fluoroquinolone antibiotics in environmental waters: sample preparation and determination

Review Article

Fluoroquinolone antibiotics inenvironmental waters: Sample preparationand determination

The aim of this review is to provide a general overview on the analytical methods proposed

in the last decade for trace fluoroquinolone (FQ) determination in environmental waters.

A large number of studies have been developed on this topic in reason of the importance

of their monitoring in the studies of environmental mobility and potential degradation

pathways. Every step of the analysis has been carefully considered, with a particular

attention to sample preparation, in relationship with the problems involved in the analysis

of real matrices. The different strategies to minimise interference from organic matter

and to achieve optimal sensitivity, especially important in those samples with lower FQ

concentrations, were also highlighted. Results and progress in this field have been

described and critically commented. Moreover, a worldwide overview on the presence of

FQs in the environmental waters has been reported.

Keywords: Analytical methods / Environmental waters / Fluoroquinoloneantibiotics / Matrix effect / SPEDOI 10.1002/jssc.200900753

1 Introduction

Recently, up to 80 pharmaceutical and personal care

products have been detected worldwide – in the low range

of nanograms up to micrograms per litre – in surface water,

groundwater, wastewater effluents [1] and also in soil at

concentrations in the microgram per kilogram and milli-

gram per kilogram ranges [2, 3]. There is concern about the

effects of the entry of these compounds into the environ-

ment, as their behaviour and that of their degradation

products are still largely unknown; potential chronic effects

of long-term and low-level exposures on environmental

organisms and on human health are suspected, in particular

regarding the effects on the endocrine system, due to the

ability of several compounds to act as hormones. Even low

concentrations could negatively affect non-target living

organisms or lead to the development of allergenic response

[4], certain is that these cause an increased bacterial

resistance, as reported in several studies [5–8].

Among different groups of pharmaceuticals, antibiotics

are of special concern: they are administrated in large

quantities to humans and animals to treat diseases and

infections and at sub-therapeutic levels for prophylactic,

metaphylactic and therapeutic purposes and as feed addi-

tives to promote growth in livestock. Antibiotics for human

use end up in wastewater coming from hospital and

municipal emissions, whereas veterinary drugs are excreted

by the animals and are released in the manure [1]. For

decades, liquid manure from livestock farming and sewage

sludge from wastewater treatment plants have been applied

to agriculture fields as a sustainable principle of nutrient

Andrea SpeltiniMichela SturiniFederica MaraschiAntonella Profumo

Department of GeneralChemistry, University of Pavia,Pavia, Italy

Received November 19, 2009Revised December 23, 2009Accepted December 24, 2009

Abbreviations: BMIm-BF4, 1-butyl-3-methylimidazoliumtetrafluoroborate; Chromabond tetracycline, C18-modifiedsilica; C2/ENV1, mixed-phase adsorbent; CIP, ciprofloxacin;

DAN, danofloxacin; DIF, difloxacin; DOC, dissolved organiccarbon; ENO, enoxacin; ENR, enrofloxacin; ENVI Chrom P,

non-ionic highly cross-linked styrene-divinylbenzenecopolymer; ENV, poly(styrene-divinylbenzene) copolymer;ENV1, hyper cross-link hydroxylated poly(styrene-divinylbenzene) copolymer; EN, polystyrene reversedphase; FD, fluorescence detection; FLU, flumequine; FQs,

fluoroquinolones; HLB, hydrophilic–lipophilic-balancedpolymers; LOM, lomefloxacin; MAX, mixed-mode stronganion exchange reverse phase; MCX, mixed-mode strongcation exchange reverse phase; MDL, method detectionlimit; MAR, marbofloxacin; MeOH, methanol; MEP, silicadivinylbenzene–vinylpyrrolidone copolymer; MIP,

molecularly imprinted polymer; MOX, moxifloxacin; MPC,

silica-based mixed-mode strong cation exchange reversephase; MQL, method quantification limit; NOM, naturalorganic matter; NOR, norfloxacin; ODS, octadecylsilylresin; OFL, ofloxacin; PPL, styrene–divinylbenzenecopolymer; SAR, sarafloxacin; SDB-2,

styrene–divinylbenzene copolymer; SPME, solid-phasemicroextraction; STPs, sewage treatment plants; Strata-X,

surface-modified styrene–divinylbenzene polymericadsorbent; TEA, triethylamine; UPLC, ultra-performance LC;

WAX, mixed-mode weak anion exchange reverse phase;

WCX, mixed-mode weak cation exchange reverse phase;

WWTPs, wastewater treatment plants

Correspondence: Professor Antonella Profumo, Department ofGeneral Chemistry, University of Pavia, via Taramelli 12, 27100Pavia, ItalyE-mail: [email protected]: 139-382-528544

& 2010 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.jss-journal.com

J. Sep. Sci. 2010, 33, 1115–1131 1115

recycling. It is well known that the main way for surface

water contamination derives from direct wastewater treat-

ment plants (WWTPs) release, but leaching from soil

cannot be excluded although strong adsorption in soil [9]

seems to indicate a substantial immobility of these

compounds [2]. However, with the rapidly increasing

knowledge about pharmaceuticals entering the environment

via liquid manure and thus probably contaminating our

feed, food and groundwater resources, there is increasing

concern about the potential risks associated with this

common practice [1].

Fluoroquinolones (FQs) are highly useful antibacterial

agents, particularly because of their broad activity spectrum

against bacteria (both Gram positive and negative) and

mycoplasma and for their good oral intake.

The target proteins of FQs are bacterial DNA gyrase

and topoisomerase IV enzymes, essential for DNA replica-

tion and transcription. FQs have a common 4-oxo-1,4-dihy-

Table 1. Molecular structures and dissociation constants (pKa) of the FQs

Antibiotics pKa 1 pKa 2 Refs. Antibiotics pKa 1 pKa 2 Refs.

CIP 5.90 8.89 [17] LOM 5.82 9.30 [28]

DAN 6.07a)–6.32b) 8.56a)–8.73b) [25] MAR 5.51b)–5.69a) 8.02a)–8.38b) [25]

DIF 5.66a)–5.80b) 7.24a)–8.26b) [25] MOX 6.4 9.5 [29]

ENO 6.32 8.62 [17] NOR 6.23 8.55 [17]

ENR 6.27 8.3 [17] OFL 5.97 8.28 [17]

FLE 5.46 8.0 [26] PEF 6.21 7.87 [26]

FLU 6.5 – [17] SAR 6.0 8.6 [17]

LEV 5.70–6.05 7.90–8.22 [27] TOS n.f. n.f. –

n.f., not found; FLE, fleroxacin; LEV, levofloxacin; PEF, pefloxacin; TOS, tosufloxacin.

a) Electrophoretic.

b) UV spectra.

J. Sep. Sci. 2010, 33, 1115–11311116 A. Speltini et al.


droquinoline skeleton, where the pharmacophore unit

consists of a pyridine ring with a carboxyl group, a piper-

azinyl group and a fluorine atom placed at positions 3, 7

and 6, respectively [10]. In Europe and in the USA, these

were introduced for human use in the mid-1980s and

approved for therapeutic treatment of livestock in the

mid-1990s [11].

Both human and veterinary FQs are expected to enter

the environment, as after oral or parenteral application;

most of these are excreted in active form [12], renally

(�80%) and via the faeces (20%). A growing interest is given

to the environmental fate of these compounds since detec-

tion of FQ residues in the natural environment has been

reported in many countries (Section 4), and a number of

analytical methods are currently available to determine

human FQs in urban wastewater, where they are only

partially removed [11, 13] and veterinary FQs in sewage

sludge and sludge-treated soil, because of their strong

adsorption properties [11]. For the time being, however, no

indicative tolerable value of antibiotics has been fixed for the

different environmental compartments [14], although in the

year 1996 the EMEA (European Agency for the Evaluation of

Medicinal products) guideline set a threshold value of

0.1 mg/kg for residues of veterinary pharmaceuticals in soil

and 0.1 mg/L for groundwater [15]. A revised guideline on

environmental impact assessment for veterinary medicinal

products has been published in the year 2008 [16].

Analytical procedures for accurate determination of FQs

in environmental waters have been overviewed. All the steps

of FQ analysis, from sample treatment to final detection,

have been evaluated, highlighting the critical points. The

analytical methods, the different solid-phase extractants and

the sample pre-treatments reported in the literature are

summarised in detail in Tables 1–4. Special attention is

given to the various adsorbent phases to be used depending

on the sample matrix and FQs to be determined, as well as

to the chromatographic conditions. Looking at these tables,

it would be easier to individuate selected studies that in fact

could quite work in relationship with the origin of the

sample to be investigated and the nature of FQs to be

determined.

2 Physical–chemical properties of FQs

Knowledge of the physical–chemical properties of these

drugs is crucial for the development of appropriate trace

analysis methods. This is especially true for those proce-

dures – in fact the majority – that are based on SPE. FQs are

molecules with a zwitterionic behaviour, as they have

functional ionisable groups. Their corresponding pKa values

are reported in Table 1. Contrary to flumequine (FLU), that

presents only the 3-carboxyl group functionality as ionisable

group, many FQs also have the N-4 of the piperazine

substituent (Fig. 1). Under what has been said, the nature of

the solvent can significantly affect the acid-base behaviour of

these compounds. From an environmental standpoint,

considering ecological pH ranges, the acid-base properties

of nitrogen atoms at positions 1, 8 and N-1 of the piperazine

ring can be neglected [17]. Instead, it is worthy of

considering their relatively good solubility in water, due to

polar groups bonded to a lipophilic core. Such structure

determines the strong interaction with natural organic

matter (NOM), as well as the adsorption on soil, due to both

hydrophobic and electrostatic interactions and hydrogen

bonds [9]. In fact, high adsorption coefficients (Kd,solid)

characterise FQs, among which ciprofloxacin (CIP), norflox-

acin (NOR) and ofloxacin (OFL) stand out, showing values

in the range 496–61 000 L/kg [17]. In confirmation of this,

strong adsorption of FQs to sewage sludge was observed,

suggesting it is the main removal pathway of such

compounds from the water stream during wastewater

treatment [13, 18] (Section 4). In aqueous media, stable

1:1 complexes with several cations, i.e. Ca21, Mg21 and

Al31, are formed by ion–dipole interaction with the 4-keto

oxygen and the ionised 3-carboxylic acid groups [3, 17]. FQs

show resistance to hydrolysis and to heat coupled to the

great chemical stability due to heterocyclic ring that makes

them highly persistent contaminants [19, 20]. Furthermore,

resistance to biodegradation is another important factor to

be considered [21], favoured by strong adsorption onto

various types of solid matrices, such as faeces, soil, sludge

and sediments [2]. In support of this, the persistence of FLU

in an artificial marine sediment was proved [22]: no

degradation occurred and moreover, preservation of anti-

bacterial activity in the sediment for at least 180 days was

observed. In contrast to thermal stability, these drugs are

known to be liable to photodegradation. Irradiation in water

leads to loss of fluoride and/or to oxidative degradation of

the amine side chain [23, 24].

3 Analytical methods for environmentaldetermination of FQs

Presented hereafter are the researches published in the last

decade that deal with the determination of FQs in

environmental samples, including urban wastewater,

sewage treatment plant (STP) and WWTP influents/

effluents, ground, natural and surface water. Besides SPE

procedure, other sample preparation methods have been

explored newly, including solid-phase microextraction

Figure 1. Ionisable groups of two FQs showing differentmolecular structures.

J. Sep. Sci. 2010, 33, 1115–1131 Sample Preparation 1117


Tab

le2.

Meth

od

sfo

rth

ed

ete

rmin

ati

on

of

FQ

sin

en

vir

on

men

tal

wate

rsa

mp

les

FQs

Sam

ple

type

SP

E/S

PM

E

sorb

ent

phas

e

Spi

ke(n

g/L)

Rec

over

y(%

)D

etec

tion

LOD

(ng/

L)R

efer

ence

CIP

,D

AN

,D

IF,

ENR

,

FLE,

LEV,

LOM

,N

OR

,

OFL

Urb

anw

aste

wat

erM

PC

50–4

00fo

rpr

imar

yan

d

16–1

33fo

rse

cond

ary

WW

TPs

efflu

ents

72–9

7FD

,

MS

/MS

45–1

35,

prim

ary

efflu

ent;

15–4

5,te

rtia

ryef

fluen

ta)

[38]

CIP

,D

IF,

ENR

,LO

M,

NO

R,O

FL,

SA

R,

TOS

Sur

face

wat

erM

PC

200

75–1

07FD

,M

S8.

6–49

[39]

CIP

,D

AN

,D

IF,

ENR

,

FLE

LOM

,S

AR

Tap

and

min

eral

wat

er

HLB

9–10

0087

–116

inta

pw

ater

;84

–99

inm

iner

alw

ater

FD1–

29[4

]

CIP

,EN

R,

NO

R,

OFL

Was

tew

ater

HLB

200–

1000

85–9

6M

S/M

S1–

8[3

3]

CIP

,N

OR

,O

FLS

urfa

ce,

grou

ndan

d

sew

age

wat

er

HLB

100

for

grou

ndan

dsu

rfac

e

wat

er;

2000

for

STP

influ

ents

;

1000

for

STP

efflu

ents

75–7

8in

grou

ndan

dsu

rfac

ew

ater

MS

1.6–

163a

)[4

7]

CIP

,EN

R,

NO

RR

iver

wat

erH

LB25

–100

78,

79,

76FD

25a

)[3

2]

CIP

,EN

R,

NO

RR

iver

and

was

te

wat

er

HLB

1000

75,

79,

72in

rive

rw

ater

;60

,

91,

49in

seco

ndar

yan

do

60in

prim

ary

WW

TPs

efflu

ents

MS

6.5–

13.2

[35]

CIP

,N

OR

,O

FLR

iver

and

urba

n

was

tew

ater

HLB

500

63,

73,

73in

rive

rw

ater

and

STP

efflu

ents

;99

,93

,93

inS

TPin

fluen

ts

FD,

UV

35in

rive

rw

ater

and

STP

efflu

ent,

100–

120

inS

TPin

fluen

ta)

[31]

CIP

,D

AN

,D

IF,

ENO

,

ENR

LOM

,N

OR

,O

FL,

SA

R

Nat

ural

wat

erH

LB50

060

–111

MS

/MS

1.2–

7.7

[21]

CIP

,EN

R,

NO

R,

SA

RC

hlor

inat

eddr

ink-

able

wat

er

HLB

5045

–67

MS

/MS

0.5–

2[3

6]

CIP

,D

AN

,D

IF,

ENR

,

FLU

,M

AR

,N

OR

,S

AR

Sea

wat

er,

grou

nd

wat

er

HLB

20–2

5010

–60

for

DA

N10

0–25

0

for

MA

R

72–9

9in

seaw

ater

;70

–96

in

grou

ndw

ater

FD0.

4–8

[34]

OFL

Riv

erw

ater

,w

aste

wat

er

HLB

––

MS

500

(was

tew

ater

)[5

7]

CIP

,O

FLS

urfa

cew

ater

HLB

-SD

B2

100

92,

66in

drin

king

wat

erM

S/M

S2–

5a

)[4

8]

CIP

Riv

erw

ater

MC

X10

102

inul

trap

ure

wat

erM

S/M

S0.

082

[41]

CIP

,O

FLW

aste

wat

erM

CX

20�

30in

min

eral

wat

erM

S1.

3–1.

8,S

TPs

efflu

ents

a)

[43]

CIP

Sur

face

wat

erM

CX

200

82.5

inul

trap

ure

wat

erM

S/M

S–

[42]

CIP

,N

OR

,O

FLS

TPs

efflu

ents

WC

X50

–100

087

–94

MS

/MS

2–10

a)

[30]

CIP

,EN

R,

NO

R,

OFL

Was

tew

ater

WC

X,

HLB

1000

80–1

00in

ultr

apur

ew

ater

MS

/MS

3.5–

5.8

(WC

X),

10–1

6.5

(HLB

)a)

[50]

CIP

,EN

R,

NO

R,

OFL

Hos

pita

lw

aste

wat

erC

2/EN

V1

2500

61,

64,

70,

78M

S/M

S15

–25a

)[1

2]

CIP

,EN

O,

LOM

,

NO

R,

OFL

STP

sef

fluen

tsC

2/EN

V1

5000

–100

0034

,34

,76

,35

,85

MS

/MS

–[4

0]

CIP

,EN

R,

FLU

,

MA

R,

NO

R,

OFL

,P

EF

Sur

face

and

grou

nd

wat

er

C18

10–1

0080

–100

MS

/MS

0.8–

4.3

[59]



Tab

le2.

Co

nti

nu

ed

FQs

Sam

ple

type

SP

E/S

PM

E

sorb

ent

phas

e

Spi

ke(n

g/L)

Rec

over

y(%

)D

etec

tion

LOD

(ng/

L)R

efer

ence

CIP

,D

AN

,D

IF,

ENR

,

LOM

,N

OR

,P

EF,

SA

R

Was

tew

ater

MEP

1000

inin

fluen

tsan

d100

in

tert

iary

WW

TPs

efflu

ents

79–1

09in

influ

ents

and

80–1

05

inte

rtia

ryW

WTP

sef

fluen

ts

FD6.

6–63

.5,

prim

ary

efflu

ents

;

1.3–

12.7

,te

rtia

ryef

fluen

ts

[44]

CIP

,EN

R,

NO

R,

OFL

Was

tew

ater

WA

X-H

LB10

0090

–129

inse

cond

ary

and

95–1

14

infin

alW

WTP

sef

fluen

ts

MS

20–4

0,fin

alef

fluen

t[2

8]

CIP

,EN

R,

NO

R,

OFL

Hos

pita

lan

dm

unic

ipal

s

was

tew

ater

SA

X-H

LB10

0–50

0075

–121

FD8.

5–85

[45]

ENR

,M

AR

Sur

face

wat

erW

AX

–HLB

2090

,11

6FD

0.7–

2[3

7]

CIP

,EN

R,

FLE,

LOM

MO

X,

NO

R,

OFL

Sur

face

wat

erH

LB,

Chr

oma-

bond

,Te

trac

y

clin

e

400

93–1

07,

83–1

07FD

1–50

[46]

CIP

,EN

R,

FLE,

LOM

,M

OX

,N

OR

,

OFL

Sur

face

and

was

tew

ater

,

sew

age

slud

ge

Chr

omab

ond

Tetr

acyc

line

400

80–1

05in

surf

ace

wat

erFD

,M

S11

–55

(FD

),0.

3–7.

5(S

IM),

0.6–

6.4

(SR

M)a

)

[49]

CIP

,D

AN

,EN

R,

FLU

LOM

,N

OR

,S

AR

Riv

erw

ater

MIP

-SP

Ead

sor-

bent

50–1

000

15–1

05FD

10–3

00[5

3]

CIP

,N

OR

,O

FL,

LOM

,EN

O

Riv

erw

ater

,w

aste

wat

erC

arbo

xen

1010

PLO

T

100

81.8

–98.

0M

S/M

S7–

29[5

5]

CIP

,EN

R,

LEV,

NO

RS

AR

Sew

age,

seaw

ater

and

grou

ndw

ater

CW

-TP

Rfib

er10

00–5

0000

81–9

2,se

wag

e;85

–112

,se

awat

er;

81–1

16,

grou

ndw

ater

FD10

–200

[54]

FLE

,fl

ero

xaci

n.

a)

LO

Q.



(SPME) and molecularly imprinted polymer (MIP) SPE

(MIP-SPE) for the determination of FQs in environmental

waters. A summary of the analytical methods is reported in

Table 2. The literature mainly refers to SPE of FQs from

water samples coupled with HPLC or CZE analysis followed

by fluorescence detection (FD) and UV or MS detection. In

effect, nowadays SPE is the most used preparation method

for the pre-treatment of liquid samples [21]. A wide range of

adsorbent phases have been tested, based on their different

chemical properties and affinity towards analytes: mixed

anionic and cationic exchange-reverse-phase, hydrophilic–

lipophilic-balanced (HLB), polymeric, copolymeric and

apolar adsorbents. Depending on the kind of cartridge,

extraction and elution were performed under acidic or

alkaline conditions, because of the zwitterionic properties

of FQs. However, the general main trouble to overcome

turned out to be the removal of interferences: as well-

known environmental waters contain variable amounts of

NOM, including humic acids, compounds naturally spread

in soil and water sediments that derive from the decom-

position of organic matter. These constitute a grave

interference in the determination of FQs, because they

involve the decrease both in the recovery rate due to FQ

NOM adsorption and in sensitivity. An approach generally

adopted to prevent the saturation of the solid phase caused

by NOM is the utilisation of appropriate sample volumes

depending on its origin [28, 30–32]. Therefore, the analytical

method to be considered must be able to minimise matrix

effect according to the physical–chemical properties of the

FQs to be determined.

The choice of a suitable eluting solution is equally

essential: in fact, the possibility to co-extract part of the

organic matter compromises chromatogram quality and

therefore separation, with consequent decrease of both

selectivity and accuracy. The ideal eluting solution should

assure quantitative FQ extraction from the SPE and, at the

same time, no elution of interferences. In this context, the

nature of adsorbent phases is fundamental to selectively

adsorb FQs, leaving organic matter in solution.

As proved [21, 33–36], pH is an important parameter,

not only in terms of efficiency in the antibiotics adsorption/

desorption, but also in regard to the enhancement or

suppression of matrix effects. For example, with a polymeric

adsorbent, a reduction of the co-extracted NOM was

achieved working at pH 6. This kind of adsorbent is known

to retain the hydrophobic portion of NOM in minor extent at

neutral pH. On the contrary, no reduction of matrix effects

was observed in finished drinkable water. This can be

justified considering that the hydrophobic fraction of NOM

is more easily removed than the hydrophilic one through

conventional drinking water treatments [36].

According to these considerations, it is evident how

significant is the optimisation of each step throughout the

analytical procedure: the choice of sample pH, the char-

acteristics of the adsorbent phase, the SPE washing solu-

tions, the composition and the pH of the eluent, the

concentration factor.

Another critical step in the sample preparation is

represented by filtration, performed to remove suspended

particles from sample: in fact most studies first filtered the

sample (Table 3), then adjusted pH before SPE, without

considering that filter could strongly influence the recovery,

both in terms of total and dissolved FQ concentration.

Consistent percentages of antibiotics were proved to be

adsorbed on hydrophilic-mixed cellulose esters and cellulose

nitrate, whereas glass fibre and nylon filters assured no

significant loss of analytes [37].

As the ability of some antibiotics to react with free

chlorine, an investigation on its role in their determina-

tion was recently developed [36]. Experiments carried out

on source and treated water demonstrated that ascorbic

acid is an effective chlorine-quenching agent that does not

affect the analysis and the stability of the antibiotics in

water.

3.1 SPE experimental procedures

A large number of methods has been developed based on

SPE followed by LC or CZE separation: an overview of the

extraction details is given in Table 3.

Golet et al. [38] were among the first to investigate the

presence of trace FQs in urban wastewater, in the region of

Zurich (Switzerland). FQs were pre-concentrated on silica-

based mixed-mode strong cation exchange reverse phase

(MPC) disk cartridge, a mixed-mode silica-based material

consisting of a special non-polar octyl-phase and benzene-

sulphonate as a strong cation exchanger, not currently

commercially available. The same adsorbent phase was also

used for the analysis in wastewater effluents and surface

samples by Nakata et al. [39], with a similar extraction

procedure.

C2/ENV1-mixed phase columns were used for extrac-

tion of pharmaceuticals from European STP effluents [40]

and from hospital sewage water [12]. Good results were

obtained [40] only for OFL and lomefloxacin (LOM) (85 and

76%), but not for CIP, enoxacin (ENO) and NOR (around

35%). This drawback was attributed to the large amount of

sodium chloride discharged in sewer that prevents the

employment of MPC cartridges, usable at very low salt

content [40]. Even if C2/ENV1 adsorbent does not show ion

exchange properties, recovery can be strongly improved by

using triethylamine (TEA) in methanol (MeOH), rather

than 100% MeOH. This effect is probably due to secondary

interactions between the drugs and the residual silanol

groups on the particles surface, particularly the ones of the

C2 adsorbent material [12].

A different adsorbent phase, the polymer-based mixed-

mode strong cation exchange reverse phase (MCX), was

employed for the extraction of various therapeutic drugs

including CIP FQ from river water [41] and surface water

[42]. In both studies, good recovery was obtained only in

ultrapure water. Certainly, the reason of the poor recovery in

the environmental samples has to be related to matrix effect



Tab

le3.

Su

mm

ary

of

the

SP

Eco

nd

itio

ns

Ads

orbe

ntS

ampl

efil

trat

ion

Sam

ple

pHP

reco

nditi

onin

gFl

owra

teex

trac

tion

(mL/

min

)

Was

hing

Elut

ion

Ref

.

MP

CC

ellu

lose

nitr

ate

(0.4

5mm

)3

2m

LM

eOH

12

mL

ultr

apur

ew

ater

(pH

3)1

–2.

5m

Lof

5%N

H3

in15

%M

eOH

[38]

MP

CC

ellu

lose

nitr

ate

(0.4

5mm

)3

8m

LM

eOH

18

mL

ultr

apur

ew

ater

(pH

3)1–

2–

4m

L5%

NH

4OH

inM

eOH

[39]

HLB

–4

5m

LM

eOH

15

mL

ultr

apur

ew

ater

––

10m

LM

eOH

1.5%

acet

icac

id[4

]

HLB

Gla

ssm

icro

fibre

(1mm

)3

6m

Lac

eton

e16

mL

MeO

H1

6m

L50

mM

Na 2

EDTA

(pH

3)

10–

3�

2m

LM

eOH

[33]

HLB

Gla

ssfib

re(0

.45mm

)10

2m

Lof

n-h

exan

e,2

mL

ofac

eton

e,10

mL

ofM

eOH

and

10m

Lof

non-

cont

amin

ated

grou

ndw

ater

(pH

10)

2–20

2m

Lof

5%M

eOH

in

2%aq

ueou

sam

mon

ia

4�

1m

LM

eOH

[47]

HLB

Gla

ssfib

re(0

.2mm

)4

5m

LM

eOH

,4

mL

ultr

apur

ew

ater

–U

ltrap

ure

wat

er(p

H4)

4m

LM

eOH

[32]

HLB

Gla

ssfib

re(0

.45mm

)3

10m

LM

eOH

,10

mL

ultr

apur

ew

ater

55

mL

ultr

apur

ew

ater

4m

LM

eOH

1%N

H3

[35]

HLB

Gla

ssm

icro

fibre

4.2

3�

2m

LM

eOH

–eth

ylac

etat

e(1

:1),

3�

2m

L

MeO

Han

d3�

2m

Lac

idifi

edul

trap

ure

wat

er(p

H4.

2),

52–

5m

L5%

MeO

H3�

2m

LM

eOH

3%N

H3

[31]

HLB

Gla

ssfib

re(0

.47mm

)7

1m

LM

eOH

,1

mL

ultr

apur

ew

ater

32

mL

wat

er–M

eOH

(95:

5)5

mL

MeO

H[2

1]

HLB

Nyl

on(0

.45mm

)3

6m

LM

eOH

,3

mL

MeO

H(0

.1%

HC

OO

H),

2�

6m

Lul

trap

ure

wat

er

52�

6m

Lul

trap

ure

wat

er4�

3m

LM

eOH

0.1%

HC

OO

H[3

6]

HLB

0.45

mmm

embr

ane

5.5

5m

LM

eOH

,10

mL

wat

er10

10m

Lw

ater

2m

L0.

01M

NaO

H–A

CN

(75:

25)

[34]

HLB

Gla

ssfib

re(1mm

),

nylo

n(0

.45mm

)

Nat

ural

5m

LM

eOH

,5

mL

deio

nise

dw

ater

105

mL

ofH

PLC

–gra

de

wat

er

2�

4m

LM

eOH

[57]

HLB

-SD

B-2

Gla

ssfib

re4

MeO

H,

ultr

apur

ew

ater

,ul

trap

ure

wat

er(p

H4)

7.5–

10U

ltrap

ure

wat

er(p

H4)

4�

1m

LM

eOH

(HLB

),4�

1m

L

MeO

H–H

CO

OH

(100

:1)

[48]

MC

XG

lass

mic

rofib

reG

F/D

(2.7mm

)

22

mL

MeO

H,

2m

Lul

trap

ure

wat

er–

1m

LH

Cl

0.1

N2

mL

MeO

H,

2m

LM

eOH

5%

NH

4OH

[41]

MC

XG

lass

mic

rofib

reG

F/D

(2.7mm

)

26

mL

MeO

H,

3m

LM

illi-

Qw

ater

and

3m

L

wat

erac

idifi

edto

pH2

20–

2m

LM

eOH

,2

mL

2%N

H4O

H

inM

eOH

,2

mL

0.2%

NaO

Hin

MeO

H

[43]

MC

XG

lass

mic

rofib

reG

F/F

(0.7mm

)

2.5

2m

LM

eOH

,2

mL

2%H

CO

OH

(pH

2.1)

42

mL

2%H

CO

OH

1m

LM

eOH

,2

mL

MeO

H5%

NH

4OH

[42]

WC

XG

F/C

(1.2mm

)3

4m

LM

eOH

,4

mL

wat

er(p

H3)

10–1

510

0m

Lw

ater

(pH

3),

5m

LM

eOH

10m

LM

eOH

–AC

N–H

CO

OH

(20:

75:5

)

[30]

C2/

ENV

10.

45mm

mem

bran

e3

5m

LM

eOH

,5

mL

MeO

H–w

ater

(50:

50)

and

5m

Lw

ater

(pH

3)

35

mL

wat

er(p

H3)

5m

LM

eOH

5%TE

A[1

2]

C2/

ENV

1–

o7

––

Ultr

apur

ew

ater

3�

5m

LM

eOH

2%TE

A[4

0]

C18

–2.

58

mM

amm

oniu

mac

etat

e(p

H2.

5)(A

)

1A

CN

–0.1

%H

CO

OH

(B):

100%

B(2

mL/

min

for

2m

in),

100%

A(2

min

)

2–

4m

L8

mM

amm

oniu

m

acet

ate

(pH

2.5)

[59]

MEP

–3

5m

LM

eOH

,10

mL

wat

er(p

H3)

1–

6m

LM

eOH

2%H

CO

OH

[44]

WA

X-H

LBG

lass

fibre

(0.5mm

)2.

56

mL

ofM

eOH

and

6m

Lof

phos

phor

ic

acid

4.38

mM

�6

–10

mL

MeO

H–H

3PO

44.

38m

M(9

5:5)

[28]

SA

X-H

LB0.

2mm

mem

bran

e4.

52

mL

MeO

H,

2m

Lci

tric

acid

(pH

4)�

32

mL

citr

icac

id(p

H

4)an

d20

mL

ultr

apur

e

wat

er(p

H4.

2)

4m

LM

eOH

[45]



that caused strong signal suppression (above 60%). However,

the analytical method was not conceived only for FQ deter-

mination; on the contrary, it should be underlined that good

results were obtained for most of the other drugs investigated

[42]. This cartridge is a mixed-mode cation exchange reverse

phase in which the strong cation exchanger sulphonic acid

groups are placed on the surface of a poly-divinylbenzene-co-

N-vinylpyrrolidone copolymer and it was also used by Casti-

glioni et al. [43] for the analysis of pharmaceuticals in Italian

wastewater. Among 30 drugs investigated, belonging to

various therapeutic categories, CIP and OFL were also

studied. Although LOQs were competitive (1.8 and 1.3 ng/L

for CIP and OFL, respectively), the method proposed did not

assure good recovery of FQs, with values of about 30%, as

highlighted in the study. According to these considerations,

they attributed the poor recovery to loss during evaporation to

dryness rather than to reduced adsorption on the phase. On

the contrary, satisfactory results have been achieved under

alkaline conditions [44].

Cation exchange resins were also tested by Lee et al. [30]

for the investigation of Canadian sewage: high recovery

(around 90%) was obtained by mixed-mode weak cation

exchange reverse-phase (WCX) resin and this is certainly an

excellent result considering that sewage represents a very

complex matrix, due to the high content of NOM.

Unlike the studies mentioned so far, numerous proce-

dures have been carried out with HLB adsorbent, a copoly-

mer containing lipophilic divinylbenzene and hydrophilic

N-vinylpyrrolidone units. Considering the nature of FQs,

which is present in both acidic and basic functional groups

that can interact with hydrophilic or lipophilic portion of the

adsorbent, HLB seems to be a very suitable material [45] and

currently it is the most common phase used for this

purpose.

As a matter of fact, Miao et al. [33] investigated the

occurrence of antimicrobials in the final effluents from

WWTPs with good recovery and competitive LOD, as well as

Ferdig et al. [46] and Prat et al. [34] in freshwater and

seawater. In this last research, the role of pH onto FQ

adsorption was investigated in the range 2–12. For most

FQs, high retention (98%) occurred in the pH range 5.5–9.

Sarafloxacin (SAR) and difloxacin (DIF) were adsorbed in

the whole pH interval, due to their additional phenyl group

that strengthened the interaction with HLB. Anyway,

working at pH ranging from 5.5 to 7, a good retention

of all analytes was possible, including the most polar

CIP and NOR. These findings were afterwards improved

by Peng et al. [31]. In acidic sample conditions, high

recovery was gained, as reported in the literature [21, 33, 46],

but at the same time, humic and fulvic acid interferences

were increased. For this reason, neutral pH resulted stra-

tegic for reducing the co-extraction of matrix components

[21]. In regard to the eluting step, three eluent reagents were

tested: MeOH solution (95% in water), MeOH with 3% of

ammonia and MeOH-THF (1:1) [31]. FQs were better

extracted by using the second one (75–85%), as also

suggested by Senta et al. [35]. These findings substantiallyTab

le3.

Co

nti

nu

ed

Ads

orbe

ntS

ampl

efil

trat

ion

Sam

ple

pHP

reco

nditi

onin

gFl

owra

teex

trac

tion

(mL/

min

)

Was

hing

Elut

ion

Ref

.

WA

X-H

LBN

ylon

(0.4

5mm

)3

5m

LM

eOH

,5

mL

ultr

apur

ew

ater

,

5m

Lul

trap

ure

wat

er(p

H3)

5m

Lul

trap

ure

wat

er2�

2.5

mL

phos

phat

ebu

ffer

(pH

3)–A

CN

(80:

20)

[37]

Chr

omab

ond

tetr

acyc

line

Gla

ssfib

re(0

.1mm

)10.

45mm

mem

bran

e

4.2

Ethy

lace

tate

1M

eOH

10.

2%ED

TA(p

H4.

2)10

5m

L0.

2%ED

TA2

mL

MeO

H–H

2O(7

5:25

)an

d

2m

LM

eOH

[46,

49]

MIP

-SP

E

adso

rben

t

0.45

mmm

embr

ane

7.5

10m

LH

EPES

buff

er0.

1M

(pH

7.5)

2.5

5m

LA

CN

–wat

er(0

.1M

HEP

ES,

pH7.

5)(1

0:90

)

1m

LM

eOH

2%TF

A[5

3]

Car

boxe

n

1010

PLO

T

Nyl

on(0

.2mm

)8

2�

40mL

wat

er1

MeO

H0.

152mL

MeO

HA

CN

/5m

Mam

mon

ium

form

ate

(pH

3)(8

5:15

)

[55]

CW

-TP

Rfib

reC

ellu

lose

acet

ate

(0.2

2mm

)

Nat

ural

––

Ultr

apur

ew

ater

60mL

PO

LE7.

5%v/

v[5

4]

SA

X,

stro

ng

an

ion

exch

an

ge

ad

sorb

en

t;C

arb

oxen

1010

PLO

T,

po

rus

carb

on

mo

lecu

lar

sieve;

CW

-TP

R,

carb

ow

ax

tem

pla

tere

sin

.



confirmed what was found previously [34]: ACN-aqueous

buffer pH 4 (89:11) gave poor recovery for all drugs with

with the exception of CIP, marbofloxacin (MAR) and NOR.

An improvement was initially reached by increasing the

content of ACN, but the best elution was gained by

the addition of 0.01 M NaOH (pH 12) to the organic solvent.

In fact, this elution proved to be more efficient in compar-

ison with pure MeOH (75–78%), acetone (around 50%)

and ACN (55–58%) tested by Vieno et al. [47] for ground-

water and surface water samples. In this work, the recovery

rate that did not exceed 80% might be related to the alkaline

pH extraction. Good recovery was also gained with acidified

MeOH [4, 36], rather than non-acidified MeOH [32, 33, 46,

47]. This has been attributed to interaction through

hydrogen bonding between the piperazynic amine group

(protonated at pH 6) and the carbonyl on the vinylpyrroli-

done in the HLB resin [36]. On the contrary, in acidified

MeOH, protons could compete with FQs, breaking down

hydrogen bonding with a subsequent efficient removal of

analyte from the solid phase. In particular, poor recovery of

polar FQs was gained by performing the elution with pure

MeOH [46].

WWTP effluents proved to be difficult to extract, with

average recovery around 60% for primary effluent [35]. For

these samples, matrix effect and ion suppression [47] were

decreased by washing the cartridge with 5% MeOH in 2%

aqueous ammonia, with similar results to 5% MeOH solu-

tion [31].

It is evident that sample origin strongly affects recovery.

At this purpose, an interesting investigation [21] pointed out

that ionic strength and organic matter content powerfully

influenced antibiotics recovery. Briefly, deionised mineral

and river samples were fortified at three concentration levels

(20, 100 and 500 ng/L). By placing the focus on river water,

recovery rate was good only for the highest spike, overall

ranging from 60 to 111%, whereas it was not satisfactory at

lower spikes ranging from 8 to 66% and 10 to 46%, except

for enrofloxacin (ENR), 105%. As explained by Tamtam etal., results obtained were different depending on the origin

of water. At spike of 20 ng/L, recovery was relatively low in

surface and deionised water and higher in mineral water.

On the contrary, for spikes of 100–500 ng/L, higher recovery

was obtained in ground and surface water in comparison

with deionised water, reasonably because organic matter

and cations compete with the reactive sites of glass

hindering FQ adsorption. Even for ‘‘cleaner’’ samples such

as drinkable water [36], SPE step must be optimised to

improve recovery (adsorption/desorption of analytes) and, at

the same time, to neutralise interferences that lead to signal

suppression.

The role of EDTA in the recovery rate is worthy of

comment. As said before, often-variable quantities of this

chelating agent were added to samples before extraction [4,

32, 33, 36, 46, 48] to improve recovery. As also FQs give

complexes with cations (i.e. calcium and magnesium [4]),

solutions of EDTA have also been used to condition [33] and

to wash the adsorbent phase [46, 49]. However, indications

are sometimes in conflict. Zorita et al. [50] found that MCX

resin is not significantly influenced by EDTA in the case of

sewage water analysis, unlike WCX, for which opposite

results were reported [30]. Recovery from HLB was also not

influenced by EDTA [50]. Anyhow, the performance of MCX

and HLB were improved by EDTA for ultrapure, surface

[30], tap and mineral water [46]. We can conclude that by

using a cation exchanger SPE, recovery increases when low

concentration of cations are present in the sample, as

expected. On the contrary, it is not so clear whether

adsorption on HLB resin is equally possible for free and

complexed FQs. Reduction of the complexing capacity of

FQs towards metal ions was got also by adjusting the

sample pH to 5.5 [34].

A polymeric-silica-based material similar to HLB,

containing silica divinylbenzene and vinylpyrrolidone

copolymer (MEP) was also tested [44]. This resin worked

very well for sample with pH around 3: in these conditions,

recovery on fortified wastewater was high (70–100%) and

reproducible. The recovery rate decreased gradually till pH

9, to increase again for values higher than 10. Like for HLB,

acidified MeOH assured the best elution.

Significant improvement in the sample cleanup – the

real challenge in such analysis – was possible using anion

exchanger adsorbents prior to traditional SPE. This method

had been already tested by Jacobsen et al. [51] for the deter-

mination of tetracycline, macrolide, sulphonamide antibiotics

and then adapted by Renew and Huang [28] for the simul-

taneous determination of FQs, sulphonamide and trimetho-

prim antibiotics in wastewater collected from effluent

downstream. Mixed-mode weak anion exchange reverse

phase (WAX) and HLB in tandem have also been utilised by

Sturini et al. [37] for the determination of MAR and ENR in

surface water, precisely river and irrigation ditches samples,

in northern Italy. The choice of WAX in tandem with HLB

was related to the high matrix effect observed by using HLB

alone, in particular referring to chromatograms quality. As a

matter of fact, better sample cleanup was gained: no anti-

biotics were in fact retained on the anion exchange cartridge

(FQs are present in the cationic form, at acidic pH), but only

a large amount of NOM, the main interference, that is

negatively charged. Antibiotics were eluted from HLB with

acidified MeOH and with phosphate buffer (pH 3)/ACN.

Acidic elution, already tested on HLB [33, 36, 47], allowed to

obtain higher FQ recovery in comparison with MeOH 100%.

The use of acidic MeOH instead of phosphate buffer [37] is

reserved to MS detection [28]. Moreover, good LODs can be

achieved avoiding evaporation of the extract to dryness and

solvent exchange [37].

An alternative SPE procedure [48] placed HLB on top of

a second cartridge, an SDB-2 adsorbent, made of styrene-

divinylbenzene units with a selectivity over a wide polarity

spectrum [52]. After sample loading, each cartridge was

separately eluted, with MeOH and acidified MeOH,

respectively.

A strong anion exchanger adsorbent was used by

Seifrtova et al. [45], in tandem with HLB. Citric acid buffer



(pH 4) was proved to be better than acetate buffer (pH 5.5)

and phosphoric acid 4.38 mM for the pre-conditioning step,

contrary to that previously found [28]. Samples spiking

(0.1–5 mg/L) gave good recovery in the range 75–121% for all

FQs investigated.

The large number of adsorbents commercially available

suggests that it is functional to have a complete overview of

their properties and of their performance, especially in

reason of the large variety of environmental matrices to be

investigated. It seems right to remark that the various SPEs

show a different behaviour in terms of trapping capacity,

recovery rate, selectivity and capability of removal of inter-

ferences.

The best performance in the sample cleanup was given

by MPC [38] (no more commercially available), strong cation

exchange reverse phase (MM1) [37] and WCX [37, 50], the

last one successfully employed for the extraction of sewage

samples [30]. Carboxylic acid silica-based weak cation

exchanger was found similar to MPC in terms of selectivity,

but with 20% lower recovery. The combination of hydro-

phobic and cation exchange properties was considered the

reason of the more specific FQ adsorption on MPC resin

[38]. HLB and octadecylsilyl resin (ODS) adsorbents did not

assure a complete removal of interferences in samples rich

in organic matter [30, 37, 50].

A better specificity was gained with MCX, although a

quantitative recovery was not possible in sewage [30],

although good results were obtained in WWTP influent/

effluent-spiked samples [44]. In addition, WCX showed good

selectivity and lower LODs than HLB. These instead

exhibited higher trapping capacity and better precision [50].

Other polymeric resins, ENV, ENV1, EN, PPL and HLB, did

not offer selectivity and enrichment efficiency for waste-

water samples [38].

Talking about recovery rate, good performance was

offered by MEP [44], MPC [38, 39], HLB and Strata-X [21, 35,

37] and chromabond tetracycline [49]; HLB was better when

compared with Nexus, MAX and non-ionic highly cross-

linked styrene-divinylbenzene copolymer (ENVI Chrom P)

[21]; MEP turned out to be the most suitable in comparison

with apolar adsorbent, but also with HLB, MAX and ENVI

Chrom P and MCX, although pleasing recovery was

achieved with the last one and on ENVI Chrom P adsor-

bents [44]. Lower extraction of antibiotics was observed on

C18 reverse-phase SPE [35, 37, 38].

3.2 MIP-SPE experimental procedures

Molecular imprinting is a technology to produce polymers

programmed to recognise a target or a class of target

molecules [53]. In SPE based on molecularly imprinted

polymer, the analyte (the template), or closely related

compounds, will remain bound to the polymer allowing

them to be subsequently eluted co-extractives free. Benito-

Pena et al. [53] demonstrated the applicability of a

urea-based MIP for the pre-concentration of seven

FQs in environmental waters followed by HPLC with

FD. SPE cartridges were packed with 150 mg of the

imprinted polymer prepared using ENR as template and

employed for extraction of river water samples: LODs were

found in the range 0.01–0.30 mg/L, comparable to those

reported for the analysis of FQs using commercial SPE

adsorbents. An important advantage is represented by the

preservation of the pre-concentration ability for at least 80

extractions and this is particularly significant considering

that most of the commercial resins once used cannot be re-

activated for a second extraction. These first results are

promising.

3.3 SPME experimental procedures

Recently, several studies have started to use SPME for FQ

determination. SPME offers some advantages with respect

to SPE, among which are little manipulation and smaller

volume of samples [54]. This technique is based on the

adsorption of the analytes on a fibre (extraction step)

followed, in the case of HPLC analysis, by desorption with

organic solvent. As reported in a recent study [55], an

automated on-line in-tube extraction was developed with

good recovery and sensitivity (Tables 2 and 3). Moreover,

today organic solvents are being replaced by surfactant

aqueous solution as desorbing agents. In this case, it is

preferable to use the acronym SPME-MD (SPME with

micellar desorption). A method based on this technique has

been published recently [54] and it was proved to be very

efficient for FQ determination. In particular, better results

were obtained when the POLE surfactant (polyoxyethylene

10 lauryl ether) was used instead of MeOH. Recovery was

good in water samples of different origin and also

competitive LODs were gained. From an analytical point

of view, a very interesting aspect is the possibility to enhance

FQ fluorescence by applying the micellar desorption that

increased the rigidity of the molecules. In fact, this

represents a great advantage in terms of sensitivity.

3.4 Chromatographic separation

As discussed in Section 3.1, the optimisation of the

sample pre-treatment constitutes a very important step

throughout the whole analytical procedure, to which

researchers have devoted many efforts. Besides extraction

and pre-concentration steps, separation is equally important

to completely take advantage of the sensitivity offered by

the detection system. This is especially important when a

large number of analytes has to be simultaneously

chromatographed. An outline of the LC conditions used is

given in Table 4. The optimisation of the HPLC process

was the subject of the study presented by Herrera-Herrera

et al. [4]. They used ionic liquids as additives in mobile

phases for the separation of seven FQs. In particular, four

room-temperature ionic liquids were investigated, differing



Table 4. Analytical details about chromatographic separation and detection

Column T

(1C)

Mobile phase Gradient Flow

rate

(mL/

min)

Detection Ref.

RP-amide C16 (250� 3 mm, 5 mm) 1

pre-column (20� 3 mm)

50 (A) acidified water (pH 2.4)

(B) ACN

5% B, linear gradient to 7% B (17 min),

isocratic elution (5 min), linear gradient to

17% B (13 min), 85% B (5 min)

0.7 FD [38]

Nucleosil RP-C18 (250� 2 mm, 5 mm)

1 pre-column (8� 3 mm)

r.t. (A) 0.1% TFA aqueous

solution, (B) ACN

12% B to 15% B (15 min) 0.3 MS/MS [38]

Based deactivated Genesis RP-

C18ec (250� 2.1 mm, 4 mm)

50 (A) 95% water 1 5% ACN 1

0.1% HCOOH, (B) ACN 0.1%

HCOOH

Isocratic 100% A (1 min), linear gradient to

10% B (4 min), to 15% B (10 min), to 20% B

(10 min), to 55% B (10 min), to 100% B (5 min)

0.2 MS/MS [40]

RP-C18 (150� 4.6 mm, 5 mm) 1 C18

guard column (10� 4 mm, 5 mm)

r.t. (A) water, (B) ACN both 0.1%

HCOOH

Linear gradient from 95% A to 50% A in 15 min 0.8 MS/MS [12]

YMC ODS-AQ S-3 (4� 50 mm) 23 (A) water pH 3–ACN (98:2),

(B) ACN

A–B (5:95), to (45:65) A–B in 25 min 0.2 MS [39]

Discovery RP-amide C16

(4.0� 50 mm)

r.t. 0.5 FD [39]

Kromasil C18 (100� 2.1 mm, 5 mm) r.t. (A) water, (B) ACN both

0.01% HCOOH

Linear gradient from 5 to 90% B in 20 min 0.2 MS/MS [59]

Luna C8 (50� 2 mm, 3 mm) r.t. (A) water 0.1% HCOOH (B)

ACN

10% (B), 100% (B) in 10 min, hold for 2 min,

to 10% (B) in 2 min, hold for 6 min

0.2 MS/MS [41]

Luna C8 (50� 2 mm, 3 mm) r.t. (A) water pH 2, (B) ACN 100% A, linear gradient to 100% B (10 min),

2 min isocratic elution, linear gradient to

100% A (2 min)

0.2 MS [43]

Zorbax SB-C8 (2.1� 150 mm,

3.5 mm)

r.t. ACN–MeOH–HCOOH–water

(6:12:0.5:81.5)

Isocratic elution 0.2 MS/MS [30]

Genesis C18 (2.1� 150 mm, 3 mm) r.t. (A) ACN, (B) 20 mM aqueous

ammonium acetate (0.1%

HCOOH, pH 4.0)

From 12 to 55% A in 8 min, to 100% in 2 min,

hold 100% A for 2 min

0.2 MS/MS [33]

Inertsil C8 (250� 4.6, 5 mm) 1

guard column

r.t. (A) 10 mM oxalic acid buffer

(pH 4)–ACN (89:11), (B) ACN

100% A for 12 min, from 12 to 25% B in 9 min,

to 45% B in 8 min; isocratic elution 32% B

for FLU

1.5 FD [34]

YMC-Pack Pro C18 (250� 4.6 mm,

3 mm) 1 YMC Pro C18 guard

column (10� 4 mm)

r.t. (A) 50 mM HCOOH–MeOH

(78.5:21.5), (B) MeOH

100% A (20 min), gradient up to 90% B 0.8 FD, MS [49]

Zorbax XDB-C18 (2.1� 50 mm,

5 mm) 1 narrow-bore guard

column (2.1� 12.5 mm, 5 mm)

(A) ACN, (B) acetic acid 1% 3% A, linear gradient to 28% A in 12 min,

to 53% in 5 min, hold 53% A for 1 min

0.25 MS [47]

Acquity C18 (50� 2.1 mm, 1.7 mm) r.t. (A) 5 mM aqueous NH4Ac/

acetic acid (pH 4.8), (B)

ACN–MeOH (2:1)

5% B for 1 min, linear gradient to 60% B in

7 min, to 90% B in 2 min, hold isocratic

for 1.5 min

0.4 MS [57]

Phenomenex SYNERGITM hydro-RP

C18 (4 mm)

r.t. – – – MS/MS [48]

Acquity BEH C18 (100� 1, 1.7 mm) 22 (A) water–MeOH–acetic acid

(94.5:5:0.5), (B) MeOH–acetic

acid (99.5:0.5)

100% A–0.2 min, 100% A–1 min, 95% A–5 min,

90% A–8 min, 80% A–10 min, 55% A–11 min,

55% A–13 min, 0% A–15 min, 0% A–16 min,

100% A–20 min, 100% A

0.07 MS/MS [42]

Chromolith performance RP-18e

column (100� 4.6 mm) 1 pre-

column (10 mm)

r.t. 0.025 M phosphoric acid

solution (pH 3)–MeOH (96:4)

Isocratic elution 2.5 FD [32]

YMC Pro C18 (150� 2.1 mm, 3 mm) 1

pre-column (10� 2 mm)

30 (A) 0.1% HCOOH aqueous

solution pH 3, (B) MeOH 0.1%

HCOOH

10% B for 2 min, 20-min linear gradient to

25% B, 5-min linear gradient to 60% B, 6-min

gradient to 80% B, 4-min gradient to 90% B, hold

90% B for 5 min

0.2 MS [35]

Zorbax eclipse XDB-C18 column

(150� 3 mm, 3.5 mm) 1 Phenomenex

guard column (3� 4 mm)

25 (A) water 0.05% HCOOH, (B)

ACN

10% B isocratic for 3 min, to 20% B at 8 min, to

35% B at 17 min, to 50% B at 22 min, to 100% B

at 25 min, hold for 10 min

0.25 FD, UV [31]



in the length of the alkyl chain on the imidazolium cation

and one ionic liquid containing tetraethylammonium, all

with the same counterion: 1-ethyl-3-methylimidazolium

tetrafluoroborate, 1-butyl-3-methylimidazolium tetrafluoro-

borate (BMIm-BF4), 1-hexyl-3-methylimidazolium tetra-

fluoroborate, 1-methyl-3-octylimidazolium tetrafluoroborate

and tetraethylammonium tetrafluroborate. It is well known

that separation of basic compounds on reverse-phase

silica-based columns is quite difficult, as a result of

strong interaction of analyte cationic sites with the anionic

silanols of the stationary phase, which causes peak tailing.

Therefore, elution at low pH is the best choice to separate

FQs, as in the acidic medium, they are in the cationic form.

TEA is the most popular silanol-suppressing agent and, for

this reason, it was initially used by Herrera-Herrera et al., in

comparison with formic and acetic acids as additives to

mobile phase. The use of TEA allowed a complete

separation of the seven drugs, but peak shape of the last

eluted FQs (SAR, DIF) was not satisfactory and much time

(about 40 min) was necessary for their separation. Similar

results were obtained with formic and acetic acids.

Increasing the length of the chain present in the ionic

liquids, a better resolution was possible. As suggested in the

literature, the repulsion between the imidazolium cation

and the ionised FQs also plays an important role in

shortening FQ retention time. Better separation, shorter

retention times and good peak shape were achieved with

BMIm-BF4.

The field of chromatographic separation was also

investigated by Kumar et al. [56], who focused their experi-

ments on comparison between C18 and RP-amide columns

evaluating different parameters: LODs, intra- and interday

RSD%, selectivity factor and resolution. Results suggest that

RP-amide is better than C18 stationary phase. In the last few

years, significant progress has been made in chromato-

graphic analyses by the introduction of ultra-performance

LC (UPLC) [21, 42, 57], associated with sub-2 mm porous

stationary phases (versus 3–5 mm HPLC). This withstands

higher pressures than traditional HPLC and allows higher

peak capacity, greater resolution, increased sensitivity and

shorter analysis time. For instance, in 10-min chromato-

graphic run, 23 [57] and 17 [21] compounds have been easily

Table 4. Continued

Column T

(1C)

Mobile phase Gradient Flow

rate

(mL/

min)

Detection Ref.

Acquity UPLC BEH C18

(100� 2.1 mm, 1.7 mm)

50 (A) water, (B) ACN both 0.01%

HCOOH

90% A for 3 min, to 70% A at 9 min, to

10% A at 10 min, hold for 2 min, to 90% A

in 1 min

0.5 MS/MS [21]

Kromasil ODS C18 (250� 4.5 mm,

5 mm)

r.t. (A) ACN, 10 mM TBAB pH 3

(B)

4% A, 8 min isocratic elution, 8 min

liner gradient to 15% B, 5 min linear

gradient to 25% B

1 FD [44]

C18 (150� 2 mm, 3 mm) 1 C18

guard column (30� 2 mm, 3 mm)

r.t. (A) water 0.1% HCOOH,

(B) ACN

Linear gradient from 10 to 100% B

in 20 min and then held for 2 min

0.2 MS/MS [36]

SB-C18 (150� 2.1 mm, 5 mm) 30 1 mM ammonia acetate,

0.007% glacial acetic acid,

10% ACN (A), ACN (B)

100% A for 2 min, to 8.5% B in 8 min,

18% in 20 min, 50% in 25 min and 10% in

30 min

0.25 MS [28]

Polaris C18-A (150� 2.0 mm,

5 mm) 1 MetaGuard Polaris

(2.0 mm, 5 mm) C18-A pre-column

25 (A) ACN, B (water 0.005%

HCOOH)

10% A for 2 min, linear gradient to

30% A in 13 min

0.2 MS/MS [50]

Nova-Pak C18 (150� 3.9 mm,

4 mm) 1 Guard-Pak C18 (4 mm)

5 mM BMIm-BF4 1 10 mM

NH4Ac (pH 3.0)–ACN (87:13)

Isocratic elution 1 FD [4]

Hypersil C18 (250 mm� 4.6 mm,

5 mm)

r.t. Phosphate buffer (pH 3) –

ACN (80:20)

Isocratic elution 1 FD [37]

AscentisTM RP-amide (150� 4.6,

5 mm)

�25 (A) citrate buffer 0.001 M (pH

4.5), (B) MeOH, (C) ACN

A–B (60:40) for 7 min, to A–C (60:40)

in 2 min

1 UV [56]

Chromolith performance RP-18e

(100� 4.6 mm) 1 pre-column 10 mm

r.t. Phosphoric acid (pH 3)–

MeOH–ACN (92:7:1)

Isocratic elution 1.2 FD [45]

Aqua C18ec (250� 4.6 mm, 5 mm)

1 RP 18 guard column

(4.0� 3 mm, 5 mm)

25 (A) 25 mM H3PO4 (pH 3), (B)

ACN, (C) MeOH

17% B 83% A for 8 min (1 mL/min), to

66% B and 15% C in 17 min (2 mL/min),

hold for 10 min

1–2 FD [53]

Water symmetry C18

(150� 3.9 mm, 4 mm)

r.t. MeOH–water (15:85), pH 2.5 Isocratic elution 1 FD [54]

Capcell Pack C8 (100� 2.1, 5 mm) 40 5 mM aqueous ammonium

formate (pH 3.0)–ACN (85:15)

Isocratic elution 0.2 MS/MS [55]

TBAB, tetrabutyl ammonium bromide.



separated, with optimal efficiency; similar retention times

(within about 15 min) were achieved [42] for the analysis of

28 different drugs. A typical UPLC chromatogram is shown

in Fig. 2.

3.5 CZE analysis

The applicability of CZE-ESI-MS/MS for the separation of

FQs was proved [58]. The process was optimised so that by

using 120 mM ammonium carbonate (pH 9.12) as buffer

solution, eight FQs have been separated within a single run

with good reproducibility.

Ferdig et al. [46] published a CZE combined with FD

method for the determination of FQs in biological and

environmental water samples. Instead of laser, a continuum

light source was used in the FD. Phosphate electrolytes were

investigated as carrier obtaining the best separation with

50 mM phosphate buffer pH 7.55 in ACN–water (40:60). By

coupling CZE with SPE (extraction details are reported in

Table 3), LOQ in the nanogram per litre range have been

achieved (Table 2).

3.6 Detection: Fluorescence and MS

Analysis of real-water samples revealed that sensitivity of

detector is heavily affected by the matrix components. The

choice of the detection system principally fell on MS or FD,

but in the literature there are conflicting opinions on this

topic. Considering the high fluorescence shown by most of

the FQs, FD was often adopted (Table 2) and was considered

suitable for its specificity and its sensitivity [44]. However,

for the analysis of sewage influents/effluents, mass detec-

tion was proved to be best suited as it provided the highest

sensitivity (LOQ 5 ng/L) and selectivity [30]. However, with

suitable cleanup, good FQ separation and LOQ [37] could

also be obtained with FD (Figs. 3 and 4). Ferdig et al. [49]

came to similar conclusions by comparing the two detectors

in the analysis of different environmental samples, obtain-

ing LOQ better by a factor of at least one order magnitude

with MS.

Nevertheless, FD was proved to be suitable for quanti-

fication of antibiotics in different types of sample (Table 2),

contrary to quadrupole TOF MS-based methods [57, 59]. The

quadrupole TOF mode offered the advantage of secure

identification of the target compounds, giving accurate mass

data and therefore optimal selectivity, but did not assure a

competitive sensitivity: method detection limits (MDLs)

were one order of magnitude higher than those achieved

using a single and a triple quadrupole instrument in MRM

mode. For this reason, the method is not suitable in ground,

river and drinking water, where lower amounts of analytes

are present.

Matrix effects induced signal suppression [12, 28, 36,

42] when MS detection was adopted or caused interfering

peaks in the FD.

Figure 3. Chromatogram of a reference mixture of (fluoro)qui-nolones. PIP, pipemidic acid [38]; FLE, fleroxacin; LEV, levoflox-acin; TOS, tosufloxacin.Figure 2. UPLC-TOF total ion chromatogram showing the

separation of 23 pharmaceutical compounds analysed inpositive ion mode including OFL (7) [57].

Figure 4. Chromatogram of a blank surface water spiked with2 mg/L of (fluoro)quinolones and pre-concentrated by SPE:fleroxacin (1); OFL (2); NOR (3); CIP (4); ENR (5); LOM (6); MOX(7); oxolinic acid (8); FLU (9) [49].



Signal suppression was evaluated [28] as the percentage

decrease in signal intensity in a wastewater matrix versusdeionised water, according to the following equation:

Signal suppression ð%Þ ¼ 1� Is � Ix

IDI

� �� 100

where Is is the antibiotic signal intensity in a sample extract

spiked with S amount, Ix the signal related to an unspiked

extract and IDI the signal in deionised water matrix

MeOH–phosphoric acid 4.38 mM (20:80), fortified with Samount of each FQs. Signal suppression was appreciable for

FQs and it caused a decrease of sensitivity: MDLs rose from

2–7 ng/L in ultrapure water to 20–50 ng/L in final effluents

and to 30–90 ng/L in secondary effluents. It is evident that

NOM is the main cause of signal suppression in wastewater,

as indicated by the good correlation between absorbance

(254 nm) and organic carbon content in water. The

concentration of organic matter was also found determinant

[21] in the processing of river water samples with dissimilar

dissolved organic carbon (DOC). No signal suppression was

registered for samples containing 2 mg/L DOC, unlike

samples containing 4 mg/L DOC. Suppression reached a

maximum of 60% for danofloxacin (DAN) and a minimum

of 30% for OFL and ENR.

No signal suppression was noticed in groundwater

extracts, whereas a signal intensity decrease was observed in

the case of STP influents, where over 40% of the signal

intensity was lost for the compounds having retention times

longer than 10 min [47]. At high ACN percentage, the more

hydrophobic components were probably the major cause of

the signal suppression.

In support of the above, low recovery and high method

quantification limits in STP effluents were found to be a

consequence of the consistent amount of the organic matter

in this kind of samples [35].

4 Environmental occurrence of FQs

The presence of these antibiotics in superficial and

municipal water is correlated to the density of population:

municipal and hospital wastewater are the main sources of

pharmaceuticals emission [45].

In fact, variable amounts of FQ antibacterial agents have

been detected in European, American and Asian surface

water. Very dissimilar concentrations are reported in the

literature [59, 32]: in the first investigation, carried out on

surface water and groundwater from Holland and the

Spanish Mediterranean area, NOR, CIP and FLU turned out

to be the most commonly detected FQs, with concentration

of 7.6, 5.5 and 5.8 ng/L, respectively.

On the contrary, in Mondego river (Portugal), ENR and

CIP were found at higher concentrations (from 67.0 to

102.5 ng/L ENR and 79.6 and 119.2 ng/L CIP). CIP and

NOR were quantified in the range 5–18 ng/L in the Glatt

river (Switzerland) [18]. Analysis of surface water in

North Rhine-Westphalia (Germany) confirmed the sporadic

presence of CIP and OFL [48], at concentrations generally

below 10 ng/L, with the exception of one single sample

(20 ng/L OFL).

Higher levels of CIP, between 14.36 and 26.15 ng/L,

were determined in river Po (Italy) [41].

The Ile-de-France region, in northern France, has been

recently investigated by Tamtam et al. [21]. Samples were

collected from Marne, Oise and Seine rivers: 10.5 ng/L OFL

and 13 ng/L NOR were detected in Oise river; NOR was also

quantified in Seine river (18.6 ng/L).

Antibiotics were also analysed in the south of Finland

[47] and in the rivers of the Croatian region, Drava, Sava and

Danube [35]. In comparison with the surveys just

mentioned, FQs were present at lower amounts. In the

Finnish rivers, NOR, OFL and CIP were proved to be

present, although mostly non-quantifiable, except CIP that

has been quantified in 25 ng/L. Otherwise, FQs in Croatia

resulted below MDLs.

Investigation in the territory between Canada and the

USA revealed that lake and river water samples collected

from Detroit, Lansing and Petoskey (Michigan) and from

western Lake Ontario (Canada) showed FQ concentrations

oLOD [39].

Recently, direct contamination of surface water, by

release of farming wastes into irrigation ditches, has been

proved [37]. MAR and ENR have been detected in rivers and

ditches in the flat land of northern Italy. Water collected

from Po and Ticino rivers contained 37 and 27 ng/L of ENR,

respectively, whereas MAR, although detectable, was not

quantifiable. Samples of irrigation ditches, located close to

farms, showed MAR around 9 ng/L and ENR between 11

and 34 ng/L.

Chinese waters have also been explored. In Pearl river,

the largest one in south China, OFL and NOR were present

at concentrations of 439 and 459 ng/L for CIP [31].

As expected, pharmaceuticals in wastewater are present

in greater quantity than in rivers and lakes, where dilution

of waste occurred [49]. WWTPs are able to eliminate a

substantial fraction in the range 70–80% of the total enter-

ing amount [38]. CIP and NOR were proved to be taken

away at different percentages: between 79 and 87% [18], 54

and 76%, and 85 and 92%, respectively [45]. About half ENR

was eliminated [45]. However, the efficiency depends on the

treatment process and, till date, biological treatment allows

the highest removal [60]. In confirmation of this, adsorption

to sewage sludge has been suggested as the main removal

pathway during secondary wastewater treatment [45], as

supported by the high concentrations of OFL (510 mg/kg),

NOR (150 mg/kg), CIP (230 mg/kg) and moxifloxacin (MOX)

(30 mg/kg) found in sewage sludge samples [49].

Two WWTPs that differed from the secondary activated

sludge treatment – UV and chlorination, respectively – have

been compared in terms of removal performance. In the

secondary effluents, drugs concentrations were comparable:

NOR was detected at less than 60 ng/L and the median

concentrations of CIP and OFL were 100–160 and

205–305 ng/L, respectively. In the tertiary effluents, FQ



concentration turned out to be very dissimilar, indicating

that a more efficient removal of FQs occurred by chlorine

treatment [28].

The incomplete elimination of antibiotics has been

extensively proved: NOR, LOM and CIP were quantified as

56–1163, 24–735 and 14–444 ng/L, respectively, in Chinese

urban wastewater collected from primary (after primary

clarification) and tertiary effluents after advanced treatment

with contact filtration [44]; in two raw wastewater samples,

NOR was quantified as 3131 and 240 ng/L, whereas in the

corresponding secondary effluents, concentration decreased

to 1035 and 54 ng/L. CIP was determined at concentration

levels of 1343 and 143 ng/L in raw samples and 427 and

46 ng/L in secondary effluents [35]; analysis of water

samples collected from several WWTPs in the territory of

Zurich (Switzerland) revealed the presence of CIP and

NOR, the most used human FQs in that area, at concen-

tration between 249–405 and 45–120 ng/L, respectively. On

the contrary, no veterinary FQs were detected [38]; OFL was

quantified in WWTP effluents in Michigan, USA [39]: its

concentration in the final effluent was approximately 50%

lower than that found in the secondary treatment effluent,

204 ng/L.

CIP, NOR and OFL are also the most diffused FQs in

the Canadian territory. They have been detected in the final

effluents of eight WWTPs presenting either primary or

secondary treatment processes and either chlorine or UV

disinfection: median concentrations varied from 0.118,

0.050 and 0.094 mg/L, respectively [33].

High amounts of pharmaceuticals were observed in raw

sewage hospital water, where CIP was in the range 3.6–101

and 0.2–7.6 mg/L OFL. Large temporal variability throughout

a 13-h monitoring time was observed as function of flow

rate, indicating that results can be strongly affected by the

sampling strategy [12].

Hospital emission is the primary source of trace FQs in

municipality wastewater [45]. Relatively, high concentra-

tions were observed, generally ranging from about a few

tens of nanogram per litre up to a ten of microgram per litre.

Hospital wastewater sampled in the area around Asahi river

and Sasagase river in Okayama city (Japan) showed

concentration of OFL ranging from 17.5 to 59.2 ng/L,

whereas OFL was quantified in the range 128.1–186.2 ng/L

in STP wastewater of that region [55].

As the affinity of FQs to sludge, important amounts of

such compounds are expected to enter the environment viaSTP effluents. It has been suggested that the main route of

transportation of human antibiotics into the environment is

via STPs [13]. In this study, a screening of human drugs was

conducted both on aqueous matrices (raw sewage water and

STP final effluents) and on solid sludge. In support of the

above, the highest quantity of substances was determined in

sludge. The water treatment applied in the considered STPs,

which included chemical removal of phosphorous, primary

clarification, active sludge treatment with nitrogen removal

and secondary clarification, was proved to significantly

reduce FQs in water. The main degree of removal for CIP,

NOR and OFL was 87% for the first two drugs and 86% for

the last.

The presence of ENO, LOM and OFL in STP effluents

was for the first time attested in the year 2003 [40], in a

monitoring campaign carried out in four European coun-

tries (France, Greece, Italy and Sweden). The survey, aimed

to the identification of more than 20 pharmaceuticals

belonging to different classes, found that OFL and LOM

were the two most abundant FQs and, in particular, OFL

was present in the range 0.12–0.58 mg/L.

Similar concentrations (150–1081 ng/L) were found in

Italian STP effluents [43], whereas CIP was in the range

27–514 ng/L.

Additional data came from southern Ontario, Canada

[30]: the three detected FQs showed concentrations ranging

from the low to the high nanogram per litre levels in all

samples, suggesting the widespread use of these anti-

bacterial agents in that area. With median concentrations of

34 ng/L in the final and 60 ng/L in the primary effluents,

NOR was usually present at lower concentration than the

other two compounds. In comparison, the median concen-

trations of CIP and OFL in these samples were 146 and

179 ng/L (for final effluent samples) and 251 and 148 ng/L

(for primary effluent samples), respectively. The higher

value for OFL found in the final effluents, in comparison

with the primary effluents, was attributed to a de-conjuga-

tion of metabolites due to microbial activity in the sewage

treatment processes.

The FQ occurrence in three Finnish STP effluents was

studied: the drugs considered, and mainly CIP, were

detected and quantified in the influents, but in the effluent

samples, they were present in concentrations near or under

the quantification limit (5.7–36 ng/L) [47]. Higher concen-

trations of antibiotics were determined in the region of

Upper Austria [49]: STP influents contained OFL 1875 ng/L,

NOR 355 ng/L, CIP 340 ng/L and MOX 2175 ng/L; the same

FQs were found in the STP effluents at somewhat lower

concentrations: OFL 240 ng/L, NOR 340 ng/L, CIP 395 ng/L

and MOX 370 ng/L. Some 20 m downstream of the sewage

plant discharge, the river water contained 28 ng/L OFL,

25 ng/L NOR, 28 ng/L CIP and 17 ng/L MOX.

5 Concluding remarks

The determination of FQ antibiotics in environmental

waters at very low concentrations (from the nanogram perlitre to the microgram per litre range) was proved to be a

hard task, being affected by strong matrix effects. These

involved drawbacks throughout the entire analytical proce-

dure, preventing an efficient sample cleanup and lowering

the sensitivity of the detection system. Independently of the

detection techniques following the HPLC separation, the

most critical step is sample preparation and, in particular,

filtration. SPE methods, the most used to pre-concentrate

and to clean the sample, are strongly influenced by the

matrix: salinity and NOM content play a significant role and



hardly affect the quality of the analysis. The choice of

appropriate sample pH proved to be crucial as well, both in

terms of cleanup and recovery. Similarly, the physical–-

chemical properties of the FQs must be considered for the

choice of the adsorbent phase. Other sample preparations

have been recently developed such as MIP-SPE and SPME.

First results seem to be promising in terms of sensitivity

and selectivity. To conclude, various methods showing good

reliability, accuracy and sensitivity are currently available,

and they represent monitoring tools to get information on

the occurrence, behaviour and fate of such compounds into

the aquatic environment, giving a precious contribute to the

understanding of their lifecycle after use.

The authors have declared no conflict of interest.

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[3] Turiel, E., Martın-Esteban, A., Tadeo, J. L., Anal. Chim.Acta 2006, 562, 30–35.

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