Fate of antibiotics during municipal water recycling treatment processes

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Available at www.sciencedirect.com

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

Review

Fate of antibiotics during municipal water recycling treatment processes N. Le-Minh a, S.J. Khan a,*, J.E. Drewes a,b, R.M. Stuetz a a b

UNSW Water Research Centre, School of Civil and Environmental Engineering, University of New South Wales, NSW 2054, Australia Advanced Water Technology Center (AQWATEC), Colorado School of Mines, Golden, CO 80401, USA

article info

abstract

Article history:

Municipal water recycling processes are potential human and environmental exposure

Received 5 January 2010

routes for low concentrations of persistent antibiotics. While the implications of such

Received in revised form

exposure scenarios are unknown, concerns have been raised regarding the possibility that

5 May 2010

continuous discharge of antibiotics to the environment may facilitate the development or

Accepted 8 June 2010

proliferation of resistant strains of bacteria. As potable and non-potable water recycling

Available online 15 June 2010

schemes are continuously developed, it is imperative to improve our understanding of the fate of antibiotics during conventional and advanced wastewater treatment processes

Keywords:

leading to high-quality water reclamation. This review collates existing knowledge with

Pharmaceutically active compounds

the aim of providing new insight to the influence of a wide range of treatment processes to

Antibiotics

the ultimate fate of antibiotics during conventional and advanced wastewater treatment.

Wastewater treatment

Although conventional biological wastewater treatment processes are effective for the

Advanced treatment

removal of some antibiotics, many have been reported to occur at 10e1000 ng L
1

Potable reuse

concentrations in secondary treated effluents. These include b-lactams, sulfonamides, trimethoprim, macrolides, fluoroquinolones, and tetracyclines. Tertiary and advanced treatment processes may be required to fully manage environmental and human exposure to these contaminants in water recycling schemes. The effectiveness of a range of processes including tertiary media filtration, ozonation, chlorination, UV irradiation, activated carbon adsorption, and NF/RO filtration has been reviewed and, where possible, semi-quantitative estimations of antibiotics removals have been provided. ª 2010 Elsevier Ltd. All rights reserved.

Contents 1. 2. 3.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical methods for determining antibiotics in wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal of antibiotics during conventional sewage treatment processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. b-Lactams . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Sulfonamides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Trimethoprim . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

* Corresponding author. Tel.: þ61 2 93855082; fax: þ61 2 93138624. E-mail address: [email protected] (S.J. Khan). 0043-1354/$ e see front matter ª 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2010.06.020

4296 4296 4297 4300 4300 4304

4296

4.

5.

1.

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

3.4. Macrolides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5. Fluoroquinolones . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.6. Tetracyclines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.7. Nitroimidazoles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.8. Other antibiotic groups . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.9. Effects of antibiotics on wastewater microbial consortia/processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fate of antibiotics during advanced treatment processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. Membrane filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Adsorptive treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1. Activated carbon . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.2. Ionic adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Chemical and photochemical oxidation processes for the removal of antibiotics . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.1. Chlorination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.2. Ozonation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.3. Ultraviolet irradiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.4. Advanced oxidation processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

Introduction

Municipal water recycling for industrial, agricultural, and non-potable municipal uses is an increasingly important component of water resources management practices in many parts of the world (Exall, 2004; Vigneswaran and Sundaravadivel, 2004; Wintgens et al., 2005). In some countries, such as the USA, Singapore, Mexico and Belgium, treated effluents are intentionally used to supplement drinking water supplies, a process known as planned indirect potable reuse (planned IPR) (Drewes and Khan, 2010; Rodriguez et al., 2009). Planned IPR is rapidly emerging as an important water supply strategy for a number of Australian cities (Khan, 2009). In Windhoek, Namibia, direct potable reuse of highly treated effluents for drinking water supply has been practiced since 1969 (du Pisani, 2006) and it is possible that other countries may adopt this strategy in the future. Pharmaceuticals including antibiotics are present in municipal sewage, largely as a result of human excretion. Many active antibiotics are not completely metabolised during therapeutic use and thus enter sewage through excretion in an unchanged form (Hirsch et al., 1999). The intentional disposal of unused drugs into the sewer (Kummerer, 2003) and veterinary use (Diaz-Cruz et al., 2003) also contribute to the quantities of antibiotics found in sewage. Discharges from veterinary clinics and runoff from agricultural applications into municipal sewers are also potential sources of veterinary antibiotics in wastewater. The reported levels of specific antibiotic drugs detected in raw sewage appear to differ between countries, possibly reflecting variable prescription practices (Miao et al., 2004) and differences in per-capita water consumption leading to various degrees of dilution (Drewes et al., 2008). Seasonal variations in sewage concentrations of antibiotics have also been reported (Alder et al., 2006). Antibiotic drugs have been identified as a particular category of trace chemical contaminants, which warrant close scrutiny (Watkinson et al., 2007). Much of the concern regarding the presence of antibiotics in wastewater and their

4307 4307 4308 4308 4308 4309 4309 4309 4312 4312 4314 4314 4314 4315 4316 4316 4316 4317 4317

persistence through wastewater treatment processes is related to concerns that they may contribute to the prevalence of resistance to antibiotics in bacterial species in wastewater effluents and surface water near wastewater treatment plants (WWTPs) (Adelowo et al., 2008; Auerbach et al., 2007; Baquero et al., 2008; Jury et al., in press). Reusing treated effluents for non-potable or potable purposes increases the range of human and environmental exposure scenarios for bacteria potentially harbouring antibiotics resistance. Accordingly, a thorough understanding of the effectiveness of treatment processes employed in water recycling schemes is warranted.

2. Analytical methods for determining antibiotics in wastewater Many antibiotics are non-volatile with high molecular weight, which tends to render them more suited to analysis by liquid chromatography (LC) rather than gas chromatography (GC) (Choi et al., 2007a). The determination of antibiotic residues by LC with spectrophotometric detection has been reported including fluorescence and ultraviolet (UV) absorbance (Choi et al., 2007a; Esponda et al., 2009; Golet et al., 2002b; Jen et al., 1998; Li et al., 2007; Peng et al., 2008). However, a survey of literature by Herna´ndez et al. (2007) revealed the impressive progress and focus on method development using liquid chromatography-mass spectrometry (LC-MS) and particularly liquid chromatography-tandem mass spectrometry (LC-MS/MS) to determine antibiotics in complex matrices such as municipal wastewater. This review also summarised the important factors affecting the analyses of different classes of antibiotics, such as pH adjustment, sample container materials, storage conditions, the addition of chelating agents, solvent types and matrix interference. Sample extraction for both clean-up and enrichment is commonly undertaken with typical concentration factors in the range 100e1000 required for the necessary low limits of detection (LOD). Although a variety of techniques have been

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

employed for extracting antibiotic, such as lyophilisation (Hirsch et al., 1998) or liquideliquid extraction (LLE) (Jen et al., 1998), the most frequently used technique is solid phase extraction (SPE) with hydrophilic e lipophilic balanced (HLB) cartridges. Fatta et al. (2007) have surveyed the use of several common SPE sorbents and confirmed that Oasis HLB was among the best performing, yielding high recoveries for multiple analyses of acid, neutral and basic analytes including antibiotics. Polar solvents such as acetone, methanol and acetonitrile have been employed for extraction and LC separation. Most recently, solid phase micro extraction (SPME) (Balakrishnan et al., 2006) and SPE coupled online to LC-MS/MS (Stoob et al., 2005) have emerged as alternative approaches for the analysis of antibiotics. Although SPME often yields poorer LODs than the conventional SPE extraction for wastewater samples, the benefits are reduction in solvent usage and significant savings of the analytical time (Fatta et al., 2007). Comparatively, online SPE methods appear to provide improved extraction recovery, reduction in costs of consumable extraction materials and analysis time and similar sensitivity to conventional offline SPE methods (Stoob et al., 2005). Regarding the analysis of antibiotics in solid matrices, a review by Kim and Carlson (2005) summarised different techniques including LLE, accelerated solvent extraction (ASE), ultrasonic solvent extraction (USE) and pressurised liquid extraction (PLE) to desorb antibiotics from wastewater sludge, manure, soil and sediment. During their method development study, Gobel et al. (2005b) compared extraction efficiencies for antibiotics in activated sludge between PLE (50% methanol 50% water) and USE (methanol and acetone). The latter technique was reported to be slightly less efficient for macrolides and trimethoprim, while the extraction efficiency for sulfonamides was significantly lower (by 20e60%) compared to PLE. Analytical methods developed to determine trace concentrations of antibiotics in wastewater and sludge are summarised in Table 1. The sensitivity, recovery and range of antibiotics included in individual methods are highly variable. This variation is largely due to the wide structural variability of these compounds and their inability to respond similarly to the same extraction and analytical procedures, differences in the use of internal standards and analytical instruments, and differences in concentration factors employed in each extraction procedure. However, most methods employing SPE-LCMS/MS have achieved good sensitivity with LODs reaching the nanogram or even sub-nanogram per litre range. Meanwhile, a few methods have been reported to analyse antibiotics in sewage sludge (Gobel et al., 2005b; Golet et al., 2002b). For these methods, LODs between 0.6 and 5.1 ng g
1 were reported for macrolides, sulfonamides and trimethoprim using PLESPE-LC-MS/MS (Gobel et al., 2005b) and 450 ng g
1 for fluorquinolones using ASE-LC-FLD (Golet et al., 2002b).

3. Removal of antibiotics during conventional sewage treatment processes The occurrence of some common antibiotics and performance of conventional WWTPs for removing them as reported in the literature are summarised in Table 2. It is apparent that removal of antibiotics during conventional wastewater

4297

treatment processes is highly variable. However, some activated sludge (AS) processes appear to be effective for the efficient removal of several of these compounds. During treatment, antibiotics can be transformed or removed from the aqueous phase by hydrolysis, biotransformation, or sorption to primary and secondary sludges. Hydrophobic chemicals are expected to occur at higher concentration in primary sludge than hydrophilic chemicals because they have a greater affinity to solids and hence concentrate in the organic-rich sewage sludge (Beausse, 2004). Pharmaceuticals can also be removed from aqueous solution onto solid particulates by ion exchange, complex formation with metal ions, and polar hydrophilic interactions (Diaz-Cruz et al., 2003). Pharmaceuticals, adsorbed to flocs, suspended solids and/or activated (microbial) sludge, will be removed from the aqueous phase by sedimentation and subsequent disposal of excess sludge. The affinity of antibiotics adsorbed to sludge has been represented by sludge sorption constants Kd (L kg
1), shown in Table 3. The greater Kd values represent the greater adsorption of the compounds to sludge. The sludge may be delivered to anaerobic digesters for stabilisation before being used in agricultural soil amendment and/or biogas generation. Those antibiotics, which are hydrophilic and highly resistant to most conventional biological treatment processes, are expected to mainly remain in the aqueous phase of the treated effluent. The operating conditions of a wastewater treatment process such as temperature, solids retention time (SRT), and hydraulic retention time (HRT), can significantly affect the removal efficiency of many pharmaceutical contaminants. While ambient temperature is not practical to control, SRT and HRT can be adjusted to some degree in order to optimise removal efficiencies. Increasing SRTs were reported to enhance the removal of several pharmaceutical compounds during aerobic biological processes (Clara et al., 2005a; Kim et al., 2005; Yasojima et al., 2006). The extended SRTs have been suggested to allow for the enrichment of slower growing bacterial species and therefore, to provide greater diversity of enzymes, some of which are capable of breaking down the pharmaceutical compounds (Jones et al., 2007). Similarly, longer HRT is expected to provide sufficient reaction time for biotransformation and sludge sorption to occur in order to reach maximum efficiency at equilibrium. Consistent with this expectation, a reduction in the removal of some pharmaceuticals during wastewater treatment processes has been observed to be generally correlated with reducing HRT (TauxeWuersch et al., 2005; Vieno et al., 2007a). These observations were made during rainfall events, resulting in increased dilution and reduced HRT. Kim et al. (2005) pointed out that during dry weather operation, shorter HRT (implying a greater substrate loading) will result in higher biomass concentration which could enhance the removal, thus compensating for shorter reaction times. Accordingly, it is difficult to generalise the expected relationship between HRT and pharmaceutical removal without accounting for the effect on biomass concentration. A number of published studies have reported SRT and HRT in conjunction with observed removal of antibiotics during membrane bioreactor (MBR) and activated sludge (AS) treatment processes as summarised in Table 4. The studies were selected to include only those based on 24 h composite sampling regimes in order to consider the specific HRT, thus

Target compounds

Sample matrices

Extraction, (sample size), pH

Solvents

Internal standards

Instruments

Recoveries (relative)

LOD (ng/L) or (ng/g)

Reference

Wastewater

SPE Oasis HLB (200 mL), pH 7.5

Methanol Acetonitrile Formic trifluoroacetic acid

Penicillin G

LC-MS/MS (þve ESI) LCQ Duo Ion Trap

>70% Amoxicillin 10%

13e18 (Inf); 8e15 (Eff)

(Cha et al., 2006)

16 Sulfonamides Trimethoprim

Wastewater

SPE Oasis HLB (250 - 500 mL) Silica gel

Methanol Formic acid

13C6 - sulfamethoazine

LC-MS/MS (þve ESI) Water QqQ

62e102%

0.02e0.2 (Inf); 0.016e0.12 (Eff)

(Chang et al., 2008)

6 Sulfnonamides Trimethoprim 5 Macrolides

Wastewater

SPE Oasis HLB (50 - 250 mL), pH 4

Methanol Formic acid Ethyl acetate

Isotope label sulfonamides 13C2-erythromycin

LC-MS/MS (þve ESI) TSQ Quantum QqQ

91e108% 30e47% 78e124%

11e68 (Inf); 1.2e9.6 (Eff) 4.5e8.1 (Inf); 0.9e2.7 (Eff) 0.36e3.9 (Inf); 0.09e2.9 (Eff)

(Gobel et al., 2004)

2 Macrolides Sulfamethoxazole Trimethoprim Ofloxacin

Wastewater

SPE Oasis HLB (100 - 200 mL)

Methanol/ acetonitrile Ammonium acetate

13C-phenacetin Carbamazepine-d10

LC-MS/MS (þve ESI) Water QqQ

40e116% 50e80% 88e111% 95e106%

6e7 (Inf); 3e6 (Eff) 42 (Inf); 20 (Eff) 25 (Inf); 10 (Eff) 43 (Inf); 43 (Eff)

(Gros et al., 2006)

Sulfnonamides Trimethoprim Macrolides

Sewage sludge

PLE, 0.2 g, pH 4 SPE Oasis HLB

Methanol Formic acid Ethyl acetate

Isotope label for sulfonamides 13C2-erythromycin for macrolides

LC-MS/MS (þve ESI) TSQ Quantum QqQ

79e106% 78% 91e142%

0.9e15 2.7e5.1 0.6e2.4

(Gobel et al., 2005b)

7 Sulfonamides

Swine wastewater

LLE, (50 mL), pH 6.6

Nicotinamide Ethyl acetate

N/A

LC-UV

86e99%

4000e15 000

(Jen et al., 1998)

2 Fluoroquinolones

Sewage sludge

ASE

Acetonitrile

LC-FLD

82e94%

450

(Golet et al., 2002b)

5 Sulfonamides

Wastewater

SPE mixed hemimicelles column, pH 2

Methanol Acetinitrile

N/A

LC-UV

89e113%

150e350

(Li et al., 2007)

5 Fluoroquinolones 3 Sulfonamides Trimethoprim

Wastewater

SPE Oasis HLB, (1000 mL), pH 2.5

Methanol Acetonitrile

Sulfamerazine Standard addition

LC-MS

90e129% 37e65% 98e109%

20e40 (Eff) 40e90 (Eff) 40e50 (Eff)

(Renew and Huang, 2004)

5 Macrolides 2 Ionophores Tiamulin

Liquid manure

LLE, (15 g), pH 8

Ethyl acetate Acetonitrile Ammonium acetate

(E)-9-[O-(2-methyloxime)]erythromycin

LC-MS/MS (þve ESI) TSQ7000 QqQ

78e94% 119% 123%

0.4e27.9 3.2e17.9 0.4

(Schlu¨sener et al., 2003)

Gentamicin

Hospital wastewater

SPE Widepores CBX, (20e50 mL), pH 7e8

Methanol Kanamycin Acetic acid Heptafluorobutyric acid

LC-MS/MS (þve ESI) API 365

107e111%

200

(Lo¨ffler and Ternes, 2003)

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

5 Beta lactams

4298

Table 1 e Summary of analytical methods for determination of antibiotics in wastewater and sludge.

Wastewater

SPE Anpel MEP, (100e500 mL), pH 3

Methanol Formic acid Acetonitrile Tetrabutyl ammonium bromide

N/A

LC-FLD

10 Sulfonamides

Wastewater

SPME Supelco CW/TRP (25 mL)

Methanol Formic acid Ammonium acetate

13C6sulfamethazine

LC-MS/MS (þve ESI) Quattro Ultima QqQ

20 Quinolones and Fluoroquinolones

Wastewater

SPE Oasis HLB (200e400 mL), pH 3 2nd SPE Water WCX

Methanol Formic acid

Norfloxacin-d5

LC-MS/MS (þve ESI) Waters Premier XE

3 Macrolides

Wastewater

SPE Oasis HLB (120 mL), pH 5

Methanol Acetonitrile Formic acid

Simatone

6 Tetracyclines 5 Sulfonamides

Wastewater

SPE Oasis HLB (120 mL), pH 3

Methanol Acetonitrile Formic acid

7 Macrolides 2 Fluoroquinolones Sulfamethoxazole Oxytetracycline Amoxycillin

Wastewater

SPE Oasis MCX (500 mL), pH 2 SPE Lichrolut EN (500 mL), pH 7

2 Fluoroquinolones 2 Macrolides Sulfamethoxazole Trimethoprim

Wastewater

5 Sulfonamides Trimethoprim

7 Beta lactams 3 Sulfonamides

(Shi et al., 2009)

9040e55300 (Inf)

(Balakrishnan et al., 2006)

1.6e50 (Inf); 0.6e50 (Eff)

(Xiao et al., 2008)

LC-MS/MS (þve ESI) 83e86% Finnigan LCQ Ion Trap

30e70 (Eff)

(Yang and Carlson, 2004)

Simatone

LC-MS/MS (þve ESI) 78e95% Finnigan LCQ Ion Trap 91e104%

40e70 (Inf); 30e50 (Eff) 40e60 (Inf); 30e40 (Eff)

(Yang et al., 2005)

Methanol Ethyl acetate Acetone Acetonitrile

Salbutamol-d3 Ibuprofen-d3

LC-MS/MS (þve ESI) API3000 QqQ

47e76% 31e32% 65% 73% 36%

0.2e1.4 (Eff) 1.3e1.8 (Eff) 1.5 (Eff) 1.2 (Eff) 2.1 (Eff)

(Castiglioni et al., 2005)

SPE Strata-X þ XC (250 mL), pH 3

Methanol Formic acid Acetonitrile

Lomefloxacin Josamycin Diaveridine

LC-MS/MS (þve ESI) TSQ Quantum QqQ

76e97% 92e100% 68% 104%

4e21 (Inf) 0.3e12 (Inf) 22 (Inf) 7 (Inf)

(Segura et al., 2007)

Wastewater

SPE Oasis HLB (50 mL), pH 4

Methanol Acetonitrile Formic acid

Sulfathiazole-d4 Sulfamethoxazole-d4

LC-MS-MS (þve ESI) TSQ Quantum QqQ

72e110% 80e103%

7e10 (Eff) 7 (Eff)

(Botitsi et al., 2007)

Wastewater

SPE Oasis HLB (250 mL), Methanol pH 3

Caffeine

UPLC-MS/MS (þve ESI) 56e93%

4.1e84 (Inf); 3.8e60 (Eff) 3.0e3.3 (Inf); 1.0 (Eff) 4.6e7.0 (Inf); 2.8e5.0 (Eff) 6.8e14 (Inf); 5.1e8.1 (Eff) 0.5e37 (inf); 0.3e26 (Eff) 2.7 (Inf); 1.1 (Eff)

(Li et al., 2009)

64e127% piromidic acid <29%

80e104%

3 Fluoroquinolones

86e105%

3 Tetracyclines

83e96%

3 Macrolides

73e93%

Trimethoprim

90e97%

Inf: influent; Eff: Effluent; SPE: Solid phase extraction; LLE: Liquideliquid extraction; ASE: Accelerated solvent extraction, PLE: Pressurised liquid extraction; SPME: Solid phase micro extraction; FLD: Fluorescence detector; UPLC: Ultra pressurised liquid chromatography; þve ESI: Positive electro-spray ionisation.

4299

100e1060

Acquity TQ QqQ

79e109%

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

8 Fluoroquinolones

4300

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

accounting for influent variability. It can be seen from Table 4, while some relationship between SRT and antibiotic removal has been reported within some well controlled studies (Clara et al., 2005b; Gobel et al., 2007; Kim et al., 2005), no clear trends are evident between these studies. This suggests that other unreported factors, such as biomass concentration and diversity as well as substrate/biomass ratios, might be important to explain the variation. MBR systems have been reported to be equal to or more effective in removing several antibiotics compared to AS systems of similar SRT and HRT, possibly due to higher biomass concentration in the MBRs. Higher biomass concentration implies the reduced sludge loading (i.e., ratio between substrate and sludge concentration), which was reported to enhance biotransformation of several antimicrobial compounds (Gobel et al., 2007). Unfortunately, it is not possible to interpret this data in further detail by consideration of treatment design and operational parameters since there are only few studies which comprehensively presented operating parameters of the MBR processes. One very good example, which confirms the importance of these parameters and illustrates how this data could be provided, is given by Gobel et al. (2007).

3.1.

b-Lactams

b-Lactams, such as penicillins and cephalosporins, are narrow spectrum antibiotics, which are highly effective against the Gram-positive genera Streptococcus, Gonococcus, and Staphylococcus (Todar, 2002). These antibiotics act as bacteriostatics by inhibiting bacterial peptidoglycan cell wall synthesis (Marzo and Dal Bo, 1998). The four-membered ring, which all b-lactam drugs feature, is a strained, cyclic amide that is highly susceptible to chemical or enzymatic hydrolysis (Deshpande et al., 2004). The hydrolysed b-lactam drugs result in an inactive product when the ring is broken. The degradation of b-lactam antibiotics such as penicillin, takes place under acidic and alkaline conditions or by reactions with weak nucleophiles, such as water or metal ions (Aksu and Tunc, 2005; Hou and Poole, 1971). Alternatively, penicillin can be enzymatically hydrolysed by b-lactamase enzyme via the same way as acid hydrolysis. b-lactamases are the widespread enzymes in bacteria, and are produced by many species to inactivate the pharmacological effects of the beta-lactam antibiotics (Neu, 1992). Historically, in order to make ‘orally available’ penicillins, some chemical modifications were developed to reduce the susceptibility of penicillins to acid hydrolysis in stomach, which resulted in the availability of acid-resistant penicillins including amoxicillin and ampicillin (Sneader, 2005). However, since these drugs are still susceptible to enzymatic hydrolysis, they do not appear to persist during sewage treatment (Table 2). The study by Li et al. (2008) focusing on Penicillin G reported concentration of 153 mg L
1 in raw sewage and 1.68 mg L
1 in treated effluent, respectively, revealing that Penicillin G had undergone partial transformation during the anaerobic, aerobic and hydrolysis processes at the WWTP. A number of studies have reported that the presence of blactams in treated wastewater samples are generally not detected or only at very low concentrations, despite them being among the most commonly prescribed antibiotics

(Cahill et al., 2004; Cha et al., 2006; Costanzo et al., 2005; Hirsch et al., 1999; Watkinson et al., 2007; Zuccato et al., 2005). Although b-lactam antibiotics have been reported to dominate the overall antibiotic concentration in some sewage influents, they tend to be significantly reduced in concentrations during biological processes (Watkinson et al., 2007). The analysis of five b-lactams in WWTP influent revealed that while chloxacillin and oxacillin were observed in three out of 72 influent samples (at concentration less than 20 ng L
1), none of these blactams were detected in effluent samples (Cha et al., 2006). The significant removal (>96%) of cephalexin from 2000 ng L
1 to 78 ng L
1 has been reported through conventional WWTP processes in Australia (Costanzo et al., 2005). Similarly, Morse and Jackson (2004) concluded that amoxicillin, a representative b-lactam drug, is quite susceptible to microbial degradation and therefore is not likely to remain in significant concentration after biological treatment systems.

3.2.

Sulfonamides

Sulfonamides and trimethoprim are bacteriostatic agents that synergistically target and inhibit two pathway steps in bacterial folic acid synthesis (Masters et al., 2003; Skold, 2001). Folate derivatives are essential cofactors in the biosynthesis of purines, pyrimidines and bacterial DNA in all living cells. Therefore, blocking this pathway inhibits the production of reduced folates and eventually the synthesis of nucleic acid, which in turn affects bacterial growth. When combined, sulfonamides and trimethoprim afford an effective treatment against a variety of potential bacterial infections. Sulfonamides are not completely metabolised during use and are excreted via urine into sewage, partly as unchanged parent compounds and partly as metabolites (Gobel et al., 2005a; Hirsch et al., 1999). The major metabolites of sulfonamides entering sewage are biologically inactive N4-acetylated products, for which transformations back to the active parent compounds during sewage treatment has been reported (Gobel et al., 2005a). This phenomenon may have led to apparent negative removal of some sulfonamides, particularly sulfamethoxazole, during biological wastewater treatment (Gobel et al., 2007; Karthikeyan and Meyer, 2006). Sulfamethoxazole is among the most frequently detected sulfonamides in municipal sewage (Brown et al., 2006; Choi et al., 2007a; Gobel et al., 2007; Levine et al., 2006; Yang et al., 2005). However, concentrations of this drug in WWTP influents and effluents vary significantly, depending on antibiotic consumption patterns and the types of wastewater treatment processes employed. For example, sulfamethoxazole was reported with concentrations as high as 7.91 mg L
1 in sewage influent in China, where the compound is one of the top 15 pharmaceuticals sold (Peng et al., 2006). Sulfamethoxazole removal efficiencies by conventional WWTPs have been reported to range from
279% to 100% (Table 2). Sulfamethoxazole’s acetylated metabolite, N4acetylsulfamethoxazole usually accounts for greater than 50% of an administered dose in human excretion (Gobel et al., 2004) and can occur in WWTP influents at concentrations 2.5e3.5 times higher than concentrations of the parent compound (Gobel et al., 2007). Despite occurring at high concentration in the raw influent, N4-acetylsulfamethoxazole does not appear to partition significantly into sludge (Gobel et al., 2005a).

Table 2 e Occurrence of common antibiotics in WWTPs. Analytes/Location

Main treatment practices

Removal Efficiency (%)

Reference

Cephalexin Australia Australia China Taiwan

2000 5600 670e2900 1563e4367

80 <(2) 240e1800 10e994

Activated sludge system (AS) AS Chemical enhanced/Secondary treatment Secondary treatments/UV or chlorination

96a e 9 to 89b 36 to 99.8b

(Costanzo et al., 2005) (Watkinson et al., 2007) (Gulkowska et al., 2008) (Lin et al., 2009)

Amoxicillin Australia

280

<(3) 30

AS

e

(Watkinson et al., 2007)

Cloxacillin Australia USA

<(1)e320 <(13)e15

<(1) <(9)

AS Secondary treatment/Chlorination

e e

(Watkinson et al., 2007) (Cha et al., 2006)

Penicillin G Australia

<(2)

<(2)

AS

e

(Watkinson et al., 2007)

Penicillin V Australia

160

80

AS

e

(Watkinson et al., 2007)

Sulfamethoxazole USA Korea China Croatia Switzerland Mexico USA Sweden Spain Spain Sweden Austria China Taiwan

1090 450 5450e7910 590 230e570 390 <(50)e1250 20 580 0.60 <(80)e674 24e145 10e118 179e1760

210 <(30) <100 0.39 210e860 0.31 <(50)e210 70 250 NA <(80)e304 18e91 9e78 47e964

AS/chlorination AS AS/filtration/chlorination AS AS/Sand filtration AS AS AS/Chemical P removal AS AS Chemical P removal/AS AS AS or Chemical enhanced/UV or chlorination Secondary treatments/UV or chlorination

81b >93a >98a 33a e 20b 18e100b e 67b 57b 42b
279 to 66b 34e63b 26e88b

(Yang et al., 2005) (Choi et al., 2007a) (Peng et al., 2006) (Gros et al., 2006) (Gobel et al., 2005a) (Brown et al., 2006) (Karthikeyan and Meyer, 2006) (Bendz et al., 2005) (Carballa et al., 2004) (Carballa et al., 2005) (Lindberg et al., 2005) (Clara et al., 2005b) (Xu et al., 2007) (Lin et al., 2009)

N4-sulfamethoxazole Switzerland

850e1600

<(20)e180

AS/Sand filtration

e

(Gobel et al., 2005a)

Sulfathiazole Korea

10 570

180

AS

98a

(Choi et al., 2007a)

Sulfamethazine USA Korea

150 4010

<(30) <(30)

AS/chlorination AS

>80a >99a

(Yang et al., 2005) (Choi et al., 2007a)

Sulfadimethoxine USA Korea

70 460

<(30) <(30)

AS/chlorination AS

> 57a >93a

(Yang et al., 2005) (Choi et al., 2007a) (continued on next page)

4301

Effluent conc. (ng/l)

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Influent conc. (ng/l)

Analytes/Location

Influent conc. (ng/l)

Effluent conc. (ng/l)

Main treatment practices

Removal Efficiency (%)

Reference

5100e5150 <(1)e72

<(150) <(1)e36

AS/filtration/chlorination AS or Chemical enhanced/UV or chlorination

> 97a 50b

(Peng et al., 2006) (Xu et al., 2007)

Sulfamerazine Korea

1530

< (30)

AS

> 98a

(Choi et al., 2007a)

Trimethoprim Croatia Switzerland Mexico USA Sweden UK Sweden China Taiwan

1172 210e440 0.59 0.14e1.10 80 213e300 99e1300 120e320 259e949

290 20e310 180 <(50)e550 40 218e322 66e1340 120e230 203e415

AS AS AS AS AS/chemical P removal Trickling filter/AS/UV Chemical P removal/AS Chemical enhanced/Secondary treatment Secondary treatments/UV or chlorination

75a 64b 70b 50 to 100b 49b
3b 3b
17 to 62b 22e56b

(Gros et al., 2006) (Gobel et al., 2005a) (Brown et al., 2006) (Karthikeyan and Meyer, 2006) (Bendz et al., 2005) (Roberts and Thomas, 2006) (Lindberg et al., 2005) (Gulkowska et al., 2008) (Lin et al., 2009)

Doxycycline USA Korea Sweden

210 220 <(64)e2480

70 30 <(64)e915

AS/chlorination AS Chemical P removal/AS

67b 86a 70a

(Yang et al., 2005) (Choi et al., 2007a) (Lindberg et al., 2005)

Tetracycline USA Korea USA China Taiwan

200 110 240e790 96e1300 46e234

< (30) <(0.03) <(50)e160 180e620 16e38

AS/chlorination AS AS Chemical enhanced/Secondary treatment Secondary treatments/UV or chlorination

>85a >73a 68 to 100b
88 to 73b 66e90b

(Yang et al., 2005) (Choi et al., 2007a) (Karthikeyan and Meyer, 2006) (Gulkowska et al., 2008) (Lin et al., 2009)

Chlortetracycline USA Korea

270 970

60 40

AS/chlorination AS

78a 96a

(Yang et al., 2005) (Choi et al., 2007a)

Oxytetracycline Korea

240

<(30)

AS

>88a

(Choi et al., 2007a)

Ciprofloxacin USA Sweden Australia Switzerland Sweden China

<(50)e310 90e300 90 320e570 320 80

<(50)e60 <(6)e60 130 60e90 31.5 27

AS Chemical P removal/AS AS AS/Fe flocculation AS/Chemical P removal/sand filtration Secondary treatment

22e100b 87b
44a 83a 90b 66a

(Karthikeyan and Meyer, 2006) (Lindberg et al., 2005) (Costanzo et al., 2005) (Golet et al., 2003) (Zorita et al., 2009) (Xiao et al., 2008)

Norfloxacin Switzerland Sweden Sweden

340e520 66e174 18

40e60 <(7)e37 <(5.5)

Act. Sludge/Fe flocculation Chemical P removal/AS AS/Chemical P removal/sand filtration

88a 87b >70a

(Golet et al., 2003) (Lindberg et al., 2005) (Zorita et al., 2009)

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Sulfadiazine China China

4302

Table 2 (continued).

110e460 54e263 339

85e320 27e85 85

Chemical enhanced/Secondary treatment AS or Chemical enhanced/UV or chlorination Secondary treatment


20 to 78b 50 to 82b 75a

(Gulkowska et al., 2008) (Xu et al., 2007) (Xiao et al., 2008)

<(43) 3520e5560 470 <(6)e287 22.5 80e368 115e1274 1208

<(43) <(80)e740 110 <(6)e45 10 41e165 53e991 503

AS AS/filtration/chlorination AS Chemical P removal/AS AS/Chemical P removal/sand filtration AS or Chemical enhanced/UV or chlorination Secondary treatment/UV or chlorination Secondary treatment

e >85b 77b 86b 56b 40 to70b <88b 58a

(Gros et al., 2006) (Peng et al., 2006) (Brown et al., 2006) (Lindberg et al., 2005) (Zorita et al., 2009) (Xu et al., 2007) (Lin et al., 2009) (Xiao et al., 2008)

Erythromycin Croatia Switzerland USA UK China China Taiwan

<(20) 60e190 <(50)e1200 71e141 470e810 253e1978 226e1537

<(20) 60e110 <(50)e300 145e290 520e850 216e2054 361e811

AS AS AS or aerated lagoon Trickling filter/Act. sludge/UV Chemical enhanced/Secondary treatment AS or Chemical enhanced/UV or chlorination Secondary treatments/UV or chlorination

e e 44 to 100b
79b
12 to 19b 15e45b <56b

(Gros et al., 2006) (Gobel et al., 2005a) (Karthikeyan and Meyer, 2006) (Roberts and Thomas, 2006) (Gulkowska et al., 2008) (Xu et al., 2007) (Lin et al., 2009)

Roxithromycin Switzerland USA Austria China

10e40 1500 25e117 75e164

10e30 870 36e69 35e278

AS/sand filtration AS or aerated lagoon AS AS or Chemical enhanced/UV or chlorination

e 42a
80 to 44b 53 to 76b

(Gobel et al., 2005a) (Karthikeyan and Meyer, 2006) (Clara et al., 2005b) (Xu et al., 2007)

Clarithromycin Switzerland Japan Taiwan

330e600 492e883 59e1433

110e350 266e444 12e232

AS/sand filtration AS Secondary treatments/UV or chlorination

21b 43b <0 to 99b

(Gobel et al., 2005a) (Yasojima et al., 2006) (Lin et al., 2009)

Azithromycin Japan

199e371

88e219

AS

49b

(Yasojima et al., 2006)

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

China China China Ofloxacin Croatia China Mexico Sweden Sweden China Taiwan China

Value in the parenthesis is the limit of detection in each study. a Removal efficiencies, not reported by authors in the cited study, are calculated from the average influent and effluent concentrations which were stated in the study. b Removal efficiency, either reported by authors in the cited study or calculated from the results of each sampling case.

4303

4304

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Table 3 e Sorption constants of antibiotics to wastewater sludge.

Sulfapyridine Sulfamethoxazole

Trimethoprim Roxythromycin Azithromycin Clarithromycin Norfloxacin Ciprofloxacin Tetracycline

Log Kd

Condition

Reference

2.3e2.6 0.8e1.8 2.1e2.6 2.2e2.7 2.2e2.6 1.1e1.9 2.2e2.7 2.5e2.7 2.5e2.6 4.2 4.3 3.9

Activated Sludge Digested Sludge Activated Sludge Activated Sludge Activated Sludge Digested Sludge Activated Sludge Activated Sludge Activated Sludge Activated Sludge Activated Sludge Activated Sludge

(Gobel et al., 2005b) (Carballa et al., 2008) (Gobel et al., 2005b) (Joss et al., 2005) (Gobel et al., 2005b) (Carballa et al., 2008) (Joss et al., 2005) (Gobel et al., 2005b) (Gobel et al., 2005b) (Golet et al., 2003) (Golet et al., 2003) (Kim et al., 2005)

From the literature data, sulfonamides appear to be only partially removed by conventional WWTPs (Table 2). This partial removal may be attributed to moderate sorption to sludge and limited biodegradability. As a general rule, chemicals with logD < 2.5 are considered to have low hydrophobic sorption potential (Drewes, 2007). Therefore, sulfonamides are expected to be quite soluble and have a low potential for hydrophobic partitioning based on LogDpH6e8 between
1.8 and 1.3 (Table 5). Nonetheless, it has recently been demonstrated that the removal of sulfamethoxazole during MBR treatment is highly pH-dependant between pH 5e9 (Tadkaew et al., 2010). Since sulfonamides are neutral or negatively charged under typical WWTP operating conditions (pH 7e8) (Quiang and Adams, 2004), their binding to biomass via cation exchange with anionic sites or by metal complexation are also likely to be minimal. Consistent with this, Peng et al. (2006) observed insignificant partitioning to particulates during primary wastewater treatment processes. However, for those sulfonamides that are partially adsorbed to sludge, anaerobic sludge digestion has been shown to provide effective biotransformation (Carballa et al., 2006). In a laboratory-scale study by Sponza and Demirden (2007), a sequential upflow anaerobic sludge blanket system was reported to have a good removal performance for sulfamerazine (above 97%). Ingerslev and Halling-Sørensen (2000) have reported that while sulfonamides are ultimately biodegradable in AS systems, they exhibit a lag phase of 6e12 days at 20  C and 34e47 days at 6  C. This lag phase is the period after spiking, during which there was no significant degradation of the sulfonamides added to the AS system. However, by serial spiking tests, they observed that once bacteria in the reactors adapted to an initial introduction of four different sulfonamides, the bacterial culture rapidly degraded either a subsequent spike of the same four compounds or four other previously unexposed sulfonamides, illustrated by shorter degradation half-lives and no lag-phase, respectively. This implies that once the bacteria have adapted to degrade one sulfonamide, they may also be capable of efficiently degrading others. Some studies have observed that conventional WWTPs were effective in removing sulfamethoxazole (Choi et al., 2007a), while others appear to contradict this (Brown et al.,

2006). The contradiction may possibly be explained by the differences in WWTP operating conditions, such as SRT, HRT, and temperature. Moreover, differences in reported removal efficiencies may, in some cases, be attributed to limitations of employed mass balance techniques. For example, short-term variations of pharmaceutical loads in influent can be significant (Gobel et al., 2005a; Khan and Ongerth, 2005), thus care must be taken when comparing influent and effluent concentrations. Collecting composite samples over a period that is longer than the hydraulic retention time may improve the comparability between influent and effluent samples (Roberts and Thomas, 2006). A further complication is that sulfonamides are easily able to undergo reversible intertransformation with their respective metabolites (Gobel et al., 2007), such that if these metabolites are not considered, removal efficiencies may be underestimated or overestimated.

3.3.

Trimethoprim

Trimethoprim has been reported to occur in raw sewage of a number of countries including the USA (Karthikeyan and Meyer, 2006), Croatia (Gros et al., 2006), and Mexico (Brown et al., 2006). The presence of trimethoprim can generally be correlated to that of sulfamethoxazole since the two drugs are often administered in combination at a ratio 1:5 (Gobel et al., 2005a). Perez et al. (2005) reported that the concentration of trimethoprim in the primary effluent of a WWTP was around four times lower than that of sulfamethoxazole, which is relatively consistent with the typical medication ratio. Biodegradation experiments undertaken by HallingSorensen et al. (2000) showed the strong persistence of trimethoprim in AS batch reactors. The removal of trimethoprim during conventional biological wastewater treatment has been reported to significantly vary but is often incomplete (Brown et al., 2006; Gobel et al., 2007; Gros et al., 2006; Levine et al., 2006; Paxeus, 2004). Sorption to biomass appears to be negligible (Gobel et al., 2005a; Lindberg et al., 2005), as reflected by the low hydrophobic partitioning coefficient (LogDpH6e8) of the compound (Table 5). Only minor removal of trimethoprim during primary and secondary treatment has been reported (Gobel et al., 2005a; Perez et al., 2005). However, improved removal has been achieved by subsequent biologically active

Table 4 e Removal efficiency (%) of antibiotics by secondary treatment processes from studies reporting SRT and HRT, based on 24 h composite samples. Study

(Gobel et al., 2007)

(Golet et al., 2003) (Kobayashi et al., 2006) (Vieno et al., 2007a) (Yasojima et al., 2006)

SRT (d)

HRT (h)

NOR

CIP

AS AS AS (N) Lab MBR (S) AS MBR AS (N/DN) MBR MBR MBR AS (N/DN) AS (N/DN) AS AS/MF, UV AS (N/DN), Fe AS AS AS AS AS AS

5.6e8.2 10e12 7e16 72 2 10 10e12 16 33 60e80 21e25 20 10 12 10 9 7 7 7 8 5

15e22 9 12e20 12 2 12 5 13 13 13 31 12 7 9 9.6 12 10 7 7 6 4

50 75 80 to 87

66 79 to 86

ROX

AZI

CLA

ERY

SPY

SMX

53

15

<0

77 27 >62 <0 to 38 39 62 59 5 to 38

91

53 <0 61 <0 to 9 37 38 37 <0 to 60

<0 5 24 22 to 55

<0 to 9 57 41 88 4 to 20

<0 to 6 34 26 87 <0

<0 60 50 58 49 to 72

SMX þ AceSMX

TRI

36

<0 to 50 87 74 68 61 to 76

<0 to 14 30 34 87 <0 to 20

92 <0 >89

11 27

89 <0

15

21 20 61 29 26 41

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

(Xu et al., 2007) (Xiao et al., 2008) (Golet et al., 2002a) (Reif et al., 2008) (Clara et al., 2005b)

Main Treatment

AS: Activated sludge, N: nitrification; DN: denitrification, FBR: Fixed bed reactor, S: synthetic feed; Fe: Ferric chloride addition, MF: membrane filtration, MBR: membrane bioreactor, SRT: sludge retention time, HRT: hydraulic retention time, NOR: norfloxacin, CIP: ciprofloxacin, ROX: roxithromycin, AZI: azithromycin, CLA: clarithromycin, ERY: erythromycin, SPY: sulfapyridine, SMX: sulfamethoxazole, AceSMX: N4-aceytl sulfamethoxazole, TRI: trimethoprim. Removal efficiency <0 (less than zero) indicate that concentration of the effluent is greater than that of influent.

4305

4306

Table 5 e Acid dissociation constants and partitioning coefficients of several antibiotics. Antibiotic Category Beta-lactam

Sulfonamides

Functional groups corresponding to

Antibiotics

pKa1

pKa2

pKa4

Log Kow [5]

LogD pH6e8 [5]
1.9 to
2.8

6.88 [7]

0.65


1.9 to
2.8

5.86 6.77 7.42 5.57 7.07 6.28 8.40

1.5 0.34 0.80 0.89 0.047
0.12 0.034

1.3 to
0.10 0.30 to
0.68 0.79 to 0.16 0.49 to
0.90 0.02 to
0.76
0.23 to
1.5
0.08 to
0.16

2.829 na 3.159 3.33

0.72 to 2.4 na 1.1 to 2.8
0.65 to 2.3


1.501
0.325
1.470
0.540


4.0
2.9
4.0
3.1

1.313 1.478


1.1 to
0.95
0.90 to
0.78


0.013


0.01

0.791


0.42 to 0.73

Amoxicillin

2.4

7.4

Cephalexin

2.56

pKa1: basic amine group (-NH2)

Sulfadimethoxine Sulfamerazine Sulfamethazine Sulfamethoxazole Sulfathiazole Sulfadiazine Sulfapyridine

1.87 2.17 2.28 1.83 2.08 2.10 na

9.6 [6]

[1] [1] [1] [1] [1] [1] [2]

Macrolides

pKa1: basic dimethylamino group [-N(CH3)2]

Erythromycin Roxithromycin Clarithromycin Azithromycin

8.90 [3] 9.17 [3] 9.0 [8] 8.59 [5]

Tetracyclines

pKa1: acidic tricarbonyl group pKa2/pKa3: basic dymethylamino group acidic b-dikentone group (simultaneous dissociation)

Oxytetracycline Chlortetracycline Tetracycline Doxycycline

3.30 3.30 3.30 3.50

7.30 7.40 7.70 7.70

9.10 [3] 9.30 [3] 9.70 [3] 9.50 [3]

Fluoroquinolones

pKa1: acidic carboxylic group connected with ring 1 pKa2/pKa3/pKa4: are assigned in order to three basic nitrogen sites starting from ring 1 (nalidixic acid group) to ring 3 (fluoro group)

Ciprofloxacin Norfloxacin

3.01 3.11

6.14 6.10

8.70 8.60

Nitroimidazoles

pKa1 basic imidazole (heterocyclic aromatic ring with 3C and 2 N)

Metronidazole

2.50 [4]

Other antibiotics

pKa1 and pKa2 are assigned to basic N3 and N1 site in the pyrimidine ring

Trimethoprim

3.23

6.76 [3]

10.58 [3] 10.56 [3]

to
4.9 to
3.8 to
4.7 to
3.7

[1] (Maria and Reginald, 1993) [2] (Huber et al., 2005) [3] (Quiang and Adams, 2004) [4] (Loke et al., 2000). [5] (American Chemical Society, 2009) [6] (Delgado and Remers, 1991) [7] (Dutta et al., 1999) [8] (Gustavson et al., 1995).

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

0.61

pKa1: acidic carboxylic group pKa2: basic amine group (-NH2) pKa3: acidic hydroxide group (-OH) pKa1: acid carboxylic group pKa2: basic amine group (-NH2)

pKa2: acidic amide group (-NH-)

pKa3

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

media filtration (Gobel et al., 2005a). Previous studies have indicated that nitrification microorganisms appear to be capable of degrading trimethoprim (Batt et al., 2006; Perez et al., 2005). This suggests an important role for aerobic conditions for the biotransformation of trimethoprim. Consistent with this, removal efficiency of trimethoprim appears to be enhanced by long SRT during biological treatment, which is conducive to nitrification (Batt et al., 2007).

3.4.

Macrolides

Macrolide antibiotics, such as erythromycin, are active against most Gram-positive bacteria by binding reversibly to 50 S ribosomal subunits and inhibiting protein synthesis in microorganisms (Marzo and Dal Bo, 1998; Todar, 2002). After administration, macrolides are largely excreted into sewage in their unchanged forms at excretion rates greater than 60% (Hirsch et al., 1999). This implies that sewage entering WWTPs may contain high concentration of macrolides, particularly in countries in which macrolides are highly prescribed antibiotics. Gobel et al. (2005a,b) reported the concentration of macrolides in raw sewage from Switzerland to vary between 0.01 and 0.6 mg L
1, while Karthikeyan and Meyer (2006) found that WWTP influent in the USA can contain macrolides at concentrations as high as 1.5 mg L
1. Erythromycin is among the principal representatives of the macrolide antibiotics for clinical use (Kirst, 2002). An important difference between erythromycin and other macrolides, such as clarithromycin and roxythromycin, is the sensitivity of erythromycin to pH. Under acidic conditions, erythromycin is unstable and is transformed into an inactive anhydro-form by the loss of one H2O molecule (Gobel et al., 2004). At the ambient operational pH ranges (6.5e8) of most municipal WWTPs, erythromycin can exist in both its active original form and as the inactive erythromycin-H2O. Reported regional differences in sewage concentrations of macrolides may be a reflection of variable prescription and consumption patterns (Gobel et al., 2005a; Miao et al., 2004). In Switzerland, clarithromycin is more often detected in WWTPs effluents at higher concentrations than erythromycin-H2O and roxythromycin, which is well correlated to consumption data (Gobel et al., 2007). Conversely, in Canada erythromycinH2O is more frequently detected (and prescribed), followed by clarithromycin and roxythromycin (Miao et al., 2004). Macrolide antibiotics are often incompletely removed by conventional WWTPs (McArdell et al., 2003). Erythromycin (including erythromycin-H2O), has been removed between 43% and >99% by secondary wastewater treatment processes employing either AS or aerated lagoons (Karthikeyan and Meyer, 2006). Average removal of about 50% for macrolide antibiotics, such as clarithromycin and azithromycin, were reported from three conventional WWTPs in Japan (Kobayashi et al., 2006). Hirsch et al. (1999) reported that macrolides were found in all investigated WWTP discharges in Germany at concentrations in excess of 100 ng L
1. Studies using 24-h composite samples have revealed that the removal of macrolides by conventional AS treatment varied from
80% to 44% (Clara et al., 2005b; Gobel et al., 2007). It has been suggested that negative removals of macrolides are likely to be due to the release of these compounds from excreted bile and faeces

4307

during the biological treatment rather than the presence of deconjugable metabolites (Gobel et al., 2007). Sampling variation may also have contributed to this negative removal as reported by Clara et al. (2005b), where the collection of effluent samples was not time-adjusted to account for long HRTs. Sorption of macrolides to wastewater biomass is mainly attributed to hydrophobic interactions (Gobel et al., 2005a). This is expected due to their high Log DpH 6e8 partitioning coefficients (Table 5). Macrolides may also adsorb to biomass via cation exchange processes due to the fact that under typical wastewater conditions, many are positively charged through the protonation of the basic dimethylamino group (pKa > 8.9) and the surface of activated sludge is predominantly negatively charged (Carberry and Englande, 1983). However, in general, sorption to sludge accounts for only minor components of most macrolide drugs in WWTPs (Gobel et al., 2005a). Greater adsorption of azithromycin to biomass compared to clarithromycin has been reported (Kobayashi et al., 2006).

3.5.

Fluoroquinolones

Fluoroquinolones are antibiotics effective against several types of Gram-negative and Gram-positive bacteria (Turiel et al., 2003). These antibiotics act by inhibiting essential enzyme function for DNA production (Marzo and Dal Bo, 1998). The occurrence of fluoroquinolones in WWTP effluents has been reported in Australia, Canada, China, Italy, Mexico, Sweden, and the USA (Brown et al., 2006; Costanzo et al., 2005; Karthikeyan and Meyer, 2006; Lindberg et al., 2006; Miao et al., 2004; Zorita et al., 2009; Zuccato et al., 2005). When screening 12 human antibiotics in five WWTPs in Sweden, Lindberg et al. (2005) reported fluoroquinolones to be the most frequently detected antibiotics above analytical quantitation limits. In that study, norfloxacin and ciprofloxacin were detected in 97% and ofloxacin in 50% of the analysed samples. Fluoroquinolones appeared not be readily biodegradable in controlled batch tests (Kummerer et al., 2000). Removal efficiencies of fluoroquinolones from the aqueous phase during wastewater treatment in Sweden were reported to be 80% for norfloxacin and 78% for ciprofloxacin. Grit removal/ferrous precipitation achieved approximately 55%e58%, while AS treatment removed about 34% and 44% of norfloxacin and ciprofloxacin, respectively (Lindberg et al., 2006). A later study reported the removal of ciprofloxacin (90%), oxfloxacin (56%), and norfloxacin (70%) during AS treatment followed by chemical coagulation/flocculation (Zorita et al., 2009). The predominant removal mechanism of fluoroquinolones has been suggested by several authors to be adsorption to sludge and/or flocs rather than biodegradation (Batt et al., 2007; Golet et al., 2003; Lindberg et al., 2006; Zorita et al., 2009). A mass balance study revealed that a conventional wastewater treatment process resulted in the removal of 88e92% of fluoronoquinolones from the aqueous phase, due to the adsorption to sludge (Golet et al., 2003). They also observed that no significant removal of the compounds occurred under methanogenic conditions of the anaerobic sludge digester (i.e. biodegradation) and 75e83% of the compounds’ input mass remained in the digested sludge. Lindberg et al. (2006) reported that more than 70% of norfloxacin and ciprofloxacin passed through the treatment plant and remained in digested sludge. These findings

4308

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

implied sludge is the main reservoir of fluoroquinolones, potentially releasing the antibiotics into the environment via biosolids application to agricultural land. Sorption behaviour of fluoroquinolones is somewhat pH-dependant (Belden et al., 2007; Cardoza et al., 2005). However, the sorption of these compounds to sludge is not significantly affected by the narrow range of pH variability normally observed for WWTPs (Lindberg et al., 2006).

3.6.

Tetracyclines

The tetracyclines consist of eight related broad spectrum antibiotics, which are bacteriostatic and are active against Gram-positive and Gram-negative bacteria (Todar, 2002). Tetracyclines inhibit protein synthesis in the microorganisms by binding to the 30 S ribosome and preventing the access of aminoacyl tRNA to the acceptor site on the mRNA-ribosome complex (Marzo and Dal Bo, 1998). Tetracycline is one of the most frequently detected antibiotics in wastewater (Kim et al., 2007). Tetracyclines were reported in the raw WWTP influent in the USA at concentrations between 0.1 and 0.6 mg L
1 (Kim et al., 2005). In Canada, the remaining concentration of tetracycline in WWTP effluent was reported to be nearly 1.0 mg L
1 (Boussu et al., 2007; Miao et al., 2004). Removal efficiency of >68% has been reported for tetracycline during conventional secondary treatment (Karthikeyan and Meyer, 2006). In the USA, chlorotetracycline and doxycycline have been reported after secondary treatment and chlorination with removal efficiencies of 78% and 67%, respectively (Yang et al., 2005). Tetracyclines removal appears not to be significantly affected by changes in HRT of AS process (Kim et al., 2005). However, these authors reported a significant reduction in removal efficiency with decreased SRT from 10 days to 3 days, indicating that the change in the nature of biomass may affect the removal via solids adsorption. Controlled sorption tests have suggested that some tetracyclines have significant potential for adsorption onto biomass (Kim et al., 2005), whereas oxytetracycline is quite soluble in aqueous solutions and poorly adsorbed to biomass (Rabolle and Spliid, 2000). Despite low logDpH 6e8 values for these chemicals, non-hydrophobic mechanisms, such as ionic interactions, metal complexation, hydrogen bond formation or polarisation, likely play a significant role in the sorption of tetracyclines to many solids (Tolls, 2001). Since pH and temperature have been reported to have an effect on hydrolysis rates of tetracyclines (Loftin et al., 2008), it is possible that this mechanism may further contribute to the degradation of these chemicals in wastewater, particularly in tropical regions where temperatures are commonly above 35  C.

3.7.

Nitroimidazoles

Nitroimidazoles, such as metronidazole, are microbicidal drugs that are active against most anaerobic bacterial species (Theron et al., 2004) and a range of pathogenic anaerobic protozoa causing infections such as Giardiasis (Schneider, 1961), amoebiasis (Powell et al., 1966), and trichomoniasis (Cosar and Julou, 1959; Upcroft et al., 1999). There are only a few studies that have examined the occurrence of nitroimidazoles in sewage or WWTP effluents

since these drugs are dispensed in relatively small quantities as compared to other antibiotics. Thus, significant concentrations of nitroimidazoles are not expected to occur in sewage (Khan and Ongerth, 2004). However, Lindberg et al. (2005) observed that the concentrations of metronidazole in hospital sewage can be as high as 90.2 mg L
1, implying that higher concentrations may be found in sewage catchment receiving large contributions from hospitals. Metronidazole appears not to be readily biodegradable in laboratory-based batch experiments (Alexy et al., 2004; Ingerslev and Halling-Sørensen, 2000; Kummerer et al., 2000). Furthermore, it is relatively hydrophilic (Loke et al., 2000). As a result of these factors, metronidazole is not expected to be effectively removed during conventional wastewater treatment (Carballa et al., 2004; Khan and Ongerth, 2004).

3.8.

Other antibiotic groups

Aminoglycosides and ionophores are other antibiotic classes of interest. Aminoglycoside antibiotics are widely used in hospitals for treatment of serious human infection by gram-negative and gram-positive bacteria (Lo¨ffler and Ternes, 2003) and in veterinary medicine (Salisbury, 1995). Their antimicrobial action is by the inhibition of microorganism protein synthesis (Marzo and Dal Bo, 1998). Aminoglycosides are mostly nonmetabolised after being administered; hence they will be excreted via urine unchanged (Marzo and Dal Bo, 1998). The analysis of wastewater from a hospital in Germany revealed that the concentration of aminoglycoside antibiotic gentamicin was between 0.4 and 7.6 mg L
1 (Lo¨ffler and Ternes, 2003). There is little other information available on the occurrence and fate of aminoglycosides in wastewater and through treatment processes. However, due to their high sorption properties, it has been suggested that aminoglycoside antibiotics in wastewater would be adsorbed onto solid particles and colloidal organic matter and significantly removed from aqueous phase by filtration (Lo¨ffler and Ternes, 2003). Ionophore antibiotics are used in agricultural applications as feed additives to treat or prevent infections in poultry and livestock and as growth promoters for ruminants (Cha et al., 2005; Khan et al., 2008; Schlu¨sener et al., 2003). The pharmacological activity of ionophore antibiotics results from their ability to readily form electrically neutral psuedomacrocyclic complexes with polar mono and divalent cations in solution and transport the cations across the cell membrane (Cha et al., 2005). Several studies reported the occurrence of ionophore antibiotics such as monensin, salinomycin and narasin at concentrations up to 40 ng L
1 in surface water near the livestock feeding operations or agricultural lands (Cha et al., 2005; Kim and Carlson, 2006). A study by Watkinson et al. (2009) showed that the detection frequency and concentrations of monensin and salinomycin in wastewater were much lower than those in environmental waters. The behaviour of ionophore antibiotics through WWTP processes is little known due to the less likely occurrence of these antibiotics in domestic wastewater except where there is runoff from agricultural lands into sewers. A study by Donoho (1984) indicated that monensin is biodegradable in manure and soil where the primary degradation occurred in 33 days under aerobic

4309

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

condition, but much longer under anaerobic condition (60e70% after 10 weeks).

be more effective circumstances.

3.9. Effects of antibiotics on wastewater microbial consortia/processes

4.1.

While concentrations of antibiotics reported in domestic wastewater (in ng L
1 and mg L
1 ranges) are not considered sufficiently high to cause noticeable effects on wastewater treatment processes, a few studies have reported observed inhibition of wastewater microbial activities at elevated antibiotics concentrations (Amin et al., 2006; Gartiser et al., 2007; Ingerslev and Halling-Sørensen, 2000). For sulfonamide antibiotics, concentrations of 10e400 mg L
1 were reported to inhibit microbial activities in activated sludge by more than 20% (Ingerslev and Halling-Sørensen, 2000). A study by Amin et al. (2006) showed that the presence of erythromycin at concentration of 1 mg L
1 reduced COD removal efficiency and biogas production in anaerobic treatment by about 5%. A study of antibiotic biodegradability using the ISO closed bottle test revealed that a metronidazole concentration of 6 mg L
1 could reduce anaerobic activity by 69% (Gartiser et al., 2007). This is perhaps not surprising since an important pharmacological property of metronidazole is to target anaerobic microorganisms (Leiros et al., 2004). The presence of antibiotics in wastewater has been suspected to contribute to the development and dissemination of antibiotic resistant species since a significant number of antibiotic resistant genes have been found in WWTPs (Szczepanowski et al., 2009). Several bacteria such as fecal coliforms, E.coli and enterococci found in wastewater influent and effluent have exhibited resistance to ciprofloxacin, trimethoprim, sulfamethoxazole and vancomycin (Nagulapally et al., 2009). A study using a sequencing batch reactor by Kim et al. (2007) showed that the exposure to tetracycline at concentration of 1 mg L
1 increased concentrations and production rates of tetracycline resistant bacteria in the reactor. Nonetheless, the actual significance of low concentrations of antibiotics in WWTPs, in term of resistance propagation, is yet to be conclusively determined (Jury et al., in press).

4. Fate of antibiotics during advanced treatment processes Conventional secondary wastewater treatment processes appear to be highly variable in their ability to remove most antibiotics, with performance apparently dependent upon specific operational conditions, such as SRT. Accordingly, tertiary and advanced treatment processes may be necessary to provide further reduction of these compounds, in order to minimise environmental and human exposure. Semi-qualitative estimations of antibiotic removal by tertiary media filtration, ozonation, chlorination, UV irradiation, activated carbon adsorption, and NF/RO filtration as reported in the literature are presented in Table 6. Generally, it can be observed that the removals of most antibiotics by tertiary sand filtration and UV disinfection are poor while ozonation, chlorination, activated carbon and NF/RO filtration appear to

when

operated

under

optimum

Membrane filtration

Rejection of chemical contaminants by high-pressure membranes, such as nanofiltration (NF) and reverse osmosis (RO), is ultimately determined by complex interactions of electrostatic and other physical forces acting between a specific solute (chemical contaminant), the solution (water and other solutes present), and the membrane itself (Bellona et al., 2004; Nghiem et al., 2005). The key rejection mechanisms for these solutes are steric hindrance (size exclusion), electrostatic interactions (attractive or repulsive), and hydrophobic interaction (diffusion and partitioning) between compounds and the membrane. The nature of these forces is dependent on numerous chemical/physical properties of the solute (molecular size, pKa, polarity or hydrophobicity), solution (pH, ionic strength), and the membrane (materials, pore size). A useful guide for the classification of organic contaminants for estimations of their rejection by high-pressure membrane processes has been proposed by Bellona et al. (2004) (see Fig. 1). This diagram was derived as the result of a comprehensive review of published studies reporting rejection behaviour of a wide range of organic solutes by various commercially available membranes. Based on this rejection diagram, qualitative predictions can be derived to provide an estimation of treatment efficiencies by NF or RO for specific antibiotics as shown in Table 7. This table was developed with the assumptions of a highly negatively charged RO membrane with a nominal molecular weight cutoff (MWCO) of 100 Da at pH 7. Effective molecular width (MWd) of each compound was estimated by MMPplus software (ChemSW, 2005) based on the energy-minimised chemical structure generated by ChemBio3D Ultra (CambridgeSoft, 2007). Molecular charge states were determined according to the pKa values provided in Table 5. From the qualitative predictions (Table 7), RO treatment appears to be a promising process for the effective removal of most antibiotics. The predicted behaviours are reasonably well matched with quantitative data obtained from previous studies (Adams et al., 2002; Baumgarten et al., 2007; Dolar et al., 2009; Kosutic et al., 2007). These studies showed that >99% rejection is often achieved by RO and some NF membranes for several antibiotics including fluoroquinolones, sulfonamides, tetracyclines, and trimethoprim. A study undertaken by Li et al. (2004) on the treatment of an initially very high concentration of oxytetracycline in wastewater from pharmaceutical manufacturing industry found that RO treatment effectively reduced oxytetracycline concentration from 1000 mg L
1 to below 80 mg L
1 (>92% removal). Despite the usefulness of the membrane rejection diagram, there are some limitations in its applicability for predicting chemical behaviour in real full-scale treatment systems. MWCO is the manufacturer’s rating of the membrane ability to reject an uncharged standard compound (such as dextran or a protein) based on the molecular weight, while other organic compounds with different molecular shapes and sizes may partially permeate the membrane although their molecular weights exceeds the MWCO (Porter,

4310

Table 6 e Semi-quantitative estimations of the removal of antibiotics by tertiary and advanced treatment processes. Treatment Processes Group

Tertiary Sand Filtration Removal

Ref.

Ozonation Removal

Conditions

Chlorination Ref.

Free chlorine conc. 1.0e1.2 mg/L Removal

na

VG-E

O3 (3e5 mg/L), DOC (5.3 mg/L) pH (7.7), secondary effluent

(Dodd et al., 2006)

na

Sulfonamides

VP

(Batt et al., 2007; Gobel et al., 2007)

E

O3 (2.0e7.1 mg/L) HRT (1.5e18.0 min); pH (7e8); DOC (6.6e23 mg/L); secondary effluent and river water

(Adams et al., 2002; Huber et al., 2005; Ternes et al., 2003)

F-E

Free Chlorine (1e1.2 mg/L); pH (8); HRT (up to 1 day); Drinking water; River water

Macrolides

VP-P

(Gobel et al., 2007)

E

O3 (2e5 mg/L); HRT (4e18 min); pH (7e8); DOC (6.6e23 mg/L), secondary effluent and river water

(Huber et al., 2005; Radjenovic et al., 2009; Ternes et al., 2003)

VP

Tetracyclines

VP

(Batt et al., 2007)

E

O3 (3 mg/L), DOC (5.3 mg/L(pH (7.7), secondary effluent

(Dodd et al., 2006)

Fluoroquinolones

VP-P

(Batt et al., 2007; Golet et al., 2003)

E

O3 (3 mg/L); DOC (5.3 mg/L), pH (7.7), secondary effluent

(Dodd et al., 2006)

Nitroimidazoles Trimethoprim

VP-G*

(Batt et al., 2007)

na E

O3 (5e7.1 mg/L); HRT (Adams et al., 2002; (1.5e18 min); pH Dodd et al., 2006; (7e8); DOC Ternes et al., 2003) (7.7e23 mg/L)

Ref.

Removal

Conditions

Ref.

na

(Chamberlain and Adams, 2006; Gibs et al., 2007)

E

Free chlorine (3.5e3.8 mg/L), pH (7e8); DOC (3.0e3.5 mg/L); CaCO3 (80e307 mg/L): river water

(Westerhoff et al., 2005)

Free Chlorine (1.2 mg/ (Gibs et al., L); pH (8); HRT (1 day); 2007) CaCO3 (63 mg/L) Drinking water

E

Free chlorine (3.5e3.8 mg/L), pH (7e8); DOC (3.0e3.5 mg/L); CaCO3 (80e307 mg/L): river water

(Westerhoff et al., 2005)

E

Free Chlorine (1.2 mg/ Gibs et al., L); pH (8); HRT (1 day); 2007 CaCO3 (63 mg/L) Drinking water

na

P

Free Chlorine (1.2 mg/ Gibs et al., L); pH (8); HRT (1 day); 2007 CaCO3 (63 mg/L) Drinking water

na

Free chlorine (3.5e3.8 mg/L), pH (7e8); DOC (3.0e3.5 mg/L); CaCO3 (80e307 mg/L): river water

(Westerhoff et al., 2005)

na VG

Free Chlorine (1 mg/L); (Adams et al., HRT (>40 min); DOC 2002) (10.3 mg/L); River water

na E

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Beta-lactam

Conditions

Free chlorine conc. >3 mg/L

Treatment Processes Group

UV irradiation Typical disinfection dose Removal

Ref.

Activated Carbon Adsoprtion

RO/NF Membrane Filtration

Higher Dose (20e100 times) Removal

Conditions

Ref.

VP

(Batt et al., 2007)

na

Sulfonamides

VP-P

(Drewes et al., 2008; F-VG Le-Minh et al., 2010)

UV dose (2760e3000 (Adams et al., 2002; mJ/cm2); HRT Kim et al., 2009) (5e30 min); DOC (3.5e10.7 mg/L)

Macrolides

VP

(Drewes et al., 2008) P

UV dose (2760 mJ/ cm2); HRT (5 min); DOC (3.5 mg/L)

(Kim et al., 2009)

Tetracyclines

na

VG-E

UV dose (2760 mJ/ cm2); HRT (5 min); DOC (3.5 mg/L)

Fluoroquinolones

VP

(Vieno et al., 2007b)

VG-E

UV dose (2760 mJ/ cm2); HRT (5 min); DOC (3.5 mg/L)

Nitroimidazoles

VP

(Shemer et al., 2006)

Trimethoprim

VP

(Drewes et al., 2008; F Le-Minh et al., 2010)

Conditions

Ref.

F-VG

pH 6e7, GAC dose (20 mg/L), Co (10ug/ L), based on Freundlich isotherm, equlibrium state

(Aksu and Tunc, 2005; Dutta et al., 1999; Putra et al., 2009)

G-VG

Removal

Ref.

E

(Morse and Jackson, 2004)

pH 7.7e7.9, PAC dose (Adams et al., 2002; (20 mg/L), Co (up to Westerhoff et al., 50ug/L), DOC 2005) (3.5e10.7 mg/L), 4 h contact time; river water

G-E

(Dolar et al., 2009; Kimura et al., 2004)

VG

pH 7.9, PAC dose (20 mg/L), Co (50e100 ng/L), DOC (3.5 mg/L), 4 h contact time, river water

E

(Dolar et al., 2009)

(Kim et al., 2009)

E

pH 5.8, PAC (20 mg/ (Ji et al., 2009) L), Co (10 ug/L), based on Freundlich isotherm, equilibrium state

E

(Kosutic et al., 2007)

(Kim et al., 2009)

G-VG

PAC (50 mg/L); Co (25ug/L); 15 min contact time; MBR permeate

E

(Baumgarten et al., 2007; Dolar et al., 2009)

E

(Rivera-Utrilla et al., na pH 6e7, PAC dose (20 mg/L), Co (10ug/ 2009) L), based on Freundlich isotherm, equlibrium state

G-VG

pH 7.7e7.9, PAC dose (Adams et al., 2002; (20 mg/L), Co (up to Westerhoff et al., 50 mg/L), 2005) experiments in riverwater, DOC (3.5e10.7 mg/L), 4 h contact time

UV dose (2760e3000 (Adams et al., 2002; mJ/cm2); HRT Kim et al., 2009) (5e30 min); DOC (3.5e10.7 mg/L)

(Baumgarten et al., 2007)

VG-E

(Dolar et al., 2009; Kosutic et al., 2007)

4311

VP: very poor (<20%); P: Poor (20e45%); F: Fair (45e65%); G: Good (65e80%); VG: Very Good (80e95%); E: Excellent (>95%); na: not available.

(Westerhoff et al., 2005)

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Beta-lactam

Removal

4312

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Fig. 1 e Rejection diagram for organic micropollutants during membrane treatment (Bellona et al., 2004). MW [ molecular weight, pKa [ acid.

1990). Studies have shown that the rejection of antibiotics is better correlated to membrane pore size than to MWCO (Dolar et al., 2009; Kimura et al., 2004). Therefore, when size exclusion is the main mechanism for rejecting compounds, the comparison of effective MWd and membrane pore size distribution should be considered in order to optimise rejection estimations. Most investigations of the relationships between physicalechemical properties of solutes and membrane interactions have been conducted using unfouled ‘virgin’ membranes and thus their conclusions are unlikely to be quantitatively extendable to full-scale systems subjected to long-term operation (Agenson et al., 2003; Nghiem et al., 2004, 2005; Schafer et al., 2003; Simon et al., 2009). During normal operation, membranes are prone to fouling by the build-up of precipitated chemicals or by the growth of microbial biomass (Nghiem et al., 2006; Nghiem and Schafer, 2006; Oschmann et al., 2005). Fouling can lead to significant changes in physicochemical properties of the membrane surface, affecting the separation mechanisms including size exclusion and electrostatic interactions and thus in the way in which the membranes interact with water and solutes (Nghiem et al., 2009; Nghiem and Hawkes, 2007). In many cases, fouling is regarded as a hindrance since it decreases membrane permeability and thus requires elevated pressures to maintain operational flux. However, some investigations reveal that fouling can also lead to improved rejection of many solutes (Drewes et al., 2006; Schafer et al., 1994; Xu et al., 2006). This observation is believed to be due to increased negative surface charge leading to increased electrostatic rejection of ionic species; along with simultaneously increased adsorptive capacity for non-ionic solutes (Xu et al., 2006). Furthermore,

membrane degradation due to exposure to residual chlorine can impact rejection of some antibiotics (Simon et al., 2009). A further limitation of the membrane rejection diagram is that it does not account for the role of electrostatic attraction (as opposed to repulsion). The surfaces of most NF/RO membranes (made of polyamide or cellulose acetate) are negatively charged under neutral conditions due to the presence of ionisable sulphonic and carboxylic functional groups (Nghiem et al., 2005; Schafer et al., 2003; Simon et al., 2009). Negatively charged solutes are rejected mainly via electrostatic repulsion; while the positively charged species are removed by combination of attractive electrostatic interaction with the membrane surface and Donnan equilibrium (Schaep et al., 2001; Verliefde et al., 2008).

4.2.

Adsorptive treatment

4.2.1.

Activated carbon

Adsorptive treatment with activated carbon can be used for removing many hydrophobic pharmaceuticals from water (Snyder et al., 2003). The removal effectiveness of the activated carbon adsorptive treatment system depends on the properties of the adsorbent (specific surface area, porosity, surface polarity, and physical shape of the material) and the characteristics of the compound (shape, size, charge and hydrophobicity). Adsorption mechanisms consist of the chemical (electrostatic interaction) and physical bindings of molecules to the surface of an adsorbent. The latter is often more important due to the capability to form multi-layer bindings (Snyder et al., 2003). In fact, it was recently reported that the greatest removal of amoxicillin by activated carbon

Table 7 e Qualitative rejection prediction based on LogKow*, MW, and charge state for antibiotics by reverse osmosis assuming MWCO [ 100, pH [ 7, negatively charge reverse. Name

MWd (nm)

Charge State at pH 7

Dominant rejection mechanisms

Amoxicillin Cephalexin Sulfadimethoxine Sulfamerazine Sulfamethazine Sulfamethoxazole Sulfathiazole Sulfadiazine Sulfapyridine Ciprofloxacin Norfloxacin Erythromycin

365 347 310 264 278 253 255 250 249 331 319 734

1.32 1.39 1.59 1.47 1.47 1.43 1.31 1.36 1.33 1.23 1.18 1.59

72% zwitterionic; 28% negative 57% zwitterionic; 43% negative 7% neutral; 93% negative 37% neutral; 63% negative; 72% neutral; 28% negative 4% neutral; 96% negative 54% neutral; 45% negative 16% neutral; 84% negative 96% neutral; 4% negative zwitterionic and positive zwitterionic and positive 99% positive

Roxithromycin

837

2.18

99% positive

Clarithromycin

748

1.61

99% positive

Azithromycin

749

1.18

97% positive

Oxytetracycline Chlortetracycline Doxycycline Tetracycline Metronidazole Trimethoprim

460 479 444 444 171 290

1.31 1.56 1.50 1.62 1.01 1.42

Zwitterionic and negative Zwitterionic and negative Zwitterionic and negative Zwitterionic and negative 100% neutral 63% neutral; 37% positive

dominantly electrostatic repusion and size exclusion dominantly electrostatic repusion and size exclusion dominantly electrostatic repulsion and less from size exclusion mainly electrostatic repuslion, and less from size exclusion mainly steric hindrance; and less from electrostatic repusion dominantly electrostatic repusion setric hindrance and electrostatic repulsion dominantly electrostatic repusion; and steric hindrance dominantly steric hindrance steric hindrance, electrostatic attraction steric hindrance, electrostatic attraction electrostatic attraction, mainly steric hindrance and hydrophobic interaction electrostatic attraction, mainly steric hindrance and hydrophobic interaction electrostatic attraction, mainly steric hindrance and hydrophobic interaction electrostatic attraction, mainly steric hindrance and hydrophobic interaction dominantly electrostatic repusion dominantly electrostatic repusion dominantly electrostatic repusion Dominantly electrostatic repusion Mainly steric hindrance Steric hindrance, electrostatic attraction

Qualitative rejection prediction from rejection diagram very high very high very high very high moderate very high moderate very high moderate very high very high moderate to high, but depends on partitioning and diffusion moderate to high depends on partitioning and diffusion moderate to high, but depends on partitioning and diffusion moderate to high, but depends on partitioning and diffusion very high very high very high very high moderate moderate

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

MW

*LogKow values from Table 5. MW: Molecular weight; MWd: Molecular width; MWCO: Membrane molecular weight cut-off.

4313

4314

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

was achieved under pH conditions corresponding to a zero net charge on the activated carbon surface (Putra et al., 2009). It has been known for many years that adsorption to activated carbon can be used to purify antibiotics for commercial production (Bansal and Goyal, 2005) and for the treatment of poisoning and overdose of drugs (Cooney, 1995). In water and wastewater treatment, several studies on the removal of antibiotics by activated carbon have been reported (Adams et al., 2002; Choi et al., 2008; Putra et al., 2009; Rivera-Utrilla et al., 2009). With dosages between 10 mg L
1 and 20 mg L
1 powdered activated carbon (PAC), the concentrations of several antibiotics in river water have been reduced by 49%e 99% after 4 h contact time (Adams et al., 2002; Westerhoff et al., 2005). After 1 day contact time, similar removals of sulfonamide and tetracycline antibiotics from river water have been achieved with the lower dosage of 1 mg L
1 PAC (Choi et al., 2008). The sorption efficiencies of antibiotics to activated carbon may be significantly altered by several factors, such as the types of activated carbon used, the initial concentrations of target compounds and the pH, temperature and dissolved organic carbon (DOC) concentration of the solution (Aksu and Tunc, 2005; Choi et al., 2008; Dutta et al., 1999; Ji et al., 2009; Rivera-Utrilla et al., 2009). A number of previous studies have reported the prediction of activated carbon adsorption capacity for antibiotics including amoxicillin, penicillin, tetracycline, and nitromidazole using Freundlich or Langmuir isotherms (Aksu and Tunc, 2005; Dutta et al., 1999; Ji et al., 2009; Putra et al., 2009; RiveraUtrilla et al., 2009). The capacity of activated carbon to adsorb a particular compound can, to some extent, be predicted based on the ‘hydrophilic’ or ‘hydrophobic’ nature of the chemical (Snyder et al., 2003). The hydrophobic (non-polar) or hydrophilic (polar) properties of antibiotics can be determined from their LogD (or pKa-adjusted Log Kow) values (Table 5). It has been reported that non-polar antibiotics with Log Kow > 2, may be effectively removed with activated carbon by hydrophobic interaction (Snyder et al., 2003). However, the adsorption of more polar or charged compounds to activated carbon is much more difficult to predict due to the additional effects of polar interactions and ion exchange (Snyder et al., 2003). A study undertaken by Westerhoff et al. (2005) examined the relationship between activated carbon capacity and the Log Kow of 62 different micropollutants including some antibiotics such as sulfamethoxazole, erythromycin-H2O and trimethoprim. The study showed that the degree of removal of non-volatile neutral compounds after contact by PAC (5 mg L
1, 4 h contact time) can be reasonably well predicted by the following relationship: [percentage removal] ¼ 15  [Log Kow] þ 27%. However, the removal of protonated and deprotonated compounds did not follow this trend (Westerhoff et al., 2005). Accordingly, caution must be exercised in the prediction of activated carbon removal efficiency by Log Kow values and the effect of solution pH as well as pKa values of the compounds must be carefully considered in order to derive LogDpH instead. LogD calculation takes into account both ionisation constant (pKa) and log Kow of various species formed in solution at different pH. Therefore, the use of LogD is more appropriate to estimate the partitioning properties of all compounds including ionisable species than log Kow (Bhal et al., 2007).

4.2.2.

Ionic adsorption

Many antibiotics, including tetracyclines and sulfonamides are often present in negatively charged form at normal operating pH conditions (Quiang and Adams, 2004). Therefore, the use of ionic treatment processes may be effective for the removal of these anionic micropollutants (Robberson et al., 2006). A study by Choi et al. (2007b) revealed that the anionic MIEX resin is effective for the removal of fourteen antibiotics from the sulfonamide and tetracycline classes, which are in zwitterionic and anionic forms at pH 7. Accordingly, significant removal of 10 mg L
1 sulfonamides and tetracyclines spiked in 0.8 mg L
1 DOC water was achieved with the addition of 3.6e13 mL L
1 MIEX. Ion exchange is the main mechanism in the ionic treatment for negatively charged antibiotics though the removal of antibiotics by agglomeration could also occur in the presence of metal oxides and natural organic matter (Boyer and Singer, 2005; Choi et al., 2007b). To the knowledge of the authors, no full-scale treatment of ionic adsorption for antibiotics has been reported in the scientific literature. One explanation for this apparent lack of interest may be a lack of cost-effectiveness since the presence of other organic contaminants in the aqueous stream can compete with targeted antibiotics for the ion exchange sites and reduce the antibiotic removal efficiency. Regardless of the efficiency, ion exchange based processes are not targeting neutral compounds and additional processes would be required to provide a more comprehensive removal of the full spectrum of antibiotics. Full-scale studies are required to determine the optimal configuration and operating conditions of adsorptive systems, which are effective and economically feasible for antibiotics removal.

4.3. Chemical and photochemical oxidation processes for the removal of antibiotics Oxidative processes may be used to transform any organic constituents of wastewaters, which appear to be both biologically recalcitrant and poorly rejected by membranes or activated carbon. Strong chemical oxidants, such as ozone (von Gunten, 2003), potassium permanganate (Adam et al., 2004; Chen et al., 2005) and chlorine (Chamberlain and Adams, 2006), have been shown to be effective for the transformation of various chemical contaminants in water.

4.3.1.

Chlorination

Chlorination can inactivate active chemical compounds via one of two general mechanisms. One possibility is by chorine substitution or addition reactions, which may alter active functional groups (Crain and Gottlieb, 1935). Alternatively, chlorine radicals may oxidise (break down) the target compound such as antibiotic drugs into smaller molecules, which may or may not possess the active properties (Crain and Gottlieb, 1935). The effective removal of antibiotics by chlorination from drinking water requires sufficient free chlorine concentration and contact time. With the use of free chlorine at 1.0 mg L
1 (as Cl2), 90% removal has been reported with contact times greater than 16 min for most sulfonamides and greater than 40 min for trimethoprim in river water (Adams et al., 2002). A

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

study by Gibs et al. (2007) on the persistence of 98 pharmaceuticals, including 23 antibiotics of sulfonamide, tetracycline, macrolide and quinolone classes in chlorinated drinking water also revealed that the presence of free chlorine in drinking water is an effective means for the transformation of some pharmaceutical compounds during distribution. With a free chlorine concentration of 1.2 mg L
1 and an initial spiked concentration of 0.5 mg L
1 for each antibiotic in drinking water, reductions of >99% for tetracyclines, 50%e80% for sulfonamides, 42% for trimethoprim, 30%e40% for fluoroquinolones, and less than 10% for macrolides, respectively, were observed after a single day’s contact time, and complete removals were achieved after 10 days (Gibs et al., 2007). At the higher free chlorine concentration of 3.5e3.8 mg L
1, removals of 90% to >99% were achieved for sulfamethoxazole, trimethoprim and erythromycin in river water after 24 h contact time (Westerhoff et al., 2005). Accordingly, although some antibiotics may be more resistant to chlorination than others, they appear to gradually degrade over time in the presence of free chlorine. In addition, the optimum dosage and contact time may increase with increased concentration of solids and organic matter in the water. Reactivity of sulfonamides with HOCl has been observed to be in the following descending order: sulfadiamethoxine > sulfathiazole > sulfamethazine > sulfamerazine > sulfamethoxazole > sulfamethizole (Chamberlain and Adams, 2006). Alkaline pH (>8) inhibits the removal of sulfonamides by chlorine oxidation (Chamberlain and Adams, 2006; Gibs et al., 2007). Removals of sulfonamides and macrolides using chloramination were minimal under typical drinking water conditions (Chamberlain and Adams, 2006). Huber et al. (2005) observed fast reactions between ClO2 oxidation and antibiotics, such as sulfonamides and macrolides, in natural water. Roxithromycin and sulfamethoxazole exhibited strong pH dependence in the oxidation with ClO2 (higher reactivity at pH > 7). Similar to HClO, ClO2 oxidises roxithromycin and sulfamethoxazole at specific functional groups with high electron densities, such as neutral tertiary amines and aniline. Based on the laboratory studies, rapid and substantial transformation of trimethoprim to a wide range of chlorinated and hydroxylated products is expected to occur under typical conditions of wastewater and drinking water chlorination (Dodd and Huang, 2007). Although a number of studies have demonstrated the ability of chlorination to reduce antibiotics concentrations in drinking water, it is still too early to confirm the relative importance of chlorine and its derivative products in degrading antibiotics in water and wastewater treatment. Indeed, the major concern for treating pharmaceuticals via chlorination is the formation of chlorinated byproducts since these may be more harmful than their parent compounds (von Gunten et al., 2006). In past decades, the use of chlorine for water treatment has attracted concern since the reaction of chlorine with natural organic matter is well known to produce harmful chlorinated byproducts. Data and research regarding the ultimate fate of these pharmaceuticals treated by chlorination processes and whether they are degraded to harmless metabolites or transformed into potentially more toxic contaminants is needed (Glassmeyer and Shoemaker, 2005).

4.3.2.

4315

Ozonation

Several previous studies have reported the effective treatment of ozonation for removal of antibiotics in water and wastewater effluents (Adams et al., 2002; Huber et al., 2005; Ternes et al., 2003). The study by Adams et al. (2002) showed that ozonation removed more than 95% of several sulfonamides and trimethoprim from river water within 1.3 min contact time at an ozone dose of 7.1 mg L
1 Huber et al. (2005) also observed that using ozonation at doses >2 mg L
1 oxidised 90% e >99% of sulfonamides and macrolides in secondary wastewater effluents. Second order rate constants for 14 fluoroquinolones, sulfonamides, b-lactams, macrolides and trimethoprim by ozone and hydroxyl radicals have been reported (Dodd et al., 2009). As the result of this study, the authors suggested that 99% removal of antibiotics from river water and wastewater effluents should be achievable at typical ozone doses used for disinfection (i.e. 5e10 mg L
1 ozone with DOC of 5e23 mg L
1). During ozone treatment, oxidative degradation of organic chemicals can occur either by direct reaction with molecular ozone (O3) or indirectly via hydroxyl radicals (Staehelin and Hoigne, 1985). During wastewater ozonation, many antibiotics, including sulfonamides, macrolides, fluoroquinolones and tetracyclines, have been shown to be predominantly transformed via direct reaction with ozone (Dodd et al., 2006) whereas cephalexin, penicillin, and N4-acetyl sulfamethoxazole were transformed to a large extent by hydroxyl radicals (Dodd et al., 2006). The relative dominance of the actual oxidative pathway will depend on the ratio of molecular ozone and hydroxyl radicals, the corresponding reaction kinetics, and presence of organic matter (Elovitz et al., 2000, von Gunten, 2003). Ozone and/or hydroxyl radicals deactivate bactericidal properties of antibiotics by attacking or modulating their pharmaceutically active functional groups, such as N-etheroxime and dimethylamino groups of macrolides (Dodd et al., 2009; Lange et al., 2006), aniline moieties of sulfonamides (Huber et al., 2005), thioether groups of penicillins, unsaturated bonds of cephalosporin and the phenol ring of trimethoprim (Dodd et al., 2009). The good removal (>90%) by ozonation was observed for those compounds with electron-rich aromatic systems, such as hydroxyl, amino (e.g. sulfamethoxazoles), acylamino, alkoxy and alkyl aromatic compounds, as well as those compounds with deprotonated amine (e.g. erythromycin, ofloxacin and trimethoprim) and nonaromatic alkene groups since these key structural moieties are highly amendable to oxidative attack (Dickenson et al., 2009). However, the major concern for the use of ozone for antibiotic oxidation is the potential transformation to products that remain biologically active and resistant to further ozonation. Dodd et al. (2009) reported that the ozonation products of b-lactam antibiotics (i.e. penicillin and cephalexin) are still biologically active after primary oxidation reactions, but could be further deactivated by hydroxyl radicals or ozone if the concentration of residual ozone is sufficient. However, in the case of roxithromycin, the primary ozonation products have the bactericidal dimethylamino groups preserved and are quite persistent to further degradation at very high ozone doses (Radjenovic et al.,

4316

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

2009). Desethylene ciprofloxacin is an ozone byproduct of ciprofloxacin, for which formation can be significantly affected by solution pH (De Witte et al., 2009).

4.3.3.

Ultraviolet irradiation

Ultraviolet (UV) irradiation can be used to degrade some organic chemicals in water (Rosenfeldt and Linden, 2004). Degradation is governed by UV energy absorption and quantum yield of that compound (Kim et al., 2009). The DOC concentration, UV dose and contact time are also important factors governing the removal efficiency. Negligible removal of antibiotics from WWTP secondary effluents through UV disinfection process with a typical dose of 30e80 mJ cm
2 has been often reported (Batt et al., 2007; Drewes et al., 2008; Le-Minh et al., 2010). Although sulfonamide antibiotics such as sulfamethoxazole are more prone to photolytic degradation (Boreen et al., 2004), the poor removal (25e50%) of the compound is still observed during UV disinfection process (Drewes et al., 2008), due to the presence of DOC in the treated effluent which highly competes for the limited UV radiation energy at typical disinfection dose. The degradation of antibiotics tends to be effective only at very high UV radiation dose, about 20e100 times higher than the typical disinfection dose for wastewater effluent (Adams et al., 2002; Kim et al., 2009). It has been recently reported that at high UV doses of nearly 3000 mJ/cm2 and DOC of 2.5e4 mg L
1, a 5 min contact time was required to achieve >90% removal for sulfamethoxazole and norfloxacin, while 15 min contact time was required for tetracycline (Kim et al., 2009).

antibiotics were removed from tertiary treated wastewater by the AOP system of 7 mg/L ozone, 3.5 mg/L H2O2 and 2 min contact time. The degradation of metronidazole using photochemical oxidations including UV, UV/H2O2, H2O2/Fe2þ (Fenton), and UV/H2O2/Fe2þ (photo-Fenton) has been investigated (Shemer et al., 2006). This study showed that while the removal of metronidazole was negligible solely by UV irradiation (a dose of 600 mJ/cm2 and 5 min retention), the addition of ferrous ions significantly enhanced the removal efficiency. However, all samples in this study were spiked deionised water samples; and the effectiveness of photochemical oxidations for treating metronidazole in wastewater is expected to be reduced because of the increased turbidity and interference of matrix compounds. The overall extent of oxidation for any AOP is dependent on the contact time and the concentration of scavengers in the water (i.e., non-target oxidisable species). Typically, dissolved organic carbon (DOC) and carbonate/bicarbonate are the most important scavengers in drinking waters. High concentrations of DOC and carbonate/bicarbonate can render mineralisation of micropollutants quite inefficient (von Gunten, 2003). The presence of natural organic material can initiate the formation of hydroxyl radical whereas the humic acid and bicarbonate can scavenge the radicals (Drewes et al., 2008). However, pre-treatment processes, such as GAC or RO, significantly reduce DOC concentrations, thus enhancing oxidation efficiency.

5. 4.3.4.

Conclusion

Advanced oxidation processes

Oxidative degradation can occur either by direct reaction with the applied oxidant, or via the production of highly reactive secondary species, most commonly, hydroxyl radicals (OH), which is one of the most powerful oxidants known (Metcalf and Eddy, 2007). Processes that promote the enhanced formation of hydroxyl radicals are generally referred to as advanced oxidation processes (AOPs). UV radiation is commonly used to promote the formation of hydroxyl radicals. This can be achieved by a number of methods including photocatalysis with titanium dioxide (TiO2) (Egerton et al., 2006; Murray and Parsons, 2006) or by direct reaction of hydrogen peroxide (H2O2) (Rosenfeldt and Linden, 2004). Other systems, such as Ozone/H2O2 and UV/Ozone are also considered AOPs since they promote the formation of hydroxyl radicals. Significant research has been conducted in the past to investigate the performance of AOPs for removing pharmaceutical contaminants in treated wastewater (Drewes et al., 2008; Rosenfeldt and Linden, 2004; Westerhoff et al., 2005). From these studies, AOPs were found to be effective treatment processes for removing the selected pharmaceutical contaminants and to provide the improved removal efficiency compared to UV radiation or ozonation alone. Although few studies on the performance of AOPs for oxidising antibiotics are available, AOPs also appear to be effective for oxidising antimicrobial contaminants. The study using lab-scale and full-scale ozone/H2O2 treatment systems showed that the concentrations of erythromycin, ofloxacin, sulfamethoxazole and trimethoprim in tertiary treated wastewater were significantly reduced (Drewes et al., 2008). More than 90% of these

Antibiotic pharmaceuticals, administered for human medication, represent a broad range of chemical classes including b-lactams, sulfonamides, macrolides, fluoroquinolones, tetracyclines, and nitroimidazoles, as well as the important compound, trimethoprim. This diversity of chemical classes is equally represented by variability of physical, chemical and biochemical properties, resulting in variable susceptibility to physical, chemical and biological treatment processes. b-lactams are characterised by the fact that they are generally very quickly degraded during conventional wastewater treatment processes. This observed behaviour is presumed to be a function of susceptibility to chemical and biochemical hydrolysis of the b-lactam ring. However, observed behaviours of other groups of antibiotics are much more difficult to characterise due to varying removal efficiencies reported from studies undertaken in different parts of the world. The most reasonable interpretation appears to be that specific design and operating conditions of individual WWTPs is an important determinative factor for antibiotic chemical removal. Unfortunately, insufficient information is currently available to thoroughly assess these differences. Inconsistencies in research objectives and focuses may also partially explain differences in reported removal efficiencies from previous studies. In fact, many previous studies often calculated removal efficiencies by the difference in an antibiotics’ concentration in treated effluent from that of the influent, but not taking into account the potential intertransformation between antibiotics and their metabolites in wastewater, particularly in the case of sulfonamide classes.

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Furthermore, the occurrence of some enzymatic degradation is known to significantly enhance the biodegradation of some antibiotics. Such enzymatic degradation is governed by several factors, such as HRT, SRT and operational temperature. Many studies did not specify, or well define, these operational factors during their investigation, rendering results difficult to compare. In spite of the variable removal of antibiotics during conventional WWTP processes, many of these chemicals are routinely observed in secondary treated effluents within the range of 10e500 ng L
1. Whether these concentrations are of any public health significance is currently unknown, however the simple fact that they have been observed and reported has been a cause of scientific and public concern. Accordingly, for many water recycling applications e especially those involving a high degree of personal contact or are designed to augment potable supplies - the efficacy of advanced water treatment processes for antibiotics removal is of interest. Comparatively little information is available regarding the effectiveness of advanced treatment processes including membrane filtration, adsorptive treatment (activated carbon and ionic adsorption), or chemical and photochemical oxidation. Nonetheless, with the aid of a few basic physicalchemical molecular properties, it has been possible to qualitatively characterise various antibiotics in terms of susceptibility to some of these processes. While there appears to be no ‘silver bullet’ for the removal of all residual antibiotics under typical operational conditions, there is a strong implication that judicious selection of advanced treatment processes can be used for the effective removal of these compounds to levels unlikely to be detectable by current analytical procedures. It is recommended that more comprehensive studies are required to fill knowledge gaps in the behaviour of antibiotics under conventional sewage treatment and advanced treatment processes. Future research should include a dedicated focus on the adsorption and fate of the antibiotics into the sewage biomass, the potential formation of pharmacologically active or more toxic metabolites and degradation products during treatment processes. Furthermore, future studies should control and report all basic treatment plant operational parameters since these are essential for later comparison or assessments.

Acknowledgements This research was supported under the Australian Research Council’s Discovery Projects funding scheme (project number DP0558029).

references

Adam, M.L., Comfort, S.D., Morley, M.C., Snow, D.D., 2004. Remediating RDX-contaminated ground water with permanganate: laboratory investigations for the Pantex perched aquifer. Journal of Environmental Quality 33 (6), 2165e2173. Adams, C., Wang, Y., Loftin, K., Meyer, M., 2002. Removal of antibiotics from surface and distilled water in conventional

4317

water treatment processes. Journal of Environmental Engineering e ASCE 128 (3), 253e260. Adelowo, O.O., Fagade, O.E., Oke, A.J., 2008. Prevalence of coresistance to disinfectants and clinically relevant antibiotics in bacterial isolates from three hospital laboratory wastewaters in Southwestern Nigeria. World Journal of Microbiology & Biotechnology 24 (9), 1993e1997. Agenson, K.O., Oh, J.I., Urase, T., 2003. Retention of a wide variety of organic pollutants by different nanofiltration/reverse osmosis membranes: controlling parameters of process. Journal of Membrane Science 225 (1e2), 91e103. Aksu, Z., Tunc, O., 2005. Application of biosorption for penicillin G removal: comparison with activated carbon. Process Biochemistry 40 (2), 831e847. Alder, A.C., Bruchet, A., Carballa, M., Joss, A., Loffler, D., McArdell, C.S., Miksch, K., Omil, F., Tuhkanen, T., Ternes, T. A., 2006. In: Ternes, T.A., Joss, A. (Eds.), Human Pharmaceuticals, Hormones and Fragrances. IWA Publishing, London, pp. 15e54. Alexy, R., Ku¨mpel, T., Ku¨mmerer, K., 2004. Assessment of degradation of 18 antibiotics in the closed bottle test. Chemosphere 57, 505e512. American Chemical Society, 2009. SciFinder Scholar. Amin, M.M., Zilles, J.L., Greiner, J., Charbonneau, S., Raskin, L., Morgenroth, E., 2006. Influence of the antibiotic erythromycin on anaerobic treatment of a pharmaceutical wastewater. Environmental Science & Technology 40 (12), 3971e3977. Auerbach, E.A., Seyfried, E.E., McMahon, K.D., 2007. Tetracycline resistance genes in activated sludge wastewater treatment plants. Water Research 41 (5), 1143e1151. Balakrishnan, V.K., Terry, K.A., Toito, J., 2006. Determination of sulfonamide antibiotics in wastewater: a comparison of solid phase microextraction and solid phase extraction methods. Journal of Chromatography A 1131 (1e2), 1e10. Bansal, R.C., Goyal, M., 2005. Activated Carbon Adsorption. CRC Press, Boca Raton, USA. Baquero, F., Martinez, J.L., Canton, R., 2008. Antibiotics and antibiotic resistance in water environments. Current Opinion in Biotechnology 19 (3), 260e265. Batt, A.L., Kim, S., Aga, D.S., 2006. Enhanced biodegradation of iopromide and trimethoprim in nitrifying activated sludge. Environmental Science & Technology 40 (23), 7367e7373. Batt, A.L., Kim, S., Aga, D.S., 2007. Comparison of the occurrence of antibiotics in four full-scale wastewater treatment plants with varying designs and operations. Chemosphere 68 (3), 428e435. Baumgarten, S., Schroder, H.F., Charwath, C., Lange, M., Beier, S., Pinnekamp, J., 2007. Evaluation of advanced treatment technologies for the elimination of pharmaceutical compounds. Water Science and Technology 56 (5), 1e8. Beausse, J., 2004. Selected drugs in solid matrices: a review of environmental determination, occurrence and properties of principal substances. TrAC-Trends in Analytical Chemistry 23 (10e11), 753e761. Belden, J.B., Maul, J.D., Lydy, M.J., 2007. Partitioning and photodegradation of ciprofloxacin in aqueous systems in the presence of organic matter. Chemosphere 66 (8), 1390e1395. Bellona, C., Drewes, J.E., Xu, P., Amy, G., 2004. Factors affecting the rejection of organic solutes during NF/RO treatmenta literature review. Water Research 38 (12), 2795e2809. Bendz, D., Paxeus, N.A., Ginn, T.R., Loge, F.J., 2005. Occurrence and fate of pharmaceutically active compounds in the environment, a case study: Hoje River in Sweden. Journal of Hazardous Materials 122 (3), 195e204. Bhal, S.K., Kassam, K., Peirson, I.G., Pearl, G.M., 2007. The rule of five revisited: applying log D in place of log P in drug-likeness filters. Molecular Pharmaceutics 4 (4), 556e560. Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in the aquatic environment: sulfa drugs

4318

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

containing five-membered heterocyclic groups. Environmental Science & Technology 38 (14), 3933e3940. Botitsi, E., Frosyni, C., Tsipi, D., 2007. Determination of pharmaceuticals from different therapeutic classes in wastewaters by liquid chromatography-electrospray ionization-tandem mass spectrometry. Analytical and Bioanalytical Chemistry 387 (4), 1317e1327. Boussu, K., Kindts, C., Vandecasteele, C., Van der Bruggen, B., 2007. Surfactant fouling of nanofiltration membranes: measurements and mechanisms. Chemphyschem 8 (12), 1836e1845. Boyer, T.H., Singer, P.C., 2005. Bench-scale testing of a magnetic ion exchange resin for removal of disinfection by product precursors. Water Research 39 (7), 1265e1276. Brown, K.D., Kulis, J., Thomson, B., Chapman, T.H., Mawhinney, D.B., 2006. Occurrence of antibiotics in hospital, residential, and dairy effluent, municipal wastewater, and the Rio Grande in New Mexico. Science of the Total Environment 366 (2e3), 772e783. Cahill, J.D., Furlong, E.T., Burkhardt, M.R., Kolpin, D., Anderson, L.G., 2004. Determination of pharmaceutical compounds in surface-and ground-water samples by solidphase extraction and high-performance liquid chromatography-electrospray ionization mass spectrometry. Journal of Chromatography A 1041 (1e2), 171e180. CambridgeSoft, 2007. ChemBioOffice2008. Carballa, M., Omil, F., Lema, J.M., Llompart, M., Garcia-Jares, C., Rodriguez, I., Gomez, M., Ternes, T., 2004. Behavior of pharmaceuticals, cosmetics and hormones in a sewage treatment plant. Water Research 38 (12), 2918e2926. Carballa, M., Omil, F., Lema, J.M., Llompart, M., Garcia-Jares, C., Rodriguez, I., Gomez, M., 2005. Behavior of pharmaceuticals and personal care products in sewage treatment plant of northwest Spain. Water Science & Technology 52 (8), 29e35. Carballa, M., Omil, F., Alder, A.C., Lema, J.M., 2006. Comparison between the conventional anaerobic digestion of sewage sludge and its combination with a chemical or thermal pre-treatment concerning the removal of pharmaceuticals and personal care products. Water Science and Technology 53 (8), 109e117. Carballa, M., Fink, G., Omil, F., Lema, J.M., Ternes, T., 2008. Determination of the solid-water distribution coefficient (Kd) for pharmaceuticals, estrogens and musk fragrances in digested sludge. Water Research 42 (1e2), 287e295. Carberry, J., Englande, A., 1983. Sludge Characteristics and Behavior. Martinus Nijhoff Publishers, Boston, The Hague, Dordrecht, Lancaster. Cardoza, L.A., Knapp, C.W., Larive, C.K., Belden, J.B., Lydy, M.J., Graham, D.W., 2005. Factors affecting the fate of ciprofloxacin in aquatic field systems. Water, Air and Soil Pollution 161 (1e4), 383e398. Castiglioni, S., Bagnati, R., Calamari, D., Fanelli, R., Zuccato, E., 2005. A multiresidue analytical method using solid-phase extraction and high-pressure liquid chromatography tandem mass spectrometry to measure pharmaceuticals of different therapeutic classes in urban wastewaters. Journal of Chromatography A 1092 (2), 206e215. Cha, J.M., Yang, S., Carlson, K.H., 2005. Rapid analysis of trace levels of antibiotic polyether ionophores in surface water by solid-phase extraction and liquid chromatography with ion trap tandem mass spectrometric detection. Journal of Chromatography A 1065 (2), 187e198. Cha, J.M., Yang, S., Carlson, K.H., 2006. Trace determination of beta-lactam antibiotics in surface water and urban wastewater using liquid chromatography combined with electrospray tandem mass spectrometry. Journal of Chromatography A 1115 (1e2), 46e57. Chamberlain, E., Adams, C., 2006. Oxidation of sulfonamides, macrolides, and carbadox with free chlorine and monochloramine. Water Research 40 (13), 2517e2526.

Chang, H., Hu, J.Y., Asami, M., Kunikane, S., 2008. Simultaneous analysis of 16 sulfonamide and trimethoprim antibiotics in environmental waters by liquid chromatography-electrospray tandem mass spectrometry. Journal of Chromatography A 1190 (1e2), 390e393. ChemSW, 2005. Molecular Modelling Pro Plus. Norgwyn Montgomery Software Inc., North Wales, PA. Chen, X.G., Xiao, B.D., Liu, J.T., Fang, T., Xu, X., 2005. Kinetics of the oxidation of MCRR by potassium permanganate. Toxicon 45 (7), 911e917. Choi, K.J., Kim, S.G., Kim, C.W., Kim, S.H., 2007a. Determination of antibiotic compounds in water by on-line SPE-LC/MSD. Chemosphere 66 (6), 977e984. Choi, K.J., Son, H.J., Kim, S.H., 2007b. Ionic treatment for removal of sulfonamide and tetracycline classes of antibiotic. Science of the Total Environment 387 (1e3), 247e256. Choi, K.J., Kim, S.G., Kim, S.H., 2008. Removal of tetracycline and sulfonamide classes of antibiotic compound by powdered activated carbon. Environmental Technology 29 (3), 333e342. Clara, M., Kreuzinger, N., Strenn, B., Gans, O., Kroiss, H., 2005a. The solids retention timeea suitable design parameter to evaluate the capacity of wastewater treatment plants to remove micropollutants. Water Research 39 (1), 97e106. Clara, M., Strenn, B., Gans, O., Martinez, E., Kreuzinger, N., Kroiss, H., 2005b. Removal of selected pharmaceuticals, fragrances and endocrine disrupting compounds in a membrane bioreactor and conventional wastewater treatment plants. Water Research 39 (19), 4797e4807. Cooney, D.O., 1995. Activated Charcoal in Medical Applications. Dekker, New York. Cosar, C., Julou, L., 1959. The activity of 1-(2-hydroxyethyl)-2methyl-5-nitroimidazole (R.P. 8823) against experimental Trichomonas vaginalis infections. Ann Inst Pasteur (Paris) 96 (2), 238e241. Costanzo, S.D., Murby, J., Bates, J., 2005. Ecosystem response to antibiotics entering the aquatic environment. Marine Pollution Bulletin 51 (1e4), 218e223. Crain, F.T., Gottlieb, S., 1935. A study of the mechanism of the effect of chlorine on the biochemical oxygen demand. Transactions of the Kansas Academy of Science 1903 (38), 145e148. De Witte, B., Dewulf, J., Demeestere, K., Van Langenhove, H., 2009. Ozonation and advanced oxidation by the peroxone process of ciprofloxacin in water. Journal of Hazardous Materials 161 (2e3), 701e708. Delgado, J.N., Remers, W.A., 1991. Wilson and Gisvold’s Textbook of Organic Medicinal and Pharmaceutical Chemistry. Lippincott, Philadelphia. Deshpande, A.D., Baheti, K.G., Chatterjee, N.R., 2004. Degradation of beta-lactam antibiotics. Current Science 87 (12), 1684e1695. Diaz-Cruz, M.S., Lopez de Alda, M.J., Barcelo, D., 2003. Environmental behavior and analysis of veterinary and human drugs in soils, sediments and sludge. TrAC-Trends in Analytical Chemistry 22 (6), 340e351. Dickenson, E.R.V., Drewes, J.r.E., Sedlak, D.L., Wert, E.C., Snyder, S.A., 2009. Applying surrogates and indicators to assess removal efficiency of trace organic chemicals during chemical oxidation of wastewaters. Environmental Science & Technology 43 (16), 6242e6247. Dodd, M.C., Buffle, M.-O., von Gunten, U., 2006. Oxidation of antibacterial molecules by aqueous ozone: moiety-specific reaction kinetics and application to ozone-based wastewater treatment. Environmental Science & Technology 40 (6), 1969e1977. Dodd, M.C., Kohler, H.P.E., Von Gunten, U., 2009. Oxidation of antibacterial compounds by ozone and hydroxyl radical: elimination of biological activity during aqueous ozonation

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

processes. Environmental Science & Technology 43 (7), 2498e2504. Dodd, M.C., Huang, C.H., 2007. Aqueous chlorination of the antibacterial agent trimethoprim: reaction kinetics and pathways. Water Research 41 (3), 647e655. Dolar, D., Kosutic, K., Pavlovic, D.M., Kunst, B., 2009. Removal of emerging contaminants of industrial origin by NF/RO e a pilot scale study. Desalination and Water Treatment-Science and Engineering 6 (1e3), 197e203. Donoho, A.L., 1984. Biochemical studies of the fate of monensin in animals and in the environment. Journal of Animal Science 58 (6), 1538e1539. Drewes, J.E., 2007. Analysis, fate and removal of pharmaceuticals in the water cycle. In: Petrovic, M., Barcelo, D. (Eds.), Wilson & Wilsons. Elsevier, Amsterdam, pp. 427e446. Drewes, J.E., Bellona, C., Oedekoven, M., Xu, P., Kim, T.U., Amy, G., 2006. Rejection of wastewater-derived micropollutants in high-pressure membrane applications leading to indirect potable reuse. Environmental Progress 24 (4), 400e409. Drewes, J.E., Sedlak, D.L., Snyder, S., Dickenson, E., 2008. Development of Indicators and Surrogates for Chemical Contaminant Removal during Wastewater Treatment and Reclamation e Final Report. WaterReuse Foundation, Alexandra, VA. Drewes, J.E., Khan, S., 2010. Water reuse for drinking water augmentation. In: Edzwald, J. (Ed.), Water Quality and Treatment Handbook, Sixth ed. McGraw-Hill. Dutta, M., Dutta, N.N., Bhattacharya, K.G., 1999. Aqueous phase adsorption of certain beta-lactam antibiotics onto polymeric resins and activated carbon. Separation and Purification Technology 16 (3), 213e224. Egerton, T.A., Christensen, P.A., Kosa, S.A.M., Onoka, B., Harper, J.C., Tinlin, J.R., 2006. Photoelectrocatalysis by titanium dioxide for water treatment. International Journal of Environmental Pollution 27 (1e3), 2e19. Elovitz, M.S., Von Gunten, U., Kaiser, H.P., 2000. Hydroxyl radical/ ozone ratios during ozonation processes. II. The effect of temperature, pH, alkalinity, and DOM properties. OzoneScience & Engineering 22 (2), 123e150. Esponda, S.M., Padron, M.E.T., Ferrera, Z.S., Rodriguez, J.J.S., 2009. Solid-phase microextraction with micellar desorption and HPLC-fluorescence detection for the analysis of fluoroquinolones residues in water samples. Analytical and Bioanalytical Chemistry 394 (4), 927e935. Exall, K., 2004. A review of water reuse and recycling, with reference to Canadian practice and potential: 2. Applications. Water Quality Research Journal of Canada 39 (1), 13e28. Fatta, D., Nikolaou, A., Achilleos, A., Meric, S., 2007. Analytical methods for tracing pharmaceutical residues in water and wastewater. TrAC-Trends in Analytical Chemistry 26 (6), 515e533. Gartiser, S., Urich, E., Alexy, R., Kummerer, K., 2007. Anaerobic inhibition and biodegradation of antibiotics in ISO test schemes. Chemosphere 66 (10), 1839e1844. Gibs, J., Stackelberg, P.E., Furlong, E.T., Meyer, M., Zaugg, S.D., Lee Lippincott, R., 2007. Persistence of pharmaceuticals and other organic compounds in chlorinated drinking water as a function of time. Science of the Total Environment 373 (1), 240e249. Glassmeyer, S.T., Shoemaker, J.A., 2005. Effects of chlorination on the persistence of pharmaceuticals in the environment. Bulletin of Environmental Contamination and Toxicology 74 (1), 24e31. Gobel, A., McArdell, C.S., Suter, M.J.F., Giger, W., 2004. Trace determination of macrolide and sulfonamide antimicrobials, a human sulfonamide metabolite, and trimethoprim in wastewater using liquid chromatography coupled to

4319

electrospray tandem mass spectrometry. Analytical Chemistry 76 (16), 4756e4764. Gobel, A., Athomsen, A., McArdell, C.S., Joss, A., Giger, W., 2005a. Occurrence and sorption behavior of sulfonamides, macrolides, and trimethoprim in activated sludge treatment. Environmental Science Technology 39 (11), 3981e3989. Gobel, A., Thomsen, A., McArdell, C.S., Alder, A.C., Giger, W., Thei, N., Loffler, D., Ternes, T.A., 2005b. Extraction and determination of sulfonamides, macrolides, and trimethoprim in sewage sludge. Journal of Chromatography A 1085 (2), 179e189. Gobel, A., McArdell, C.S., Joss, A., Siegrist, H., Giger, W., 2007. Fate of sulfonamides, macrolides, and trimethoprim in different wastewater treatment technologies. Science of the Total Environment 372 (2e3), 361e371. Golet, E.M., Alder, A.C., Giger, W., 2002a. Environmental exposure and risk assessment of fluoroquinolone antibacterial agents in wastewater and river water of the Glatt Valley Watershed, Switzerland. Environmental Science & Technology 36 (17), 3645e3651. Golet, E.M., Strehler, A., Alder, A.C., Giger, W., 2002b. Determination of fluoroquinolone antibacterial agents in sewage sludge and sludge-treated soil using accelerated solvent extraction followed by solid-phase extraction. Analytical Chemistry 74 (21), 5455e5462. Golet, E., Xifra, I., Siegrist, H., Alder, A.C., Giger, W., 2003. Environmental exposure assessment of fluoroquinolone antibacterial agents from sewage to soil. Environmental Science Technology 37 (15), 3243e3249. Gros, M., Petrovic, M., Barcelo, D., 2006. Development of a multiresidue analytical methodology based on liquid chromatography-tandem mass spectrometry (LC-MS/MS) for screening and trace level determination of pharmaceuticals in surface and wastewaters. Talanta 70 (4), 678e690. Gulkowska, A., Leung, H.W., So, M.K., Taniyasu, S., Yamashita, N., Yeunq, L.W.Y., Richardson, B.J., Lei, A.P., Giesy, J.P., Lam, P.K.S., 2008. Removal of antibiotics from wastewater by sewage treatment facilities in Hong Kong and Shenzhen, China. Water Research 42 (1e2), 395e403. von Gunten, U., 2003. Ozonation of drinking water: part I. Oxidation kinetics and product formation. Water Research 37 (7), 1443e1467. von Gunten, U., Janex-Habibi, M.-L., Ternes, A.T., Weber, L., 2006. In: Ternes, T.A., Joss, A. (Eds.), Human Pharmaceutical, Hormones and Fragrances: The Challenge of Micropollutants in Urban Water Management. IWA Publishing, London, pp. 293e322. Gustavson, L.E., Kaiser, J.F., Edmonds, A.L., Locke, C.S., DeBartolo, M.L., Schneck, D.W., 1995. Effect of omeprazole on concentrations of clarithromycin in plasma and gastric tissue at steady state. Antimicrobial Agents and Chemotherapy 39 (9), 2078e2083. Halling-Sorensen, B., Lutzhoft, H.C.H., Andersen, H.R., Ingerslev, F., 2000. Environmental risk assessment of antibiotics: comparison of mecillinam, trimethoprim and ciprofloxacin. Journal of Antimicrobial Chemotherapy 46 (Suppl. 1), 53e58. Herna´ndez, F., Sancho, J.V., Iba´n˜ez, M., Guerrero, C., 2007. Antibiotic residue determination in environmental waters by LC-MS. TrAC Trends in Analytical Chemistry 26 (6), 466e485. Hirsch, R., Ternes, T.A., Haberer, K., Mehlich, A., Ballwanz, F., Kratz, K.-L., 1998. Determination of antibiotics in different water compartments via liquid chromatography-electrospray tandem mass spectrometry. Journal of Chromatography A 815 (2), 213e223. Hirsch, R., Ternes, T., Haberer, K., Kratz, K.-L., 1999. Occurrence of antibiotics in the aquatic environment. The Science of the Total Environment 225 (1e2), 109e118.

4320

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Hou, J.P., Poole, J.W., 1971. Beta-lactam antibiotics e their physicochemical properties and biological activities in relation to structure. Journal of Pharmaceutical Sciences 60 (4), 503e532. Huber, M.M., Gobel, A., Joss, A., Hermann, N., Loffler, D., McArdell, C.S., Ried, A., Siegrist, H., Ternes, T.A., von Gunten, U., 2005. Oxidation of pharmaceuticals during ozonation of municipal wastewater effluents: a pilot study. Environmental Science & Technology 39 (11), 4290e4299. Ingerslev, F., Halling-Sørensen, B., 2000. Biodegradation properties of sulfonamides in activated sludge. Environmental Toxicology and Chemistry 19 (10), 2467e2473. Jen, J.F., Lee, H.L., Lee, B.N., 1998. Simultaneous determination of seven sulfonamide residues in swine wastewater by highperformance liquid chromatography. Journal of Chromatography A 793 (2), 378e382. Ji, L.L., Chen, W., Duan, L., Zhu, D.Q., 2009. Mechanisms for strong adsorption of tetracycline to carbon nanotubes: a comparative study using activated carbon and graphite as adsorbents. Environmental Science & Technology 43 (7), 2322e2327. Jones, O.A.H., Voulvoulis, N., Lester, J.N., 2007. The occurrence and removal of selected pharmaceutical compounds in a sewage treatment works utilising activated sludge treatment. Environmental Pollution 145 (3), 738e744. Joss, A., Keller, E., Alder, A.C., Gobel, A., McArdell, C.S., Ternes, T., Siegrist, H., 2005. Removal of pharmaceuticals and fragrances in biological wastewater treatment. Water Research 39 (14), 3139e3152. Jury, K.L., Khan, S.J., Vancov, T., Stuetz, R.M., Ashbolt, N.J. Are sewage treatment plants promoting antibiotic resistance? Critical Reviews in Environmental Science & Technology, in press. Karthikeyan, K.G., Meyer, M.T., 2006. Occurrence of antibiotics in wastewater treatment facilities in Wisconsin. USA. Science of the Total Environment 361 (1e3), 196e207. Khan, S.J., 2009. In: Van den Hoven, T., Kazner, C. (Eds.), TECHNEAU: Safe Drinking Water from Source to Tap. IWA Publishing, London, UK, pp. 445e456. Khan, S.J., Roser, D.J., Davies, C.M., Peters, G.M., Stuetz, R.M., Tucker, R., Ashbolt, N.J., 2008. Chemical contaminants in feedlot wastes: concentrations, effects and attenuation. Environment International 34 (6), 839e859. Khan, S.J., Ongerth, J.E., 2004. Modelling of pharmaceutical residues in Australian sewage by quantities of use and fugacity calculations. Chemosphere 54 (3), 355e367. Khan, S.J., Ongerth, J.E., 2005. Occurrence and removal of pharmaceuticals at an Australian sewage treatment plant. Water 32 (4), 80e85. Kim, S.-C., Carlson, K., 2005. LC-MS2 for quantifying trace amounts of pharmaceutical compounds in soil and sediment matrices. TrAC Trends in Analytical Chemistry 24 (7), 635e644. Kim, S.-C., Carlson, K., 2006. Occurrence of ionophore antibiotics in water and sediments of a mixed-landscape watershed. Water Research 40 (13), 2549e2560. Kim, S., Eichhorn, P., Jensen, J.N., Webber, A.S., Aga, D., 2005. Removal of antibiotics in wastewater: effect of hydraulic and solid retention times on the fate of tetracycline in the activated sludge process. Environmental Science Technology 39 (15), 5816e5823. Kim, S., Jensen, J.N., Aga, D.S., Weber, A.S., 2007. Tetracycline as a selector for resistant bacteria in activated sludge. Chemosphere 66 (9), 1643e1651. Kim, I., Yamashita, N., Tanaka, H., 2009. Performance of UV and UV/H2O2 processes for the removal of pharmaceuticals detected in secondary effluent of a sewage treatment plant in Japan. Journal of Hazardous Materials 166 (2e3), 1134e1140. Kimura, K., Toshima, S., Amy, G., Watanabe, Y., 2004. Rejection of neutral endocrine disrupting compounds (EDCs) and

pharmaceutical active compounds (PhACs) by RO membranes. Journal of Membrane Science 245 (1e2), 71e78. Kirst, H.A., 2002. In: Schonfeld, W., Kirst, H.A. (Eds.), Macrolide Antibiotics. Birkhauser Verlag, Basel, pp. 1e14. Kobayashi, Y., Yasojima, M., Komori, K., Suzuki, Y., Tanaka, H., 2006. Removal characteristics of human antibiotics during wastewater treatment in Japan. Water Practice & Technology 1 (3), 1e9. Kosutic, K., Dolar, D., Asperger, D., Kunst, B., 2007. Removal of antibiotics from a model wastewater by RO/NF membranes. Separation and Purification Technology 53 (3), 244e249. Kummerer, K., 2003. Significance of antibiotics in the environment. Journal of Antimicrobial Chemotherapy 52 (1), 5e7. Kummerer, K., Al-Ahmad, A., Mersch-Sundermann, V., 2000. Biodegradability of some antibiotics, elimination of the genotoxicity and affection of wastewater bacteria in a simple test. Chemosphere 40 (7), 701e710. Lange, F., Cornelissen, S., Kubac, D., Sein, M.M., von Sonntag, J., Hannich, C.B., Golloch, A., Heipieper, H.J., Moder, M., von Sonntag, C., 2006. Degradation of macrolide antibiotics by ozone: a mechanistic case study with clarithromycin. Chemosphere 65 (1), 17e23. Le-Minh, N., Coleman, H.M., Khan, S.J., van Luer, Y., Trang, T.T.T., Watkins, G., Stuetz, R.M., 2010. The application of membrane bioreactors as decentralised systems for removal of endocrine disrupting chemicals and pharmaceuticals. Water Science and Technology 61 (5), 1081e1088. Leiros, H.K.S., Kozielski-Stuhrmann, S., Kapp, U., Terradot, L., Leonard, G.A., McSweeney, S.M., 2004. Structural basis of 5nitroimidazole antibiotic resistance e the crystal structure of NimA from Deinococcus radiodurans. Journal of Biological Chemistry 279 (53), 55840e55849. Levine, A.D., Meyer, M.T., Kish, G., 2006. Evaluation of the persistence of micropollutants through pure-oxygen activated sludge nitrification and denitrification. Water Environment Research 78 (11), 2276e2285. Li, S.Z., Li, X.Y., Wang, D.Z., 2004. Membrane (RO-UF) filtration for antibiotic wastewater treatment and recovery of antibiotics. Separation and Purification Technology 34 (1e3), 109e114. Li, J.D., Cai, Y.Q., Shi, Y.L., Mou, S.F., Jiang, G.B., 2007. Determination of sulfonamide compounds in sewage and river by mixed hemimicelles solid-phase extraction prior to liquid chromatography-spectrophotometry. Journal of Chromatography A 1139 (2), 178e184. Li, D., Yang, M., Hu, J., Zhang, Y., Chang, H., Jin, F., 2008. Determination of penicillin G and its degradation products in a penicillin production wastewater treatment plant and the receiving river. Water Research 42 (1e2), 307e317. Li, B., Zhang, T., Xu, Z.Y., Fang, H.H.P., 2009. Rapid analysis of 21 antibiotics of multiple classes in municipal wastewater using ultra performance liquid chromatography-tandem mass spectrometry. Analytica Chimica Acta 645 (1e2), 64e72. Lin, A.Y.C., Yu, T.H., Lateef, S.K., 2009. Removal of pharmaceuticals in secondary wastewater treatment processes in Taiwan. Journal of Hazardous Materials 167 (1e3), 1163e1169. Lindberg, R.H., Wennberg, P., Johansson, M.I., Tysklind, M., Andersson, B.A.V., 2005. Screening of human antibiotic substances and determination of weekly mass flows in five sewage treatment plants in Sweden. Environmental Science & Technology 39 (10), 3421e3429. Lindberg, R., Olofsson, U., Rendahl, O., Tysklind, M., Andersson, B.A.V., 2006. Behavior of fluoroquinolones and trimethoprim during mechanical, chemical, and active sludge treatment of sewage water and digestion of sludge. Environmental Science Technology 40 (3), 1042e1048.

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Lo¨ffler, D., Ternes, T.A., 2003. Analytical method for the determination of the aminoglycoside gentamicin in hospital wastewater via liquid chromatography-electrospray-tandem mass spectrometry. Journal of Chromatography A 1000 (1e2), 583e588. Loftin, K.A., Adams, C.D., Meyer, M.T., Surampalli, R., 2008. Effects of ionic strength, temperature, and pH on degradation of selected antibiotics. Journal of Environmental Quality 37 (2), 378e386. Loke, M.L., Tjornelund, J., Halling-Sorensen, B., 2000. Determination of the distribution coefficient (logKd) of oxytetracycline, tylosin A, olaquindox and metronidazole in manure. Chemosphere 48 (3), 351e361. Maria, C.R., Reginald, F.C., 1993. Capillary electrophoresis separation of sulphonamides and dihydrofolate reductase inhibitors. Journal of Microcolumn Separations 5 (3), 207e215. Marzo, A., Dal Bo, L., 1998. Chromatography as an analytical tool for selected antibiotic classes: a reappraisal addressed to pharmacokinetic applications. Journal of Chromatography A 812 (1e2), 17e34. Masters, P.A., O’Bryan, T.A., Zurlo, J., Miller, D.Q., Joshi, N., 2003. Trimethoprimesulfamethoxazole revisited. Archives of Internal Medicine 163 (4), 402e410. McArdell, C.S., Molnar, E., Suter, M.J.F., Giger, W., 2003. Occurrence and fate of macrolide antibiotics in wastewater treatment plants and in the Glatt Valley watershed. Switzerland. Environmental Science Technology 37 (24), 5479e5486. Metcalf, Eddy, 2007. Water Reuse: Issues, Technologies, and Applications. McGraw Hill, New York. Miao, X.S., Bishay, F., Chen, M., Metcalfe, C.D., 2004. Occurrence of antimicrobials in the final effluents of wastewater treatment plants in Canada. Environmental Science Technology 38 (13), 3533e3541. Morse, A., Jackson, A., 2004. Fate of amoxicillin in two water reclamation systems. Water Air and Soil Pollution 157 (1e4), 117e132. Murray, C.A., Parsons, S.A., 2006. Preliminary laboratory investigation of disinfection by-product precursor removal using an advanced oxidation process. Water and Environment Journal 20 (3), 123e129. Nagulapally, S.R., Ahmad, A., Henry, A., Marchin, G.L., Zurek, L., Bhandari, A., 2009. Occurrence of ciprofloxacin-, trimethoprim-sulfamethoxazole-, and vancomycin-resistant bacteria in a municipal wastewater treatment plant. Water Environment Research 81 (1), 82e90. Neu, H.C., 1992. The crisis in antibiotic resistance. Science 257 (5073), 1064e1073. Nghiem, L.D., Schafer, A.I., Elimelech, M., 2004. Removal of natural hormones by nanofiltration membranes: measurement, modeling, and mechanisms. Environmental Science & Technology 38 (6), 1888e1896. Nghiem, L.D., Schafer, A.I., Elimelech, M., 2005. Pharmaceutical retention mechanisms by nanofiltration membranes. Environmental Science & Technology 39 (19), 7698e7705. Nghiem, L.D., Oschmann, N., Schafer, A.I., 2006. Fouling in greywater recycling by direct ultrafiltration. Desalination 187 (1e3), 283e290. Nghiem, D.L., Vogel, D., Khan, S.J., 2009. Characterising humic acid fouling of nanofiltration membranes using bisphenol A as a molecular indicator. Water Research 43 (4), 863e881. Nghiem, L.D., Hawkes, S., 2007. Effects of membrane fouling on the nanofiltration of pharmaceutically active compounds (PhACs): mechanisms and role of membrane pore size. Separation and Purification Technology 57 (1), 176e184. Nghiem, L.D., Schafer, A.I., 2006. Fouling autopsy of hollow-fibre MF membranes in wastewater reclamation. Desalination 188 (1e3), 113e121.

4321

Oschmann, N., Nghiem, L.D., Schafer, A.I., 2005. Fouling mechanisms of submerged ultrafiltration membranes in greywater recycling. Desalination 179 (1e3), 215e223. Paxeus, N., 2004. Removal of selected non-steroidal antiinflammatory drugs (NSAIDs), gemfibrozil, carbamazepine, beta-blockers, trimethoprim and triclosan in conventional wastewater treatment plants in five EU countries and their discharge to the aquatic environment. Water Science and Technology 50 (5), 253e260. Peng, X., Wang, Z., Kuang, W., Tan, J., Li, K., 2006. A preliminary study on the occurrence and behavior of sulfonamides, ofloxacin and chloramphenicol antimicrobials in wastewaters of two sewage treatment plants in Guangzhou, China. Science of the Total Environment 371 (1e3), 314e322. Peng, X., Tan, J., Tang, C., Yu, Y., Wangt, Z., 2008. Multiresidue determination of fluoroquinolone, sulfonamide, trimethoprim, and chloramphenicol antibiotics in urban waters in China. Environmental Toxicology and Chemistry 27 (1), 73e79. Perez, S., Eichhorn, P., Aga, D.S., 2005. Evaluating the biodegradability of sulfamethazine, sulfamethoxazole, sulfathiazole and trimethoprim at different stages of sewage treatment. Environmental Toxicology and Chemistry 24 (6), 1361e1367. du Pisani, P.L., 2006. Direct reclamation of potable water at Windhoek’s Goreangab reclamation plant. Desalination 188 (1e3), 79e88. Porter, M.C., 1990. Handbook of Industrial Membrane Technology. Noyes Publications, Park Ridge, New Jersey, USA. Powell, S.J., Wilmot, A.J., Macleod, I., Elsdon-Dew, R., 1966. Metronidazole in amoebic dysentery and amoebic liver abscess. The Lancet 2 (7477), 1329e1331. Putra, E.K., Pranowo, R., Sunarso, J., Indraswati, N., Ismadji, S., 2009. Performance of activated carbon and bentonite for adsorption of amoxicillin from wastewater: mechanisms, isotherms and kinetics. Water Research 43 (9), 2419e2430. Quiang, Z., Adams, C., 2004. Potentiometric determination of acid dissociation constants (pK(a)) for human and veterinary antibiotics. Water Research 38 (12), 2874e2890. Rabolle, M., Spliid, N.H., 2000. Sorption and mobility of metronidazole, olaquindox, oxytetracycline and tylosin in soil. Chemosphere 40 (7), 715e722. Radjenovic, J., Godehardt, M., Petrovic, M., Hein, A., Farre, M., Jekel, M., Barcelo, D., 2009. Evidencing generation of persistent ozonation products of antibiotics roxithromycin and trimethoprim. Environmental Science & Technology 43 (17), 6808e6815. Reif, R., Suarez, S., Omil, F., Lema, J.M., 2008. Fate of pharmaceuticals and cosmetic ingredients during the operation of a MBR treating sewage. Desalination 221 (1e3), 511e517. Renew, J.E., Huang, C.H., 2004. Simultaneous determination of fluoroquinolone, sulfonamide, and trimethoprim antibiotics in wastewater using tandem solid phase extraction and liquid chromatography-electrospray mass spectrometry. Journal of Chromatography A 1042 (1e2), 113e121. Rivera-Utrilla, J., Prados-Joya, G., Sanchez-Polo, M., FerroGarcia, M.A., Bautista-Toledo, I., 2009. Removal of nitroimidazole antibiotics from aqueous solution by adsorption/bioadsorption on activated carbon. Journal of Hazardous Materials 170 (1), 298e305. Robberson, K.A., Waghe, A.B., Sabatini, D.A., Butler, E.C., 2006. Adsorption of the quinolone antibiotic nalidixic acid onto anion-exchange and neutral polymers. Chemosphere 63 (6), 934e941. Roberts, P.H., Thomas, K.V., 2006. The occurrence of selected pharmaceuticals in wastewater effluent and surface waters of

4322

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

the lower Tyne catchment. Science of the Total Environment 356 (1e3), 143e153. Rodriguez, C., Van Buynder, P., Lugg, R., Blair, P., Devine, B., Cook, A., Weinstein, P., 2009. Indirect potable reuse: a sustainable water supply alternative. International Journal of Environmental Research and Public Health 6 (3), 1174e1209. Rosenfeldt, E.J., Linden, K.G., 2004. Degradation of endocrine disrupting chemicals bisphenol A, ethinyl estradiol, and estradiol during UV photolysis and advanced oxidation processes. Environmental Science & Technology 38 (20), 5476e5483. Salisbury, C.D.C., 1995. In: Oka, H., Nakazawa, H., Harada, K.E., MacNeil, J.D. (Eds.), Chemical Analysis for Antibiotics Used in Agriculture. AOAC International, Toronto. Schaep, J., Vandecasteele, C., Mohammed, A.W., Bowen, W.R., 2001. Modelling the retention of ionic components for different nanofiltration membranes. Separation and Purification Technology 22e23, 169e179. Schafer, A.I., Kalinowski, J., Puhler, A., 1994. Increased fertility of Corynebacterium glutamicum recipients in intergeneric matings with Escherichia coli after stress exposure. Applied and Environmental Microbiology 60 (2), 756e759. Schafer, A.I., Nghiem, L.D., Waite, T.D., 2003. Removal of the natural hormone estrone from aqueous solutions using nanofiltration and reverse osmosis. Environmental Science and Technology 37 (1), 182e188. Schlu¨sener, M.P., Bester, K., Spiteller, M., 2003. Determination of antibiotics such as macrolides, ionophores and tiamulin in liquid manure by HPLC-MS/MS. Analytical and Bioanalytical Chemistry 375 (7), 942e947. Schneider, J., 1961. Treatment of giardiasis (lambliasis) by metronidazole. Bulletin de la Sociiti. de Pathologie Exotique 54, 84e95. Segura, P.A., Garcia-Ac, A., Lajeunesse, A., Ghosh, D., Gagnon, C., Sauve, S., 2007. Determination of six anti-infectives in wastewater using tandem solidphase extraction and liquid chromatography-tandem mass spectrometry. Journal of Environmental Monitoring 9 (4), 307e313. Shemer, H., Kunukcu, Y.K., Linden, K.G., 2006. Degradation of the pharmaceutical metronidazole via UV, Fenton and photoFenton processes. Chemosphere 63 (2), 269e276. Shi, L., Zhou, X.F., Zhang, Y.L., Gu, G.W., 2009. Simultaneous determination of 8 fluoroquinolone antibiotics in sewage treatment plants by solid-phase extraction and liquid chromatography with fluorescence detection. Water Science and Technology 59 (4), 805e813. Simon, A., Nghiem, L.D., Le-Clech, P., Khan, S.J., Drewes, J.E., 2009. Effects of membrane degradation on the removal of pharmaceutically active compounds (PhACs) by NF/RO filtration processes. Journal of Membrane Science 340 (1e2), 16e25. Skold, O., 2001. Resistance to trimethoprim and sulfonamides. Veterinary Research 32 (3e4), 261e273. Sneader, W., 2005. Drug Discovery: a History. John Wiley & Sons, West Sussex, England. Snyder, S.A., Westerhoff, P., Yoon, Y., Sedlak, D.L., 2003. Pharmaceuticals, personal care products, and endocrine disruptors in water: implications for the water industry. Environmental Engineering Science 20 (5), 449e469. Sponza, D.T., Demirden, P., 2007. Treatability of sulfamerazine in sequential upflow anaerobic sludge blanket reactor (UASB)/ completely stirred tank reactor (CSTR) processes. Separation and Purification Technology 56 (1), 108e117. Staehelin, J., Hoigne, J., 1985. Decomposition of ozone in water in the presence of organic solutes acting as promoters and inhibitors of radical chain reactions. Environmental Science & Technology 19 (12), 1206e1213. Stoob, K., Singer, H.P., Goetz, C.W., Ruff, M., Mueller, S.R., 2005. Fully automated online solid phase extraction coupled directly

to liquid chromatography-tandem mass spectrometry: quantification of sulfonamide antibiotics, neutral and acidic pesticides at low concentrations in surface waters. Journal of Chromatography A 1097 (1e2), 138e147. Szczepanowski, R., Linke, B., Krahn, I., Gartemann, K.H., Gutzkow, T., Eichler, W., Puhler, A., Schluter, A., 2009. Detection of 140 clinically relevant antibiotic-resistance genes in the plasmid metagenome of wastewater treatment plant bacteria showing reduced susceptibility to selected antibiotics. Microbiology-Sgm 155, 2306e2319. Tadkaew, N., Sivakumar, M., Khan, S.J., McDonald, J.A., Nghiem, L.D., 2010. Effect of mixed liquor pH on the removal of trace organic contaminants in a membrane bioreactor. Bioresource Technology 101 (5), 1494e1500. Tauxe-Wuersch, A., De Alencastro, L.F., Grandjean, D., Tarradellas, J., 2005. Occurrence of several acidic drugs in sewage treatment plants in Switzerland and risk assessment. Water Research 39 (9), 1761e1772. Ternes, T.A., Stuber, J., Herrmann, N., McDowell, D., Ried, A., Kampmann, M., Teiser, B., 2003. Ozonation: a tool for removal of pharmaceuticals, contrast media and musk fragrances from wastewater? Water Research 37 (8), 1976e1982. Theron, M.M., Janse van Rensburg, M.N., Chalkley, L.J., 2004. Nitroimidazole resistance genes (nimB) in anaerobic Grampositive cocci (previously Peptostreptococcus spp.). Journal Antimicrobial Agents Chemotherapy 54 (1), 240e242. Todar, K., 2002. In: Todar, K. (Ed.), Todars Online Textbook of Bacteriology. Tolls, J., 2001. Sorption of veterinary pharmaceuticals in soils: a review. Environmental Science and Technology 35 (17), 3397e3406. Turiel, E., Bordin, G., Rodrı´guez, A.R., 2003. Trace enrichment of (fluoro)quinolone antibiotics in surface waters by solid-phase extraction and their determination by liquid chromatographyultraviolet detection. Journal of Chromatography A 1008 (2), 145e155. Upcroft, J.A., Campbell, R.W., Benakli, K., Upcroft, P., Vanelle, P., 1999. Efficacy of new 5-nitroimidazoles against metronidazole-susceptible and -resistant Giardia, Trichomonas, and Entamoeba spp. Antimicrobial Agents and Chemotherapy 43 (1), 73e76. Verliefde, A.R.D., Cornelissen, E.R., Heijman, S.G.J., Verberk, J.Q.J.C., Amy, G.L., Van der Bruggen, B., van Dijk, J.C., 2008. The role of electrostatic interactions on the rejection of organic solutes in aqueous solutions with nanofiltration. Journal of Membrane Science 322 (1), 52e66. Vieno, N., Tuhkanen, T., Kronberg, L., 2007a. Elimination of pharmaceuticals in sewage treatment plants in Finland. Water Research 41 (5), 1001e1012. Vieno, N.M., Harkki, H., Tuhkanen, T., Kronberg, L., 2007b. Occurrence of pharmaceuticals in river water and their elimination a pilot-scale drinking water treatment plant. Environmental Science & Technology 41 (14), 5077e5084. Vigneswaran, S., Sundaravadivel, M., 2004. In: Saravanamuthu, V. (Ed.), Wastewater Recycle, Reuse and Reclamation. Eolss Publishers, Oxford, UK. Watkinson, A.J., Murbyc, E.J., Costanzo, S.D., 2007. Removal of antibiotics in conventional and advanced wastewater treatment: implications for environmental discharge and wastewater recycling. Water Research 41 (18), 4164e4176. Watkinson, A.J., Murby, E.J., Kolpin, D.W., Costanzo, S.D., 2009. The occurrence of antibiotics in an urban watershed: from wastewater to drinking water. Science of the Total Environment 407 (8), 2711e2723. Westerhoff, P., Yoon, Y., Snyder, S.A., Wert, E.C., 2005. Fate of endocrine-disruptor, pharmaceutical, and personal care product chemicals during simulated drinking water treatment processes. Environmental Science & Technology 39 (17), 6649e6663.

w a t e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 2 9 5 e4 3 2 3

Wintgens, T., Melin, T., Schafer, A., Khan, S., Muston, M., Bixio, D., Thoeye, C., 2005. The role of membrane processes in municipal wastewater reclamation and reuse. Desalination 17 (178), 1e11. Xiao, Y., Chang, H., Jia, A., Hu, J.Y., 2008. Trace analysis of quinolone and fluoroquinolone antibiotics from wastewaters by liquid chromatography-electrospray tandem mass spectrometry. Journal of Chromatography A 1214 (1e2), 100e108. Xu, P., Drewes, J.E., Kim, T.U., Bellona, C., Amy, G., 2006. Effect of membrane fouling on transport of organic contaminants in NF/RO membrane applications. Journal of Membrane Science 279 (1e2), 165e175. Xu, W.H., Zhang, G., Li, X.D., Zou, S.C., Li, P., Hu, Z.H., Li, J., 2007. Occurrence and elimination of antibiotics at four sewage treatment plants in the Pearl River Delta (PRD), South China. Water Research 41 (19), 4526e4534. Yang, S., Carlson, K.H., 2004. Solid-phase extraction-highperformance liquid chromatography-ion trap mass spectrometry for analysis of trace concentrations of macrolide

4323

antibiotics in natural and waste water matrices. Journal of Chromatography A 1038 (1e2), 141e155. Yang, S., Cha, J., Carlson, K., 2005. Simultaneous extraction and analysis of 11 tetracycline and sulfonamide antibiotics in influent and effluent domestic wastewater by solid-phase extraction and liquid chromatography-electrospray ionization tandem mass spectrometry. Journal of Chromatography A 1097 (1e2), 40e53. Yasojima, M., Nakada, N., Komori, K., Suzuki, Y., Tanaka, H., 2006. Occurrence of levofloxacin, clarithromycin and azithromycin in wastewater treatment plant in Japan. Water Science & Technology 53 (11), 227e233. Zorita, S., Martensson, L., Mathiasson, L., 2009. Occurrence and removal of pharmaceuticals in a municipal sewage treatment system in the south of Sweden. Science of the Total Environment 407 (8), 2760e2770. Zuccato, E., Castiglioni, S., Fanelli, R., 2005. Identification of the pharmaceuticals for human use contaminating the Italian aquatic environment. Journal of Hazardous Materials 122 (3), 205e209.