antibióticos

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Journal of Environmental Management 92 (2011) 2304e2347

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Journal of Environmental Management

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Review

Degradation and removal methods of antibiotics from aqueousmatrices e A review

Vera Homem, Lúcia Santos*

LEPÆ, Departamento de Engenharia Química, Faculdade de Engenharia da Universidade do Porto, Rua Dr. Roberto Frias, 4200-465 Porto, Portugal

a r t i c l e i n f o

Article history:Received 1 November 2010Received in revised form2 May 2011Accepted 18 May 2011Available online 16 June 2011

Keywords:AntibioticsEmergent pollutantsDegradation/removal processes

* Corresponding author.E-mail address: [email protected] (L. Santos).

0301-4797/$ e see front matter � 2011 Elsevier Ltd.doi:10.1016/j.jenvman.2011.05.023

a b s t r a c t

Over the past few years, antibiotics have been considered emerging pollutants due to their continuousinput and persistence in the aquatic ecosystem even at low concentrations. They have been detectedworldwide in environmental matrices, indicating their ineffective removal from water and wastewaterusing conventional treatment methods. To prevent this contamination, several processes to degrade/remove antibiotics have been studied. This review addresses the current state of knowledge concerningthe input sources, occurrence and mainly the degradation and removal processes applied to a specificclass of micropollutants, the antibiotics. In this paper, different remediation techniques were evaluatedand compared, such as conventional techniques (biological processes, filtration, coagulation, flocculationand sedimentation), advanced oxidation processes (AOPs), adsorption, membrane processes andcombined methods. In this study, it was found that ozonation, Fenton/photo-Fenton and semiconductorphotocatalysis were the most tested methodologies. Combined processes seem to be the best solution forthe treatment of effluents containing antibiotics, especially those using renewable energy and by-products materials.

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1. Introduction

The presence of pharmaceutical compounds, namely antibiotics,in the ecosystem has been known for almost 30 years. However, itwas only in mid-1990s, when the use of these compounds waswidespread and new analytical technologies were developed, thattheir presence became an emerging concern (Lissemore et al., 2006;Hernando et al., 2006; Bound and Voulvoulis, 2006). Residues ofhuman and veterinary antibiotics were detected in a multiplicity ofmatrices (Ternes, 1998; Hirsch et al., 1999; Lindsey et al., 2001;Sacher et al., 2001; Díaz-Cruz et al., 2003; Jacobsen et al., 2004;Batt et al., 2006; Brown et al., 2006; Cha et al., 2006; Kim andCarlson, 2006; Díaz-Cruz and Barceló, 2007; Bailón-Pérez et al.,2008; Feitosa-Felizzola and Chiron, 2009; Minh et al., 2009; Mom-pelat et al., 2009). The introduction of these compounds into theenvironmental through anthropogenic sources can constitutea potential risk for aquatic and terrestrial organisms. Althoughpresent at vestigial levels, antibiotics may cause resistance inbacterial populations,making them, in the near future, ineffective inthe treatment of several diseases (Schwartz et al., 2003, 2006;Baquero et al., 2008; Rosenblatt-Farrel, 2009; Martínez, 2009).

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Review articles about the input, occurrence and effects of anti-biotics in the environment (Kemper, 2008; Kümmerer, 2009) andabout the analytical methodologies for determination of thesekinds of compounds in aqueous matrices (Petrovi�c et al., 2005; Haoet al., 2007) have been recently published. To prevent environ-mental matrices contamination, several processes to degrade/remove antibiotics have been studied. According to the authors’knowledge, four review articles on the oxidation technologies forthe removal of several pharmaceuticals were published (Ikehataet al., 2006; Esplugas et al., 2007; Sharma, 2008; Klavarioti et al.,2009), as well as a general review of antibiotics in the aquaticenvironment, which refers possible disposal methodologies(Kümmerer, 2009). However, the authors did not find a reviewarticle concerning different types of methodologies to the antibi-otics removal. Therefore, the aim of this work is to review, evaluateand compare different developed processes for the degradation andremoval of antibiotics in aqueous matrices.

2. Background

2.1. Antibiotics classification

Traditionally antibiotics are defined as chemical compoundsthat eradicate or inhibit the growth of other microorganisms

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Table 1Principal classes of antibiotics.

Class Core structure

Aminoglycosides

Consists of 2 or more amino sugars joined bya glycoside linkage to a hexose nucleus ofthe drug. The structure of these antibiotics isderived from the these two molecules:

OH

OH

NH2

R

H2N

HO

NHHO

HO

O

HN

OHOH3C

HO

O

R

OH2C

HO

OH

NHCH3

HONH

NH2

H2NNH

Anthracyclines They are structurally glycosylated tetracyclines.

O

O

R2

R1

R3

OH

R4

R7

R6

R5

b-Lactams

CarbapenemsThey are structurally very similar to the penicillins,but the sulphur atom of the structure has beenreplaced by a carbon atom.

NO

COOR2

NH

R1

O CH3

CH3

CephalosporinsThey possess a cephem nucleus to which 2 sidechains are linked: one esterifies the carbamategroup (R1) and the other is linked to the nucleus (R3).

N

S

OCOOR2

R3

NH

R1

O

MonobactamsIn these compounds, the b-lactam ring is alone andnot fused to another ring.

NR5

O

R2 R3R4N

H

R1

PenicillinsConsists of a thiazolidine ring connected to ab-lactam ring, to which a side chain is attached.

N

S

OCOOR2

NH

R1

O CH3

CH3

GlycopeptidesThey are composed for carbohydrate moieties(glycans) covalently attached to theside chains of an amino acid.

O O

OR1 R2

R4R3O R5

NH

NHN

NH

HN

NH

O

O

O

O

O

R10

R11

HO OR9

NH

R8O

OO

O

R12

R13O

R14 OR15

R16

R6

R7

Imidazoles

They are heterocyclic compounds of 5 memberdi-unsaturated ring structure with 2 nitrogenatoms at nonadjacent positions. If R2]NO2

than it is a nitroimidazole compound. N

NH3C R2

R1

LincosamidesThey are a small family of antibiotics that havecarbohydrate-type structures.

NCH3

R1

NH

O

O

HOOH

OH

SR4

CH3R3

R4

(continued on next page)

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Table 1 (continued ).


Macrolides They are highly substituted monocyclic lactonewith 1 or more saccharides glycosidicallyattached to hydroxyl groups. The lactone ringsare usually 12, 14 or 16-membered.

O

O

O

CH3

CH3

CH3

HC

R1

H3C

CH3

R2O

O

CH3

N(CH3)2HO

O

O

O

CH3

CH2

CH2CH3HO

CH3O

H2C

H3C

OH3C

R3O

N(CH3)2

OH

O

O

CH3HO

R1O

R2O

O

O

CH3

OHHO

CH3

O

R1

O

H3C CH3

H3CR2

CH3R3

H3CH2C O

O

CH2R4

OH

H3CO

CH3

N(CH3)2HO

O

CH3

OH

H3C

R5O

Polyethers

They are characterized by multiple tetrahydrofuranand tetrahydropyran rings connected byaliphatic bridges, direct CeC linkages, or spiro linkages. Otherfeatures include a free carboxyl function, manylower alkyl groups, and a variety of functional oxygen groups.

O

O

H3C CH3

OH

OH

H3C

O

H3CH2C

O

CH3

O

HO CH3

CH3

H3C

O CH3

HOOC

Polypeptides They are polymers formed from the linkageof a-amino acids.

QuinolonesTheir structure contains 2 fused rings with acarboxylic acid and a ketone group.If R4]F, then it is a fluoroquinolone compound. N

OH

O

R1

OR5R4

R3R2

Quinoxalinederivative

Their structure contains a benzene ringand a pyrazine derivative ring.

N+

N+ RO-

O-

SulfonamidesThey are characterized by sulfonyl group connectedto an amine group.

SNH

R1

O O

NH

R2

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TetracyclinesThese antibiotics contain an octrahydronaphtacenering skeleton, consisting of 4 fused rings.

R1

OH

R4

OHOH

O O

R2 R3N

CH3H3C

OH

CONHR5Other antibiotics

ChloramphenicolIt contains a nitrobenzene moiety connected toa propanol group as well as an amino groupbinding a derivative of dichloroacetic acid.

O2N

HN

Cl

OH

OHO

Cl

Mitomycins

They have a unique chemical structure, inwhich 3 different functional groups e aziridine,carbamate and quinone e are arranged arounda pyrro[1,2-a]indole nucleus.

N

O

O

H3C

R1 R2

R3

O

NH2O

TrimethoprimIt is a diaminopyrimidine, a structural analogueof the pteridine moiety of folic acid.

N

N

NH2

H2N

OCH3

OCH3

OCH3


(Marzo and Bo, 1998). However, the term “antibiotic” has beenexpanded for antibacterial, antiviral, antifungal and antitumourcompounds. Most of these substances have a microbial origin, buttheymay be also semi-synthetic or totally synthetic. Antibiotics canbe divided into several classes, according to different criteria:spectrum, mechanism of action or chemical structure. In Table 1 arepresented the principal classes of antibiotics, according to theirchemical structure (Marzo and Bo, 1998; Bannister, 2000; Lindneret al., 2000; McGregor, 2000; Ponsford, 2000; Roberts, 2000;Southgate and Osborne, 2000; Cavalleri and Parenti, 2001; Kirst,2001; Mor, 2001; Kadow et al., 2002; Brimble, 2003; Weidner-Wells and Macielag, 2003; Sum, 2004; Ohno et al., 2010). Thisclassification will be used later in the efficiency comparison of theremoval/degradation methodologies.

2.2. Sources of antibiotics in the environment

In these last years, the use of antibiotics in veterinary andhuman medicine was widespread (annual consumption of 100000e200 000 tons) and consequently, the possibility of watercontaminationwith such compounds increased (Xu et al., 2007). Asmentioned above, human and veterinary antibiotics have beendetected in different matrices. These pollutants are continuallydischarged into the natural environment as parent compounds,metabolites/degradation products or both forms by a diversity ofinput sources as shown in Fig. 1.

When dispersed in the fields as fertilizer, manure can contam-inate soil and consequently surface and groundwater through run-off or leaching (Kemper, 2008; Farré et al., 2008; Díaz-Cruz et al.,2003). Similarly, human antibiotics are introduced into the

environment through excretion (urine and faeces), entering in thesewer network and reaching the wastewater treatment plants(WWTPs). Most of WWTPs are not designed to remove highly polarmicropollutants like antibiotics (Xu et al., 2007). Therefore, they canbe transported to surface waters and reach groundwater afterleaching. Ultimately, the contaminated surface waters can enter inthe drinking water treatment plants (DWTPs), which are also notprepared to remove these compounds, reaching the water distri-bution systems. The sludge produced in WWTPs is applied in thesoil fertilization and may cause the same problems as the use ofmanure, as explained above. Another important contaminationsource is the direct release of veterinary antibiotics through theapplication in aquaculture. Improper disposal of unused/expireddrugs, which are directly discharged in the sewage network ordeposited in landfills, waste effluents from manufacture or acci-dental spills during manufacturing or distribution can also beconsidered as significant points of contamination (Mompelat et al.,2009; Díaz-Cruz et al., 2003).

2.3. Occurrence in the environment

In the last years, the presence of antibiotics in environmentalmatrices has been investigated. Actually, the first reported case ofwater contamination (surface water) by antibiotics was in Englandin 1982, when Watts et al. detected macrolides, tetracyclines andsulphonamides in a river at concentrations of 1 mg/L (Sarmah et al.,2006). After this case, several studies about the occurrence ofantibiotic residues in aquatic ecosystems have been reported:surface waters (Constanzo et al., 2005; Ferdig et al., 2005; Brownet al., 2006; Cha et al., 2006; Kim and Carlson, 2006; Xu et al.,

Fig. 1. Origin and principal contamination routes of human and veterinary antibiotics.


2007; Smith et al., 2007; Tamtam et al., 2008; Feitosa-Felizzola andChiron, 2009; Watkinson et al., 2009), groundwaters (Sacher et al.,2001; Batt et al., 2006; Xu et al., 2007), sea waters (Xu et al., 2007;Minh et al., 2009), drinking water (Watkinson et al., 2009; Yiruhanet al., 2010), WWTPs effluents (McArdell et al., 2003; Ferdig et al.,2005; Brown et al., 2006; Cha et al., 2006; Seifrtová et al., 2008;Watkinson et al., 2009; Minh et al., 2009) and hospital wastewa-ters (Kümmerer, 2001; Lindberg et al., 2004; Ferdig et al., 2005;Brown et al., 2006; Batt et al., 2007; Seifrtová et al., 2008;Watkinson et al., 2009). Antibiotics have also been detected interrestrial matrices and biosolids (Jacobsen et al., 2004; Ferdig et al.,2005; Pei et al., 2006; Kim and Carlson, 2006; Kinney et al., 2006;Jones-Lepp and Stevens, 2007; Feitosa-Felizzola and Chiron, 2009).

Usually, antibiotics are detected in the higher mg/L range inhospital effluents, lower mg/L range in municipal wastewater andng/L in surface, sea and groundwater. Moreover, it has been verifiedthat sediments from agriculture-influenced rivers have higherantibiotic concentrations than the overlying water matrix or thanother sediments from rivers located far from agricultural areas. Thisindicates the possibility of run-off contamination from farmland(Kümmerer, 2009).

Soil analyses have also revealed the presence of measurableantibiotic residues in this matrix for several months, following theapplication of manure or sludge as fertilizers. Positive findings ofthese residues have also been reported in vegetables and cerealssuch as carrots, lettuces, green onions, cabbages, cucumbers andcorn (Migliore et al., 2003; Kumar et al., 2005; Boxall et al., 2006;Grote et al., 2007; Dolliver et al., 2007; Shenker et al., 2011).

The accumulation and persistence of antibiotics in the envi-ronment can produce harmful effects, either in aquatic or terres-trial ecosystems, even in the low concentrations levels, in whichthey are detected (ng/L to mg/L for water matrices and low-medium mg/kg for sediments). The extensive and indiscriminateuse of these compounds in human and veterinary medicine andtheir continual introduction into the environmental matrices mayexplain such bioaccumulation and pseudo-persistence. The highlypolarity and non-volatile nature of most antibiotics prevent theirescape from these matrices (Hernando et al., 2006). The physico-chemical properties of each antibiotic (e.g., molecular structure,size and shape, solubility and hydrophobicity) will define theirdistribution in the environmental matrices (solids or water)(Kemper, 2008). In addition to these facts, they are suspected to beresponsible for the production of resistant microorganisms,causing serious problems of public health, namely difficulties

in treating pathologies and imbalance of microbial ecosystems(Bailón-Pérez et al., 2008).

So far, legal limits have been established for antibiotics in food(4e1500 mg/kg formilk and 25e6000 mg/kg for the other food stuffsof animal origin (European Union, 1990)), but there is no legislationapplied to environmental matrices.

3. Remediation processes e an overview

As mentioned above, most conventional WWTPs or DWTPs arenot designed for the treatment of wastewaters containing highlypolar contaminants (Xu et al., 2007). Therefore, practical andeconomical solutions must be achieved in order to reduce the dailyamounts of antibiotics discharged into the environment.

A wide range of chemical and physical methodologies fororganic compounds removal can be employed, for example,chemical oxidation and biodegradation (destructive methods),adsorption, liquid extraction and membrane techniques (non-destructive processes). Depending on the pollutant concentrationin the effluent and the cost of the process, different methods can bechosen.

An overview of the works published in international journalsbetween 2000 and 2010 in this area is presented in Table 2.

3.1. Conventional treatments

The biological processes, filtration and coagulation/flocculation/sedimentation are the most used in conventional wastewatertreatment plants (Adams et al., 2002; Göbel et al., 2007;Stackelberg et al., 2007; Vieno et al., 2007; Arikan, 2008).

In the biological systems, activated sludge technology is widelyused, especially in industrial effluent treatments. The methodconsists of the organic compounds degradation in activated sludgetanks, with aerobic or anaerobic systems, by monitoring continu-ously temperature and the chemical oxygen demand (COD). Thehigh toxicity of many contaminants prevents the application of thisprocess in effluents with high pollutants concentration, since theyare recalcitrant and toxic to the microorganisms (Britto and Rangel,2008). However, this methodology can be applied to large effluentflow rates. (Eckenfelder, 2000).

The filtration is the removal of the solids, specially suspendedmatter, by passing the wastewater through a granular media (sand,coal, diatomaceous earth, granular activated carbon). Particles maybe removed by interstitial straining, but smaller particles must be

Table 2Summary of the removal/degradation processes applied in treatment of environmental matrices contaminated with antibiotics.

Antibiotic Concentration Matrix Treatment Operating conditions Analytical methods Results and comments References

1. AminoglycosidesSpectinomycin 1 mM Distilled water Ozonation 0.06e0.10 mM O3

pH¼ 2e9Absorbance at 260 nm - 2nd order reaction rates

increased withincreasing pH.

- Fast degradation aroundneutral pH.

- In 10 s, a totaldegradation wasachieved.

Qianget al. (2004)

2. AnthracyclinesEpirubicin 17.9 mg/L Distilled water Electrochemical

oxidationTwo Pt/Ir electrodesNaCl as electrolyte100 mA current

HPLC-UV, microbiologicalassay with S. aureus,cytoxicity and mutagenicassays

- Electrolysis maindegraded epirubicin andeliminated its cytotoxicity,mutagenicity andmicrobiological activityafter 360 min.

Hiroseet al. (2005)

3. b-LactamsAmoxicillin 5.0� 10�4 M Deionised water Ozonation 1.6� 10�4 M O3

pH¼ 2.5e7.2TOC, HPLC-DAD - 90% removal after 4 min

and 18% mineralisationafter 20 min.

- Low degree of mineralisation,even for long treatment times.

Andreozziet al. (2005)

Amoxicillin COD¼ 80� 103 mg/LTOC¼ 18,925 mg/L

Industry plantwastewater

Combination ofFenton oxidationwith two-stagereverse osmosis

Fenton:10 g/L H2O2

0.74 g/L Fe2þ

Reverse osmosis:Polyamide membrane cellsArea¼ 155 cm2

BOD5, COD, TOC - After liquideliquid extractionwith dichloromethane toremove solvents and otherorganic compounds, the TOCand COD removal efficiencywere around 50%.

- Fenton oxidation improvesthe degradation of organiccompounds before thetwo-stage reverse osmosissystem (TOC removal of 38%).

- After reverse osmosis 11% TOCremoval was achieved andbiodegradability was enhanced(COD:BOD5¼ 4:1).

- After the combined treatmentthe overall TOC removalefficiency was 99.7%.

Zhanget al. (2006)

Amoxicillin 42 mg/L Spiked STPeffluent

Photo-Fenton Black light at 365 nmand solar irradiation1.0e2.0 mM H2O2

0.20 mM ferrioxalateor Fe(NO3)3pH¼ 2.5

TOC, DOC, HPLC-DAD - Amoxicillin degradation wasnot influenced by the type ofirradiation or by the matrix.Its degradation was enhancedin the presence of ferrioxalate.

- The increase of the H2O2

concentration improved theefficiency of oxidation.

- After 10 min of irradiationa total degradation wasachieved.

Trovóet al. (2008)


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Amoxicillin 500 mg/LCOD¼ 790 mg/L

Simulatedwastewater

Photo-Fenton UV light (6 W)at 365 nmH2O2/Fe2þ molarratio¼ 10e100pH¼ 2e4

HPLC-DAD, COD,BOD5, DO

- The maximumbiodegradability ratio(>0.40) was achieved atH2O2/COD molar ratio¼ 2,H2O2/Fe2þ molar ratio¼ 50 andpH¼ 3e3.5, after 30e45 minof reaction. Under theseconditions, complete degradationwas achieved in 1 min(TOC removal¼ 71%).

- The study indicated thatphoto-Fenton can be used asa pre-treatment for improvementof amoxicillin biodegradability.

Elmolla andChaudhuri(2009)

Amoxicillin 300 mg/L Distilled waterWastewater

Adsorption onactivated carbon andbentonite

pH¼ 2e70.1e3.5 g adsorbentT¼ 30 �C

UV at 230 nm - The amoxicillin solution indistilled water was used tomodel the adsorption process.

- Both Langmuir and Freundlichmodels well fit the data.

- The kinetic were fittedby pseudo 2nd order model.

- In real wastewater matrices,both activated carbon (95%)and bentonite (88%) havehigh removal efficiencies.

Putraet al. (2009)

Amoxicillin 1e100 mg/L Distilled water Semiconductorphotocatalysis

pH¼ 3e9LP UV at 365 nmand solar radiation0.1e0.7 g/L TiO2 orTiO2 doped withC and Fe

COD - Amoxicillin degradationunder solar radiationproceeded about 3 timesfaster than under artificial UV.

- The maximum photocatalyticdegradation was achieved inneutral pH with 37% Cdoped catalyst (85% removal).

- By-products determinedby LCeMS.

Klausonet al. (2010)

Amoxicillin 20 mg/L Distilled water Removal usingmetallic iron

ZVI or Fe�

(metallic iron)0.5e2 g/L iron

LCeMS - Kinetic studies demonstratedthat this process followed1st order decay.

- Antibiotics removal wasattributed to the ruptureof the b-lactam ring,adsorption onto ironcorrosion products andco-precipitation with ironcorrosion products.

- Complete removal after3 h of contact with ZVI.

- By-products determinedby LCeMS.

Ghauchet al. (2009)Ampicillin

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Amoxicillin 104 mg/L Distilled water Fenton pH¼ 2e4H2O2/COD molarratio¼ 1.0e3.5H2O2/Fe2þ molarratio¼ 2e150

HPLC-DAD, COD,BOD5, TOC, DOC

- Under the optimalconditions (H2O2/Fe2þ

molar ratio¼ 10, pH¼ 3) itwas achieved the completedegradation of theantibiotics in 2 min.

- The biodegradability wasimproved from 0 to 0.37in 10 min and COD and DOCdegradation were 81.4%and 54.3%, respectivelyin 60 min.

- Fenton process was effectivein the treatment of solutionscontaining these antibiotics.


Ampicillin 105 mg/LCloxacillin 103 mg/L

Amoxicillin 104 mg/L Distilled water Photo-Fenton pH¼ 2e4H2O2/COD molarratio¼ 1.0e2.5H2O2/Fe2þ molarratio¼ 10e150UV light (6 W)at 365 nm

HPLC-DAD, COD,BOD5, TOC, DO

- Under the optimal conditions(H2O2/COD¼ 1.5, H2O2/Fe2þ

molar ratio¼ 20, pH¼ 3) itwas achieved that thecomplete degradation ofthe antibiotics in 2 min.

- The biodegradability wasimproved from 0 to 0.4and COD and DOCdegradation were 80.8%and 58.4%, respectivelyin 50 min.

- Mineralisation of organiccarbon occurred.



Amoxicillin 104 mg/L Distilled water Semiconductorphotocatalysis

UV light (6 W) at 365 nm0.2e2.0 g/L ZnOpH¼ 5e11


- The optimal conditions forcomplete degradation ofantibiotics were 0.5 g/LZnO, irradiation time180 min and pH¼ 11.

- Under these conditions,complete degradationoccurred and COD andDOC removal were 23.9and 9.7%, respectively.



Amoxicillin 104 mg/L Distilled water Semiconductorphotocatalysis

UV light (6 W) at 365 nm0.5e2.0 g/L TiO2

50e300 mg/L H2O2

pH¼ 3e11


- No significant degradationoccurred by 300 minof UV irradiation.

- At pH¼ 5 and 1.0 g/LTiO2, 50% degradationwas achieved for allcompounds (81%DOC removal).

- Addition of 100 mg/LH2O2 at pH¼ 5 and1.0 g/L TiO2 resulted incomplete degradationafter 30 min and 40%mineralisation after 24 h.




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Amoxicillin 1.60 mg/L Distilled water Chlorination ClO2/antibiotic molarratio¼ 0.25e2.00

HPLCeMS - ClO2 reacts with penicillin,amoxicillin and cefadroxil.

- It is possible to achieve totaldegradation of penicillinafter 2 h of reaction(ClO2/penicillin> 1.00)and amoxicillin and cefadroxilafter 1 min of reaction(ClO2/antibiotic> 1.50).

- It was detected degradationmetabolites.

Navalonet al. (2008)Cefadroxil 1.89 mg/L

Penicillin 25 mg/L

Ampicillin 20 mg/L Distilled water FentonPhoto-Fenton

pH¼ 2.3e5.7230e570 mM H2O2

53e87 mM Fe2þ

UV light (20 W)at 365 nm

HPLC-UV, TOC, COD - Under the optimizedconditions (pH¼ 3.7, 87mM Fe2þ, 373 mM H2O2

for Fenton and pH¼ 3.5, 87mM Fe2þ, 454 mM H2O2 forphoto-Fenton) a completedegradation was reached.

- It was achieved a highermineralisation (50% TOCremoval) with photo-Fentonthan Fenton (20% TOCremoval).

- The degradation productsnot present in antibacterialactivity.

Rozaset al. (2010)

Ceftriaxone COD¼ 250e1400mg/L

Formulationwastewater

Ozonation 3 g/(h L) O3

pH¼ 3e110e100 mM H2O2

BOD5, COD, TOC,Absorbanceat 254 nm

- COD removal was increasedwith increasing pH from3 to 7. The addition of H2O2

had no advantage for CODremoval kinetics over thedirect ozonation.

- Biodegradability representedin terms of BOD5/CODwas increased.

- After 60 min, 95% degradationwas achieved (45%TOC removal).

Balcio�gluand Ötker(2003)

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Penicillin COD¼ 830 mg/L Formulationeffluent

Ozonation Ozonation:2500 mg/(L h) O3

pH¼ 2.5e12.0

Perozonation:2500 mg/(L h) O3

pH¼ 10.52e40 mM H2O2

BOD5, COD,absorbance at344 and 274 nm

- Degradation increased withthe increase of pH and inthe presence of H2O2

up to 20 mM.- Removal efficiencies variedbetween 10 and 56% forozonation and 30 and83% for perozonation.

- Both processes followed1st order kinetics withrespect to CODremoval rates.

- Biodegradability ratio,BOD5/COD, increased by afactor of 6 and 23 forozonation and perozonation,respectively for 20 min.

- Treated effluent wassubjected to biologicalactivated sludge treatment.Biological COD removalrates and efficiencies wereappreciably improved, inparticular with perozonation.

Arslan-Alatonet al. (2004)


OzonationDirect and indirectphotolysisFenton andFenton-likePhoto-Fenton andphoto-Fenton-like

Ozonation:2760 mg/(L h) O3

pH¼ 3e11.5

Photolysis:LP UV at 254 nmpH¼ 70e40 mM H2O2

Photo- andFenton:LP UV at 254 nmpH¼ 320 mM H2O2

1 mM Fe2þ or Fe3þ

BOD5, COD, TOC,HPLC-DAD

- Ozonation was pH-dependentand the highest COD andTOC removals occurred inalkaline conditions(after 60 min of treatment,the maximum CODremoval was 49%).

- Photolysis proved to be aless effective method.

- Comparison betweenFenton and Fenton-likeprocesses.

- Relatively higher CODand TOC removal rateswere obtained with Fe2þ/H2O2

when compared with Fe3þ/H2O2

(61% and 46% COD removal,respectively).

- The presence of UV light onlyslightly improved the treatmentperformance.

- Poor improvement ofbiodegradability.

Arslan-Alatonand Dogruel(2004)


Ozonation 2760 mg/(L h) O3

pH¼ 3e11BOD5, COD, TOC,respirometricexperiments

- Degradation increased withthe increase of pH (6% atpH¼ 3 to 34% at pH¼ 11).

- Biodegradability ratio,BOD5/COD, increased uponozonation for 40 min.TOC abatement remained low.

- Biological treatabilitystudies shown the poorbiodegradability of this effluent.

Cokgoret al. (2004)


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Penicillin COD¼ 49,913mg/L

Pharmaceuticalwastewater

MicrowaveenhancedFenton-like

Microwavepower¼ 100e500 WpH¼ 1e11Radiationtime¼ 2e10 min3200e19,000mg/L H2O2

2000e8000mg/L Fe2(SO4)3

BOD5, COD, TOC,HPLC-UV, absorbanceat 254 nm

- Operating parameters wereinvestigated and the optimalcondition were: microwavepower 300 W, radiationtime 6 min, initial pH 4.42,H2O2 dosage 1300 mg/Land Fe2(SO4)3 4900 mg/L.

- Within the present conditions,the COD removal was 57.53%,TOC removal >40%, 55.06%degradation.

- Microwave enhancedFenton-like reaction hadsuperior treatment efficiency.

Yanget al. (2009)

Penicillin G COD¼ 600 mg/L Formulationeffluent

Photo-Fenton-likeFenton-like

Black light (125 W)5e40 mM H2O2

0.1e5 mM Fe3þ

pH¼ 3

BOD5, COD, TOC - In optimal conditions (1.5 mMFe3þ, 25 mM H2O2), it wasobtained 56% COD and 42%TOC removal after 30 minfor photo-Fenton-likeand 44% COD and 35% TOCremoval for Fenton-like.

- Both processes followed1st order kinetics withrespect to COD removal rates.

- Photo-Fenton-like processover dark-Fenton-like wasmore effective in termsof their effect onbiodegradabilityimprovement.

Arslan-Alatonand Gurses(2004)

Penicillin G COD¼ 200e600mg/L

Formulationeffluent

Ozonation 600e2600mg/L O3

pH¼ 3e12

COD, TOC - Ozonation followed 1storder kinetics.

- COD removal was increasedwith increasing pH and ozonedose and with the decreaseof initial COD value.

- At pH¼ 7, 1800 mg/L O3

and after 1 h, a TOC removalof 36% and a COD removalof 37% were achieved.

Arslan-Alatonand Coglayan(2005)

Penicillin G COD¼ 600 mg/L Formulationeffluent

Ozonation Ozonation:1800 mg/(h L) O3

pH¼ 7e12

Perozonation:1800 mg/(h L) O3

pH¼ 7e1210 mM H2O2

BOD5, COD, toxicityto D. magna,activated sludgeinhibition test

- COD removal was increasedwith increasing pH and inthe presence of H2O2.

- Biodegradability wasimproved by perozonationat pH¼ 7, but thispre-treatment did notcompletely remove theecotoxicity, leading toserious inhibition thetreatment of activatedsludge.

- The COD removal efficiencyincreased from 37% after1 h at pH¼ 7 for ozonationto 76% to perozonation.

Arslan-Alatonand Coglayan(2006)

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Penicillin V COD¼ 250e1400 mg/L Formulationwastewater

Ozonation Ozonation:3 g/(h L) O3

pH¼ 3e11

Perozonation:3 g/(h L) O3

pH¼ 70e100 mM H2O2

BOD5, COD, TOC,Absorbance at 254 nm

- COD removal was increasedwith increasing pH from3 to 7. The addition of H2O2

had no advantage for CODremoval kinetics over thedirect ozonation.

- Biodegradability representedin terms of BOD5/COD wasincreased.

- After 60 min, 40% degradationwas achieved (40%TOC removal).


4. GlicopeptidesBleomycin 44.6 mg/L Distilled water Electrochemical

oxidationTwo Pt/Ir electrodesNaCl as electrolyte100 mA current

HPLC-UV,microbiologicalassay with S. aureus,cytoxicity andmutagenic assays

- Electrolysis slightly degradedand eliminated cytotoxicity,mutagenicity andmicrobiological activity ofthis antibiotic.

Hiroseet al. (2005)

5. ImidazolesDimetridazole 10e30 mg/L Distilled water,

natural watersand wastewater

Simultaneousapplicationof ozonation andadsorption

pH¼ 2e90.25e0.50 g/Lactivated carbon

HPLC-DAD, TOC,Toxicity testswith V. fischeri

- The ozonation degradationswere higher than 90% and10e20% of TOC removal.

- Ozonation generates highlytoxic oxidation by-products.

- The presence of activatedcarbon during theozonation produces anincrease in the removalrate, a reduction in thetoxicity of oxidationby-products and areduction of around30% in the TOC.

Sánches-Poloet al. (2008)Metronidazole

RonidazoleTinidazole

Dimetridazole 100e600 mg/L Distilled water,naturalwaters andwastewater

Adsorption/bioadsorptionon activatedcarbon

T¼ 25 �CpH¼ 2e110e0.1 M NaCl1 g/L activated carbon

Absorbanceat 320 nm

- The pH of the mediumand the electrolyteconcentration did notinfluence the adsorptionremoval.

- Antibiotics were notdegraded by themicroorganisms usedin the biological treatment.

- The presence of thesemicroorganisms duringthe adsorption increasestheir adsorption/bioadsorption on theactivated carbon.

Rivera-Utrillaet al. (2009)Metronidazole


Dimetridazole 150 mg/L Distilled water Adsorption onactivated carbons

T¼ 25 �C0.2e1 g/L activatedcarbonpH¼ 7

Absorbance at320, 308 and 317 nm

- 90% removal was achievedwith 1 g/L of activatedcarbon.

- 2nd order kinetic fitssuitably theexperimental data.

Méndez-Díazet al. (2010)Metronidazole



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Levamisole 10 mg/L Pharmaceuticalwastewater

Reverse osmosisNanofiltration

Reverse osmosis:XLE and HR95PPmembranes

Nanofiltration:NF90 and HLDesal membranes

HPLC-DAD - The removal of the antibioticby reverse osmosis and thetight nanofiltrationmembrane is acceptably high(rejection factors> 0.9).

Ko�suti�cet al. (2007)

Metronidazole 1 mg/L Deionised water Direct and indirectphotolysisFentonPhoto-Fenton

Photolysis:LP UV at 254 nmMP UV at200e400 nmpH¼ 60e50 mg/LH2O2

Photo- and Fenton:LP UV at 254 nmpH¼ 3.529.4 mM H2O2

2.94e11.76mM Fe2þ

HPLC-PDAD, absorbanceat 220e230 nm

- photo-degradation exhibitedpseudo 1st order kinetics.

- MP irradiation was moreeffective than LP.

- Direct photolysis (6e12%removal) was less effectivethan UV/H2O2 oxidation(58e67% removal).

- Fenton oxidation followed2nd order kinetics and therate was increased withhigh Fe2þ concentrations.

- An increase in the removalefficiency and in the reactionrate occurred whenphoto-Fenton (74e94%removal) is compared toFenton oxidation (53e76%removal).

Shemeret al. (2006)

6. LincosamidesLincomycin 1 mM Distilled water Ozonation 0.06e0.10 mM O3

pH¼ 2e9Absorbanceat 260 nm

- 2nd order reaction ratesincreased with increasing pH.

- Total degradation wasachieved after 1 s.

- Fast degradation aroundneutral pH.

Qianget al. (2004)

Lincomycin 10e75 mM Distilled water Semiconductorphotocatalysiscoupled withnanofiltration

TiO2 catalyst0.2 g/L catalystMembraneDL2540C andDK2540C

TOC, HPLC-UV - The photo-oxidationfollowed 1st order kinetics.

- Lincomycin was successfullyoxidized by photocatalysis.

- Filtration allowed theseparation of thephotocatalyst particles fromlincomycin and itsdegradation products frompermeate.

Augugliaroet al. (2005)

Lincomycin 10e50 mg/L Distilled water Semiconductorphotocatalysis

MP UV (125W)pH¼ 6.0TiO2 (100% anataseor anatase/rutile¼ 4/1)1 and 0.4 g/L ofcatalyst,respectively

HPLC-UV, TOC,Absorbance at200e500 nm

- Degradation followedpseudo 1st order kinetic rate.

- After 5 h, 20% of lincomycinwas photolytic degraded. Inthe presence of TiO2 morethan 98% of drug wasoxidized within about 2 h.

- Using TiO2 (anatase/rutile)as catalyst, 60% TOC removalwas achieved, but a lesssignificant mineralisationwas observed using100% anatase.

Addamoet al. (2005)

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Lincomycin 25e50 mg/L Distilled water Electrochemicaloxidation

- stainlesssteel cathode

- Ti/Pt, graphite,Ti/IrO2/Ta2O5

or 3D GAC asanode

- Na2SO4 or NaClas electrolyte

Absorbance at200e400 nm,COD, iodometrictitration,voltammetricanalyses

- Lincomycin was hardlyoxidized (CODremoval¼ 30%) withslow overall kineticsdue to difficultdeprotonation, aprerequisite for thefollowing electrontransfer step.

- The electro-oxidationwas found to occurwith first order kinetics.

- Different anodeshad been tested.

Jara et al.(2007)

Lincomycin 25 mg/L Distilled waterand wastewater

Photo-Fenton Black light(lmax¼ 365 nm)and solar radiationpH¼ 2.50.20 mM ferrioxalate,FeSO4, Fe(NO3)31.0e10 mM H2O2

TOC, DOC,HPLC-DAD

- The degradation wasimproved whenferrioxalate wasemployed in comparisonto Fe(NO3)3 and FeSO4.

- After 8 min of irradiationthe antibiotic was totallyremoved in the presenceof ferrioxalate, whilewhen using Fe(NO3)3,20 min were necessaryto achieve the samedegradation.

- 94% TOC removal wasobserved after 60 minirradiation when usingferrioxalate, while 21%TOC removal wasachieved using Fe(NO3)3.

Bautitz andNogueira(2010)

7. MacrolidesAvilamycin COD¼ 7000 mg/L Pharmaceutical

wastewaterAnaerobic process Up-flow anaerobic

reactorpH¼ 6.5e7.8V¼ 11 L dividedinto four 2.75L stagesRetentiontime¼ 2e4 days

HPLC-UV, COD - The staged designfacilitated efficienttreatment of thisantibiotic wastewater.

- COD reduction was70e75%.

- An average of 95%tylosin reduction wasachieved, indicating thatthis antibiotic could bedegraded efficiently inthe anaerobic reactorsystem.

Chelliapanet al. (2006)Tylosin


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Azithromycin COD¼ 360 or590 mg/L

Two differentWWTP effluents

Primary treatment:ScreeningAerated grit removalClarifier

Secondary treatment:Activated sludgeMembrane reactor(1) or fixed-bedreactor (2)ClarifierSand filtration

Activated sludgetreatment withV¼ 5600 m3,t¼ 15 hMembrane reactorwith V¼ 16 m3,t¼ 13 hSand filter withV¼ 288 m3,t¼ 25 minActivated sludgetreatment withV¼ 9100 m3,t¼ 31 hFixed-bed reactorwith V¼ 1500 m3,t¼ 1 hSand filter withV¼ 360 m3,t¼ 6e8 h

LCeMS - In treatment plant 1, apre-treatment withactivated sludgewas studied.

- Bioreactor consisted of astirred anaerobiccompartment and adenitrification andnitrification cascade.

- Similar eliminationswere observedin the secondarytreatment oftwo conventionalactivated sludgesystems and afixed-bed reactor.

- Varying results wereobtained for theinvestigated macrolides.

Göbelet al. (2007)Clarithromycin

ErythromycinRoxithromycin

Azithromycin 2 mg/L Spiked WWTPeffluent

Ozonation 2 columnsoperating in series0.5e5 mg/L O3

pH¼ 7

LCeMS - Degradation followed2nd order kinetics.

- 90e99% of degradationfor O3 doses >2 mg/L.

- Suspended solids haverevealed a minor effecton the removal efficiency.

Huberet al. (2005)Clarithromycin

Roxithromycin

Clarithromycin 0.2e0.6 mg/L Spiked STPeffluent

Ozonation 5e15 mg/L O3 LCeMS - Ozonation revealedappropriate to oxidizethese compounds andinactivate relevantmicroorganisms(degradation between 76and 92%).

Terneset al. (2003)Erythromycin

Roxithromycin

Clarithromycin 1� 10�4 M Distilled water Ozonation 1� 10�5 M O3

pH¼ 3.2e4.4LCeMS, absorbanceat 260 nm inhibitionto P. putida

- Ozonation followed a 2ndorder reaction rate and therate was increased withincreasing pH.

- Reaction by-products weredetermined and they wereless inhibitory than theparent drug.

Langeet al. (2006)

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Erythromycin 10 ng/L DWTP effluent ClarificationChlorinationGAC filtration

Clarification:2 clarification tanksoperating in parallelFeCl3 as coagulantpH¼ 4.5e5.5Microsand injectionContact time¼ 15e20 min

Chlorination:Addition of NaClOTime¼ 200e300 min

GAC filtration:8 filter banks operatingsimultaneouslyContact time¼ 1.5e3 min

LCeMS - Clarification decreased theconcentration of thisantibiotic with a 47%removal. The moderateremoval of this hydrophiliccompound from the waterphase during clarificationmay be explained byferric chloride coagulation,which results in base oracid hydrolysis.

- After clarification, thewater samples weresubmitted to adisinfection process. Inthis process 92% removalwas achieved.

- Finally, chlorinated waterfrom disinfection processwas passed through GACfilters. A total removalwas obtained.

Stackelberget al. (2007)

Erythromycin 40 mg/L Spiked distilledwaterPharmaceuticalwastewater

Ozonation pH¼ 3e11H2O2/O3 molarratio¼ 0.5e18

LCeMS - The addition of H2O2

accelerated the degradation(optimum molar ratioH2O2/O3 of 5).

- The wastewater ozonationresulted in 97% removalafter 10 min and completedegradation in 20 min forall the compounds.

Lin et al.(2009)Tylosin

Roxithromycin 0.5 mM Spiked lake, riverand well water

Ozonation 0.1e2 mg/L O3

pH¼ 8HPLC-UV - Oxidation followed 2nd

order kinetics.- Water matrix affectsozone stability, radicalformation and scavenging.

- % Removal efficiencies> 90%.

Huberet al. (2003)

8. QuinolonesCiprofloxacin 0.15 mM Ultrapure water Photo-Fenton using

a heterogeneouscatalyst

HP UV at 280and 260 nmpH¼ 3e100e60 mM H2O2

Fe-Lap-RD catalyst0e1.5 g/L catalyst

TOC, COD, HPLC-DAD - The degradation wasimproved with the increaseof the H2O2 concentrationand catalyst loading. Anoptimum pH of 3 wasachieved.

- At optimal conditions acomplete conversion and57% of mineralisation wasachieved within 30 min.

- photo-Fenton followedpseudo 1st order kinetics.

Bobuet al. (2008)


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Ciprofloxacin 23e136 mM Deionised water Ozonation Ozonation:660e3680 mg/L O3

pH¼ 3e10T¼ 6.0e62.0 �C

Perozonation:2500 mg/L O3

pH¼ 3e10T¼ 27.5 �C2e990 mM H2O2

HPLC-UV - Ciprofloxacin ozonation andperozonation could be welldescribed by 1st orderkinetics.

- Highest degradation ratewas obtained at the highestozone concentration and thelowest drug concentration(95% degradation reachedbetween 60 and 75 min).No effect of temperaturewas found.

- In the perozonationexperiments, the additionof small amounts of H2O2

(2e50 mM) increased theciprofloxacin degradation.

- Reaction by-products weredetermined.

De Witteet al. (2009)

Ciprofloxacin 30 ng/L Surface riverwater

Coagulation/Flocculation/SedimentationSand filtrationOzonationGranular activatedcarbon filtrationDirect photolysis

Coagulation:Fe2(SO4)394e200 g/m3

coagulantpH¼ 5.0

Sand filtration:Surfaceload¼ 12 m/hpH¼ 7.3

Ozonation:3 columnsin series1 mg/L O3

Contact time¼ 30 minpH¼ 7.3

GAC filtration:2 step processChemviron Filtrasorb carbonContact time¼ 36 min

Direct photolysis:UV dose of 250 J/m2

LCeMS - Coagulation had not beenfound to affect the removalof norfloxacin and ofloxacin.30% of ciprofloxacin wasremoved from the effluent.

- Rapid sand filtration andgranular activated carbonfiltration didn’t removesignificantly thesepharmaceuticals (w10%).Drugs concentration wasnot affected by directphotolysis.

- Ozonation was anefficient technique for theelimination of most of thestudied compounds.Ciprofloxacin was onlyreduced by anaverage of 16%.

Vienoet al. (2007)Norfloxacin

Ofloxacin

Enrofloxacin COD¼ 250e1400mg/L

Formulationwastewater

Ozonation Ozonation:3 g/(h L) O3

pH¼ 3e11Perozonation:pH¼ 70e100 mM H2O2

BOD5, COD, TOC,Absorbance at 254 nm

- COD removal wasincreased with increasingpH from 3 to 7. Theaddition of H2O2 had noadvantage for COD removalkinetics over the directozonation.

- Biodegradability representedin terms of BOD5/COD wasincreased.

- After 60 min, 95% removal(45% TOC) was achieved.


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Enrofloxacin 50e200 mg/L Distilled/deionisedwater

Adsorptioncoupled withozonation

pH¼ 5e10T¼ 28e50 �CZeolite as adsorbent0.75 g adsorbentContact time¼ 24 h1430 mg/h O3

Contact time¼ 10e30 min

Absorbance at 271,272 and 275 nm

- Enrofloxacin removal wasincreased by decreasing pHand increasing temperature.

- Adsorption equilibrium datawere fitted by Langmuirisotherm model. 80% ofenrofloxacin was adsorbedon zeolite.

- Ozone at sufficientconcentration was able todecompose totallyenrofloxacin adsorbed onzeolite, but it caused thechange of zeolite porestructure.

Ötker andAkmehmet-Balcio�glu(2005)

Enrofloxacin 10 mg/L Pharmaceuticalwastewater



Nanofiltration:NF90 and HL Desalmembranes

HPLC-DAD - The removal of theantibiotic by reverseosmosis and the tightnanofiltration membraneis acceptably high(rejection factors> 0.972).


Enrofloxacin 1580 mg/L Syntheticwastewater

Conductive-diamondelectrochemicaloxidationOzonationFenton oxidation

Conductive-diamond

electrochemical

oxidation:Si-boron dopeddiamond as anodeStainless steel as cathodeT¼ 35 �C

Ozonation:0.5 L/min O3 flow rateT¼ 25 �CFenton

oxidation:pH¼ 3FeSO4.7H2O as catalyst

COD, TOC - The three processes canreduce the organic contentof synthetic wastewaterpolluted with enrofloxacin.

- Conductive-diamondelectrochemical oxidationis the more efficienttechnology in terms ofmineralisation, but not onCOD removal, which is moreefficiently achieved byozonation.

- The high efficiency interms of oxidant-useobtained by Fentonoxidation during the initialstages indicates that thisprocess is very efficientin the removal of theantibiotic, but it rapidlyleads to the formationof refractory compounds.

Guineaet al. (2009)


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Enrofloxacin 158e790 mg/L Ultrapure water Anodic oxidation withelectrogenerated H2O2

Electro-FentonPhotoelectron-FentonSolarphotoelectron-Fenton

Pt or boron-dopeddiamond (BDD) as anodeCarbon-PTFE cathode12 ml/min O2

UVA lightwith lmax¼ 360 nm0.05 M Na2SO4

0.1e0.5 mM Fe2þ

pH¼ 3.0T¼ 35 �C

TOC, HPLC-DAD - All procedures are lesspotent using Pt as anode.

- In the stirred tank reactorusing BDD as anode, itwas achieved:

(i) 67% TOC removal inanodic oxidation withelectrogenerated H2O2

(ii) 78% TOC removal inElectro-Fenton

(iii) 96% TOC removal inphotoelectron-Fenton

(iv) 97% TOC removal in solarphotoelectron-Fenton

- In the batch recirculation flowreactor using BDD as anode, itwas achieved:

(i) 28% TOC removal in anodicoxidation with electrogeneratedH2O2

(ii) 45% TOC removal inElectro-Fenton

(iii) 86% dolarphotoelectron-Fenton

Guineaet al. (2010)

Flumequine 19.1e95.7 mM Distilled water Semiconductorphotocatalysis

UV lamp (20 W)pH¼ 3e100.5e1.5 g/L TiO2

0.17e0.83 mM H2O2

Absorbance at331 nm, HPLC,TOC, GCeMS,antibacterialtest with E. coli

- Under optimized conditions(pH¼ 6, absence of H2O2 andlow titania load), the timerequired to completely eliminateflumequine was 30 min.

- Mineralisation after 60 minirradiation was around 80%.

- The oxidation products arenot biologically active.

- The method has the advantagethat after short period oftreatment, the by-productscould be treated by conventionalbiological systems.

Palominoset al. (2008)

Nalidixic acid 45 mg/L Distilled water Solar photo-Fentoncombined withbiological treatment

Solar photo-Fenton:- pH¼ 2.6e2.8- 20 mg/L Fe2þ

- 200e400 mg/L H2O2

- solar UV power

Biological treatment:- activated sludge- 500 L/h operation flux

HPLC-UV, TOC,DOC, toxicitytests withV. fischeri

- After 200 min, the totaldegradation of the antibioticwas achieved.

- 90% of mineralisation wasachieved after 400 min.

- Photo-Fenton successfullyenhanced the biodegradability.

- The global efficiency in thecombined solar photo-Fentonand immobilized biomass reactorsystem operated in batch modewas 95% of DOC elimination, ofwhich 33% was accomplished bythe solar photo-Fenton treatmentand 62% by the biologicaltreatment.

Sirtoriet al. (2009)

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Ofloxacin 5e10 mg/L Ultrapure water Semiconductorphotocatalysiscoupled withnanofiltration

LP UV (125 W)TiO2 catalyst1 g/L catalystMembrane NTR7410,PAN GKSS HV3IT,N30F and NF PES 10

Absorbance at230e280 nm

- Degradation followed 1st orderkinetics.At pH¼ 6, 54% of the antibioticwas adsorbed on the catalyst.66% degradation was achievedby photocatalysis.Filtration separated thecatalyst particles, but notthe reaction by-productsfrom the permeate.

Molinariet al. (2006)

Ofloxacin 25e50 mg/L Distilled water Electrochemicaloxidation

Electrolytic cell:- stainless steelplate as cathode

- Ti/Pt, graphite,Ti/IrO2/Ta2O5 or3D GAC as anode

- Na2SO4 or NaCl aselectrolyte

- 1.5e400 A/m2 current

Absorbance at200e400 nm,COD, iodometrictitration,voltammetricanalyses

Different anodes had beentested.Ofloxacin was efficientlyoxidized (99.995%) over alltested anodes.The electro-oxidation wasfound to occur with1st order kinetics.

Jara et al.(2007)

9. QuinoloxalineDerivative

Carbadox 50 mg/L Spiked deionised/distilled waterand surface riverwater

Coagulation/Flocculation/SedimentationExcess lime/soda ashsofteningPowderedactivated carbonsorptionChlorinationOzonationDirectphotolysisIon exchangeReverseosmosis

Coagulation:Al2(SO4)3, Fe2(SO4)30e170 mg/L coagulantTreatment time¼ 3 h21

Excess lime/soda

ash softening:Lime¼ 232 mg/Las CaCO3

Soda ash¼ 191mg/L as CaCO3

pH¼ 11.3

Sorption:Calgon WPH PulvPAC adsorbent0e50 mg/L PACContact time¼ 4 h

Chlorination:1.0 mg/L Cl2(pH¼ 7.5)

Ozonation:7.1 mg/L O3 (pH¼ 7.5)

Direct photolysis:LP UV at 254 nm(pH¼ 7.5)

Ion exchange:0.66 g strong-acid cationand strong-base anionresins (pH¼ 7)

Reverse osmosis:Cellulose acetatemembrane

HPLC-UV Coagulation/Flocculation/Sedimentation, excess lime/soda ash softening, directphotolysis and ion exchangewere all relatively ineffectivemethods of antibiotic removal.The percent removal of thisantibiotic was greater than90% for PAC dosageof 50 mg/L.Oxidation with both ozoneand chlorine at typical doseswas effective in the removalof the studiedantibiotics (>90%)Reverse osmosis was effectiveat the removal of thestudied compound withrejection levels greater than90%. However, this process isusually not economical.

Adamset al. (2002)


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-

-

-

-

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-

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-

-



10. SulfonamideSulfacetamideSulfadiazineSulfamethoxazoleSulfathiazole

0.1 mM Distilled water Semiconductorphotocatalysis

UV at 366 nmTiO2 catalyst2.5 g/L catalyst

HPLC-UV, BOD5,TOC, microbiologicalassay with C. vulgaris

- All drugs were completelydegraded after 300 min ofirradiation.

- TOC removal varied between30 and 70%.

- Intermediate products aremore biodegradables and lesstoxic than the parentcompounds.

Baranet al. (2006)

SulfachloropyridazineSulfadimethoxineSulfamerazineSulfamethazineSulfamethoxazoleSulfamonomethoxineSulfathiazole

10 mg/L Deionised waterand spiked STPeffluent

Ion exchange Self-decomposition:pH¼ 7.0T¼ 25 �CTreatment time¼ 15 days

MIEX� resin treatment:pH¼ 7.0T¼ 25 �C0.5e5.0 mL/L resinTreatment time¼ 24 h

HPLC-DAD - Self-decomposition was slowand considerable amounts ofantibiotics (4.1e7.3 mg/L) stillremained after 15 days.

- MIEX� treatment was effectivefor removal of these drugsthrough ion exchange(w90% removal), but organicinterference was observed.This treatment was ineffectivefor removal of sulfamerazineand sulfamethazine.

- Compounds with strongerelectronegativity were moreeasily removed by MIEX�

treatment.

Choiet al. (2007)

SulfachloropyridazineSulfadimethoxineSulfamerazineSulfamethazineSulfathiazole

50 mg/L Spiked deionised/distilled waterand surface riverwater

Coagulation/Flocculation/SedimentationExcess lime/soda ashsofteningPowderedactivatedcarbon sorptionChlorinationOzonationDirect photolysisIon exchangeReverse osmosis


Excess lime/soda ash

softening:Lime¼ 232 mg/Las CaCO3

Soda ash¼ 191 mg/Las CaCO3

pH¼ 11.3


Chlorination:1.0 mg/L Cl2 (pH¼ 7.5)




Reverse osmosis:Cellulose acetate membrane

HPLC-UV - Coagulation/Flocculation/Sedimentation, excess lime/soda ash softening, directphotolysis and ion exchangewere all relatively ineffectivemethods of antibiotic removal.

- The percent removal of thisantibiotic was greater than90% for PAC dosageof 50 mg/L.

- Oxidation with both ozoneand chlorine at typical doseswas effective in the removalof the studiedantibiotics (>90%)

- Reverse osmosis waseffective at the removal ofthe studied compound withrejection levels greater than90%. However, this process isusually not economical.

Adamset al. (2002)

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SulfachloropyridazineSulfadimethoxineSulfamerazineSulfamethazineSulfamethizoleSulfamethoxazoleSulfathiazole

10 mg/L Distilled waterspiked withcalcium chloride,humic acidand NaCl

Nanofiltration NF 200 membranes(14.6 cm2 area)pH¼ 7T¼ 20 �C

LCeMS - The sulphonamideantibiotic groupshowed slowadsorption kinetic andrelatively little massadsorption, rising to only11e20% after 90 min.

- The organic matter, salinityand kind of antibiotic affectedmembranes rejection.

Koyuncuet al. (2008)

SulfadiazineSulfamethozaxoleSulfapyridineSulfathiazole

2 mg/L Spiked WWTPeffluent

Ozonation 2 columns operatingin series0.5e5 mg/L O3

pH¼ 7

LCeMS - Degradation followed 2ndorder kinetics.

- 90e99% of degradation forO3 doses > 2 mg/L.

- Suspended solids haverevealed a minor effect on theremoval efficiency.

Huberet al. (2005)

SulfadiazineSulfadimethoxineSulfamerazineSulfathiazole

15 mg/L Distilled water Semiconductorphotocatalysis

UV at 340e400 nmTiO2 catalyst200 mg/L catalystT¼ 50 �C

LC-conductimeter,LCeMS

- After 30 min of irradiation,sulfadimethoxine andsulfathiazole were completelydegraded. 80 and 90% removalefficiencies were achieved forsulfadiazine and sulfamerazine,respectively.

- LC/MS was used to identifythe intermediates productsand the reaction mechanism.

Calzaet al. (2004)

SulfadiazineSulfaguanidineSulfamethazine

10 mg/L Pharmaceuticalwastewater




HPLC-DAD - The removal of the antibioticby reverse osmosis and thetight nanofiltration membraneis acceptably high (rejectionfactor> 0.989).


SulfadimethoxineSulfamethazineSulfamethoxazole

40 mg/L Spiked distilledwater andpharmaceuticalwastewater

Ozonation pH¼ 3e11H2O2/O3 molarratio¼ 0.5e18

LCeMS - The addition of H2O2

accelerated the degradation(optimum molar ratioH2O2/O3 of 5).

- The ozonation of thepharmaceutical wastewaterresulted in completedegradation in 20 min forall the compounds.

Linet al. (2009)

Sulfamethazine 10e70 mg/L Distilled water Semiconductorphotocatalysis

UV at 350e400 nmpH¼ 4.80e800 mg/L H2O2

TiO2 (anatase/rutile¼ 3.6/1or 100% anatase),TiO2Na or ZnO catalyst0e4 g/L catalyst

Absorbance at260 nm, TOC, DOC

- Degradation followed pseudo 1st order kinetic rate.

- The increase in the catalystquantity and the presence ofH2O2 enhanced thedegradation rate.

- TiO2 is more effective than ZnOfor drug removal andmineralisation.

- With 1 g/L TiO2 it was achieved95% degradation (120 min)and 85% DOC removal(350 min).

- With 1 g/L ZnO it was achievedtotal degradation (120 min) and30% DOC removal (350 min).

Kaniouet al. (2005)


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Sulfamethazine 50 mg/L Deionisedwater

Photo-Fenton Sunlight lamp (400e580 nm)pH¼ 3176e1024 mg/L H2O2

12e68 mg/L Fe2þ

HPLC-DAD, TOC,bacterial tests

- To achieve maximum TOCremoval (60%) and totaldegradation it was necessaryto use 600 mg/L H2O2,50 mg/L Fe2þ.

- After 2 min, total degradationof sulfamethazinewas achieved.

- It has been established thattoxicity increases during its1st reaction stages.

Pérez-Moyaet al. (2010)

SulfamethizoleSulfamethoxazoleSufamoxoleSulfathiazoleSulfisoxazole

100 mM Distilled waterand spikedriver water

Direct photolysisFenton

Direct photolysis:Natural sunlight irradiationor MP UV (175W)pH¼ 2.5e11

Fenton:pH¼ 330% H2O2

40 mM Fe2þ

HPLC-UV, LCeMS - Direct photolysis was foundto be highly pH-dependent.

- Sulfamethoxazole andsulfisoxazole werephotodegraded most readilyin acidic media, whereassulfamethizole andsulfathiazole were degradedin basic. Sulfamoxole wasdegraded nonphotochemicallyin aqueous solutions at allpH values.

- Sulfonamides drugs wereall degraded by Fentonreagent.


Boreenet al. (2004)

Sulfamethoxazole 0.5 mM Spiked lake,river and wellwater

Ozonation 0.1e2 mg/L O3

pH¼ 8HPLC-UV - Oxidation followed 2nd

order kinetics.- Water matrix affects ozonestability and radical formation.

- % Removal efficiencies > 90%.

Huberet al. (2003)

Sulfamethoxazole 0.2e0.6 mg/L Spiked STPeffluent

Ozonation 5e15 mg/L O3 LCeMS - Ozonation revealedappropriate to oxidize thiscompound and inactivaterelevant microorganisms(92% degradation).

Terneset al. (2003)

Sulfamethoxazole 5e500 mM Deionised waterspiked withnatural organicmatter andbicarbonates

Semiconductorphotocatalysis

UV at 324e400 nmpH¼ 3e11TiO2 (anatase/rutile¼ 9/1or 100% anatase, 100%rutile) catalyst0.01e1 g/L catalyst

HPLC-PDA, DOC - Sulfamethoxazole degradationwas influenced by the initialdrug concentration, catalystphase and concentration, pHand water matrix.

- Degradation followed pseudo1st order kinetic rate.

- A mixed phase was morereactive than pure rutileor anatase.

- A total degradation wasachieved after 60 min.

Hu et al.(2007)

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Sulfamethoxazole 25e200 mg/L Distilled water Semiconductorphotocatalysis

UV at 240e310 nmpH¼ 2e11TiO2 catalyst0e2 g/L catalyst

HPLC-PDA, DOC,BOD5, COD, TOC

- Degradation and mineralisationwere increased, increasing thecatalyst dose (360 min, 2.0 g/LTiO2, pH¼ 5, 90% degradationand 40% TOC removal).

- pH didn’t influencedsignificantly the drugremoval, but TOC removalsuffered an enhancementwith pH increase.

- Low biodegradabilityof the effluent.


Abellánet al. (2007)

Sulfamethoxazole 200 mg/L Distilled water Photo-Fenton Black light e 365 nm50e1000 mg/L H2O2

10 mg/L Fe2þ

pH¼ 2.8

BOD5, COD, TOC,HPLC-UV, toxicitytest to V. fischeri,oxygen uptake

- Drug degradation,mineralisation andbiodegradability wereimproved with the increaseof H2O2 concentration.

- Using H2O2< 400 mg/L, theprocess producedintermediates with enoughacute detoxicity and noinhibition effects.

- With 400 mg/L, a TOC removalof 50% was achieved, aswell as a COD removal of 75%.

Gonzálezet al. (2007)

Sulfamethoxazole 30 ng/L DWTP effluent ClarificationChlorinationGAC filtration

Clarification:2 clarification tanksoperating in parallelFeCl3 as coagulantpH¼ 4.5e5.5Microsand injectionContact time¼ 15e20 min

Chlorination:Addition of sodiumhypochloriteContact time¼ 200e300 min

GAC filtration:8 filter banks operatingsimultaneouslyContact time¼ 1.5e3 min

LCeMS - Clarification decreased theaverage concentration of thisantibiotic with a 33% removal.

- After clarification, the watersamples were submitted to adisinfection process. With thisprocess 100% removalwas achieved.

Stackelberget al. (2007)


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SulfamethoxazoleSulfapyridine

COD¼ 360or 590 mg/L


Primary treatment:ScreeningAerated grit removalClarifier

Secondary treatment:Activated sludgeMembrane reactor (1)or fixed-bed reactor (2)ClarifierSand filtration

Treatment plant 1:Activated sludge treatmentwith V¼ 5600 m3, t¼ 15 hMembrane reactorwith V¼ 16 m3, t¼ 13 hSand filter with V¼ 288 m3,t¼ 25 min

Treatment plant 2:Activated sludge treatmentwith V¼ 9100 m3, t¼ 31 hFixed-bed reactorwith V¼ 1500 m3, t¼ 1 hSand filter withV¼ 360 m3, t¼ 6e8 h

LCeMS - In treatment plant 1, apre-treatment with activatedsludge was studied.

- Bioreactor consisted of astirred anaerobic compartmentand a denitrification andnitrification cascade.

- Similar eliminations wereobserved in the secondarytreatment of two conventionalactivated sludge systems anda fixed-bed reactor.

- Varying results were obtainedfor the investigatedsulfonamides. Thesecompounds were detected insecondary effluents.

Göbelet al. (2007)

Sulfamethoxazole 200 mg/L Distilled water Ozonation 0.15e1.5 g/L O3

pH¼ 3e11HPLC-UV, TOC, COD,BOD5, absorbanceat 254 nm, toxicityto V. fischeri

- Ozonation followed 2ndorder kinetics.

- Process performanceincreased with increasingpH. Temperature had nosignificant effect ondegradation.

- After 60 min of ozonationwith 1.5 g/L O3, it wasachieved a complete drugdegradation with an increasein biodegradability, and lowmineralisation (TOCremoval w18%). Ecotoxicityremained practicallyunchanged.

- By-products were identifiedby means of LCeMS.

Dantaset al. (2008)

Sulfamethoxazole 50 ng/L Spikedgroundwater

Nanofiltration andreverse osmosis

Polyamide thin-filmcomposite membranes

Reverse osmosis:486 m3/h feed flow356 m3/h permeate flowTwo parallel stages (240þ 120membranes)

Nanofiltration:360 m3/h feed flow234 m3/h permeate flowTwo parallel stages(186þ 90 membranes)

HPLCeMS - Excellent overall performanceof both nanofiltration and reverseosmosis was noted, with highrejection percentages (>95%).

Radjenovi�cet al. (2008)

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Sulfamethoxazole 10 mg/L Distilled waterDistilledwaterþ nitrate(10 and 20 mg/L)Seawater

Photolysis Solar simulator (l< 290 nm) HPLC-UV, LCeMS,DOC, toxicity testto V. fischeri andD. magna

- The presence of nitrate indistilled water did notaffect the degradation rate.

- 98% of the initialconcentration in distilledwater was removed after30 h of irradiation.

- No DOC variation wasobserved, demonstratingthe formation ofintermediates.

- In the seawater solutions,a slower kinetic wasobserved, where onlya slight reduction in theantibiotic concentration(14%) occurred after 7 h.

- LC/MS was used to identifythe intermediates productsand the reactionmechanism.

- High toxicity of thephoto-transformationproducts generated.

Trovóet al. (2009)

Sulfamethoxazole 50 mg/L Distilled waterSeawater

Solar photo-Fenton Lamp at 290 nmpH¼ 2.5e9.030e210 mg/L H2O2

2.6e10.4 mg/L Fe2þ

HPLC-UV, HPLCeMS,DOC, toxicity tests

- Degradation andmineralisation werestrongly hindered inseawater compared todistilled water.

- The increase in ironconcentration showeda slight improvementon the pollutant degradationand mineralisation rate.

- The increase of H2O2

concentration up to 120 mg/Lin distilled water reduced thesample toxicity.

- The DOC removal obtainedwas about 80% for distilledwater and 50% for seawater.

Trovóet al. (2009)


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11. TetracyclinesChlortetracyclineDemocyclineDoxycyclineMeclocyclineMinocyclineOxytetracyclineTetracycline

10 mg/L Deionised waterand spiked STPeffluent

Ion exchange Self-decomposition:pH¼ 7.0T¼ 25 �CTreatment time¼ 15 days

MIEX� resin treatment:pH¼ 7.0T¼ 25 �C0.5e5.0 mL/L resinTreatment time¼ 24 h

HPLC-DAD - Self-decomposition was slowand considerable amountsof antibiotics (6.1e7.8 mg/L)still remained after 15 days.An exception was minocyclineand oxytetracycline, whichwere reduced to 0.7and 2.2 mg/L, respectively.

- MIEX� treatment waseffective for removal ofthese drugs through ionexchange (removal> 80%),but organic interferencewas observed.

- Compounds with strongerelectronegativity were moreeasily removed byMIEX� treatment.

Choi et al.(2007)

Chlortetracycline COD¼ 52,240e49,100 mg/L

Manure slurry Anaerobicdigestion

pH¼ 7.5T¼ 35 �CTreatment time¼ 33 days

LCeMS - Drug concentration decreasedapproximately 75%.

Arikan(2008)

ChlorotetracyclineDoxycyclineOxytetracyclineTetracycline

10 mg/L Distilled waterspiked withcalcium chloride,humic acidand NaCl

Nanofiltration NF 200 membranes(14.6 cm2 area)pH¼ 7T¼ 20 �C

LCeMS - After 90 min, the degradationof the tetracyclines rangedbetween 50 and 80%.

- The organic matter, salinityand kind of antibiotic affectedmembranes rejection.

Koyuncuet al. (2008)

ChlortetracyclineDemeclocyclineDoxycyclineMeclocyclineMinocyclineOxytetracyclineTetracycline

10 mg/L (adsorptionstudies)100 mg/L(coagulation)

Spiked syntheticand river water

CoagulationAdsorption withactivated carbon

Coagulation:PACl (5e60 mg/L)Contact time¼ 5 min

GAC filtration:Calgon F400 andCoconut-based carbon

LCeMS, DOC - In coagulation, the removalefficiency was improved withthe increasing of PACl.

- Coagulation could remove43e94% of the drugs fromthe synthetic water (40 mg/Lcoagulant), but in the riverwater the removal efficiencieswere slightly deteriorated(44e67%) due to organicinterference.

- With activated carbon, morethan 68% of tetracyclineswere removed.

Choi et al.(2008)

ChlortetracyclineOxytetracyclineTetracycline

20e110 mM Distilled water Adsorption withaluminium oxide

0.8e3.5 g/L Al2O3 HPLC-DAD, LCeMS - Rapid adsorption occurs inthe first 3 h (40 mMantibiotic, 1.78 g/L Al2O3,pH¼ 5, T¼ 22 �C).

- The %adsorption was 43%tetracycline, 57%chlortetracycline and 44%oxytetracycline.

Chen andHuang(2010)

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Oxytetracycline 1000 mg/L(COD¼ 10000 mg/L)

Pharmaceuticalwastewater

Reverse osmosisUltrafiltration

Reverse osmosis:- NTR-7450 membrane- NTR-7459 membrane- Area¼ 155 cm2

- T¼ 21e23 �C- Operationalpressure¼ 1.8 MPa

- Volume reduction¼ 3.5

Ultrafiltration:- Stirred cell- 3,10,30,50 K Dacut-off membranes

- Operationalpressure¼ 0.3 MPa

- Volume reduction¼ 10

Absorbance at 480 nm,COD, TOC

- Using reverse osmosisthe organic content inthe permeate was decreasedfrom COD¼ 10000 mg/L toless than 200 mg/L(98% removal).

- The oxytetracycline wasreduced from 1000 mg/Lto lower than 80 mg/L (87.5%removal) and was concentratedmore than 3 timesin the retentate.

- With additional treatmentof ultrafiltration by 3Kmembranes, the antibioticrecovery ratio was higherthan 60% and the purityhigher than 80%.

Li et al.(2004)

Oxytetracycline 10 mg/L Pharmaceuticalwastewater




HPLC-DAD - The removal by reverse osmosisand the tight nanofiltrationmembrane is acceptably high(rejection factor> 0.990).


Oxytetracycline 10e40 mg/L Ultrapure water Direct photolysis MP UV at 365 nmpH¼ 4e9

HPLC-UV, TOC, toxicityto P. phosphoreum

- Oxytetracycline photolysisfollowed 1st ordermodel kinetics.

- After 240 min of irradiation,an inhibition rate of 47%was achieved with 90%degradation and 14%TOC removal.

Shaojunet al. (2008)

Oxytetracycline 100e200 mg/L Ultrapure water Ozonation 11 mg/L O3

pH¼ 3e11HPLC-DAD, COD, BOD5,toxicity to V. fischeri

- Performance increasedwith increasing pH (100%degradation after 20 min).BOD5/COD was higherthan 0.3-biodegradableeffluent.

- The toxicity resultsindicated that by-products(5e30 min) were moretoxic than the parentcompound.

Li et al.(2008)

Tetracycline 10e50 mg/L Distilled water Semiconductorphotocatalysis

MP UV (125 W)pH¼ 6.0TiO2 (100% anatase oranatase/rutile¼ 4/1)1 and 0.4 g/L of catalyst,respectively

HPLC-UV, TOC,Absorbance at200e500 nm

- Degradation followedpseudo 1st orderkinetic rate.

- After 5 h, 70% of tetracyclinewas photolytic degraded. Inthe presence of TiO2 morethan 98% of drug wasoxidized within about 2 h.

- Using TiO2 (anatase/rutile)as catalyst, tetracycline wastotally mineralised, whereasonly 50% was observed using100% anatase.

Addamoet al. (2005)


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Tetracycline 40 mg/L Deionised water Semiconductorphotocatalysis

UV at 254, 365 nm andsolarium device at300e400 nmTiO2 catalyst0.5e1 g/L catalyst

HPLC-UV, BOD5, COD,TOC, microbiologicalassay with S. aureus

- Degradation followed 1storder kinetics.

- Degradation and TOCremoval were influencedby the light source(UV 254 nm>solarium>

UV 365 nm).- With 0.5 g/L TiO2 andafter 120 min itwas achieved:UV 254 nm: 100%degradation, 90%TOC removalSolarium: 100% degradation,70% TOC removalUV 365 nm: 50% degradation,10% TOC removal

- By-products did not showantibacterial and were morebiodegradable than theparent compound.

Reyeset al. (2006)

Tetracycline 24 mg/L Spiked STPeffluent, surfaceand deionisedwater

Photo-Fenton Black light (15 W)and solar irradiation1e10 mM H2O2

0.20 mM ferrioxalateor Fe(NO3)3pH¼ 2.5

TOC, HPLC-UV - photo-Fenton underartificial or solar irradiationwas very efficient, achievingtotal degradation afterapproximately 1 minirradiation.

- Under black light irradiation,higher efficiency wasobtained using iron nitrate,while solar degradationis favoured by the useof ferrioxalate.

Bautitz andNogueira(2007)

Tetracycline 10e40 mg/L Ultrapure water Direct photolysis ML UV at 365 nmpH¼ 6

HPLC-UV, TOC,toxicity to V. fischeri

- Photolysis followed 1storder kinetics.

- Upon 300 min of irradiation,only 15% reduction of TOCoccurred in spite of quickconversion of 73% oftetracycline, suggestingthat a majority of this drugtransformed into intermediateproducts without completemineralisation.

- Toxicity increased withirradiation. Reactionby-products were studied.

Jiao et al.(2008)

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Tetracycline 20 mg/L Distilled water Semiconductorphotocatalysis

Xe lamp (300e800 nm)TiO2 and ZnO as catalyst0.5e1.5 g/L TiO2

0.2e1.5 g/L ZnOpH¼ 3e10 (TiO2) orpH¼ 6e11 (ZnO)

Absorbance at360 nm and 380 nm,TOC, antibacterialtest with S. aureus

- The optimal conditionsfound were 1.5 g/L andpH¼ 8.7 for TiO2 and 1.0 g/Land pH¼ 11 for ZnO.

- After 15 min irradiation, inthe absence of catalyst, 80%degradation was achieved.After 15 min of irradiation,the antibiotic was completelyremoved with TiO2 and morethan 50% was mineralised(total removal of theantibacterial activity).In the presence of ZnO,degradation was achievedafter 10 min. andmineralisation after 60 min.

Palominoset al. (2009)

12. Other antibioticsChloramphenicol 10e80 mg/L Deionised water Semiconductor

photocatalysisUV at 320e400 nmpH¼ 50e600 mg/L H2O2

TiO2 (anatase/rutile¼ 3.6/1or 100% anatase), TiO2Na(100% anatase) or ZnOcatalyst0e4 g/L catalystT¼ 3e57 �C

Absorbance at276.5 nm, TOC,antimicrobialactivity to E. coli

- Pseudo 1st order kinetic rateincreased with increasing drugconcentration, catalyst loadingand H2O2 concentration.

- Temperature slightly affectedthe photo-degradation process.

- TiO2 (anatase/rutile¼ 3.6/1)and ZnO catalysts seem to bethe most efficient.

- It was achieved a completeelimination of the antibioticactivity after 90 min, with70% of mineralisation.

Chatzitakis(2008)

Mitomycin C 17.9 mg/mL Distilled water Electrochemicaloxidation

Two Pt/Ir electrodesNaCl as electrolyte100 mA current

HPLC-UV,microbiologicalassay with S. aureus,cytoxicity andmutagenic assays

- Electrolysis slightly degradedand eliminated cytotoxicity,mutagenicity andmicrobiological activity ofthis antibiotic.

Hiroseet al. (2005)


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Trimethoprim 50 mg/L Spiked deionised/distilled waterand surfaceriver water

Coagulation/Flocculation/SedimentationExcess lime/soda ashsofteningPowderedactivatedcarbon sorptionChlorinationOzonationDirectphotolysisIon exchangeReverse osmosis


Excess lime/soda

ash softening:Lime¼ 232 mg/Las CaCO3

Soda ash¼ 191 mg/Las CaCO3

pH¼ 11.3


Chlorination:1.0 mg/L Cl2 (pH¼ 7.5)




Reverse osmosis:Cellulose acetatemembrane

HPLC-UV - Coagulation/Flocculation/Sedimentation, excesslime/soda ash softening,direct photolysis and ionexchange were allrelatively ineffectivemethods of antibioticremoval.

- The percent removal ofthis antibiotic wasgreater than 90% forPAC dosage of 50 mg/L.

- Oxidation with bothozone and chlorine attypical doses waseffective in the removalof the studiedantibiotics (>90%)

- Reverse osmosis waseffective at the removalof the studied compoundwith rejection levelsgreater than 90%.However, this processis usually not economical.

Adamset al. (2002)

Trimethoprim 0.2e0.6 mg/L Spiked STP effluent Ozonation 5e15 mg/L O3 LCeMS - Ozonation revealedappropriate to oxidizethis compound andinactivate relevantmicroorganisms(85% degradation).

Terneset al. (2003)

Trimethoprim 10 mg/L Pharmaceuticalwastewater




HPLC-DAD - The removal of theantibiotic by reverseosmosis and the tightnanofiltration membraneis acceptably high(rejection factor> 0.888).


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Trimethoprim COD¼ 360or 590 mg/L


Primary

treatment:ScreeningAerated gritremovalClarifier

Secondary treatment:Activated sludgeMembranereactor(1) or fixed-bedreactor (2)ClarifierSand filtration

Treatment plant 1:Activated sludgetreatment withV¼ 5600 m3, t¼ 15 hMembrane reactorwith V¼ 16 m3,t¼ 13 hSand filter withV¼ 288 m3,t¼ 25 min

Treatment plant 2:Activated sludgetreatment withV¼ 9100 m3, t¼ 31 hFixed-bed reactorwith V¼ 1500 m3, t¼ 1 hSand filter withV¼ 360 m3, t¼ 6e8 h

LCeMS - In treatment plant 1, apre-treatment withactivated sludgewas studied.Bioreactor consistedof a stirred anaerobiccompartment and adenitrification andnitrification cascade.Similar eliminationswere observed in thesecondary treatmentof two conventionalactivated sludgesystems and afixed-bed reactor.Slight eliminationof up to 20% wasobserved in conventionalactivated sludgetreatment and thefixed-bed reactor.

Göbelet al. (2007)

Trimethoprim 50 mg/L Distilled water Adsorption onpowered andgranularactivatedcarbon

pH¼ 4e10T¼ 25 �C0e3 g/L adsorbent

UV at 278 nm The adsorption isothermon both activated carbonscould be fitted withToth equation.The powered activatedcarbon was moreefficient than granularactivated carbon toremove the trimethoprim.However the separationfrom aqueous solutionwas not easy, so thatgranular activatedcarbon was used instead(% removal> 90% after30 min at pH¼ 4 and2.0 g/L adsorbent).

Kimet al. (2010)

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-

-

-

-

-


transported to the surfaces of the media, where an attachmentmechanism such as electrostatic attraction, chemical bond oradsorption retains the particles (Eckenfelder, 2007). This processhas the disadvantage of not degrading the contaminant, butconcentrating it in the solid phase, generating a new waste.

Coagulation/flocculation/sedimentation employs chemicals toenhance the solids sedimentation, pollutants precipitation andcolloids formation and subsequent settle out. The most usedchemicals are lime, alum, iron salts and polymers (Eckenfelder,2007). These techniques require a subsequent treatment toremove the pollutants (in a coagulated form) from the effluents.

Over the last ten years, several conventional techniques havebeen tested in antibiotics removal from environmental matrices.When the pollutants toxicity against the microorganisms used inthe biological systems is very low, this method continues to be thebest approach. Chelliapan et al. (2006) and Arikan (2008) studiedthe application of anaerobic processes in the removal of macro-lides and tetracyclines, respectively. In these two cases, a reduc-tion of 90% for macrolides and 75% for tetracyclines was achieved.As mentioned above, this removal does not always occur. Göbelet al. (2007) tested the traditional primary and secondary treat-ments used in the WWTPs in matrices containing macrolides,sulphonamides and trimethoprim. In this case slight eliminations(z20%) were verified. Adams et al. (2002), Stackelberg et al.(2007) and Vieno et al. (2007) investigated the efficiency ofsome physicochemical methods, such as clarification, coagulation/flocculation/sedimentation and filtration. They were applied tomacrolides, sulphonamides, quinolones, quinoxaline derivativesand trimethoprim, resulting in low removals (maximum removalof about 30%).

Due to the low efficiencies of these methodologies and some-times the inability of their use, new alternatives have emerged.

3.2. Oxidation processes

3.2.1. ChlorinationDue to its low cost, chlorine gas or hypochlorite have been

frequently applied in the disinfection of drinking water treatmentplants. They are currently used as a post-treatment, in order tomaintain a disinfectant residual level in the distribution systems(Acero et al., 2010). However, some studies also refer the chemicaloxidation using chlorinated species in wastewater treatments. Theapplication of this technique for treating water containing phar-maceuticals before the application of biological treatments hasbeen employed in order to oxidize them to readily biodegradableand less toxic compounds.

From the chlorinated species, hypochlorite has the higheststandard oxidation potential (E0¼1.48 V), followed by chlorine gas(E0¼1.36 V) and chlorine dioxide (E0¼ 0.95 V) (Sharma, 2008).Chlorine gas hydrolyses in water according to the reaction:

Cl2þH2O/HOClþ Cl�þHþ (1)

Hypochlorous acid (HOCl) is a weak acid that dissociates inaqueous solutions in hypochlorite (ClO�) and Hþ. For pH> 4,Cl2 hydrolysis is almost complete and consequently hypochlorousacid and hypoclorite are the main chlorine species (Acero et al.,2010). However, among the different aqueous chlorine species,hypochlorous acid is the major reactive form during water treat-ment. Due to its oxidizing power and its chemical structure,hypochlorous acid can react with organic compounds throughoxidation reactions, addition reactions to unsaturated bonds orelectrophilic substitution reaction (Deborde and von Gunten,2008). In fact, this species reacts with aromatic rings, neutralamines and double bonds, producing halogenated organic

compounds, some of themwith potentially dangerous carcinogenicactivity (trihalomethanes and haloacetic acids) (Acero et al., 2010;Navalon et al., 2008).

Chlorine dioxide has been used as an alternative to other chlo-rine species because in its reactionwith organic compounds it doesnot form trihalomethanes. Besides that, it is more selective than theother oxidants and reacts with micropollutants through a one-electron exchange reaction e radical reaction (Navalon et al.,2008). It is a free stable radical that does not react with aromatic,hydrocarbons, carbohydrates and molecules containing primaryand secondary amines, aldehydes and ketones. Nonetheless, itreacts with molecules containing phenolic and tertiary aminogroups (Huber et al., 2005a,b; Sharma, 2008).

Few articles of antibiotics degradation using this techniquewere found. Navalon et al. (2008) studied the oxidation of threeb-lactams (amoxicillin, cefadroxil and penicillin G) with chlorinedioxide. They concluded that penicillin reacted sluggishly withClO2, whereas amoxicillin and cefadroxil were highly reactive(both have a phenolic group). The authors also studied the influ-ence of ClO2 dose and pH on the process. They concluded that theClO2 reacted stoichiometrically with these antibiotics and theinfluence of pH was directly related to the compound structure(e.g. the reactivity of chlorine dioxide towards penicillin wasenhanced with the pH decrease, while for amoxicillin and cefa-droxil an opposite situation occurred). A total degradation wasachieved after 2 h for penicillin and 1 min for the othercompounds. Although, degradation metabolites were detected,their toxicity was not discussed.

Adams et al. (2002) also studied the degradation of sulfon-amides, trimethoprim and carbadox at a concentration level of50 mg L�1 with 1.0 mg L�1 of Cl2. They also concluded that oxidationwith chlorine was effective in the removal of the studied antibiotics(>90%). However, they verified that natural organic matter influ-enced the oxidation process, comparing the reaction rates in riverand distilled water. The slower reaction rates in river water sug-gested that organic matter may complex or otherwise interact withthe studied compounds, reducing reactivity. The authors alsodetected the formation of chlorinated by-products, which shouldhave higher toxicity than the original compounds. A similarconclusionwas obtained by Stackelberg et al. (2007), which studiedthe degradation of macrolides and sulphonamides, using NaClO.

From the bibliographic research done, the authors conclude thatthis technique seems to be efficient in the degradation of antibioticspresent in matrices with low loads of organic matter, such asdrinking water. In addition, the degradation rates are also influ-enced by pH. This technique has been replaced by advancedoxidation processes in order to avoid the formation of halogenatedspecies, which are potentially carcinogenic.

3.2.2. Advanced oxidation processesThe recalcitrant nature of the effluents containing antibiotics

residues interferes in the elimination of these compounds bytraditional biological treatments. In these cases, one alternative isto apply advanced oxidation processes (AOPs).

AOPs are oxidative methods based on the generation of inter-mediate radicals, the hydroxyl radicals (OH�), which are extremelyreactive and less selective than other oxidants (e.g. chlorine,molecular ozone,.). Its standard oxidation potential (E0¼ 2.8 V) isgreater than the conventional oxidants, making them extremelyeffective in the oxidation of a great variety of organic compounds(Hernandez et al., 2002; Bautitz and Nogueira, 2007). These radicalsare produced from oxidizing agents such as ozone (O3) or hydrogenperoxide (H2O2), often combined with metallic or semiconductorcatalysts and/or UV radiation. In these processes, it is expected thatorganic compounds are oxidized to less refractory intermediate


species (less toxic andmore biodegradables) or evenmineralised toCO2 and H2O. Sometimes, the produced metabolites are potentiallymore dangerous than the parent compound (Dantas et al., 2008).

Examples of AOPs include ozonation, Fenton, photo-Fenton,photolysis, semiconductor photocatalysis and electrochemicalprocesses.

3.2.2.1. Ozonation. The ozone is a strong oxidant (E0¼ 2.07 V)capable to act direct or indirectly. For a direct oxidation withmolecular ozone (in this case, this is not a AOP method), it isrequired that the study-compounds have carbonecarbon doublebonds, aromatic bonds or nitrogen, phosphorous, oxygen orsulphur atoms (Ikehata et al., 2006) since it only reacts selectivelywith nucleophilic molecules (Stockinger et al., 1995). Otherwise,the decomposition of ozone in water to form hydroxyl radicalsoccurs through the following mechanism (Andreozzi et al., 1999),where hydroxide ions initiate the reaction:

O3þOH�/O2þHO2� (2)

O3þHO2�/HO2

� þO3�� (3)

HO2�/HþþO2

�� (4)

O2�� þO3/O2þO3

�� (5)

O3�� þHþ/HO3

� (6)

HO3�/OH� þO2 (7)

According to reactions (2) and (3) the initiation of ozonedecomposition can be artificially accelerated by increasing thepH value.

Side reaction (8) is a fast process and plays an important role inwaters with low dissolved organic carbon and alkalinity (Gunten,2003), since it can reduce the oxidation capacity of the system:

OH� þO3/HO2� þO2 (8)

This technique has the advantage of being applied when theflow rate and/or composition of the effluents are fluctuating.However, the high cost of equipment and maintenance, as well asenergy required to supply the process, constitutes some of thedisadvantages. Mass transfer limitations are also a relevant factor tobe considered in the oxidation process with ozone. These systemsrequire the transfer of ozone molecules from gas phase to liquidphase, where the attack on the organic molecules occurs. In manycases, the ozone consumption rate per unit of volume can be sohigh that themass transfer is the limiting step, reducing the processefficiency and increasing the operating costs (Britto and Rangel,2008). In addition, the ozonation performance is affected by thepresence of organic matter, suspended solids, carbonate, bicar-bonate and chlorine ions and also by pH and temperature(Andreozzi et al., 1999; Gunten, 2003).

Several studies have been developed about the ozonationapplied to waters contaminated with antibiotics. For example,Andreozzi et al. (2005), Balcio�glu and Ötker (2003), Arslan-Alatonet al. (2004), Cokgor et al. (2004) and Arslan-Alaton and Caglayan(2005, 2006) studied the degradation of b-lactams using thistechnique. They concluded that although high removal efficiencieswere achieved (COD removals> 50%), the degree of mineralisationwas low (z20%), even for long treatment times. All the authorsstudied the pH influence on the process. They all concluded thatdegradation rates increased with increasing pH, as a consequenceof enhanced ozone decomposition rates into free radicals,

improving the mass transfer rates. This is a critical point in theozonation process. If the pH value is not well controlled, a decreaseof pH during the process will occur due to the accumulation ofcarboxylic acids. These will affect the reaction rate and its mecha-nism and also the absorption rates of ozone.

The same study was performed for other antibiotic classes, suchas lincosamides (Qiang et al., 2004), macrolides (Ternes et al., 2003;Huber et al., 2003, 2005a,b; Lange et al., 2006; Lin et al., 2009),quinolones (Balcio�glu and Ötker, 2003; De Witte et al.,2009), sulfonamides (Huber et al., 2003, 2005a,b; Ternes et al.,2003; Dantas et al., 2008; Lin et al., 2009) and tetracyclines (Liet al., 2008), as can be seen in Table 2. Overall, it was found thatfor all these classes studied degradation above 76% occurred,accompanied by low total organic carbon removal and a slightincrease in the effluents biodegradability. Lower degradations wereachieved for the b-lactam antibiotics. The results from the ecotox-icity of treated effluents are not consensual. Some authors suggestthat the metabolites produced are less inhibitory than the parentcompound (Lange et al., 2006), others find that ecotoxicity remainspractically unchanged (Dantas et al., 2008) and there are stillauthors who verified an increase in the effluent toxicity (Li et al.,2008). Therefore, this is an issue that cannot be generalised anddepends on the compound to be oxidized.

In order to improve the performance of this kind of treatment itis possible to combine ozone with UV irradiation, hydrogenperoxide or catalysts. In the first case, the photolysis of ozonewithin aqueous solutions produces directly hydrogen peroxide,which initiates the further decomposition of residual ozone intohydroxyl radicals by the following mechanism (Hernandez et al.,2002):

O3 þ H2O/hnO2 þ H2O2 (9)

H2O2/HO2�þHþ (10)

O3þHO2�/O3

�� þHO2� (11)

O3�� þHþ/O2þHO� (12)

The homolytic cleavage of the hydrogen peroxide by UV lightproduces also hydroxyl radicals:

H2O2/hn2HO$ (13)

The UV light used in this process can degrade some compounds bydirect photolysis or can excite the micropollutants moleculesmaking them more susceptible for hydroxyl radical attack.

Other possibility to enhance the ozonation performance iscombining the O3 with H2O2 e perozonation. The mechanism forthe formation of hydroxyl radicals is similar to those presented forUV/O3, but in this case, hydrogen peroxide is added from anexternal source. The reaction mechanism is described byHernandez et al. (2002):

H2O2 þH2O%H3Oþ þ HO�

2 (14)

O3þHO�/HO2�þO2 (15)

O3þHO2�/HO� þO2

�� þO2 (16)

The production of hydroxyl radicals by perozonation can alsooccur through reactions (17) and (18):


O3þO2��/O3

�� þO2 (17)

O3�� þH2O/HO� þHO�þO2 (18)

This process can be used in turbid effluents, which do not occurwith the previous system. This technique has been investigatedover the past ten years. Balcio�glu and Ötker (2003) studied thedegradation by perozonation of beta-lactams and quinolonesantibiotics, concluding that perozonation brought no advantageover the ozonation. However, other authors verified that theaddition of small amounts of hydrogen peroxide increased theremoval efficiency (up to 15%) and the effluents biodegradability(Arslan-Alaton and Caglayan, 2006; De Witte et al., 2009), but theecotoxicity was not totally removed. However, the addition of highconcentrations of H2O2 will not improve the process efficiency,since it may act as a free radical scavenger.

Ozonation could be applied to fluctuating flow rates andcompositions. However, this process is limited by mass transferissues. So, in comparison with other oxidative methods, it requiresa greater amount of oxidant to treat the same pollutant load. Theresults obtained in different studies show that although highdegradation efficiencies are achieved, the mineralisation degree islow and the ecotoxicity of effluents remains or even increase. Inaddition, this methodology is extremely pH-dependent, requiringa control over thework range. For these reasons and due to the highcost of equipment and the energy required to supply the process,this methodology does not seem to be adequate for the contami-nated water treatment.

3.2.2.2. Fenton and photo-Fenton. In the 1890s, Henry John Horst-man Fenton developed the Fenton’s reagent, a solution of hydrogenperoxide and ferrous ions, which has strong oxidizing properties(Gan et al., 2009). The Fenton’s oxidation can occur in homoge-neous or heterogeneous systems, although the first one has been,until now, the most used.

In the homogeneous oxidation, the Fenton’s reagent consists ofa hydrogen peroxide solution and an iron salt catalyst (ferrous orferric ions) in acidic medium. From this reagent, hydroxyl radicalsare formed through a radical mechanism. The main steps of thereaction mechanism are (Andreozzi et al., 1999; Arslan-Alaton andGurses, 2004; Britto and Rangel, 2008):

Fe2þþH2O2/ Fe3þþOH�þOH� (19)

Fe3þ þ H2O2%Hþ þ FeðHO2Þ2þ (20)

Fe(HO2)2þ/ Fe2þþHO2� (21)

FeOH2þþH2O2/ Fe(OH)(HO2)þþHþ (22)

Fe(OH)(HO2)þ/ Fe2þþHO2� þOH� (23)

OH� þ organic substance/H2Oþ degradationproducts/ CO2þH2O (24)

One way to increase the oxidation process efficiency is itsconjugation with UV radiation e photo-Fenton (González et al.,2007; Trovó et al., 2008, 2009; Elmolla and Chaudhuri, 2009a,b;Bautitz and Nogueira, 2010). The use of radiation can increase theefficiency of this process mainly due to the regeneration of ferrousion and the extra production of hydroxyl radicals by the photolysisof ferric complexes (Eq. (25)):

FeOH2þ/hnFe2þ þ HO� (25)

The production of hydroxyl radicals via direct H2O2/UV photol-ysis (slow reaction) is also possible. The use of solar radiationconstitutes an advantage since it decreases significantly the overallcosts of treatment.

The performance of these processes is affected mainly by pH,temperature, catalyst, hydrogen peroxide and target-compoundconcentration. In fact, the pH value is an extremely importantvariable in the efficiency of Fenton and photo-Fenton processes. ForpH values below 3, the Fenton’s reaction (Eq. (19)) is severelyaffected, causing the reduction of hydroxyl radicals in solution.Hydrogen peroxide is more stable at low pH, due to the formationof oxonium ions (H3O2

þ), which improves its stability and,presumably, greatly reduces its reactivity with ferrous ions (Elmollaand Chaudhuri, 2009a,b). Some authors also suggest that at low pHvalues the amount of soluble iron Fe3þ decreases, inhibiting theradical OH� formation. On the other hand, at pH 1e2 an inhibition ofthe hydroxyl radical formation exists, due to Hþ ions scavenging(Lucas and Peres, 2006):

HO� þHþþ e�/H2O (26)

For pH values above 4, the precipitation of oxyhydroxidesoccurs, inhibiting both the regeneration of the active species Fe2þ

and the formation of hydroxyl radicals (El-Desoky et al., 2010).Besides that, an excessive pH elevation promotes the HO2

� ionsformation and the scavenging of OH� radicals by carbonate andbicarbonate ions. This narrow pH range of operation constitutesone disadvantage, as well as, the common necessity to recover thedissolved catalyst. The heterogeneous system fills these gaps, sincethe catalyst is immobilized in a heterogeneous matrix, allowing towork in all pH range and to recover the catalyst from the treatedeffluent (Bobu et al., 2008).

Usually the increase in temperature affects positively the Fentonand photo-Fenton processes because an increase of kinetic energyoccurs and consequently, the reaction rate also increases. However,it is also possible to occur acceleration in the hydrogen peroxidedecomposition process (Eq. (27)), decreasing the amount availablefor reaction.

2H2O2/ 2H2OþO2 (27)

A decrease in the process efficiency can also take place if anexcess of hydrogen peroxide is used. The recombination of hydroxylradicals (Eqs. (28) and (29)) and the reaction between them and thehydrogen peroxide (Eq. (30)) may explain this fact.

HO� þHO�/H2O2 (28)

HO� þHO2�/H2OþO2 (29)

HO� þH2O2/HO2� þH2O (30)

This kind of system is attractive because it uses low-costreagents, iron is abundant and a non toxic element and hydrogenperoxide is easy to handle and environmentally safe.

As shown in Table 2, several studies have been conducted on theapplicability of these two techniques to different antibiotic classes:b-lactams (Arslan-Alaton and Dogruel, 2004; Arslan-Alaton andGurses, 2004; Trovó et al., 2008; Elmolla and Chaudhuri, 2009a,b;Rozas et al., 2010), imidazoles (Shemer et al., 2006), lincosamides(Bautitz and Nogueira, 2007), quinolones (Bobu et al., 2008; Guineaet al., 2009), sulphonamides (González et al., 2007; Trovó et al.,2009; Pérez-Moya et al., 2010), tetracyclines (Bautitz andNogueira, 2007).


In all these research papers, it can be seen that although Fentonproduces good results (degradation efficiency above 53%, CODremoval> 44%, TOC removal> 20% and a slightly increase in biode-gradability), the photo-Fenton seems to be more efficient (degrada-tion efficiency above 74%, COD removal> 56%, TOC removal> 50%).Comparing at the same conditions the dark-Fenton process withphoto-Fenton, it is clear that the latter method conducts to highdegradation rates, with an improvement on biodegradability and onmineralisation levels. Pérez-Moya et al. (2010) also concluded in theirtests with photo-Fenton, that toxicity increased during the firstreaction stages, gradually decreasing over the time. The oppositeconclusion was drawn by Guinea et al. (2009), who studied theenrofloxacin (quinolone) degradation by Fenton oxidation. Theyconcluded that the initial stages of the process were very efficient,but the process quickly led to the formationof refractorycompounds,increasing the effluent toxicity.

Overall, the presence of UV light in the Fenton process (photo-Fenton) seems to improve the treatment performance. However,photo-Fenton is generally inapplicable to wastewaters with highorganic matter content (high concentrations of COD, such asmunicipal, hospital and antibiotics manufacturing wastewaters),since the turbidity prevents the penetration of UV radiation.Although Fenton process produces lower removal efficiencies andmineralisation, it seems potentially applicable to treat thesematrices. Nevertheless, Fenton and photo-Fenton are both appli-cable to matrices with low COD concentrations, but waters withhigh ions concentration (e.g. seawater) cannot be treated by thesemethods because Cl�, NO3

�, CO32� and HCO3

� are OH� scavengers.As mentioned above, in both cases, it is important to control the pHoperation range in order to prevent the sludge formation (oxy-hydroxides precipitates).

3.2.2.3. Photolysis. The photolysis is the decomposition or disso-ciation of chemical compounds caused by natural or artificial light.Two photo-induced processes are commonly applied: direct andindirect photolysis. In the first case, the organic compounds absorbUV light and may react with the constituents of the water matrix orsuffer self-decomposition (Boreen et al., 2004; Giokas and Vlessidis,2007; Shaojun et al., 2008; Jiao et al., 2008; Trovó et al., 2009).Indirect photolysis involves the photo-degradation by photo-sensitizers like oxygen and hydroxyl or peroxyl radicals (Arslan-Alaton and Dogruel, 2004; Giokas and Vlessidis, 2007). Theseoxidants should be produced by photolysis of humic and inorganicsubstances present in the water matrices or by external addition ofhydrogen peroxide (homolytic cleavage of hydrogen peroxide,producing the hydroxyl radicals) or even ozone. Although bothdirect and indirect processes can occur simultaneously, indirectphotolysis plays the most important role in the half-life of thecontaminants (Giokas and Vlessidis, 2007).

The photolysis performance depends on the absorption spec-trum of the target-compound, radiation intensity and frequency,H2O2 and O3 concentration (if used) and type of matrix (Kümmerer,2009). Natural waters have different substances that may eitherinhibit or enhance the process by scavenging (organic matter) orgenerating oxidant species (humic and inorganic substances likedissolved metals). This technique has proved to be less effectivethat the other ones, in which radiation is combined with hydrogenperoxide, ozone or catalysts. Photolysis under solar irradiation,instead of using mercury vapour lamps (l< 280 nm), has beenconsidered one promising method to degrade antibiotics in naturalaquatic environment (Jiao et al., 2008).

From the results summarised inTable 2, it seems that this processis extremely dependent on the chemical structure of the compound.Only the photo-sensitive compounds are easily degraded. Shemeret al. (2006) studied the photo-degradation of metronidazole

(imidazole) and they concluded that 6e12% removal was achieved.As expected, they conclude that using UV/H2O2 the removalincreased for 58e67%. Shaojun et al. (2008) and Jiao et al. (2008)studied the degradation of tetracyclines, an antibiotic group verylight sensitive. They obtained high removals (about 80%), but verylow TOC removal (14%), which prove the production of intermediatecompounds. In their study, they also found that the toxicity of thetreated effluent was higher than the original one. Shaojun et al.(2008) tested the influence of dissolved organic matter, particularlyhumic acids inphotolysis treatment. They proved that photolysiswasenhanced by low concentrations of these compounds. However, forrelatively high concentrations there was an inhibitory effect, sincehumic acids behaved as irradiation filters. Other authors studied thedegradationofdifferent antibiotics belonging toother classes, suchasquinolonesandsulphonamides, achievingvery lowremovals (Adamset al., 2002; Arslan-Alaton and Dogruel, 2004) or very long reactiontimes to obtain a high degradation (Trovó et al., 2009).

Comparing this method with the others described so far, this isrelatively ineffective in treating aqueous matrices contaminatedwith antibiotics. This technology is only applicable to wastewatercontaining photo-sensitive compounds and waters with low CODconcentrations (e.g. river, drinking waters).

3.2.2.4. Semiconductor photocatalysis. The semiconductor photo-catalysis started after the discovery of the photo-induced splittingof water on TiO2 electrodes. Later, researchers found that illumi-nated semiconductor particles could catalyse a wide range of redoxreactions of organic and inorganic substrates (Fujishima et al.,2007).

In semiconductor photocatalysis, the reactions of oxidativedegradation require the presence of three basic components:a catalytic photo-sensitive surface (typically an inorganic semi-conductor, such as titanium dioxide), a source of photon energy anda suitable oxidizing agent (Calza et al., 2004; Addamo et al., 2005;Baran et al., 2006; Reyes et al., 2006; Hu et al., 2007; Abellán et al.,2007; Palominos et al., 2008; Klauson et al., 2010; Elmolla andChaudhuri, 2010a,b,c). The principle of this methodology involvesthe activation of a semiconductor (typically TiO2 due to its highstability, good performance and low cost) by artificial or sunlight.A semiconductor is characterized by valence and conduction bands,and the area between them is the band gap. The absorption ofphotonswith energy higher than the band gap energy results in thepromotion of an electron from the valence to the conduction band,with a concomitant hole generation in the valence band (Andreozziet al., 1999; Melo et al., 2009).

TiO2/hnTiO2

�e� þ hþ� (31)

These holes have a very high oxidation potential, which isenough to generate hydroxyl radicals from the water molecules orhydroxide ions adsorbed on the semiconductor surface (Andreozziet al., 1999):

TiO2(hþ)þH2Oads/ TiO2þHO ads� þHþ (32)

TiO2(hþ)þHO�/ TiO2þHO ads� (33)

The formed electrons can reduce the dissolved oxygen, creatinga superoxide radical ion (O2

��), which subsequently in convertedinto H2O2 (Eq. (34)e(36)):

TiO2(e�)þO2/ TiO2þO2�� (34)


O2�� þH2O/HO2

� þHO� (35)

2HO2�/H2O2þO2 (36)

The hydrogen peroxide also acts as an electron receptor,generating extra hydroxyl radicals, as follows:

TiO2(e�)þH2O2/ TiO2þHO�þHO� (37)

The adsorbed substrate (RXads) may be directly oxidized byelectron transfer:

TiO2(hþ)þ RXads/ TiO2þ RXads�þ (38)

Overall, the mechanism of semiconductor photocatalysis can bedivided into five main steps: (1) transfer of reactants in the fluidphase to the surface; (2) adsorption of the reactants; (3) reaction inthe adsorbed phase; (4) desorption of the products and (5) removalof products from the interface region.

Recent studies have shown that the degradation mechanismdoes not occur exclusively by means of hydroxyl radicals, but alsothrough other radical species derived from oxygen. The processperformance is affected by catalyst concentration, wavelength,radiation intensity, pH and water matrix. This method has theadvantage of being applied at ambient conditions, with an energygain when the sunlight is used as irradiation source. However, theindustrial applicability of this technology has been widely dis-cussed due to the difficulty of radiation penetration in an aqueoussolution containing a suspension of opaque fine particles and thedifficulty to remove the catalyst at the end of the process (Britto andRangel, 2008). This phase separation process can be timeconsuming and costly. As a result, filtration and re-suspension ofphotocatalyst powder should be avoided. To overcome theseproblems, studies on semiconductors immobilization have beenperformed. Diverse materials have been studied as a TiO2 support.From the practical point of view, the ideal support should havestrong adherence to the catalyst, a high specific surface area, strongadsorption affinity towards the contaminants and the attachmentprocess should not affect the catalyst reactivity (Shan et al., 2010).Studies about the attachment of photocatalyst particles ontosupports that are easily removable, such as glass, silica gel, metal,ceramics, polymer, zeolite, alumina clays, activated carbon havebeen developed. The semiconductors immobilized systems havethe disadvantage of suffering from mass transfer limitation due tothe reduction in specific surface when compared with the tradi-tional systems (Shan et al., 2010).

Diverse authors studied the application of this method todifferent antibiotic classes and they concluded that semiconductorphotocatalysis was very efficient. Klauson et al. (2010) and Elmollaand Chaudhuri (2010a,b,c) studied the application of this tech-nique to b-lactam antibiotics, concluding that degradations above50% occur, associated with high removal dissolved organic carbon(z80%). The sulfonamides degradation was also studied by severalauthors (Calza et al., 2004; Kaniou et al., 2005; Baran et al., 2006;Abellán et al., 2007; Hu et al., 2007). They concluded that highremovals can be achieved using this methodology (>80%) and theyare often accompanied by important mineralisation levels (40e70%TOC removal). The produced intermediate compounds were lesstoxic andmore biodegradable than the parent compounds. Addamoet al. (2005), Reyes et al. (2006) and Palominos et al. (2009) studiedthe tetracycline degradation. As in the previous cases, the degra-dation rates were high (>98%), as well asmineralisation (>50%). Thesame conclusions were withdrawn by Addamo et al. (2005),Palominos et al. (2008) and Chatzitakis et al. (2008), who studiedlincosamides, quinolones and chloramphenicol.

In terms of removal efficiency, this methodology seems to bepromising for the treatment of effluents with low loads of organicmatter (river, groundwater and drinking water). Although semi-conductor photocatalysis has been studied for decades andnumerous papers have been published, this technology has neverbeen practically applied to water/wastewater treatment due to itsvery low value of electrical energy per order.

3.2.2.5. Electrochemical processes. The electrochemical treatmentsare interesting processes to remove toxic organic compounds,applying an effective, versatile, cost-effective, ease and clean tech-nology (Hirose et al., 2005; Jara et al., 2007; Panizza and Cerisola,2009). In the electrochemical processes, the oxidation occurs overanodes (graphite, TiO2, Ti-based alloys, Ru or Ir oxides, boron-dopeddiamond) in the presence of an electrolyte. Pollutants can bedestroyed electrochemically by a direct anodic oxidation, wherepollutants are first adsorbed on the anode surface and thendestroyed through the anodic electron exchange. On the other hand,if the molecules are degraded in the liquid bulk with mediation ofelectroactive species (such as metallic redox couples e Ag(II), Fe(III),Ce (IV), Mn (III) or strong oxidants as H2O2, O3, persulfate, percar-bonate, perphosphate and chlorinated species), which act asintermediaries for electrons transference between the electrode andthe organic compounds, the reaction is classified as indirect (Chianget al., 1995; Panizza and Cerisola, 2009). The process selectiondepends on the nature and structure of the electrode material,experimental conditions and electrolyte composition. Usually thiskind of system is used to prevent electrode fouling.

The efficiency of the direct oxidation depends on the electrodecatalytic activity, diffusion rates of the compounds towards theactive sites of the anode and the applied current, while the indirectoxidation strongly depends on the diffusion rate of secondaryoxidants into the solution, temperature and pH (Saracco et al.,2000; Jara et al., 2007).

According to the authors’ best knowledge, only two papers werewritten on the applicability of electrochemical oxidation to anti-biotics. Hirose et al. (2005) studied the epirubicin (anthracycline),bleomycin (glycopeptides) and mitomycin C degradation andconcluded that only epirubicin was mostly removed. The otherstudy was developed by Jara et al. (2007), who tested the linco-mycin (lincosamide) and ofloxacin (quinolone) degradation. Thefirst compound was hardly oxidized (30%), while the other one wastotally removed (>99%). As in the previous case, no studies weredeveloped about effluent mineralisation.

These kinds of processes seem suitable to treat toxic wastewa-ters with high concentrations of both antibiotics and COD (such asmanufacturing wastewaters). However, the applicability of thistechnology is limited to small flow rates. Moreover, the highoperating cost of the reactor is also a drawback.

3.3. Adsorption processes

Adsorption processes are widely used in industry to removeorganic contaminants. The term adsorption is commonly used todescribe the tendency of the molecules in fluid phase have to adheretoa solid surface. The forcefieldcreates a regionnear the solid surface,whose potential energy is low, resulting in the molecular densityincrease near the surface. The adsorption phenomenon involves thefollowing steps: (i) solute transport in the bulk e adsorbate move-ment by the stagnant liquid film surrounding the adsorbent, (ii) filmdiffusion e adsorbate transport along the film, (iii) pores diffusion e

adsorbate diffusion through the porous structure to the active sites(molecular diffusion in the pore and/or in the adsorbent surface), (iv)adsorption e interaction between adsorbate and porous structure.Depending on the nature of the forces involved, the process can be


divided in chemical adsorption or physical adsorption. In the last one,the forces are relatively weak, involving mainly van der Waals inter-actions. In the cases of chemical adsorption, the electron transfer andchemical bonds formation between adsorbate and solid surfaceoccurs. These interactions are strong and more specific than thoseexistent in physical adsorption and they are obviously limited to themonolayer coverage (Ruthven, 2000).

This technique has the advantage of removing the analytesinstead of producing potentially more dangerous metabolites(Putra et al., 2009; Rivera-Utrilla et al., 2009). However, this processdoes not enable the effective removal of the contaminants, but onlyits transference to a new phase, where they are more concentrated.

The adsorption efficiency is related with the adsorbent proper-ties, namely surface area, porosity and pore diameter (Estevinhoet al., 2007). It is also important to notice that the process effec-tiveness is dependent not only on the trace compounds of concernor even the adsorbent materials, but also on dissolved naturallyoccurring organic matter present in any natural water. A majoreffect of the organic matter is its direct competition for the avail-able adsorption surface/sites (Qui et al., 2007).

The most used adsorbents are granular activated carbons (GACs),but their high cost and difficulty of regeneration are disadvantages(Crisafully et al., 2008). Therefore, the interest for alternativeadsorbents grows up with the purpose of finding new low-costadsorbents, as by-products or waste materials from industrial oragricultural processes. Hazelnuts (Pehlivan and Altun, 2008; Bulutand Tez, 2007; Kazemipour et al., 2008), coconut (Crisafully et al.,2008), walnut (Kazemipour et al., 2008), almond shells (Pehlivanand Altun, 2008; Bulut and Tez, 2007; Estevinho et al., 2008;Ardejani et al., 2008; Kazemipour et al., 2008) and others havebeen used for this purpose applied to different contaminants. Someof these adsorbents require a previous activation treatment (such aschemical or thermal activation) in order to increase their surfaceareas and consequently, the adsorption efficiency. Batch systemsremain the most studied processes. However, only continuousprocesses (packed columns) should be more similar to thoseemployed for treatments at industrial scale and will provide a realinsight into the applicability of this technology.

Although adsorption is a well-known process, in the past tenyears the study of this technology applied to antibiotics removalhas not beenmuch extended. Adams et al. (2002) andMéndez-Díazet al. (2010) studied the batch adsorption on activated carbon ofimidazoles and sulphonamides with trimethoprim, respectively. Inthese two studies about 90% removal was achieved. A similar studywas developed by Kim et al. (2010), but they investigated whetherbatch or continuous adsorption of trimethoprim, obtaining alsoremovals above 90%. Putra et al. (2009) compared the adsorptioncapacity of activated carbon and bentonite, using amoxicillin(a beta-lactam antibiotic). As in the previous cases, high removalefficiencies were achieved (95% for activated carbon and 88% forbentonite). Chen and Huang (2010) analysed the adsorption ofthree tetracyclines antibiotics on aluminium oxide. They concludedthat these compounds were adsorbed (>50%) and besides that,they suffer structural transformations along the process. Therefore,aluminium oxide was capable to catalyse structural trans-formations, phenomenon that was not recognized previously.

These studies revealed that adsorption continues to be aneffective method to remove antibiotics from aqueous effluents.Unlike some of the processes mentioned so far, adsorption can beapplied to waters containing either high levels of organic matter, orhigh antibiotic concentrations. However, in this process only occursthe contaminant transference from the liquid to the solid phase,producing a new solid residue, where the contaminant is concen-trated. This solid waste should be subsequently treated (e.g incin-eration). In the author’s opinion, a great publication lack on

adsorption with low-cost alternative materials and continuoussystems (packed columns) exists.

3.4. Membrane processes

The membrane processes are increasingly used as separationprocesses. However, this technology does not enable the removal ordegradation of the contaminant, but only its transference toa new phase (the membrane), where it is present in a moreconcentrated form.

3.4.1. Reverse osmosis, nano and ultrafiltrationThe reverse osmosis constitutes one of the membrane processes

(Li et al., 2004; Ko�suti�c et al., 2007; Radjenovi�c et al., 2008). Thisdiffusion method is usually applied to remove large molecules andions from liquid effluents, applying pressure to the solution on oneside of a selective semipermeable membrane. The contaminantsare retained in the pressurized side of the membrane and the cleaneffluent passes to the other side. In reverse osmosis the flow occursagainst the concentration gradient. This technology is efficient toreduce high levels of dissolved salts, but has limitations in theremoval of organic compounds.

Reverse osmosis membrane separations are mostly governed bythe properties of the membrane used in the process, which dependon its chemical nature and physical structure (porosity, mechanicalresistance, etc.). Therefore, these membranes should be resistant tochemical and microbial attack, mechanical and structural stableover long operating periods (Meyer et al., 2003). So, polymericmembranes are chosen.

This kind of process does not require thermal energy, but onlyan electrically driven feed pump; they are simple and have highenergy efficiency. However, reverse osmosis has disadvantages. Themembranes can be easily fouled or damaged and they are suscep-tible to be attacked by oxidizing agents. As mentioned above, thesmall pores in the membrane block large molecules, but smallchemicals can pass through the porous membrane. For this reason,carbon filters are often used as a complimentary technique toreverse osmosis (Binnie et al., 2002). Besides that, reverse osmosisis a slow process when compared to other techniques.

As reverse osmosis, nanofiltration and ultrafiltration aremembrane processes (Li et al., 2004; Ko�suti�c et al., 2007; Koyuncuet al., 2008; Radjenovi�c et al., 2008). They are cross-flow filtrations,in which the process takes place on a selective separation layerformed by an organic semipermeable membrane. Nano and ultrafil-tration membranes are usually charged (carboxylic groups, sulfonicgroups) and, asa result, ion repulsion is themain factor in thisprocess.The driving force of the separation process, like in reverse osmosis, isthe pressure difference between the feed and the filtrate side at theseparation layer of the membrane (Ko�suti�c et al., 2007). These tech-niques are able to remove smaller molecules. The size of the retainedmolecules represents the difference between these two processes.

There are different studies about reverse osmosis, nano andultrafiltration applied to the antibiotics removal. In most studies,the percentage removal obtained for differentmembrane types washigher than 90% for all the antibiotic classes studied (Adams et al.,2002; Ko�suti�c et al., 2007; Li et al., 2004; Radjenovi�c et al., 2008).Koyuncu et al. (2008) obtained the lowest values for the removal oftetracyclines (50e80%) and sulphonamides (11e20%).

As in adsorption, these techniques produce a new solid residue(membrane), where the contaminant is concentrated. So far, thesetechniques have been mostly used in combination with othermethodologies. Reverse osmosis, nano and ultrafiltration aresensitive processes to temperature (this parameter significantlyaffects feed pump pressure, the hydraulic flux balance betweenstages and solubility of the dissolved salts in the effluent), organic

Fig. 2. (A) Most studied antibiotic classes (results according to Table 2) and (B) most prescribed antibiotic classes in 2010.


material occurring naturally in the water matrices and theconcentration of the dissolved salts. The occurrence of highconcentration levels of these compounds can cause membranestructure deterioration or fouling.

3.4.2. Ion exchangeIon exchange is a process in which cations or anions in a liquid

medium are exchanged with cations or anions on a solid sorbent. Inthis process the cations are exchanged with other cations, anionswith other anions, and electroneutrality is maintained in bothphases (Choi et al., 2007). The exchange membranes can be clas-sified in anion or cation exchangers. The first one contains posi-tively charged groups (eNH3

þ, eNRH2þ, eNR2Hþ, eNR3

þ) fixed tothe membrane and allows the passage of anions, but reject cations.The cation exchange membranes contain negatively chargedgroups (eSO3

�, eCOO�, ePO32�) and allow the passage of cations,

rejecting anions. Depending on their preparation, the membranescould be classified into homogeneous or heterogeneous. If thecharged groups are chemically bonded to the membrane matrix,the membrane is homogeneous. On the other hand, if the chargedgroups are physically mixed with the membrane matrix, they areheterogeneous (Xu, 2005). The most used membranes are thepolymeric (styrenic and acrylic resins) because they usually haveboth chemical and mechanical stability and great permselectivity(Dickert, 2007). Ion exchange membranes can also be preparedfrom inorganic materials such as zeolites, betonite or phosphatesalts. However, these membranes are too expensive, have badelectrochemical properties and have frequently large pores (Xu,2005; Nagarale et al., 2006).

The ion exchange systems have been used to improve waterquality. Nevertheless, this kind of technique has disadvantages suchas thenecessityofbackwashingand regeneration.Anotherproblem isthe appearance of fouling, which constitutes most of times an irre-versible fixation of organic materials to the resin (Üstün et al., 2007).One advantage is that ion exchange is a reversible process, allowingextended use of the adsorbent resin before replacement is necessary.In this process, semi-continuous columns are often used instead ofbatch systems. As mentioned above, ion exchange systems have thedisadvantage of requiring resin regeneration. For this reason, thesystems are usually designedwithmultiple units inparallel to ensurea continuous flow, when one or more columns require regeneration.

The authors found only two papers using ion exchange asa removal method of antibiotics. Adams et al. (2002) studied theapplicability of a polymeric resin to remove trimethoprim, carba-dox and sulphonamides, concluding that this methodology wasineffective. On the other hand, Choi et al. (2007) also studied theuse of other polymeric resin for the removal of sulphonamides andtetracyclines. They obtained high removal efficiencies (90% forsulphonamides and >80% for tetracyclines), but some organicinterferences were detected.

In the context of antibiotics removal, ion exchange is a techniquerarely used. Besides ion exchange consists of a phase transferprocess (production of a new residue), this method is only effectiveif the antibiotics to be removed possess ionisable groups in itsstructure.

3.5. Combined processes

Considering that treatment processes must be applied indus-trially, it is necessary to study the process integration to maximizethe treatment performance. Therefore, combined processes havebeen developed.

In some cases, the degradation or removal processes can/shouldnot be applied alone. This is the case of biodegradation becausemost microorganisms are sensible to the toxic pollutants. Thus,AOPs have been applied as a pre-treatment step, in which thepollutants are oxidized to by-products that are easily biodegradableand less toxic, preventing the death of microorganisms that arepresent in the subsequent biological treatments (Tekin et al., 2006).The same situation occurs with the reverse osmosis, which isapplied in combinationwith carbon filters. The use of adsorption aspre-treatment, with subsequent treatment by AOPs is also frequent(Klavarioti et al., 2009).

Zhang et al. (2006) studied the combination of Fenton withreverse osmosis in the removal of amoxicillin. First of all, the authorsused a liquideliquid extraction to remove several organiccompounds (with this step 50% TOC removal was achieved). Afterthat, Fenton oxidation was performed, improving the degradation(TOC removal of 38%). Finally, the authors used a reverse osmosissystem, reaching 11% TOC removal and an enhancement in theeffluent biodegradability. The overall TOC removal was 99%.Sánchez-Polo et al. (2008) investigated the simultaneous applicationof ozonation and adsorption in the removal of imidazoles. Witha simple ozonation the authors achieved 90e100% degradation and10e20% mineralisation and they also concluded that by-productsgenerated were highly toxic. The presence of activated carbonduring the ozonation process besides increasing the removal ratealso reduces the by-products toxicity and the TOC removal in about30%. This combined process enables the treatment of watermatriceswith a high content of organic matter (municipal wastewater),which would not be possible if ozononation was applied alone.

Ötker and Akmehmet-Balcio�glu (2005) also studied thecombination of these two techniques for the removal of enro-floxacin (quinolones). 80% of the compound was adsorbed and theozonation was able to completely degrade enrofloxacin adsorbedon zeolite. Augugliaro et al. (2005) studied the removal of linco-mycin (lincosamide) by semiconductor photocatalysis coupledwithnanofiltration. Lincomycin was successfully oxidized (100% degra-dation) and the filtration allowed separation of the photocatalystparticles and the degradation products from permeate. A similar

Fig. 3. Comparison between the most studied methods (A) and the most applied methodology for each class (B) between the years 2000 and 2010.


studywas performed byMolinari et al. (2006) with quinolones. 66%degradation was achieved by photocatalysis, while filtration onlyseparates the catalyst particles from the treated effluent. Sirtoriet al. (2009) studied the degradation of quinolones by solarphoto-Fenton combined with biological treatment. The globalefficiency of dissolved organic carbon removal was 95%, of which33% for solar photo-Fenton and 62% for biological treatment. Totaldegradation of antibiotic was achieved, with 90% mineralisation.The application of photo-Fenton as a preliminary method allowsincreasing the effluent biodegradability, enabling the subsequentuse of biological treatment.

Despite that combinedmethods are not a very commonpractice,they are one of the most powerful processes for the antibioticsremoval from the environment.

4. Assessment of the remediation techniques

Remediation techniques are employed for the removal ofcontaminants from environmental media, for the general protectionof human health and ecosystems. As mentioned above, Table 2provides an overview of the works published in international jour-nals in this area, illustrating the most studied antibiotic classes.Based on this research work, it was verified that b-lactams andsulfonamides antibiotics classes were the most studied (Fig. 2A).However, in accordance with the European Surveillance ofAntimicrobial Consumption (2010) and Muller et al. (2007) themost prescribed antibiotics in Europe were the b-lactams, lincosa-mides and macrolides (Fig. 2B).

A similar study was conducted in order to ascertain whichmethods of degradation and removal were most studied and fromthese, which were more applied to each class of antibiotics (Fig. 3).

Ozonation, Fenton/photo-Fenton and semiconductor photo-catalysis were themost methodologies tested so far. They were also

the most applied to the b-lactam class (the mainly prescribedantibiotics).

5. Conclusions

In the last years, the presence and fate of antibiotics in envi-ronmental matrices have received a special attention by thescientific community. These compounds are persistent and resis-tant to biodegradation, accumulating in the environment. Even atlow concentration levels, in which they are detected, they canproduce harmful effects either in aquatic or terrestrial ecosystems.For these reasons, several degradation/removal processes havebeen studied to solve environmental contamination issues.

Most conventional treatments applied in WWTPs and DWTPs(such as coagulation, flocculation, sedimentation and filtration)were unsuccessful in the removal of these compounds (Vieno et al.,2007; Adams et al., 2002; Göbel et al., 2007), requiring the devel-opment of new efficient methodologies. Due to the recalcitrantnature of the effluents containing antibiotics residues, the appli-cation of the advanced oxidation processes (AOPs) emerge as analternative. Actually, ozonation and Fenton’s oxidation are themosttested methodologies. Although ozonation has the advantage ofbeing applied to fluctuating flow rates and compositions, the highcost of equipment and the energy required to supply the processconstitutes major drawbacks. Several studies report that thistechnique is effective in the antibiotics removal. However, low ratesof mineralisation are achieved even for long treatment times andthe ecotoxicity of the treated effluents remains practicallyunchanged or evenworse, indicating the production of metabolitesthat are more toxic than the parent compounds (Balcio�glu andÖtker, 2003; Cokgor et al., 2004; Andreozzi et al., 2005; Dantaset al., 2008; Li et al., 2008). This method has also been applied tothe most prescribed class of antibiotics, obtaining the same


conclusions (Balcio�glu and Ötker, 2003; Arslan-Alaton et al., 2004;Cokgor et al., 2004; Andreozzi et al., 2005; Arslan-Alaton andCaglayan, 2005, 2006). As discussed above, besides ozonation,Fenton’s oxidation is one of most studied AOPs. When applied asa homogeneous process (more common case), the production ofoxyhydroxides precipitates (if the pH operation range is not wellcontrolled) and the necessity to recover the dissolved catalystconstitute disadvantages. This is another process often applied tothe group of beta-lactam antibiotics, especially when combinedwith UV irradiation (photo-Fenton). In these cases, a completedegradation was achieved, accompanied by an increase in the TOCremoval (mineralisation degree) and an improvement of theeffluent biodegradability (Trovó et al., 2008; Elmolla andChaudhuri, 2009a,b). Therefore, this seems to be a promisingmethod for antibiotic elimination.

Adsorption is another process that has been reported as analternative to oxidation techniques, though not widely applied tothe more prescribed antibiotics. In all studies, this technique wasvery efficient (removals above 80%). However, it has the disad-vantage of producing a new residue. Most of the studies usedactivated carbon, a high cost adsorbent material. In the authors’opinion, there is a publication lack on adsorption with low-costalternative materials, including the agriculture by-products (withor without pre-treatment), that have already been described aseffective for the removal of other type of micropollutants (hazel-nuts, coconuts, walnut, almond shells, apricot stone,.).

Despite that, the combined methods are not a very commonpractice, they are one of the most powerful processes for theantibiotics removal from the environment, reducing drastically thetoxicity of treated effluents. As previously indicated, an AOP fol-lowed by biological treatment or by a membrane or even by anadsorption process is the most usual combined process (Augugliaroet al., 2005; Ötker and Akmehmet-Balcio�glu, 2005; Zhang et al.,2006). These methods are not usually used due to theircomplexity, high operating costs and most of the time due to theirimpracticability of being used in a continuous mode.

In this research, many studies have utilized high initialconcentrations of antibiotics, which are far from those found inenvironmental matrices. This situation occurs because mostanalytical methods applied to quantify the antibiotics have highdetection limits, preventing its detection in the low levels. There-fore, most results in the literature can only illustrate specific situ-ations, such as the treatment of industrial pharmaceutical effluents.In the authors’ opinion, the results obtained could be eventuallydifferent if the concentrations were close to the values detected inthe ecosystem.

From a practical point of view, the combined processes would bethe best solution for the treatment of effluents containing antibi-otics, including those that can use renewable energy resources topower the processes or by-products materials.

Acknowledgements

The authors wish to thank the Fundação para a Ciência e a Tec-nologia (FCT), Portugal, for financial support (SFRH/BD/38694/2007).

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