Application of advanced MS techniques to analysis and identification of human and microbial...

Trends Trends in Analytical Chemistry, Vol. 26, No. 6, 2007

Application of advanced MStechniques to analysis andidentification of human andmicrobial metabolites ofpharmaceuticals in the aquaticenvironmentSandra Perez, Damia Barcelo

This overview reviews current knowledge on identification, presence and

distribution of drug metabolites in the environment. We compare the met-

abolic routes of environmentally relevant pharmaceuticals in the human

body to those of biotransformations brought about by mixed consortia of

microorganisms in natural and engineered wastewaters and surface waters.

We describe the characteristics of modern mass-spectrometric instrumenta-

tion coupled to liquid-chromatographic separation techniques and review

their application in profiling metabolites. We set out analytical methodol-

ogies for quantitative analysis of drugs and their degradates in sewage

and surface waters and give results of monitoring obtained from their

application.

ª 2007 Elsevier Ltd. All rights reserved.

Keywords: Environmental analysis; Identification; LC-MS; Liquid chromatography; Mass

spectrometry; Metabolite; Microbial degradation; Pharmaceutical; Sewage

Sandra Perez,

Damia Barcelo*

Department of Environmental

Chemistry, IIQAB-CSIC c/Jordi

Girona, 18-26,

08034 Barcelona,

Spain

*Corresponding author.

Tel.: +34 934 006 100x435;

Fax: +34 932 045 904;

E-mail: [email protected]

494

1. Introduction

Pharmaceutically active compounds havecaptured the attention of the scientificcommunity because such pollutants resultnot primarily from manufacturing butfrom widespread, continual use in humanand veterinary clinical practice. The bio-logical activity of these compounds canlead to adverse effects in aquatic ecosys-tems and potentially have an impact ondrinking-water supplies [1].

In the human body, pharmaceuticals canbe transformed to one or more metabolitesand excreted as a mixture of parent com-

0165-9936/$ - see front matter ª 2007 Elsev

pound and metabolites, in which the parentcompound is often the minor component.However, some drugs are poorly metabo-lized and are excreted unchanged. Thedegree of metabolism depends on a numberof parameters, including age, gender andethnicity, the constitution of the patientand the time of administration. Drug-druginteractions caused by enzyme induction orinhibition, as well as enhanced metabolismdue to previous exposure, can also influ-ence the pharmacokinetics of drugs [2].

Both the parent compound and themetabolites enter the aquatic environmentonce they are excreted from the humanbody. Monitoring studies in the environmenthave demonstrated the discharge of phar-maceuticals and their metabolites throughmunicipal wastewater-treatment plants(WWTPs). Although unchanged drugs canundergo biochemical transformationsduring sewage treatment, some studiesindicate that the absence of pharmaceuticalcompounds in treated water does not nec-essarily imply their complete removal. Inmost instances, human drugs are metabo-lized in the body to more polar compoundsthat are more likely to pass through theWWTP. In some cases, pharmaceuticals andtheir human metabolites can be microbiallydegraded in the activated sludge treatment.

ier Ltd. All rights reserved. doi:10.1016/j.trac.2007.05.004

mailto:[email protected]

Trends in Analytical Chemistry, Vol. 26, No. 6, 2007 Trends

Knowledge of the formation of stable metabolites inWWTPs is also important in order to understand theenvironmental fate of the parent compound. Once in theenvironment, these compounds can be transported anddistributed in rivers, streams, and possibly further bio-degraded. For most pharmaceuticals and their biotrans-formation products, these pathways in the aquaticenvironment are largely unknown, and investigationsinto their occurrence in environmental compartmentsare still rare.

Studies have been carried out to investigate their fatenot only in surface waters, but also in sediment and soilenvironments. By nature, most pharmaceuticals are de-signed to be at least moderately water-soluble and topossess half-lives in the human body in the range of hours.Because human and microbial degradates will generallycoexist with their parent compounds in the environment,indicators that summarize all the information on parentsubstances and degradates would be important instru-ments for decision-making and assessment [3].

Recent advances in liquid chromatography with massspectrometry (LC-MS) have allowed detection of drugresidues at ultra-trace levels in environmental samples[4,5]. Compared with gas chromatography (GC)-MS,LC-MS can determine polar analytes without the needfor prior derivatization. This advantage of LC-MS isparticularly attractive when simultaneously analyzingcompounds belonging to structurally distinct groupswhose determination by GC-MS would involve morethan one derivatization reaction.

LC coupled to tandem MS (LC-MS2) affords superiorperformance in terms of sensitivity (i.e. signal-to-noiseratio) and selectivity compared to single-quadrupole MSinstruments (LC-MS).

The application of ion trap (IT)-MS or, more recently,hybrid quadrupole linear ion trap (QqLIT)-MS, some-times in combination with accurate mass measurementsobtained on time-of-flight (ToF) instruments, hasadvanced the identification of drug metabolites inbiological and environmental samples.

This review summarizes studies performed on micro-bial transformation of pharmaceuticals in settingssimulating aquatic environments. We present strategiesfor identifying unknown metabolites originating fromhuman metabolism and microbial degradation based onMS techniques. We review the analysis of these metab-olites and their parent compounds that are released intothe aquatic environment with particular emphasis onremoval efficiencies in WWTPs.

2. Identification with MS

2.1. Advances in MSChemical identification of a biotransformation product isa part of the analytical chemistry that covers assignation

of an analytical signal to a chemical structure that hasundergone some structural transformations. In the fieldof analytical chemistry, quantitative analysis haseclipsed qualitative analysis. Many publications aboutthe occurrence of pharmaceuticals and, to a lesser extenttheir human metabolites, in the environment have beenpublished (see Section 3). LC-MS using atmosphericpressure ionization (API), atmospheric pressure chemicalionization (APCI) and electrospray ionization (ESI) hasconsiderably changed the analytical methods used todetermine polar compounds in aqueous environmentalsamples. ESI has become the most important ionizationtechnique in MS for on-line coupling with LC in analysisand identification of low molecular-mass molecules [6].With single-quadrupole MS instruments, structuralinformation of drug metabolites can be obtained byinducing in-source fragmentation, but collision-induceddissociation (CID) in triple-quadrupole (QqQ) and ITinstruments offers higher selectivity. Modern instru-mentation, including hybrid QqLIT-MS with MS3 capa-bilities, high-sensitivity LITs for MSn experiments andhybrid QqToF-MS machines for accurate mass mea-surements (precision in the low-ppm range) of productions, are powerful tools for identifying biotransformationproducts. Accurate mass measurements at enhancedmass resolution have recently been achieved on a novelQqQ instrument with hyperbola-faced rods. This allowsthe instrument to be operated at improved mass reso-lution while maintaining high transmission efficiency[7]. Irrespective of the structural information afforded bythe distinct mass analyzers, valuable information can beobtained by the retention behavior of transformationproducts on the chromatographic system that is com-monly coupled to the mass spectrometer (e.g., oxidationsyield products eluting earlier than the parent compoundon reversed-phase columns).

IT-MS uses three electrodes to trap ions in a smallvolume. A mass spectrum is obtained by changing theelectrode voltages to eject the ions from the trap. Theadvantages of IT-MS include compact size, relativelyinexpensive instrumentation, the ability to trap andaccumulate ions to increase the signal-to-noise ratio of ameasurement and MSn capabilities. The MSn feature isparticularly attractive for identifying drug metabolitesbecause sequential fragmentation identifies fragmenta-tion pathways that, in many cases, are not as obvious inproduct-ion spectra generated on QqQ or QqToF instru-ments [8]. However, due to the small trapping volume,IT-MS has limited capacity for ion storage, and overfill-ing the IT results in deterioration in the mass spectrumand loss of the dynamic response range due to spacecharging [4]. This phenomenon can become a criticalfactor when ion ratios serve as criteria for compoundidentification in analytical guidelines established byregulatory agencies. Whereas for QqQ instrumentsvariations are typically within 10–15%, the variability in

http://www.elsevier.com/locate/trac 495


IT-MS can be as high as 30%. Trapping ions can also beperformed in LITs that have two major advantages overconventional ITs: a larger ion-storage capacity and ahigher trapping efficiency. The QTRAP system is basedon a QqQ in which the third quadrupole (Q3) can beoperated either as normal quadrupole or in the LITmode. In the LIT mode, the trapped ions are ejectedaxially in a mass-selective fashion and are detected bythe standard detector of the system [7]. QqLIT systemsrequire considerably higher investment than QqQinstruments.

ToF instruments measure the mass-dependent time ittakes ions of different mass-to-charge ratios to move fromthe entrance of the analyzer, where they are orthogo-nally accelerated in a pulsed fashion, to the detector. Full-scan sensitivity, high-mass resolution and mass accuracyprovided by ToF-MS are especially suited to identification

Figure 1. General approach for the identification of biotransformation pro

496 http://www.elsevier.com/locate/trac

of degradates. Mass errors below 2 mDa or 5 ppm areachievable on modern ToF instruments allowing positiveidentification [5]. Even more powerful in terms of con-firmatory analysis are hybrid QqToF-MS systems withwhich MS2 experiments provide fragmentation informa-tion together with accurate mass measurements ofproduct ions (precision in the low-ppm range) [9]. Analternative to hybrid ToF systems is the recently laun-ched LTQ Orbitrap that combines conventional LIT-MSwith an Orbitrap mass analyzer. This system providesoutstanding mass accuracy, mass resolution and reliablehigh-sensitivity MSn performance. Fourier transform MSis also a very high-resolution technique with whichmasses can be determined with very high accuracy.However, the high cost of these instruments currentlyrestricts the circle of potential customers to laboratoriesin the pharmaceutical industry [10].

ducts by mass spectrometry (MSn and accurate mass experiments).


Although these MS techniques are available in manylaboratories, there is little published data on the frag-mentation behavior of pharmaceuticals and metabolites.Ion masses are important for elucidating the structure ofthe biotransformation products because they produceuseful information about the fragmentation pattern ofthe molecule, so Fig. 1 proposes an approach to identi-fying biotransformation products with MSn experiments(LC-[Qq]LIT-MS) and/or confirmation by accurate massmeasurements (LC-[Qq]ToF-MS). The first step in iden-tifying biotransformation products is to acquire MS2 andMSn spectra of the unchanged drug in order to elucidatethe fragmentation pathway of the molecule. Comparisonof the MSn fragment ions of the drug with those obtainedfrom the biotransformation product provides valuableclues in determining the structure (e.g., which parts ofthe molecule undergo modification in the degradation ofthe parent compound and which remain unchanged). Ifnecessary, derivatization of functional groups or H/D-exchange experiments can be performed. Accurate massmeasurements are valuable in unambiguous identifica-tion of postulated structures of biotransformation prod-ucts by calculating elemental compositions. ToF-MS andQqToF-MS have played important roles as high-resolu-tion mass spectrometers.

As with most advanced technologies, the experienceand the skills of the scientists operating them are cru-cial in generating meaningful, reliable results. Even theapplication of the most sophisticated techniques will notbe able to provide the expected outcomes if dataacquisition and interpretation are not done properly.This is particularly true for highly selective MS fordetection of target analytes only. The proper choices ofionization mode, scan range, and fragmentation con-ditions are decisive in metabolism studies. It is alsonecessary to take into account that accurate quantifi-cation of metabolites (e.g., in laboratory degradationstudies of pharmaceuticals) requires authentic stan-dards. Unfortunately, these are not available for themajority of degradates.

2.2. Human metabolism of environmentally relevantdrugsAfter achieving systemic circulation, most drugs aresubject to biotransformation in the human body with theliver being the major site of metabolism. These conver-sions frequently entail the loss of pharmacologicalactivity and an increase in hydrophilicity, therebypromoting elimination. Conversely, some drugs, so-called prodrugs, require metabolic activation in order torelease the biologically active compound, as is the casefor the easily cleavable esters, clofibrate, fenofibrate andenalapril, which form clofibric acid, fenofibric acid andenalaprilat, respectively (Table 1).

The metabolism of drugs in the human body can bedivided into Phase I and Phase II reactions (Fig. 2). The

first type comprises hydrolytic cleavages (see previousexamples), along with oxidations (e.g., aliphatichydroxylation of ibuprofen and diclofenac, ring oxida-tion of propranolol, epoxidation of carbamazepine,N-oxidation of trimethoprim, N-oxidation of acetamino-phen), reductions (e.g., ketone reduction in prednisone),alkylations (e.g., O-methylation of norepinephrine) anddealkylations (e.g., O-demethylation of naproxen anddiazepam, O-deethylation of phenacetin). The secondtype refers to conjugation reactions, in which a usuallypolar group or molecule is transferred to the parent drugor a metabolite formed previously in a Phase I reaction.The most prominent Phase II reaction is glucuronida-tion, comprising transfer of glucuronic acid to phenols,aliphatic hydroxyl, carboxyl (e.g., ketoprofen, clofibricacid, diclofenac), thiol, amine, and hydroxylaminogroups. Less frequent are sulfation (e.g., of hydroxyl-ated diclofenac), N-acetylation (e.g., of sulfamethoxa-zole) and amino-acid conjugation (e.g., of salicylic acidwith glycine).

The extent of drug metabolism in the human bodyvaries greatly among the pharmaceutical compounds,ranging from undetectable (e.g., X-ray contrast mediaiopromide and diatrizoate) to almost complete biotrans-formation (e.g., carbamazepine and diazepam).

2.3. Microbial degradates in the aquatic environmentAlthough an extensive body of literature has beenpublished on transformation of pharmaceuticals in thehuman body, the microbial degradability of these com-pounds as well as their degradation pathways in theenvironment have rarely been examined. In this section,we summarize the available studies on the identificationof microbial degradates of pharmaceuticals in waste-water and surface water (Table 1).

The majority of the studies deal with identifying deg-radation products of anti-inflammatory, analgesics andblood-lipid regulators, since they are the drugs used mostwith antimicrobials. Three metabolites of anti-inflam-matory ibuprofen were identified in biodegradationexperiments with activated sludge from a municipalWWTP in oxic and anoxic conditions [11]. After a lag of1 day, approximately 75% of the initial ibuprofen wasdegraded under oxic conditions but only about 22% wasdegraded under anoxic conditions. After on-line meth-ylation with trimethylsulfonium hydroxide, the metab-olites were identified by GC-IT-MS as hydroxy-ibuprofen,carboxy-hydratropic acid (mainly in anoxic conditions)and carboxy-ibuprofen (both oxic and anoxic). Thesemetabolites are identical to those formed during humanmetabolism (Table 1). Identification of the ibuprofenmetabolites was accomplished by recording electron-impact MSn spectra of their methyl esters to monitor themass transitions reported in the literature. The threeisomers of hydroxyl-ibuprofen (1-hydroxy-ibuprofen,2-hydroxy-ibuprofen and 3-hydroxy-ibuprofen) could


Table 1. Drug metabolites of human and/or microbial origin

Pharmaceutical Structure Human metabolites identified in

biological samples

Ref. Human metabolites

detected in the

environment

Ref. Identified microbial

degradates

Ref.

Ibuprofen [IBU]OH

O

OH

O Hydroxy-ibuprofen [13,44,45] Hydroxyibuprofen [15,30] Activated sludge: [14,31]

Carboxy-ibuprofen Carboxyibuprofen Hydroxyibuprofen

Carboxy-hydratropic acid Carboxy-ibuprofen

Conjugated ibuprofen Carboxy-hydratropic acid

Phenylbutazone

N

N

O

O

Oxyphenbutazone [46] not found [32] not reported

Hydroxylated P.

Metamizole

(=Dipyrone) N

NN

O

S

O

O

OH

N

N

O

4-Aminoantipyrine [47–49] 4-Aminoantipyrine [28,32] not reported

Acetaminoantipyrine Acetaminoantipyrine

Formylaminoantipyrine Formylaminoantipyrine

Antipyrine

(=Phenazone)

4-Hydroxyantipyrine

3-Hydroxymethylantipyrine

Norantipyrine (Desmethyl-A.)

ClofibrateO

O

O

Cl

Clofibric acid [CFA] (+conjugate) [50] Clofibric acid [33] not reported

Bezafibrate

NH

COOH

O

Cl

Glucuronide [51] not reported Activated sludge: [12]

Hydroxylated B. 4-chlorobenzoic acid

Fenofibrate

O

O ClO

O Fenofibric acid (+conjugate) [52] Fenofibric acid [33] not reported

Glucuronide of benzhydrol

Line missing

Tren

ds

Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

498

http

://ww

w.elsevier.co

m/lo

cate/trac

Acetylsalicylic

acid [ASA]O OH

O

O

Salicylic acid (+conjugate) [53] Salicylic acid [15] not reported

Gentisic acid Gentisic acid

o-Hydroxyhippuric acid

Diclofenac [DCF]

HN

Cl

Cl

O

OH4 0-Hydroxydiclofenac [54] not reported not reported

5-Hydroxydiclofenac

3 0-Hydroxydiclofenac

4 0,5-Dihydroxydiclofenac

3 0-Hydroxy-4 0-

methoxydiclofenac

Conjugates of diclofenac and

metabolites

KetoprofenO

COOH

Glucuronide [55] not reported Activated sludge: [12]

3-(Hydroxyl-carboxy-

methyl) hydratopic acid

3-(Keto-carboxymethyl)-

hydratopic acid

Naproxen [NAP] O

COOH

O

O-Desmethyl-naproxen [16] not reported Activated sludge: [12]

O-Desmethyl-naproxen

Carbamazepine

[CBZ]N

OH2N10,11-Dihydro-10,11-

epoxycarbamazepine

[56–58] 10,11-Dihydro-10,11-

epoxycarbamazepine

[35] not reported

10,11-Dihydro-10,11-

dihydroxycarbamazepine

10,11-Dihydro-10,11-

dihydroxycarbamazepine

10,11-Dihydro-10-

hydroxycarbamazepine

2-Hydroxycarbamazepine


10-Hydroxycarbamazepine+

conjugates




Diazepam [DZM]

N

NO

Cl

Desmethyldiazepam [59] Desmethyldiazepam [36] not reported

Atrovastatin[AVT]

N NH

O

HO

F

HO

HO

O

p-Hydroxyatorvastatin [60] p-Hydroxyatorvastatin [37]

o-Hydroxyatorvastatin o-

Hydroxyatorvastatin lactone

o-Hydroxyatorvastatin

A. lactone

(continued on next page)

Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

Tren

ds

http

://ww

w.elsevier.co

m/lo

cate/trac499

Table 1 (continued)

Pharmaceutical Structure Human metabolites identified in

biological samples

Ref. Human metabolites

detected in the

environment

Ref. Identified microbial

degradates

Ref.

Simvastatin

O

OO

OHO Simvastatin hydroxy acid [61] Simvastatin hydroxy acid [37] not reported

Fluoxetine [FLX]

O

HN

CF3

Norfluoxetine (Desmethyl-F.) [62] Norfluoxetine [37] not reported

Sulfamethoxazole

[SMX] NH2 S

O

O

NH

O

N

CH3

N4-Acetylsulfamethoxazole [63] N4-Acetylsulfamethoxazole [38,39] not reported

Glucuronide

Trimethoprim

(TMP)

N

N

NH2

O

O

O

NH2

a-Hydroxytrimethoprim [64,65] not reported Activated sludge: [19]

O-Desmethylated T. a-Hydroxy-trimethoprim

N-oxide of T. Hydroxylated trimethoprim

Line missing

Tren

ds

Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

500

http

://ww

w.elsevier.co

m/lo

cate/trac

Methotrexate

N

NN

NN

HN

O

NH2

NH2

OH

HO

7-Hydroxymethotrexate [66] not reported 7-Hydroxymethotrexate [20]

MestranolOH

O

17a-Ethynylestradiol [67–70] Estrone [40] Activated sludge: [21]

17a-

Ethynylestradiol

17b-Estradiol-

valerate

2-Hydroxy-E. (+conjugates)

Estradiol (+conjugates)

Estrone (+conjugates)

Estriol (+conjugates)

16a-Hydroxyestrone Mestranol fi 17a-

ethynylestradiol

17b-Estradiol fi estrone

OH

HO

OH

HO

Enalapril [ENL]

N

O

HO

O

NH

O O

Enalaprilat [71] Enalaprilat [41] not reported

Iopromide [IOP]

II

I

HN OH

OH

O

NH

O

O

N OHO

CH3 OH not metabolized not reported Activated sludge: [23,25]

Carboxyiopromide

Dicarboxylated iopromide

Bis-dehydroxy-iopromide

5-Amino-N,N 0-bis(2,3-

dihydroxypropyl)-2,4,6-

triiodo-N-

methylisophthalamide

Diatrizoate [DTZ]

II

I

NH

NH

O

OHO

O

not metabolized not reported Activated sludge: [24,26]

3-Acetylamino-5-amino-

2,4,6-triiodobenzoate

Water-sediment system:

3-Acetylamino-5-amino-

2,4,6-triiodobenzoate

3,5-Diamino-2,4,6-

triiodobenzoate

Cocaine

O

O

N O

O

HBenzoylecgonine [72] Benzoylecgonine [42] not reported

Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

Tren

ds

http

://ww

w.elsevier.co

m/lo

cate/trac501

Figure 2. Simplified scheme of drug metabolism in the human body.


not be differentiated unequivocally based on their massspectra alone, but the most likely metabolite wasbelieved to correspond to 2-hydroxy-ibuprofen.Together, the metabolites did not account for more than10% of the initial concentration of ibuprofen.

Quintana et al. [12] also investigated the biodegrad-ability of ibuprofen in sewage sludge, demonstrating thatit was degraded only co-metabolically. Biodegradation ofibuprofen started after a lag of about 5 days and wascomplete after 22 days. Two isomers of hydroxy-ibu-profen were tentatively identified as the major degrada-tion products by LC-(-)ESI-MS2, showing ions of thedeprotonated molecule at m/z 221.

In mammals, hydroxylation at the primary carbon,yielding 1-hydroxy-ibuprofen, was observed, which isthe first step of oxidation of ibuprofen towards carboxy-ibuprofen [13]. However, carboxy-ibuprofen was notfound in the degradation experiment [12].

In biofilm reactors, the degradability of ibuprofen andclofibric acid (a human metabolite of clofibrate and ablood-lipid regulator) was studied by spiking river waterat a concentration of 100 lg/L [14]. The disappearanceof the pharmaceuticals was monitored with GC-IT-MSafter derivatization with trimethylsulfonium hydroxide.The study indicated that ibuprofen was readily degradedin these settings. Two metabolites were detected andidentified as hydroxy-ibuprofen and carboxy-ibuprofen.The methylated hydroxy-ibuprofen showed two charac-teristic fragments from b-cleavage at m/z 178([M-COOCH3]+Æ) and m/z 119 ([M-C3H7]+), whereas themethylated carboxy-ibuprofen showed three character-istic ions at m/z 264 ([M]+Æ), m/z 205 ([M-COOCH3]+Æ)


and m/z 145 ([M+-2 COOCH3]+Æ). These spectra were inagreement with those reported for the human metabo-lites of ibuprofen (authentic standards were not avail-able). Both metabolites were observed to be furtherdegraded in the biofilm reactors. The authors pointed outtwo differences between human and microbial metabo-lism of ibuprofen:� In humans, carboxy-ibuprofen is formed later than

hydroxy-ibuprofen and tends to persist, but, in thebiofilm reactors, the formation and subsequent degra-dation of these metabolites occurred in the oppositeorder; and,

� Ibuprofen occurs as two enantiomers; one of which ispharmacologically active and readily metabolized inhumans, but, in environmental samples, this isomerwas the dominant form detected.In the reactors, the non-pharmacologically-active

isomer was degraded more rapidly than the pharmaco-logically active one. These results imply that the phar-macologically active isomer may be of greater potentialenvironmental concern. Unlike ibuprofen, clofibric acidwas stable over the entire incubation time of up to 3weeks. Its stability and high polarity are believed to bethe major reasons for the widespread contamination ofsurface, ground and even potable water with clofibricacid in countries with high prescription rates of clofi-brate.

In batch experiments with suspended activated sludge,the degradability of acetylsalicylic acid was studied [15].The initial concentration of the compound (10 lg/L)decreased significantly after a lag of 6 h, and, after 24 hand 72 h, elimination of about 70% and >99%,


respectively, was observed. No metabolites could be de-tected with GC-MS in any sample taken during theincubation time.

Quintana et al. [12] reported the degradation of anti-inflammatories and blood-lipid regulators (ketoprofen,naproxen, diclofenac and bezafibrate) in activated sludgefrom a municipal WWTP. For co-metabolic transforma-tion, milk powder was added to the test medium, while, ina second experimental design, the pharmaceutical com-pound served as the sole carbon source. Of the fourpharmaceuticals, ketoprofen was the only compoundbiotransformed in the reactor without milk powder.During ketoprofen degradation, two novel metabolites[3-(hydroxy-carboxy-methyl) hydratropic acid (metabo-lite A) and its oxidative form 3-(keto-carboxymethyl)-hydratropic acid (metabolite B)] were identified byLC-diode array detection (DAD) and LC-(-)ESI-MS2

(Table 1). The UV spectrum of metabolite A had only oneabsorption band around 220 nm, and the MS2 spectrumwas characterized by a deprotonated molecule at m/z223, which initially lost two CO2 molecules to formm/z 179 and m/z 135 followed by the loss of H2O to formm/z 117. These fragmentations suggested the presence oftwo carboxy groups and one aliphatic hydroxy group inthe metabolite. The UV spectrum of metabolite B wascharacterized by two absorption bands, indicating higherconjugation than metabolite A. In addition, MS2 frag-mentation from the precursor ion (m/z 221) was char-acterized by the loss of two CO2 molecules (m/z 177 andm/z 133) and subsequently CO (m/z 105). Based on thesemetabolites, it was proposed that ketoprofen was degradedaccording to the pathway known for biphenyls, biphenylethers and related compounds. Metabolite B remainedstable throughout the experimental period of 28 days. Inthe experiment conducted with an additional carbonsource to identify co-metabolic degradation, ketoprofenand diclofenac were not degraded. Primary degradation ofbezafibrate was complete within 5 days but mineralizationwas not achieved. One metabolite (4-chlorobenzoic acid)was detected. The isotopic pattern of the protonatedmolecule of this compound (m/z 155), with about 30%intensity for the M+2 signal (m/z 157), indicated amonochlorinated compound. The product-ion spectrumshowed the loss of CO2 (m/z 111), which was proposed tocorrespond to a carboxy group in the structure. Based onthese data, the microbial hydrolysis of the amide bond wasproposed. Rapid disappearance of this metabolite wasconfirmed. As for the degradation of naproxen, about 60%of biotransformation occurred within 28 days of theexperiment, producing one detectable metabolite. It wasidentified as O-desmethyl-naproxen based on its deproto-nated molecule (m/z 215) and the product-ion spectrumthat indicated the presence of one carboxy group. Thismethyl-ether cleavage is a well-documented metabolicpathway in mammals [16]. This metabolite was unstableunder the experimental conditions.

The persistence of clofibric acid in a pilot sewage-sludge plant and biofilm reactors operated under oxic oranoxic conditions, was confirmed in a study by Hebereret al. [17].

Regarding the biodegradation of antimicrobials, Perezet al. [18] reported the behavior of trimethoprim in alaboratory biodegradation reactor filled with activatedsludge collected from different stages of a municipalWWTP: primary treatment; activated sludge; nitrifyingactivated sludge; and, final effluent after disinfection.Trimethoprim displayed high resistance to microbialdegradation in the sewage from the primary treatmentand the activated sludge treatment. However, primarydegradation of this compound was completed within 3days in the sewage from the nitrification process.

In a subsequent work, the same group described theidentification of two microbial degradation products oftrimethoprim produced by nitrifying activated sludgebacteria in a small-scale laboratory batch-reactor usingLC-IT-MS and LC-QqToF-MS as tools for structure eluci-dation [19]. Trimethoprim was quickly degraded, gen-erating two detectable metabolites (Table 1). While onemetabolite corresponded to a-hydroxy-trimethoprim,which had been described as one of the major humanmetabolites, for the second degradate, a two-fold oxida-tion of the aromatic ring within the diaminopyrimidinesubstructure was postulated. Regarding the fragmenta-tion pattern of the parent drug, trimethoprim m/z 291,obtained with LC-(+)ESI-IT-MS, all of the fragment ionswith m/z values between 230 and 276 were based on astructure comprising the benzene ring, the bridgingmethylene group, as well as the intact diaminopyrimidinethat did not undergo any transformation. Two furtherfragment ions, m/z 181 and m/z 123, were observed tooccur from cleavage of the protonated molecule eitherside of the central methylene group. Metabolite a-hy-droxy-trimethoprim had a molecular weight of 16 Dagreater than trimethoprim, confirming that the antibiotichad undergone an oxidative transformation. The IT-MSspectrum merely displayed two fragment ions: a basepeak at m/z 289, and a less intense ion at m/z 274. Theneutral loss of 18 Da corresponding to the expulsion ofwater was strongly favored over other fragmentationprocesses, such as cleavages from the methoxy groups, asobserved for trimethoprim. As for the fragment ion m/z274 differing from m/z 289 by 15 Da, it was assigned tothe loss of a methyl radical analogous to the generation ofm/z 276 from the protonated trimethoprim. Regardingthe second metabolite of trimethoprim, its molecularweight was 34 Da higher relative to trimethoprim, indi-cating oxidation or addition of a functional group.Detection of the base peak ion at m/z 181, similarlyobserved in the product ion profile of the protonated tri-methoprim, indicated that the trimethoxybenzeneportion of the molecule had not undergone any struc-tural modifications. Further fragment ions [m/z 308 (loss



of ammonia), m/z 307 (loss of water) and m/z 290(concomitant loss of ammonia and water)] indicated thathydroxylation and oxidation of trimethoprim occurred inthe pyrimidine ring. Accurate mass measurements ofboth metabolites were provided by QqToF-MS, achievingabsolute mass errors of <5 mDa for all fragment ions.This allowed confirmation of the postulated elementalcompositions. The study emphasized the potential ofnitrifying activated sludge bacteria for breaking down anenvironmentally relevant pharmaceutical that is other-wise poorly degradable by a bacterial communityencountered in conventional activated sludge.

The biodegradation of anti-tumorals cisplatin, cyclo-phosphamide, cytarabine, 5-fluorouracil, and metho-trexate in OECD tests at concentrations in the mg/L rangewas monitored by LC-DAD [20]. The results showed thatcisplatin and cyclophosphamide were not biodegradable,but cytarabine and 5-fluoracil were biodegradable to 70%and 100%, respectively. The primary biodegradation ofmethotrexate reached a level of about 95% after 8 days.On the second day, a metabolite started to appear and wasidentified by UV and IR spectra as the toxic, persistentdegradation product, 7-hydroxymethotrexate. This com-pound is also a human metabolite in chemotherapeutictreatment with high doses of methotrexate. In the OECDtest, 7-hydroxymethotrexate did not undergo furtherdegradation.

Ternes et al. studied the biodegradability of estrogens[21]. They investigated the degradability of naturalestrogens (17b-estradiol, estrone and estriol), contra-ceptives (17a-ethynylestradiol and mestranol) and twoestrogen glucuronides in aerobic batch-reactors loadedwith diluted activated sludge from a municipal WWTP.They spiked the test medium at two different concen-trations, 1 mg/L and 1 lg/L. The experiments weremonitored by GC-IT-MS and revealed that naturalestrogen 17b-estradiol was oxidized to estrone in thebatch experiments and was further eliminated inapproximate linear time dependence without leavingbehind detectable degradates. In the human body, 17b-estradiol is rapidly oxidized to estrone, which can befurther oxidized to estriol. In [21], 16a-hydroxy-estronewas rapidly eliminated without detection of degradationproducts. Contraceptive 17a-ethinylestradiol was per-sistent under the selected aerobic conditions, in line withits persistence in WWTPs. However, mestranol wasrapidly eliminated and small portions of 17a-ethinyl-estradiol were formed by demethylation.

After administration to humans, pro-drug mestranol isconverted into 17a-ethynylestradiol by O-demethyla-tion. Here, contraceptive 17a-ethynylestradiol is mainlyeliminated in the form of conjugates, whereas othermetabolic transformations occur but to a lesser degree.In contact with sludge, the two glucuronide derivates of17b-estradiol (17b-estradiol-17-glucuronide and 17b-estradiol-3-glucuronide) were cleaved so 17b-estradiol


was released. In conclusion, the predominant presence ofestrone in WWTP effluents and rivers is presumably aresult of the relatively high stability of estrone inWWTPs, the cleavage of glucuronide conjugates fromboth estrone and 17b-estradiol and the oxidation of 17b-estradiol to estrone.

Identification of the metabolites was accomplished bycomparison with reference standards. Using nitrifyingactivated sludge [22], of the four estrogens, 17b-estra-diol was most easily degraded via the estrone pathway,supporting the findings reported by Ternes et al. [21].

Concerning the degradability of X-ray contrast agents,Perez et al. [23] reported the biotransformation ofiopromide in conventional activated sludge and nitrifyingactivated sludge collected from a municipal WWTP.Identification of the four detected metabolites was per-formed with LC-IT-MS. In a batch-reactor containingmixed liquor from conventional activated sludge, threemetabolites produced upon oxidation of the primaryalcohols to carboxylic acids in the side chains of iopro-mide (m/z 791) were identified. Two metabolites (m/z806) carried one carboxy group in either side chain,while the third had one carboxy groups in each sidechain (m/z 820). Furthermore, in a batch-reactor withmixed liquor from the nitrification tank of a WWTP, onemetabolite (m/z 760) was identified as a dehydroxylationproduct. As to the fragmentation pattern of iopromide,selection of the protonated molecule (m/z 792) showedfive main fragment ions [m/z 774 (loss of water), m/z 701and m/z 687 (cleavage of the amide bond in different sidechains), and m/z 573 and m/z 559 formed from the lattertwo by loss of HI]. Key fragment ions for identifying thetwo metabolites at m/z 806 (14 Da higher than iopro-mide) and at m/z 820 (28 Da higher than iopromide)were loss of water (m/z 788 and m/z 802) and neutralloss of one HCOOH (46 Da) for m/z 806 and two HCOOHmolecules for the metabolite with m/z 820. Regarding themass spectrum of the metabolite at m/z 760 (32 Dalighter than iopromide), it showed the formation of m/z685 and m/z 671, attributed to the breakage of the amidebonds in the side chains, and m/z 557 and m/z 543,corresponding to further losses of HI. Compared to theMS2 spectrum of iopromide, the mass shift by �16 Da inthe fragment ions m/z 685 and m/z 671 indicated thatboth side chains had undergone structural changesduring biotransformation. The absence of a neutral lossof water from [M + H]+in the MS2 mass spectrum of m/z760 (in contrast to the MS2 spectra of iopromide and themetabolites m/z 806 and m/z 820) was a strong indicatorthat dehydroxylation of the secondary hydroxy groupoccurred in the side chain. H/D-exchange experimentsand derivatization of the carboxy groups to methyl estersprovided definitive confirmation for the identities of themetabolites. Overall, the iodinated ring of iopromide re-mained intact during the biodegradation processes, incontrast to the outcomes reported in [24].

Table 2. Analytical methodologies and occurrence of selected drugs and their human metabolites in aqueous environmental samples

Parent drug

Metabolites

Matrix Analytical procedure MRM transition in MS2 Accuracy and precision Occurrence Ref.

SPE3 sorbent (sample pH)

elution solvent.

Chromatography

Ibuprofen SW1 and

WW2C18 (pH 2) CH3OH. – R4: Ibu-OH5 :63% WW7

e : Ibu-OH: 6.7 lg/L [30]

Hydroxyibuprofen GC-MS after derivatization Ibu-CX6: 74%

Carboxyibuprofen

Ibuprofen

Hydroxyibuprofen

Carboxyibuprofen

Clofibric acid

WW and

SeW8

Oasis HLB; hexane, EtOAc and

CH3OH.

GC-MS after derivatization of

clofibric acid to its methyl ester

– R: 70–100%

IDL9: 8–15 pg

MQL10:

0.24–0.69 ng/L

WWi11: Ibu-OH + Ibu-CX (2.95–29.84 ng/L)>

ibuprofen (0.6–1.66 ng/L); clofibric acid

(0–0.17 ng/L)

WWe: Ibu-OH (0.59–1.13 ng/L)

> Ibu-CX (0–1.27 ng/L); clofibric acid (0–0.11 ng/L)

SeW: Ibu-CX (0.7–5.3 ng/L)> Ibu-OH (0.42–1.5 ng/L)

[29]

Phenylbutazone GW12, SW

and WW

Isolute C18 (pH 7–7.5) Phenylbutazone: [M + H]+ fi 160 RGW: 66–82% 4-aminoantipyrine: WWi: 0.78 lg/L [32]

Oxyphenbutazone CH3OH. Oxyphenazone: [M + H]+ fi 160 RSW: 15–52% WWe: 0.36 lg/L SW: 0.13–0.63 lg/L

Metamizole LC-(+)ESI-MS2 4-aminoantipyrine: RWWi: 30–49%

4-Aminoantipyrine [M + H]+ fi 160 RWWe: 15–19%

Phenylbutazone MQL

WW: 25–250 ng/L

3 Phenazone-type drugs

6 metabolites

GW, SW and

WW

C18; CH3OH.

LC-(+)APCI-MS2

AMDOPH13: [M + H]+ fi 72, 222

AMPH14: [M + H]+ fi 92, 65

DP15: [M + H]+ fi 113, 69

PDP16: [M + H]+ fi 155, 113

AAA17: [M + H]+ fi 228, 83

FAA18: [M + H]+ fi 214, 83

DMAA19: [M + H]+ fi 113, 111

Phenazone: [M + H]+ fi 56, 147

Propyphenazone: [M + H]+ fi 56,

189

RGW,SW and WW:87–117%

except for DMAA

(Rsw:17% RGW and WW:

72–133% )

MQL: 10–20 ng/L

µg/L WW WW Ee Ei SW

AMDOPH 0.71 0.73 0.39 0.35 0.30

AMPH 0.15 0.19 0.15 0.13 0.16

DP nd20 nd 0.07 0.11 0.04

PDP 0.04 0.03 nd nd nd AAA 8.8 7 0.83 1.01 1.0

FAA 1.9 2.0 0.67 0.51 1.0

DMAA nd nd nd nd nd

Phenazone 0.45 0.41 0.50 0.48 0.37

Propyphenazone 0.18 0.17 0.14 0.08 0.11

[28]

Clofibric acid

Fenofibric acid

Acetylsalicylic acid

Salicylic acid

o-Hydroxy-hippuric

acid

Gentisic acid

SW, DW23

and WW

C18 and LiChrolut EN (pH 2)

CH3OH.

GC-MS

R: 40–90%

MQL: 5 ng/L–0.2 lg/L

WW: 50–250 (GC-MS)

SW: 5–20 (GC-MS)

DW: 1–10 (GC-IT-MS)

µg/L WW WW SW

Acetylsalicylic 3.2 0.50 ↓ ng/L

Salicylic acid 54 nd 0.14

o-Hydroxy- 4.6 nd nd

Gentisic acid 6.8 nd nd hippuric acid

acid

[15]


Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

Tren

ds

http

://ww

w.elsevier.co

m/lo

cate/trac505

Table 2 (continued)

Parent drug

Metabolites

Matrix Analytical procedure MRM transition in MS2 Accuracy and precision Occurrence Ref.

SPE3 sorbent (sample pH)

elution solvent.

Chromatography

Clofibrate

Clofibric acid

Etofibrate

Fenofibric acid

Acetylsalicylic acid

Salicylic acid

o-Hydroxy-hippuric

acid

Gentisic acid

SW, DW and

WW

GC-MS, GC-IT-MS and LC-ESI-

MS2

– µg/L WWi WWe SW

Clofibric acid nd 1.6

Fenofibric acid nd 1.2

Salicylic acid 54 0.14 4.1

Gentisic acid 0.59 1.2

[33]

Clofibric acid DW, SW and

WW

SDB-XC (pH 2) CH3OH, CH2Cl2/

CH3OH. GC-MS

R: 44%

IDL: 3 ng/L

MDL: 0.8 ng/L

SW: Clofibric acid: 103 ng/L [34]

Carbamazepine

5 metabolites

SW, WW Oasis HLB (pH 7) CH3OH

[Optimization of stationary

phases]. LC-(+)ESI-MS2

CBZ24 : [M + H]+ fi 194

CBZ-EP25: [M + H]+ fi 180

CBZ-DiOH26: [M + H]+ fi 253

CBZ-2OH27: [M + H]+ fi 210

CBZ-3OH28: [M + H]+ fi 210

CBZ-10OH29: [M + H]+ fi 237

RSW: 96–103%

RWWi: 84–102%

RWWe: 91–104%

IDL: 0.8–4.8 pg

SW: CBZ-DiOH (2.2 ng/L),

CBZ (0.7 ng/L)

WWi: 5 metabolites and

CBZ (8.5–1571 ng/L)

WWe: 5 metabolites and

CBZ (9.3–1325 ng/L 2.2 ng/L)

[35]

Diazepam

Desmethyldiazepam

Clofibric acid

WWe Oasis MCX (pH 1.5/2.0) CH3OH,

CH3OH (+2% NH3), CH3OH

(+0.2% NaOH). LC-(+/-)ESI-MS2

Diazepam: [M + H]+ fi 193, 154

Desmethyldiazepam:

[M + H]+ fi 165, 140

Clofibric acid: [M-H]�fi 127, 85

R: 81–92%

IQL31: 16–70 pg

MQL: 0.36–1.08 ng/L

Diazepam: n.d.

Desmethyldiazepam: 1–62 ng/L

Clofibric acid: 0.5–82 ng/L

[36]

Atorvastatin

p-Hydroxyatorvastatin

o-Hydroxyatorvastatin

Simvastatin

Simvastatin hydroxyl

acid

Fluoxetine

Norfluoxetine

DW, SW and

WW

HLB; CH3OH/MTBE

(10/90).

LC-(+/-)ESI-MS2

Atorvastatin: [M + H]+ fi 440

p-Hydroxyatorvastatin:

[M + H]+ fi 440

o-Hydroxyatorvastatin:

[M + H]+ fi 440

Simvastatin:

[M-H2O-H]� fi 115

Simvastatin hydroxy acid:

[M-H]� fi 319

Fluoxetine: [M + H]+ fi 44

Norfluoxetine: [M + H]+ fi 134

RDW: 100–106%

RSW: 99–103%

RWWi: 96–101%

RWWe: 96–102%

MDL: 0.25–0.5 ng/L

ng/L WWi WWe SW DW

Atorvastatin 201 <0.50 7.3 <0.25

p-Hydroxy-

atorvastatin

280 <1.0 9.2 <0.50

o-Hydroxy-

atorvastatin

196 <1.0 6.9 <0.50

Simvastatin <0.25 <0.50 <0.25 <1.0

Simvastatin

hydroxy acid

10 <0.50 0.74 <0.25

Fluoxetine 17 25 2.6 <0.50

Norfluoxetine 9.9 3.9 1.3 <0.50

[37]

Line missing

Tren

ds

Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

506

http

://ww

w.elsevier.co

m/lo

cate/trac

Sulfamethoxazole

N4-Acetylsulfa-

methoxazole

Clofibric acid

SW and WW Strata X (pH 3) CH3OH

[Stationary phase optimization].

LC-(+/-)ESI-IT-MS

Sulfamethoxazole: [M + H]+ fi 188

N4-Acetylsulfamethoxazole:

[M + H]+ fi 236

Clofibric acid: [M-H]� fi 127

Column R: 56–120%

MQL: 50 ng/L

N4-acetylsulfamethoxazole WWe: 690–2200 ng/L

Downstream SW: 240 ng/L

[38]

Sulfamethoxazole

N4-Acetyl-

sulfamethoxazole

WW Oasis HLB (pH 4) CH3OH/EtOAc

(1:1), CH3OH (+1 % NH3). LC-

(+)ESI-MS2

N4-Acetylsulfamethoxazole:

[M + H]+ fi 134, 198

Sulfamethoxazole:

[M + H]+ fi 156, 108

R: 91–105%

MQL: 11–212 ng/Lng/L 1º effl. 2º effl. 3º effl.

WWTP1

SMX32 343 344 352

N4-SMX23 518 86 82

WWTP2SMX 641 352 352

N4-SMX 943 nd 71

[39]

Estrogenic compounds

3 metabolites

SW and WW Tandem LiChrolute EN

and RP-C18(pH 3) acetone.

GC-IT-MS

17b-Estradiol: M+ fi 326, 285

Estrone: M+ fi 257, 244

17a-Ethynylestradiol: M+ fi 231, 193

Mestranol: M+ fi 349, 193

16a-Hydroxyestrone: M+ fi 244, 230

17b-Estradiol-17-valerate: M+ fi 326,

297

RSW: 41–90%

RWW: 56–82%

MQL: SW:

0.5–1 ng/L

WW: 1–2 ng/L

µg/L WWi WWe SW

17-Estradiol 0.015 0.003 nd

Estrone 0.027 0.070 1.6

17-Ethinylestradiol Nr34 0.015 nd

Mestranol nr 0.004 nd

16-Hydroxyestrone nr 0.005 nd

[21]

Enalaprilat SW HLB; CH3OH, CH3OH (+TCA).

LC-(+)ESI-MS

– > 80% SW: 0.046 lg/L [41]

Cocaine

Benzoylecgonine

SW and WW Oasis HLB (pH 2) CH3OH,

CH3OH (+2 % NH3).

LC-(+)ESI-MS2 and LC-ESI-IT-MS

Cocaine: [M + H]+ fi 105, 182

Benzoylecgonine: [M + H]+ fi 105,

168

>90

MDL:

Cocaine: 0.12 ng/L

Benzoylecgonine: 0.06

ng/L

ng/L WWi SW

Cocaine 42–120 1.2

Benzoylecgonine 390–750 25

[42]

1SW, Surface water.2WW, Wastewater.3SPE, Solid-phase extraction.4R, Recovery.5Ibu-OH, Hydroxy-ibuprofen.6Ibu-CX, Carboxy-ibuprofen.7WWe, Wastewater effluent.8SeW, Sea water.9IDL, Instrumental detection limit.10MQL, Method quantification limit.11WWi, Wastewater influent12GW, Groundwater.13AMDOPH, 1-acetyl-1-methyl-2-dimethyloxamoyl-2-phenylhydrazide.14AMPH, 1-acetyl-1-methyl-2-phenylhydrazide.


Tren

ds

inA

nalytical

Chem

istry,V

ol.

26,

No.

6,

2007

Tren

ds

http

://ww

w.elsevier.co

m/lo

cate/trac507

15D

P,

1,5

-dim

ethyl

-1,2

-deh

ydro

-3-p

yraz

olo

ne.

16PD

P,

4-(

2-m

ethyl

ethyl

)-1,5

-dim

ethyl

-1,2

-deh

ydro

-3-p

yraz

ole

).17A

AA

,A

ceta

min

oan

tipyr

ine.

18FA

A,

Form

ylam

inoan

tipyr

ine.

19D

MA

A,

Dim

ethyl

amin

ophen

azone.

20n.d

.,not

det

ecte

d.

21E i

;en

han

ced

trea

ted

influen

t.22E e

,En

han

ced

trea

ted

effluen

t.23D

W,

Dri

nki

ng

wat

er.

24C

BZ

,C

arbam

azep

ine.

25C

BZ

-EP,

10,1

1-d

ihyd

ro-1

0,1

1-e

poxy

carb

amaz

epin

e.26C

BZ

-DiO

H,

10,1

1-d

ihyd

ro-1

0,1

1-d

ihyd

roxy

carb

amaz

epin

e.27C

BZ

-2O

H,

2-h

ydro

xyca

rbam

azep

ine.

28C

BZ

-3O

H,

3-h

ydro

xyca

rbam

azep

ine.

29C

BZ

-10O

H,

10,1

1-d

ihyd

ro-1

0-h

ydro

xyca

rbam

azep

ine.

30M

DL,

Met

hod

det

ecti

on

lim

it.

31IQ

L,In

stru

men

tal

quan

tifica

tion

lim

it.

32SM

X,

Sulf

amet

hoxa

zole

.33N

4-S

MX

,N

4-a

cety

lsu

lfam

ethoxa

zole

.34n.r

.,not

report

ed.

Tab

le2

(continued

)



In another study [25], the degradation of iopromide insewage sludge, including the identification of a metab-olite with cleaved amide bonds (5-amino-N,N 0-bis(2,3-dihydroxypropyl)-2,4,6-triiodo-N-methyliso-phthalamide)was described using LC-DAD. Iopromide was eliminatedto more than 80% after a lag of 31 days with concom-itant formation of the metabolite. Details of the metab-olite identification were not reported.

For the study of the degradability of 5-amino-N,N 0-bis(2,3-dihydroxypropyl)-2,4,6-triiodo-N-methyliso-phthalamide, the compound was spiked at a concen-tration of 97 mg/L into a surface-water-model system[25]. It took 23 weeks for the compound to disappear,probably because the resulting algal community inhib-ited the growth of bacteria necessary for degradation bynutrient competition. The adsorbable organic iodine(AOI) decreased only marginally, indicating that theresulting intermediates were still iodinated organics,though far more hydrophilic than the parent compound.

The biodegradability of the ionic X-ray contrast agentdiatrizoate was described in sewage sludge [26]. Nodegradation occurred in this medium following the OECDguideline 303A, whereas, in the modified Zahn-Wellenstest (OECD 302B), the substrate was biotransformed into2,4,6-triiodo-3,5-diamino-benzoic acid, corresponding tode-acylation of the parent compound. This compoundwas characterized with HPLC-UV, 1H nuclear magneticresonance (NMR) spectroscopy, IR spectroscopy and fast-atom bombardment MS.

Further studies [24,27] showed that the biodegrad-ability of diatrizoate in activated sludge batch-reactorswas poor. It appeared that this compound could bedegraded only under very special conditions. By con-trast, in water-sediment batch-reactors, partial degra-dation of diatrizoate and formation of two metaboliteswere observed [24] and isolated from the test mediumby HPLC. It could be demonstrated that no deiodinationhad taken place, but the presence of free amino groupscould be proved by 1H-NMR after separation onthin-layer plates. The metabolites were 3-acetylamino-5-amino-2,4,6-triiodobenzoate and 3,5-diamino-2,4,6-triiodobenzoate.

3. Analysis and occurrence in the aquaticenvironment

Drug metabolites deserve special attention as environ-mental pollutants, because they may occur at higherconcentrations than the unchanged pharmaceuticalsand their determination is essential to understand theenvironmental fate of the parent compound. The devel-opment of robust, reliable analytical methodologies fortheir determination faces several challenges. As bio-transformation products are usually more polar than theparent compound, in some cases, they cannot be


determined using conventional SPE sorbents for extrac-tion. In order to improve the retention on such materi-als, the addition of derivatizing agents to the aqueoussample was proposed [28]. A further challenge is theinstability of some Phase II metabolites, in particular,glucuronides can be cleaved by microorganisms to yieldeither the intact drug or a Phase I metabolite. The lack ofcommercially-available standards adds to the problem ofquantitative metabolite analysis. Regarding the behaviorand the fate of drug metabolites in the environment, weneed to consider that, in some instances, the microbialdegradates are identical in structure with humanmetabolites, making it difficult do determine the origin ofa specific biodegradation product (Table 1). Besides therequirements in the extraction protocol, the methods formetabolite detection have to guarantee sensitive,unequivocal identification of the target analytes. This isroutinely accomplished by employing modern MS tech-niques (Table 2).

Few works have reported simultaneous determinationof human metabolites and their parent compounds inthe environment. Next, we review the study of thesecompounds in the aquatic environment. Table 2 lists theanalytical methods and the occurrence of the humanmetabolites and their parent compounds.

First, we present the determination of anti-inflamma-tory, analgesics and blood-lipid regulators and theirmetabolites in the aquatic environment. The develop-ment of a methodology for the determination of clofibricacid, a human metabolite of clofibrate and etofibrate, andof ibuprofen along with two metabolites (hydroxy-ibuprofen and carboxy-ibuprofen) was described forwastewater and seawater [29]. The extraction was car-ried out on Oasis HLB followed by GC-MS analysis (Table2). Limits of quantification (LOQs) for the entire methodwere in the range of 0.24–0.69 ng/L with recoveries of70–100%. Sewage samples were collected from twoWWTPs, one in Norway and one in Germany. Clofibricacid was not detected in the Norwegian samples becauseprecursor drugs clofibrate and etofibrate are not pre-scribed in Norway but the free acid was detected in thesample from the German WWTP. Hydroxy-ibuprofen andcarboxy-ibuprofen were detected in all samples investi-gated from both WWTPs and in most seawater samples.The two matrices showed characteristic patterns, withhydroxyl-ibuprofen being the major component insewage, whereas carboxy-ibuprofen was dominant inseawater samples. Due to distinct sewage-treatmenttechnologies, different patterns of ibuprofen-metabolitedistribution were found in the two WWTPs studied;in effluents from the German WWTP, carboxy-ibuprofenwas not detected, while it was frequently detected ineffluents from the Norwegian WWTP. Hydroxy-ibuprofenwas detected in all WWTP samples and was the dominantcompound in the samples from the German WWTP. Ininfluent samples, the ratio of ibuprofen to its metabolites

was identical to that in human excretions, indicating thatno significant transformation processes took place in thesewer system between excretion and WWTP.

Stumpf et al. [30] reported similar results to thosefound by Weigel et al. [29] for the occurrence of ibu-profen and its metabolites in German WWTPs. Traces ofthe hydroxy metabolite were determined in effluentsfrom all municipal sewage-treatment plants monitored,whereas carboxy-ibuprofen was nearly completelyeliminated during passage through the sewage-treat-ment plants. The method encompassed SPE, derivatiza-tion and subsequent GC-MS obtaining quantitativerecoveries for the analytes (Table 2). Stumpf et al. [30]also investigated 12 German rivers, detecting hydroxy-ibuprofen in all samples at low-lg/L levels, those beingconsiderably higher than the concentrations of the par-ent drug detected in the same rivers.

Monitoring studies in a WWTP revealed that carboxy-ibuprofen was almost quantitatively eliminated duringpassage through the facility, whereas hydroxy-ibuprofenwas hardly affected. In some instances, the concentra-tions in the effluents even exceeded those measured inthe influents. The elevated concentrations of hydroxy-ibuprofen in the effluent were explained as being formedfrom ibuprofen and/or cleavage of hydroxy-ibuprofenglucuronides during the activated sludge treatment.Zwiener et al. [31] identified hydroxy-ibuprofen, car-boxy-ibuprofen and carboxy-hydrotropic acid in bio-degradation experiments performed with activatedsludge from a municipal WWTP, demonstrating that themicrobial metabolites combined did not account formore than 10% of the initial concentration of ibuprofen.These results suggested that the major amounts ofmetabolites in sewage stemmed from human excretion.

Ternes et al. [32] reported the determination ofanti-inflammatory phenylbutazone, its metabolite oxy-phenbutazone and the metamizole metabolite 4-amino-antipyrine using SPE and LC-ESI-MS2 (Table 2). Lowrecoveries of the compounds in different matrices wereattributed to signal suppression in ESI. The use of sur-rogate standard 10,11-dihydrocarbamazepine did notallow compensation for analyte losses. Of the threeanalytes, only 4-aminoantipyridine was detected inwastewater influent and effluent samples at concentra-tions of 0.78 lg/L and 0.36 lg/L, respectively, corre-sponding to 54% elimination. The trace amounts of4-aminoantipyridine were probably due to excretionsfrom treated humans.

In a different work [28], the determination of anal-gesic drugs (4-dimethylamino-phenazone, phenazoneand propylphenazone) along with some of their humanmetabolites [(1-acetyl-1-methyl-2-dimethyloxamoyl-2-phenylhydrazide, 1-acetyl-1-methyl-2-phenylhydrazide,acetaminoantipyrine, 1,5-dimethyl-1,2-dehydro-3-pyr-azolone [DP] and 4-(2-methylethyl)-1,5-dimethyl-1,2-dehydro-3-pyrazole [PDP], formylaminoantipyrine)] was



described using SPE and LC-(+)APCI-MS2. To allow forefficient extraction of the metabolites DP and PDP fromthe water using a conventional C18 sorbent, the authorsprepared the water samples by a simple in situ deriva-tization with acetic anhydride in basic media in order todecrease the polarity and to increase the molecularweight of these analytes. DP and PDP were convertedinto the N-acetyl derivatives and could be extractedalong with the other unchanged target analytes. Quan-titative measurements were performed by LC-APCI-MS2

with internal calibration using dihydrocarbamazepine assurrogate. Although ESI led to higher peak intensitiesthan APCI, APCI was preferred because it providedmatrix-independent ionization, resulting in recoveries ofabout 100% (Table 2).

Unacceptable recovery for dimethylaminophenazone insurface-water samples (17%) might have resulted fromfast degradation processes occurring during samplepreparation. The fate of the compounds in municipalWWTP, in an enhanced effluent-treatment facility and insurface waters was investigated [28]. In the WWTP,seven target compounds were detected (Table 2).Concentrations above 1 lg/L were determined for acet-aminoantipyrine and formylaminoantipyrine. Dimethyl-aminophenazone was not detected above its LOQ.

In a further case study, effluents from a WWTP weredischarged into a canal and reached the enhanced-treat-ment facility equipped with a phosphate removal stage. Inthe effluent from this facility, high levels of 1-acetyl-1-methyl-2-dimethyloxamoyl-2-phenylhydrazide, 1-acetyl-1-methyl-2-phenylhydrazide and formylaminoantipyrinewere reported (Table 2). However, the concentrations ofphenazone, propyphenazone and acetaminoantipyrinewere up to 40% lower after treatment. In the effluents,phenazone metabolite 1,5-dimethyl-1,2-dehydro-3-pyr-azolone was measured at concentrations slightly higherthan in the influents of the enhanced facility. In thereceiving surface waters, ng/L levels of the compoundswere still detectable.

Ternes et al. [15] described a methodology for deter-mining analgesic acetylsalicylic acid and its metabolites(salicylic acid, o-hydroxyhippuric acid and gentisic acid)as well as clofibric acid and fenofibric acid in sewage,river and drinking water. The method comprised SPEusing C18 and LiChrolut EN followed by methylation ofthe carboxy groups with diazomethane, acetylation ofphenolic hydroxy groups with acetanhydride/triethyl-amine and determination by GC-MS. In wastewatereffluents and in river water, the method LOQs were aslow as 5 ng/L for the parent drugs and down to 0.2 lg/Lfor o-hydroxyhippuric acid and gentisic acid. Salicylicacid, the major metabolite of acetylsalicylic acid, wasdetected in raw sewage of the WWTP with averageconcentration of 54 lg/L over 6 days. Gentisic acid ando-hydroxyhippuric acid were present in the sewage ataverage concentrations of 4.6 lg/L and 6.8 lg/L,


respectively. In the discharge of the WWTP, no residuesof metabolites of acetylsalicylic acid were detectable, butthe parent compound was present at a concentration of0.5 lg/L. Removal rates of the three investigatedmetabolites therefore exceeded 98%. These results werein agreement with the findings of a degradation study inbatch reactors, demonstrating that acetylsalicylic acidwas almost completely eliminated within 24 h [15].Acetylsalicylic acid was detected in several river-watersamples. Only a few WWTP effluents showed concen-trations of acetylsalicylic-acid metabolites above theLOQ. However, salicylic acid was frequently detected insurface waters in the low-ng/L range, whereas gentisicacid and o-hydroxyhippuric acid were rarely found(Table 2) [15].

With the methodologies developed, Ternes et al.[15,32] examined the occurrence of several metabolites(clofibric acid, fenofibric acid, salicylic acid, o-hydrox-yhippuric acid and gentisic acid) and their parentcompounds (clofibrate, fenofibrate and acetylsalicylicacid) in the environment [33]. Non-polar lipid regulatorsclofibrate, etofibrate and fenofibrate were generally notdetected in WWTP effluents but their polar metabolites,clofibric acid and fenofibric acid, were present at con-centration levels of 1.6 lg/L and 1.2 lg/L, respectively.In rivers and streams, their levels were in the ng/Lrange. In the human body, complete hydrolysis of theoriginal drugs to clofibric and fenofibric acids occursrapidly after administration. The principal excretionproducts are glucuronides of the acidic metabolites(60%) and very small proportion of non-conjugatedacids (<10%). As the concentrations of clofibric andfenofibric acids in environmental samples were high, itappeared that the glucuronides were at least partiallycleaved during the passage through WWTP, leading toan increase in concentration.

For the metabolites of analgesic drug acetylsalicylic,no detectable amounts or sporadic concentrations werefound in effluents. Although the metabolite, salicylicacid, was detected in surface waters at maximum levelsof 4.1 lg/L, this high concentration might have been aresult of the use of this product as an antioxidant agent.

Boyd et al. [34] reported the analysis of clofibric acidfrom wastewater, surface water and untreated drinkingwater samples from the U.S.A. and Canada by SPE usinga polar SDB-XC Empore disk. Derivatization with N,O-bis(trimethylsilyl)-trifluoroacetamide in the presence oftrimethylchlorosilane was used to enhance thermalstability of clofibric acid, which thermally degraded inthe GC injection port, and to reduce the polarity of thesemetabolites in order to facilitate their GC-MS analysis.The method limit of detection (LOD) was 0.8 ng/L andthe recovery 44% (Table 2). Clofibric acid was detectedin only one surface water sample at a concentration of103 ng/L. The absence of clofibric acid in other surfacewaters could be attributed to the declining use of


clofibrate in the U.S.A. The authors also studied the re-moval of this compound in a drinking-water facility.Clofibric acid was totally removed by non-conventionalwater-treatment processes that included chlorination,ozonation and dual media filtration [34].

Concerning the determination of antiepileptics. Miaoand Metcalfe [35] developed a quantitative method for thesimultaneous determination of carbamazepine and five ofits human metabolites [57] (i.e. 10,11-dihydro-10,11-epoxycarbamazepine, 10,11-dihydro-10,11-dihydroxy-carbamazepine, 2-hydroxycarbamazepine, 3-hydroxy-carbamazepine and 10,11-dihydro-10-hydroxycarbam-azepine) in sewage and surface waters (Table 2). Themethod relied on an SPE procedure followed by separationand detection with LC-(+)ESI-MS2. Recoveries on OasisHLB exceeded 80% for all compounds in both spiked watermatrices. Ion suppression by interfering matrix compo-nents was assessed in WWTP influent, effluent and surfacewater. Ion suppression was most severe in untreatedsewage and the least critical in surface water. In WWTPinfluent samples, only 13–42 % of the signals detected in amatrix-free standard were observed for the analytes andinternal standard 10,11-dihydrocarbamazepine. Car-bamazepine and the five studied metabolites were detectedin WWTP influent and effluent at ng/L levels. Theconcentrations of 10,11-dihydro-10,11-dihydroxycarb-amazepine in influent and effluent were 1572 ng/L and1325 ng/L, respectively, being about three times that ofcarbamazepine. The metabolite concentrations weresimilar in influent and effluent samples analyzed, indi-cating little removal during sewage treatment. The parentcompound showed a similar behavior.

Concerning the determination of psychiatric drugs,the development of an analytical method for the deter-mination of diazepam and desmethyldiazepam and alsothe human metabolite of blood-lipid-regulator clofibrate,clofibric acid, in wastewater, including SPE and LC-ESI-MS2, was reported [36]. The optimized SPE method re-lied on Oasis MCX with a sample pH adjusted to 1.5,yielding recoveries for the two metabolites greater than70% and achieving instrumental LODs and methodLODs of the order of low ng/L. Although the clofibricacid and desmethyldiazepam were frequently detected,diazepam was not detected in any of the wastewatereffluents.

For the determination of anti-depressant fluoxetineand its human metabolite, norfluoxetine, and of the twoblood-lipid regulators, atorvastatin and simvastatin,along with their metabolites (p-hydroxyatorvastatin, o-hydroxyatorvastatin and simvastatin hydroxy acid),extraction on Oasis HLB and LC-(+/-)ESI-MS2 were em-ployed [37]. The authors used an isotope-dilution tech-nique for every analyte to compensate for matrix effectsin the ESI source, SPE losses and instrument variability.Matrix-spiked recoveries for all compounds were 96–102%. The method LODs were 0.25–0.5 ng/L. Except for

fluoxetine and norfluoxetine, analysis of wastewater-effluent samples indicated substantial removal of thetarget compounds, possibly brought about by tertiarytreatment and chlorination employed at the WWTPsampled (Table 2). Downstream from the WWTP outlet,all compounds, except simvastatin, exhibited an increasein concentration in the receiving surface water due toeffluent discharges from a second WWTP that did notemploy chlorination.

Regarding the analysis of antimicrobials and theirmetabolites in the aquatic environment, few works havebeen reported. A method using SPE and IT-MS in con-secutive reaction monitoring (CRM) mode for sulfa-methoxazole, its metabolite N4-acetylsulfamethoxazole,and also clofibric acid (a human metabolite of a blood-lipid regulator clofibrate) was developed for wastewaterand surface waters [38].

A number of stationary phases (Isolute ENV+, OasisHLB, Oasis MCX, Isolute C8, Isolute C18, Varian BondElut C18 and Phenomenex Strata X) were compared forextractability of target compounds. The last two sorbentsproved the most effective and, eventually, Strata X wasselected as the best phase for extracting the majority ofthe selected compounds. Recoveries for sulfamethoxazole(120%), N4-acetylsulfamethoxazole (56%) and clofibricacid (83%) were determined in spiked tap water. Theoccurrence of the three compounds was investigated inWWTP samples and samples collected upstream anddownstream of the WWTP outlet. Whereas clofibric acidand sulfamethoxazole were not detected in any sample,N4-acetylsulfamethoxazole was found in the sewage andthe surface-water sample taken downstream from theWWTP (Table 2).

A second method for sulfamethoxazole and itsmetabolite, N4-acetylsulfamethoxazole, was described fordifferent wastewater matrices (primary, secondary andtertiary effluent) combining SPE on Oasis HLB and LC-(+)ESI-MS2 with isotope dilution [39]. The highestrecoveries (>90%) for the two compounds were achievedat pH 4 (Table 2). N4-acetylsulfamethoxazole was typi-cally present in large amounts in the primary effluents,but only small amounts could be found in the tertiaryeffluents. If the amount of sulfamethoxazole present asN-acetyl metabolite was neglected, the elimination ofsulfamethoxazole would have been underestimated. Bycontrast, Hilton et al. [38] reported a high level of themetabolite in treated sewage (2200 ng/L), in whichsulfamethoxazole was not detectable.

As to the determination of estrogens, a sensitivemethod for sewage and river-water samples was devel-oped for the quantitative analysis of natural estrogen17b-estradiol, its synthetic precursor (17b-estradiol-17-valerate), contraceptives 17a-ethynylestradiol and mes-tranol (prodrug of the former) and the two 17b-estradiolmetabolites (estrone and 16a-hydroxyestrone) [40].Recoveries of the analytes in spiked groundwater after


Figure 3. Correlation of extent of drug metabolism in the humanbody vs. removal efficiency in wastewater-treatment plant (WWTP)employing activated sludge treatment. Removal efficiencies corre-spond to average values found in the literature, actual values for agiven WWTP depend on plant design and operational parameters.Refer to Table 1 for compound abbreviations. Compounds in boldface are partially or completely excreted from the human body asconjugates and can undergo cleavage in WWTP to release the par-ent drug. Compound names in italics denote removal at the nitrifi-cation stage of WWTP. Data taken from various sources (see Tables1 and 2).


SPE (combination of RP-C18 and LiChrolute EN)followed by clean up, derivatization and GC-IT-MS,generally exceeded 75% except for 16a-hydroxyestrone(41%). Method LOQs in different matrices were 0.5–1ng/L (Table 2).

The confirmation provided with GC-IT-MS wasessential, because 17a-ethynylestradiol and an un-known compound exhibited exactly the same retentiontime. Natural estrogens 17b-estradiol and estrone, theestrone metabolite (16a-hydroxyestrone) and contra-ceptive 17a-ethynylestradiol were frequently detected inWWTP discharges due to their incomplete removalduring passage through the WWTP. The concentrationof 17a-ethynylestradiol was not appreciably reduced bythe wastewater treatment, while 17b-estradiol and 16a-hydroxyestrone were removed with reductions in con-centrations of 68% and 64%, respectively. The detectionof 17b-estradiol and 16a-hydroxyestrone in WWTPdischarges was unexpected in view of their behavior inbatch reactors amended with activated sludge. In thesesettings, 17b-estradiol was oxidized to estrone, whichwas further degraded to 16a-hydroxyestrone, which wasrapidly eliminated [21]. At the same time, the biodeg-radation of 17a-ethynylestradiol was very low, in linewith its persistence during activated sludge treatment inWWTP.


Regarding the occurrence of the estrogenic com-pounds in the environment, estrone was the only com-pound detected in surface waters [40]. The predominantpresence of estrone in WWTP effluents and rivers ispresumably a result of the relatively high stability ofestrone in the WWTP, cleavage of glucuronides of es-trone and 17b-estradiol, and oxidation of 17b-estradiolto estrone [21].

Enalapril and other angiotensin-converting enzymeinhibitors are used in the treatment of hypertension andsome types of chronic heart failure. In the aquaticenvironment, only one study reported by Kolpin et al.[41] determined the metabolite of enalapril (enalaprilat)in surface waters. Oasis HLB was used to enrich thesurface waters, and detection and separation were per-formed by LC-(+)ESI-MS. The recoveries exceeded 80%and the maximum concentration detected was 0.046lg/L.

Recently, Zuccato et al. [42] developed a method toanalyze drugs of abuse in the aquatic environment. Co-caine and its main human metabolite (benzoylecgonine)were determined in surface and wastewaters using anSPE method (Oasis HLB) and LC-(+)ESI-MS2 and LC-ESI-IT-MS. Recoveries were >90% for both compounds andmethod LODs were 0.06 ng/L and 0.12 ng/L for benz-oylecgonine and cocaine, respectively. They were foundin all four WWTP sampled and in surface water (RiverPo, Italy). As expected, the parent-drug levels weremuch lower than those of the metabolite with a ratio of0.15 ± 0.03 in wastewater samples, being in accordwith the metabolic pathway of cocaine in humans. Insurface waters, the cocaine/benzoylecgonine ratio wasstable over time (0.05 ± 0.02), but lower than expected,suggesting a different pattern of degradation and/orpartition for cocaine and benzoylecgonine in WWTP ascompared to river systems. This methodology was pre-sented as a ‘‘non-intrusive’’ approach to determineusage of drugs of abuse in the community. What theyfound was surprising – cocaine consumption of 40,000doses per month from this study was greater than the15,000 doses per month from official figures estimatedby current methods.

Fig. 3 attempts to correlate human-metabolism data ofenvironmentally-relevant drugs with their respectiveremoval efficiencies estimated from monitoring surveysin WWTP. For most of the compounds depicted, con-siderable metabolic conversion in the human bodytranslates into moderate-to-high removal rates of thedrug in activated sludge treatment. Exceptions are car-bamazepine and fluoxetine, which are extensivelymetabolized in the human body but tend to persistduring sewage treatment. The two X-ray contrast agentsiopromide and diatrizoate, cannot be degraded biologi-cally although they have been reported to be subject tobiotransformation in nitrifying activated sludge [43]. Ininterpreting Fig. 3, we need to take into account that


bioavailability of pharmaceuticals in the human bodyvaries greatly (i.e. a considerable portion of a compoundcan be excreted without entering systemic circulation; inthis instance, metabolic clearance in the liver is notfeasible).

4. Conclusions

After administration to humans, pharmaceuticals can betransformed into metabolites and are excreted by thebody as a mixture of parent compound and metabolites.Analytical measurements have shown that residues canreach the aquatic environment through discharges fromWWTPs. For a better understanding of the fate and theoccurrence of pharmaceuticals in the environment, it isnecessary to gather information on the identity and thedistribution of drugs and their metabolites. To this end,sensitive, selective multi-analyte methods need to bedeveloped and validated.

Whereas in the human body the major metabolicroutes are known for most pharmaceuticals – with manyfindings disseminated through scientific publications –the products of the biodegradation brought about bymicroorganisms (e.g., by activated sludge bacteria inWWTP) are known only for a very small number ofpharmacologically-active compounds. Major researchefforts are required to investigate the metabolic path-ways of all environmentally-relevant pharmaceuticals innatural and engineered systems. Degradation studies inlaboratory-scale models should be designed so that realconditions are simulated as closely as possible.

The availability of sophisticated MS instruments (e.g.,QqToF-MS, LIT-MS and QqLIT-MS) has enabled consid-erable progress to be made in characterizing biotrans-formation products of pharmaceuticals in environmentalsamples. Efficient extraction methods combined withselective MS detection allow not only determination ofultra-trace levels of parent drugs in complex matricesbut also tracking of biotransformation products.

Given the large gap to be bridged in elucidating bio-degradation pathways of pharmaceuticals in the envi-ronment, there needs to be a significant increase inresearch efforts, in which modern MS instrumentationwill undoubtedly play a key role.

Acknowledgement

The work described in this article was supported by theEU Project EMCO-INCO-CT-2004-509188) and by theSpanish Ministerio de Educacion y Ciencia Project EVITA(CTM2004-06255-CO3-01). This work reflects only theviews of the authors, and the European Community isnot liable for any use that may be made of the infor-mation contained therein. SP acknowledges a post-

doctoral contract from I3P (Itinerario Integrado deInsercion Profesional) Program, co-financed by CSIC andthe European Social Fund.

References

[1] O.A. Jones, J.N. Lester, N. Voulvoulis, Trends Biotechnol. 23

(2005) 163.

[2] B. Lemmer, Chronopharmacology – Cellular and Biochemical

Interactions, Birkhauser, New York, USA, 1996.

[3] A. Boxall, K. Fenner, D.W. Kolpin, S. Maund, Environ. Sci.

Technol. 38 (2004) 369A.

[4] G. Hopfgartner, E. Varesio, V. Tschappat, C. Grivet, E. Bourgogne,

L.A. Leuthold, J. Mass Spectrom 39 (2004) 845.

[5] F. Hernandez, O.J. Pozo, J.V. Sancho, F.J. Lopez, J.M. Marın,

M. Ibanez, Trends Anal. Chem. 24 (2005) 596.

[6] W.F. Smyth, Anal. Chim. Acta 492 (2003) 1.

[7] G. Paul, W. Winnik, N. Hughes, H. Schweingruber, R. Heller,

A. Schoen, Rapid Comm. Mass Spectrom. 17 (2003) 561.

[8] Z. Tozuka, H. Kaneko, T. Shiraga, Y. Mitani, M. Beppu,

S. Terashita, A. Kawamura, A. Kagayama, J. Mass Spectrom. 38

(2003) 793.

[9] T. Reemtsma, J. Chromatogr. A 1000 (2003) 477.

[10] A. Makarov, E. Denisov, A. Kholomeev, W. Balschun, O. Lange,

K. Strupat, S. Horning, Anal. Chem. 78 (2006) 2113.

[11] C. Zwiener, S. Seeger, T. Glauner, F.H. Frimmel, Anal. Bioanal.

Chem. 372 (2002) 569.

[12] J.B. Quintana, S. Weiss, T. Reemtsma, Water Res. 39 (2005)

2654.

[13] D.R. Kepp, U.G. Sidelmann, J. Tjornelund, S.H. Hansen,

J. Chromatogr., B 696 (1997) 235.

[14] M. Winkler, J.R. Lawrence, T.R. Neu, Water Res. 35 (2001)

3197.

[15] T.A. Ternes, M. Stumpf, B. Schuppert, K. Haberer, Vom Wasser 90

(1998) 295.

[16] U.G. Sidelmann, I. Bjornsdottir, J.P. Shockor, S.H. Hansen,

J.C. Lindon, J.K. Nicholson, J. Pharm. Biomed. Anal. 24 (2001) 569.

[17] T. Heberer, K. Reddersen, A. Mechlinski, Water Sci. Technol. 46

(2002) 81.

[18] S. Perez, P. Eichhorn, D.S. Aga, Environ. Toxicol. Chem. 24

(2005) 1361.

[19] P. Eichhorn, P.L. Ferguson, S. Perez, D.S. Aga, Anal. Chem. 77

(2005) 4176.

[20] T. Kiffmeyer, H.J. Gotze, M. Jursch, U. Luders, Fresenius� J. Anal.

Chem. 361 (1998) 185.

[21] T.A. Ternes, P. Kreckel, J. Mueller, Sci. Total Environ. 225 (1999)

91.

[22] J. Shi, S. Fujisawa, S. Nakai, M. Hosomi, Water Res. 38 (2004)

2323.

[23] S. Perez, P. Eichhorn, M.D. Celiz, D.S. Aga, Anal. Chem. 78 (2006)

1866.

[24] W. Kalsch, Sci. Total Environ. 225 (1999) 143.

[25] T. Steger-Hartmann, R. Lange, H. Schweinfurth, M. Tschampel,

I. Rehmann, Water Res. 36 (2002) 266.

[26] A. Haiß, K. Kummerer, Chemosphere 62 (2006) 294.

[27] A. Joss, S. Zabczynski, A. Gobel, B. Hoffmann, D. Loffler,

C.S. McArdell, T.A. Ternes, A. Thomsen, H. Siegrist, Water Res.

40 (2006) 1686.

[28] S. Zuehlke, U. Duennbier, T. Heberer, Anal. Chem. 76 (2004)

6548.

[29] S. Weigel, U. Berger, E. Jensen, R. Kallenborn, H. Thoresen,

H. Huhnerfuss, Chemosphere 56 (2004) 583.

[30] M. Stumpf, T.A. Ternes, K. Haberer, W. Baumann, Vom Wasser

91 (1998) 291.

[31] C. Zwiener, S. Seeger, T. Glauner, F.H. Frimmel, Anal. Bioanal.

Chem. 372 (2002) 569.



[32] T.A. Ternes, M. Bonerz, T. Schmidt, J. Chromatogr., A 938 (2001)

175.

[33] T.A. Ternes, Water Res. 32 (1998) 3245.

[34] G.R. Boyd, H. Reemstma, D.A. Grimm, S. Mitra, Sci. Total

Environ. 311 (2003) 135.

[35] X.S. Miao, C.D. Metcalfe, Anal. Chem. 75 (2003) 3731.

[36] S. Castiglioni, R. Bagnati, D. Calamari, R. Fanelli, E. Zuccato,

J. Chromatogr., A 1092 (2005) 206.

[37] B.J. Vanderford, S.A. Snyder, Environ. Sci. Technol. 40 (2006)

7312.

[38] M.J. Hilton, K.V. Thomas, J. Chromatogr., A 1015 (2003) 129.

[39] A. Gobel, C.S. McArdell, M.J.F. Suter, W. Giger, Anal. Chem. 76

(2004) 4756.

[40] T.A. Ternes, M. Stumpf, J. Muller, K. Haberer, R.D. Wilken,

M. Servos, Sci. Total Environ. 225 (1999) 81.

[41] D.W. Kolpin, E.T. Furlong, M.T. Meyer, E.M. Thurman,

E.D. Zaugg, L.B. Barber, H.T. Buxton, Environ. Sci. Technol. 36

(2002) 1202.

[42] E. Zuccato, C. Chiabrando, S. Castiglioni, D. Calamari, R. Bagnati,

S. Schiarea, R. Fanelli, Environ. Health 4 (2005) 14.

[43] A.L. Batt, S. Kim, D.S. Aga, Environ. Sci. Technol. 40 (2006) 7367.

[44] F. Von Bruchhausen, G. Dannhardt, S. Ebel, A.W. Frahm,

U. Holzgrabe, fifth ed.Hagers Handbuch der Pharmazutischen

Praxis, vol. 7–9, Springer-Verlag, Berlin, Germany, 1994.

[45] R.F. Mills, S.S. Adams, E.E. Cliffe, W. Dickinson, J.S. Nicholson,

Xenobiotica 3 (1973) 589.

[46] J. Aarbakke, O.M. Bakke, E.J. Milde, D.S. Davies, Biomed. Life Sci.

Med. 11 (1977) 359.

[47] I. Carretero, M. Vadillo, J.J. Laserna, Analyst (Cambridge, U.K.)

120 (1995) 1729.

[48] R. Schmidt, R. Brockmeyer, Vom Wasser 98 (2002) 37.

[49] H. Uchino, T. Inaba, W. Kalow, Xenobiotica 13 (1983) 155.

[50] P.J. Meffin, D.M. Zilm, J. Chromatogr. 278 (1983) 101.

[51] U. Abshagen, W. Bablok, K. Koch, P.D. Lang, H.A.E. Schmidt,

M. Senn, H. Stork, Eur. J. Clin. Pharmacol. 16 (1979) 31.

[52] A. Weil, J. Caldwell, M. Strolin-Benedetti, Drug Metab. Disp. 18

(1990) 115.

[53] Missing reference to be provided.


[54] I. Wiesenberg-Boettcher, J. Pfeilschifter, A. Schweizer, A. Sallmann,

P. Wenk, Inflamm. Res. 34 (2005) 135.

[55] R.T. Foster, F. Jamali, A.S. Russell, S.R. Alballa, J. Pharm. Sci. 77

(1988) 70.

[56] C.R. Valentine, J.L. Valentine, J. Seng, J. Leakey, D. Casciano, Cell.

Biol. Toxicol. 12 (1996) 155.

[57] K. Lertratanangkoon, M. Horning, Drug Metab. Disp. 10 (1982)

1.

[58] N.R. Kitteringham, C. Davis, N. Howard, M. Pirmohamed,

B.K. Park, J. Pharmacol. Exp. Ther. 278 (1996) 1018.

[59] M. Mandelli, G. Tognoni, S. Garattini, Clin. Pharmacokinetics 3

(1978) 72.

[60] R.L. Lins, K.E. Matthys, G.A. Verpooten, P.C. Peeters, M. Dratwa,

J.C. Stolear, N.H. Lameire, Nephrol. Dial. Transplant. 18 (2003)

967.

[61] S. Vickers, C.A. Duncan, I.W. Chen, A. Rosegay, D.E. Duggan,

Drug Metab. Dispos. 18 (1990) 138.

[62] D.T. Wong, F.P. Bymaster, E.A. Engleman, Life Sci. 57 (1995)

411.

[63] T.B. Vree, Y.A. Hekster, Pharmacokinetics of Sulfonamides

(Revised), Karger, Basel, Switzerland, 1985.

[64] D.E. Schwartz, W. Vetter, G. Englert, Arzneimittel-Forschung 20

(1970) 1867.

[65] T. Meshi, Y. Sato, Chem. Pharm. Bull. 20 (1972) 2079.

[66] S. A Jacobs, R.G. Stoller, B.A. Chabner, D.G. Johns, J. Clin. Invest.

57 (1976) 534.

[67] C. Laritzen, Natural and synthetic sexual hormones – biological

basis and medical treatment principles, in: H.P.G. Schneider,

C. Lauritzen, E. Nieschlag (Editors), Grundlagen und Klinik der

menschlichen Fortpflanzung, Walter de Gruyter, Berlin, Germany,

1988.

[68] J. Rombke, Th. Knacker, P. Stahlschmidt-Allner, Study about

Environmental Problems in Context with Drugs, F+E Vorhaben Nr

106 04 121, Umweltbundesamt, Berlin, Germany, 1996.

[69] R.E. Ranny, J. Toxicol. Environ. Health 3 (1977) 139.

[70] F.P. Guengerich, Life Sci. 47 (1990) 1981.

[71] P.A. Todd, R.C. Heel, Drugs 31 (1996) 198.

[72] J. Ambre, J. Anal. Toxicol. 9 (1985) 241.

Application of advanced MS techniques to analysis and identification of human and microbial...

Documents

Transcript of Application of advanced MS techniques to analysis and identification of human and microbial...