF occurs in the presence of nitrite (NO− 2 ). During CAS, nitrification is often incomplete, i.e. nitrite does not fully oxidize to nitrate. Thus, nitrite is free to react with DCF and could result in reduction of monitored DCF concentration. It should be noted, that nitro-DCF can degrade and release DCF in the following treatment units or in the environment. During nitrification, nitrite is rapidly further oxidized to nitrate (NO− 3 ), which does not react with DCF and form nitro-DCF. Thus, more DCF could therefore be released from the reactor compared to the CAS reactor. Some studies have reported higher elimination of DCF in MBR processes compared to CAS, whereas others have noted no difference between the treatments. Radjenovic et al. (2009) reported a higher elimination of DCF in MBR process (in average 63–66%) over the CAS process (in average of 22%). Kimura et al. (2007) noted a removal of 40% in CAS and up to 80% in MBR. Bernhard et al. (2006) compared the optimized laboratory scale MBR and CAS treatment. They reported slightly better elimination (in average of 58%) in MBR than in CAS (in average of 24%) process. On the other hand, Clara et al. (2005b)

concluded that there are no correlation between the treatment method and DCF elimination and that MBR and CAS both showed low removal of DCF. Similar results were obtained by Joss et al. (2005) who reported low (i.e. 20–40%) removal of DCF in CAS, MBR and fixed bed reactors. Also Xue et al. (2010) recorded low DCF elimination rates with all tested treatment configurations (anaerobic, anoxic, arobic and membrane). Further, Kimura et al. (2005) reported a poor removal of DCF in both CAS and two different types of MBRs, namely hybrid MBR and conventional MBR. The studies that report better performance of MBR over CAS or BNR, suggest that higher biomass content and the longer SRTs applied in MBR could be the reason for its better performance (Kimura et al., 2007; Sui et al., 2011). Higher biomass leads to a lower food to micro-organisms ratio and the relative shortage of biodegradable organic matter may force micro-organisms to metabolize more recalcitrant compounds in the sewage (Sui et al., 2011). Longer SRTs allow bacterial population to become more diversified and more capable of degrading DCF either by direct metabolism or by co-metabolic degradation via enzymatic reactions. Also, MBR sludge has a smaller particle size than activated sludge and thus a higher surface area for adsorption. Even though adsorption of DCF to sludge is low, Kimura et al. (2007) suggested that the higher microbial activity in MBR sludge enhances the biodegradation of the adsorbed DCF. A significant advantage of MBR over CAS or BNR is that micropollutant removal in general has been noted to be less sensitive to operational variables (De Wever et al., 2007). According to De Wever et al. (2007), if the micropollutant feed is stopped for some period and feed started again, MBR resumes much faster than CAS. This is probably due to the retention of bacteria required to degrade micropollutants by the MBR membrane. In CAS, the washout of these bacteria is faster and more pronounced. MBR can therefore respond more quickly to temporal fluctuations in micropollutant influent concentrations. 4.2. Effect of process parameters 4.2.1. Solids retention time (SRT) Solids retention time (also known as sludge retention time or sludge age) represents the mean residence time of micro-organisms in a reactor. Only organisms that are able to reproduce themselves during this time can be detained and enriched in the system. High SRT allow the enrichment of slowly growing autotrophic bacteria such as nitrifiers. Therefore, in BNR reactors SRT can exceed 20 days whereas in CAS processes, SRTs normally range between 5 and 15 days. It has been suggested that the presence of more diversified bacterial populations induced by longer SRTs would enhance the degradation of micropollutants either by direct metabolism or by co-metabolic degradation via enzymatic reactions. For example, it has been reported that an increase in SRT would increase the elimination of some pharmaceuticals, such as gemfibrozil, ketoprofen, clofibric acid and EE2 (Clara et al., 2005a; Kimura et al., 2007; Maeng et al., 2013; Suárez et al., 2010, 2012). For DCF, the effect of SRT is not that straightforward. Joss et al. (2005) reported no enhancement of elimination rates of DCF even when extreme SRTs (N 60 d) were applied. Similarly, Suárez et al. (2012) and Clara et al. (2005b) noted no correlation between the elimination of DCF and SRT. Reif et al. (2008) applied an SRT of 72 d in MBR treatment and noted no elimination of DCF. Majority of the studies reporting higher elimination of DCF at higher SRTs can be related to MBR processes. Bernhard et al. (2006) noted an increase in DCF elimination with increasing SRT using MBR reactor. DCF removal was 8–38% when SRT was 20–48 d, 59% at SRT of 62 d and 53% at SRT of 322 d. Kimura et al. (2007) studied the removal DCF in real WWTP applying activated sludge (AS) and in membrane bioreactor (MBR). Sludge ages were 7 d for the AS and 15 d and 65 d, respectively, for the two tested MBRs. The authors noted that the removal of all the pharmaceuticals were the highest at the MBR with the sludge age of 65 d. DCF was removed by 40% at AS and MBR (15 d) but up to 80% at MBR (65 d). The

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only exception to this pattern was presented in Fernandez-Fontaina et al. (2012), where increasing SRT to N 150 d has resulted in 70% elimination of DCF in laboratory scale AS process working under nitrifying conditions. Generally, it seems that increase in SRT may not be a feasible method to increase the elimination of DCF under normal WWTPs. It has been noted that only extremely high SRTs N 150 d may be needed to obtain efficient DCF elimination in activated sludge plants (Fernandez-Fontaina et al., 2012). Application of this high SRT values in real life CAS or BNR plant is unrealistic. 4.2.2. Hydraulic retention time (HRT) Hydraulic retention time is the residence time of the aqueous sewage in a reactor or in the entire process. Increase in HRT has been reported to increase the removal of certain pharmaceuticals, such as ibuprofen, ketoprofen, atenolol, sotalol, metoprolol, fluoxetine and roxithromycin probably due to increased contact time between the micro-organisms and the water to be treated (Fernandez-Fontaina et al., 2012; Maurer et al., 2007; Tauxe-Wuersch et al., 2005). Gros et al. (2010) concluded that pharmaceuticals that are biodegradable (high kbiol, i.e. low half-life t1/2) but have a low tendency to sorb to sludge, are the most influenced by the changes in HRT. Those micropollutants that have a half-life of less than the HRT can be efficiently removed in the biological reactor (Verlicchi et al., 2012; Gros et al., 2010). Suárez et al. (2012) estimated that a half-life of DCF for WWTPs working with biomass concentration in the range of 2–4 g/l would be 2–3.5 days. According to them, only those plants operating at high enough HRT are expected to be able to degrade it to a certain extent. In their study, the applied HRT was only 1 d and the authors suggested that the poor removal of DCF was due to this. HRTs in full scale WWTPs are normally measured in hours not in days. Therefore, increase in HRT could significantly enhance the elimination of DCF. In fact, in Clara et al. (2005a), 70% removal of DCF at a plant with HRT of 13 d was measured, whereas negligible removal was measured at a plant operating at HRT b 1.2 d. 5. Conclusions Elimination of DCF has been shown to vary greatly in laboratory, pilot and full-scale wastewater treatment processes. To some extent, the compound is shown to adsorb to primary sludge but generally biotransformation is considered as the main elimination mechanism in wastewater treatment. However, since DCF is only moderately or poorly biodegradable incomplete elimination during the conventional wastewater treatment applying CAS or BNR processes can be expected. Elimination could be enhanced by the following means: – Application of membrane bioreactor processes or attached-growth biomass process may result in higher DCF elimination. There are also other advantages that favor the use of these processes over the conventional suspended-growth biomass processes. – Increase of HRT to N 2–3 days could enhance the elimination of DCF since this would increase the contact time of water with the biomass. – Enrichment of DCF degrading micro-organisms into the bioreactor. This could be done by applying SRTs N150 d which may not be realistic at full-scale wastewater treatment plants. Also, bioaugmentation could be used, i.e. addition of cultured micro-organisms able to degrade DCF into the biological process. However, due to the deconjugation processes and formation of several biotransformation products as well as abiotic nitro-products, degradation pathways need to be further investigated to fully understand the fate of DCF in biological wastewater treatment. Further research is also needed to identify the microbial communities degrading DCF that could be used in bioaugmentation.

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In the future, DCF may be classified as priority substance with EQS values ranging from 10 to 100 ng/l. These values were proposed during the revision of EU's priority substance directive during 2012–2013. Many studies reviewed here reported DCF concentrations of N1000 ng/l in WWTP effluents. Thus, especially in the rivers that are highly influenced by WWTP effluents, DCF concentrations can exceed 100 ng/l. To summarize, low enough concentrations of DCF in the WWTP effluents may not be possible to obtain by optimizing the existing biological processes only. To enhance DCF elimination, we need to further investigate and develop effective and cost-efficient tertiary treatment methods, such as methods based on oxidation or adsorption. Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.envint.2014.03.021.

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Fate of diclofenac in municipal wastewater treatment plant - a review.

Diclofenac (DCF) is a common anti-inflammatory pharmaceutical that is often detected in waste wasters, effluents and surface waters. Recently, DCF was...
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