jem spotlight: recent advances in analysis of pharmaceuticals in the aquatic environment

14
JEM Spotlight: Recent advances in analysis of pharmaceuticals in the aquatic environment Charles S. Wong * ab and Sherri L. MacLeod b Received 3rd November 2008, Accepted 10th February 2009 First published as an Advance Article on the web 9th March 2009 DOI: 10.1039/b819464e Both ecosystem and human health rely on clean, abundant supplies of water, thus many classes of potential pollutants are regulated. In recent years, the possible risks associated with largely uncontrolled inputs of pharmaceuticals to rivers, lakes, groundwater, and coastal waters, mainly via wastewater, have been a focus of much research. During this time, our capacity to sequester, identify, and quantify pharmaceuticals in environmental matrices has improved. Devices have emerged to allow passive uptake of drugs to augment or replace laborious grab sampling. Advances in sample preparation have streamlined extraction procedures and removed interfering matrix components. New instrumental techniques have allowed faster, more accurate and sensitive detection of drugs in water samples. This review highlights all of these advances, from sample collection to instrumental analysis, which will continue to help us better understand the fate and effects of pharmaceuticals in aquatic systems. Introduction Pharmaceuticals are an emerging environmental concern. Drugs are heavily used for both humans and veterinary animals. 1 These drugs enter the waste stream by disposal and excretion, and are generally not specifically targeted by wastewater treatment, and consequently enter receiving surface waters and groundwaters. 2–6 Although the concentrations of drugs in the aquatic environment are generally well below levels leading to acute human effects, 1 chronic effects on humans and on ecosystems may be possible. These subtle effects from long-term exposure are not well characterized, and may be different for aquatic organisms compared to humans. 7 Chemical effects depend on exposure, which is controlled by environmental fate processes. Thus, it is crucial to understand the occurrence, fate, and effects of pharmaceuticals in the environment, in order to assess properly the risks these highly biologically-active chemicals may pose to human and ecosystem health. None of the above can occur if there do not first exist reliable and robust methods by which to measure pharmaceuticals in environmental matrices. Drugs are typically present in receiving waters at extremely low concentrations (i.e., ng L 1 and lower) in extremely complex matrices full of possible interfering compounds (e.g., surface water, wastewater). Most of our current understanding of the fate, transport, and effects of environmental drugs stems from grab sampling of water, brought back to the laboratory for extraction and concentration into a form suitable Sherri MacLeod and Charles Wong Charles Wong is Associate Professor and Canada Research Chair in Environmental Toxicology at the University of Winnipeg. His research interests focus on the measurement, fate and effects of emerging pollut- ants. He received the 2003 Early Career Award for Applied Ecological Research from the Society of Environmental Toxicology and Chemistry (SETAC) and the American Chemistry Council, and the 2007 SETAC Weston Environmental Solutions Award for the outstanding environ- mental chemist of the year under age 40. He holds SB and SM degrees from MIT and a PhD from the University of Minnesota, all in civil and environmental engineering. Sherri MacLeod is a PhD Candidate at the University of Alberta. Her research focuses on mass balance of pharmaceuticals in Alberta water- sheds. She has won graduate fellowships from Canada’s Natural Sciences and Engineering Research Council (2003–2007), the Alberta Ingenuity Fund (2004–2009), and the American Chemical Society’s Division of Analytical Chemistry, sponsored by DuPont (2008). She holds a BSc (Honors) in chemistry from Acadia University. a Environmental Studies Program and Department of Chemistry, Richardson College for the Environment, University of Winnipeg, Winnipeg, MB, R3B 2E9, Canada. E-mail: [email protected]. edu; Fax: +1 204-775-2114; Tel: +1 204-786-9335 b Department of Chemistry, University of Alberta, Edmonton, AB, T6G 2G2, Canada This journal is ª The Royal Society of Chemistry 2009 J. Environ. Monit., 2009, 11, 923–936 | 923 CRITICAL REVIEW www.rsc.org/jem | Journal of Environmental Monitoring Published on 09 March 2009. Downloaded by Universitat Autonoma de Barcelona on 27/10/2014 09:41:01. View Article Online / Journal Homepage / Table of Contents for this issue

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Page 1: JEM Spotlight: Recent advances in analysis of pharmaceuticals in the aquatic environment

CRITICAL REVIEW www.rsc.org/jem | Journal of Environmental Monitoring

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View Article Online / Journal Homepage / Table of Contents for this issue

JEM Spotlight: Recent advances in analysis of pharmaceuticals in theaquatic environment

Charles S. Wong*ab and Sherri L. MacLeodb

Received 3rd November 2008, Accepted 10th February 2009

First published as an Advance Article on the web 9th March 2009

DOI: 10.1039/b819464e

Both ecosystem and human health rely on clean, abundant supplies of water, thus many classes of

potential pollutants are regulated. In recent years, the possible risks associated with largely

uncontrolled inputs of pharmaceuticals to rivers, lakes, groundwater, and coastal waters, mainly via

wastewater, have been a focus of much research. During this time, our capacity to sequester, identify,

and quantify pharmaceuticals in environmental matrices has improved. Devices have emerged to allow

passive uptake of drugs to augment or replace laborious grab sampling. Advances in sample

preparation have streamlined extraction procedures and removed interfering matrix components. New

instrumental techniques have allowed faster, more accurate and sensitive detection of drugs in water

samples. This review highlights all of these advances, from sample collection to instrumental analysis,

which will continue to help us better understand the fate and effects of pharmaceuticals in aquatic

systems.

Introduction

Pharmaceuticals are an emerging environmental concern. Drugs are

heavily used for both humans and veterinary animals.1 These drugs

enter the waste stream by disposal and excretion, and are generally

not specifically targeted by wastewater treatment, and consequently

enter receiving surface waters and groundwaters.2–6 Although the

concentrations of drugs in the aquatic environment are generally

well below levels leading to acute human effects,1 chronic effects on

humans and on ecosystems may be possible. These subtle effects

Sherri MacLeod and Charles Wong

Charle

Enviro

interes

ants. H

Resea

(SET

Westo

menta

from M

enviro

Sherri

resear

sheds.

and Engineering Research Council (2003–2007), the Alberta Ingenuity

of Analytical Chemistry, sponsored by DuPont (2008). She holds a B

aEnvironmental Studies Program and Department of Chemistry,Richardson College for the Environment, University of Winnipeg,Winnipeg, MB, R3B 2E9, Canada. E-mail: [email protected]; Fax: +1 204-775-2114; Tel: +1 204-786-9335bDepartment of Chemistry, University of Alberta, Edmonton, AB, T6G2G2, Canada

This journal is ª The Royal Society of Chemistry 2009

from long-term exposure are not well characterized, and may be

different for aquatic organisms compared to humans.7 Chemical

effects depend on exposure, which is controlled by environmental

fate processes. Thus, it is crucial to understand the occurrence, fate,

and effects of pharmaceuticals in the environment, in order to assess

properly the risks these highly biologically-active chemicals may

pose to human and ecosystem health.

None of the above can occur if there do not first exist reliable

and robust methods by which to measure pharmaceuticals in

environmental matrices. Drugs are typically present in receiving

waters at extremely low concentrations (i.e., ng L�1 and lower) in

extremely complex matrices full of possible interfering

compounds (e.g., surface water, wastewater). Most of our current

understanding of the fate, transport, and effects of environmental

drugs stems from grab sampling of water, brought back to the

laboratory for extraction and concentration into a form suitable

s Wong is Associate Professor and Canada Research Chair in

nmental Toxicology at the University of Winnipeg. His research

ts focus on the measurement, fate and effects of emerging pollut-

e received the 2003 Early Career Award for Applied Ecological

rch from the Society of Environmental Toxicology and Chemistry

AC) and the American Chemistry Council, and the 2007 SETAC

n Environmental Solutions Award for the outstanding environ-

l chemist of the year under age 40. He holds SB and SM degrees

IT and a PhD from the University of Minnesota, all in civil and

nmental engineering.

MacLeod is a PhD Candidate at the University of Alberta. Her

ch focuses on mass balance of pharmaceuticals in Alberta water-

She has won graduate fellowships from Canada’s Natural Sciences

Fund (2004–2009), and the American Chemical Society’s Division

Sc (Honors) in chemistry from Acadia University.

J. Environ. Monit., 2009, 11, 923–936 | 923

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for instrumental analysis, typically by either gas chromatography

(GC) or liquid chromatography (LC), particularly coupled to

mass spectrometry (MS) and tandem mass spectrometry (MS/

MS). Many such analytical methods have been recently

reviewed,8–10 including those for specific classes of drugs, like non-

steroidal anti-inflammatory drugs,11 b-blockers,12 and antibi-

otics.13 Important considerations for a chemical monitoring plan

have likewise been discussed,14 as has the key role of sampling in

environmental analysis.15–18 While the current state of knowledge

is useful, there remain many gaps in our understanding of drugs in

aquatic ecosystems. These arise in part because of limitations in

the sampling and analytical protocols currently used.

The objective of this review is to highlight some recent advances

in the measurement of pharmaceuticals and personal care products

in the aquatic environment. In particular, we discuss innovations in

all aspects of such analytical techniques that are improvements

from traditional grab sampling, solid phase extraction (SPE) and

concentration, and liquid chromatography/tandem mass spec-

trometry (LC/MS/MS) chromatography and analysis, the method

of choice for current instrumental analysis for trace polar envi-

ronmental chemicals. We detail advances with sample collection

and handling; extraction, processing, and cleanup procedures;

analytical separations via LC; and finally instrumental detection

and analysis (Table 1). All these aspects are vital for reliable

monitoring of drugs in waters, and to our knowledge a ‘‘cradle-to-

grave’’ approach to analysis have not been covered in a single

review to date. We also discuss improvements in the analysis of

chiral pharmaceuticals and of degradation products, both of which

provide enhanced insight into the occurrence, fate, transport, and

effects of drugs in the natural and engineered waters. Given the

breadth of our approach and space limitations, we have restricted

our discussion to studies occurring within the last four years or so.

Table 1 Selected sample preparation and LC-based instrumental analysis te

Traditional

Sample PreparationSampling method Grab4,5,40,42,45–51,53,57,117

Short-term temporal composite2,38,41,58,59,80,

Sample pre-treatment Samples almost always filtered.Reagent addition (e.g., Na2EDTA)38,44–48 a

adjustment38,39,41,44,46,49,53,81,83 are analyteRecovery determination

extractionSpike-and-recovery experiments40–42,104

SPE (e.g., Oasis HLB)38,40,41,44–48,50,52,55,56,58,5

Elution Aqueous wash followed by organic solvenelution,38,40–42,44,46–48,55–59 sometimes withadjustments45,49,50,52,53

Instrumental analysisMatrix effects reduction Volatile LC additives53,102

Chromatography High performance LC5,38–42,44–50,55,57,58,80,81,8

Mass spectrometry ToF58 QqQ5,40–42,47,50,51–53,55,57,59,61,80,81,83,99,102

QLIT46,58

Orbitrap119

Overall approach Targeted analysis for drugs commonly usemetabolized/frequently found38,41,42,44–48,5

924 | J. Environ. Monit., 2009, 11, 923–936

Much has been achieved in these areas of research in that time

period, and earlier reviews cited in this work have covered prior

advancements.

Sample collection and processing

Passive samplers

Traditional water sampling for chemical monitoring consists of

‘‘grab’’, ‘‘spot’’, or ‘‘bottle’’ sampling,14 and is typically carried

out by direct fill at the surface, submerged samplers triggered at

desired depths, or portable pumps. Active sampling methods

account for much existing knowledge of these emerging

contaminants in the aquatic environment. However, these

samples are taken at only a specific time and location, and may

not necessarily be representative of chemical residues of the

water body at other times. In some cases, particularly where

continuous or episodic inputs are expected, monitoring of

chemical contamination might require a more intensive sampling

strategy. More frequent sampling or the use of automatic or

continuous samplers may provide temporally representative

data. However, such options are generally less attractive due to

increased maintenance and cost,14 including use of electrical

power precluding extensive deployment in remote regions.

Passive sampling devices, which consist of a protective housing

with a collector material that sorbs analytes, are a complemen-

tary or alternative technique to active sampling. They are cost

effective tools for qualitative and semi-quantitative screening, as

well as a quantitative measurement of dissolved phase contami-

nants including pharmaceuticals.14,19

A number of passive samplers have been developed for

optimal accumulation of particular chemical classes,18 given

chniques for pharmaceuticals in the aquatic environment

Alternative and complementary techniques

Passive sampling devices22–31

81,90

Centrifugation38,39

nd pH-dependent

Standard addition41,42,52,57,81,86

9,117 Online SPE39,79,80–83

Mixed-mode SPE49,51,53,57,68,69

MIPs72–78

SPME66,86 and LPME89–91

tpH

Online SPE39,79,80–83

UPLC51–53,56,59,99

Dilution40,53

Split flow57

Labeled internal standards40,47,48,55,59,80,112

3,90,117 UPLC51–53,56,59,99,118

HILIC129,130

Chiral stationary phases3,6,104,123,124

QToF39,49,56,60,88,117,118

d/not extensively0–53,55,57–59,81,83,90,117,127

Screening analysis49,56,58,118

Metabolites51,55,58,59,80,99,102,110,130,131

This journal is ª The Royal Society of Chemistry 2009

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differences in physical-chemical properties of pollutants.

Hydrophobic contaminants (e.g., persistent organic pollutants)

are frequently monitored with semi-permeable membrane

devices (SPMDs),20 although other samplers have also been used,

such as Chemcatchers with an appropriate receiving phase (e.g.,

n-octanol-saturated octadecyl carbon Empore� disks).21 The

Polar Organic Chemical Integrative Sampler (POCIS)22 collects

hydrophilic contaminants, including pharmaceuticals, from

wastewater treatment plant (WWTP) effluent,23,24 impacted

rivers24,25 and streams,23,26 lakes,24 constructed wetlands,27 and

estuaries.28 Empore� SDB-RPS disks in a Chemcatcher housing

serve a similar role for polar contaminants in WWTP effluent

and river water.29

Passive samplers can play an important role in facilitating

large-scale temporal and spatial monitoring programs for phar-

maceuticals. Unlike active samplers, they provide continuous

monitoring of time-weighted-average (TWA) concentrations of

target analytes. In addition, they provide a rich dataset with

potential advantages of reduced cost, time, and effort. Contin-

uous sampling enables compound detection at concentrations

below limits associated with spot sampling of both target ana-

lytes24,26 and non-target chemicals,23,27,28 as passive sampling

effectively samples more water than spot sampling to provide

more mass per sample. For example, a study of estrogenicity of

Swiss WWTP effluents and rivers initially conducted grab

sampling campaigns with highly variable results.30 Follow-up

work with POCIS, showed that the main sources of the vari-

ability were environmental factors and efficiency of the WWTP

processes. Thus, passive samplers were an effective means to

assess both WWTP efficiency and chemical loadings from

wastewater, and to understand the variability in pharmaceutical

contamination to receiving waters.

Passive samplers must be calibrated to determine analyte

sampling rates, Rs, for quantifying TWA concentrations. These

rates depend on the degree of analyte and sorbate saturation of

the sampler sequestration material. When the analyte concen-

tration in the passive sampler sorbent is well below equilibrium

levels, the sampler will collect analytes over time in a linear

fashion (Fig. 1). As the sorbent approaches saturation, a curvi-

linear sampling rate is expressed, which levels off to an equilib-

rium value at saturation (Fig. 1). Linear Rs are preferred, as

equilibrium partitioning coefficients are often difficult to measure

Fig. 1 Mass of analyte sequestered to passive sampling device over time

in passive sampling devices, based on Huckins et al.34

This journal is ª The Royal Society of Chemistry 2009

for sorbents with a high affinity for analytes, such as Oasis

hydrophilic-lipophilic balanced (HLB), a divinylbenzene-N-

vinylpyrrolidone copolymer typically used for sequestration of

many polar chemicals22. The TWA concentration is easily deter-

mined from the amount of analyte collected, divided by the

effective amount of water sampled.19 The latter is the product of

the sampler deployment duration and the Rs, generally expressed

as the effective water volume cleared of the analyte per unit time

(e.g., L d�1). Using a custom-built flow-through chamber

immersed in WWTP effluent and river water, Chemcatcher Rs for

carbamazepine, clarithromycin, and sulfamethoxazole by

Empore� disks were 0.09, 0.14, and 0.25 L d�1, respectively, at 12

to 14 �C and a flow rate of 0.03 m/s.29 Uptake was linear only for 1

d for sulfamethoxazole to 5 d for clarithromycin. As a conse-

quence, integrative sampling can only be expected for flow <0.1

m/s and/or for short deployment periods29 for characterizing

short-term contamination events such as pulse inputs.31

If longer integration periods are required, a different sampler

may be more appropriate, such as the POCIS for which linear

uptake of >30 d has been reported.22–25,28 The first reports of

laboratory based calibration for POCIS uptake of drugs included

Rs for azithromycin, fluoxetine, levothyroxine and omeprazole,22

and methamphetamine and methylenedioxymethamphetamine.23

Another laboratory based calibration experimentally determined

POCIS Rs for 25 pharmaceuticals and personal care products,

including b-blockers, selective serotonin re-uptake inhibitors

(SSRIs) used as antidepressants, non-steroidal anti-inflamma-

tory drugs (NSAIDs), and antibiotics, commonly found in

environmental waters.24 Discrepancies exist between Rs for

fluoxetine and omeprazole, the two drugs comment to both

studies, possibly due to differences in flow rate, sampler geom-

etry, and study design.22,24 These differences point out the need

for both standardized methods of calibration and caution in data

interpretation.

Researchers have attempted to discern relationships between

Rs and the physical-chemical characteristics of analytes.24,28,32

Prediction of Rs for uncalibrated analytes would thus be possible,

vastly increasing the versatility of passive samplers. In fresh-

water, increasing hydrophobicity (i.e., higher octanol-water

partition coefficients, Kow) resulted in larger values of POCIS Rs

for basic drugs at low ionic strength, but not for acidic drugs;

neither was the case in seawater.28 Other POCIS results suggested

a curvilinear relationship between log Kow and Rs,32 with sepa-

rate trends for anions, cations, and zwitterions.24 The mecha-

nisms behind these trends are as yet unclear.

In addition to chemical factors controlling Rs, environmental

parameters such as water flow rate, temperature, pH, and

biofouling may impact analyte Rs, presenting a challenge to

applying Rs from calibration work to field measurements. Drug

TWA concentrations from POCIS calibrations were generally

similar to those from spot sampling, although some TWA

concentrations were lower.24,25,28 These discrepancies could

simply be due to differences between TWA concentrations over

the entire sampler deployment period, and that measured

instantaneously during spot sampling. Alternatively, differences

in environmental conditions between calibration and deploy-

ment could also affect Rs, and hence TWA concentrations.

There is evidence that uptake of drugs to passive samplers is

boundary-layer controlled, at least under some water flow rates.

J. Environ. Monit., 2009, 11, 923–936 | 925

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For POCIS, drug Rs were often higher in flowing water (Rs from

0.030 to 2.462 L d�1) than in quiescent water (Rs from 0.007 to

0.223 L d�1).22,24 For three drugs (omeprazole, paroxetine, and

sulfisoxazole), a quiescent water Rs could not be calculated.24

Flow chamber experiments using POCIS found that uptake of

estrogenic substances was approximately doubled when flow was

increased from 0.025 to 0.37 m s�1.30 For Chemcatcher, analyte

uptake increased with increasing flow rates.29 Field-derived

POCIS Rs for endocrine disrupting compounds (EDCs) were

significantly higher (0.3 to 0.8 L d�1) than lab-derived Rs (0.036

to 0.069 L d�1), likely due to increased flow.25 All these obser-

vations suggest that mass transfer from bulk water into the

sampler sequestration phase was water-side boundary-layer

controlled. However, no differences were observed in POCIS Rs

for non-estrogenic drugs at flow rates between 0.03 to 0.12 m

s�1,24 suggesting that factors other than boundary-layer diffusion

controlled uptake at those flow rates. Likewise, Chemcatchers

consisting of bare Empore� disks had Rs an order of magnitude

higher than those that also had a protective polyethersulfone

membrane.31 This observation suggests that diffusion through

the membrane also limited mass transfer, which can be beneficial

in increasing the integration period of Chemcatchers.31 Hydro-

dynamics has a complex effect on passive Rs, as shown for

SPMDs33, but an extensive analysis of this parameter for passive

drug samplers has not been conducted to date.

The effects of temperature, salinity, and analyte concentration

on POCIS sampling rate were investigated for 17 common

pharmaceuticals.28 These results are difficult to interpret, as

replicates were not provided for each condition, and it is clear if

the water flow rate was consistent among experiments. However,

a temperature change from 15 �C to 21 �C increased Rs for some

drugs,28 likely due to an increase in analyte diffusivity and mass

transfer.22 Increasing ionic strength from 0 M (freshwater) to 0.7

M (seawater) decreased Rs for some drugs, presumably from the

‘‘salting out’’ effect.28 However, no significant change in POCIS

Rs of EDCs bisphenol A, estrone, 17-b-estradiol and 17-b-ethy-

nylestradiol was observed for ionic strength values of 0, 0.35 and

0.7 M, or for pH values between 4 and 10.25 Analyte concen-

tration had no effect on POCIS Rs for drugs28 or EDCs.25

Membrane biofouling may lower passive Rs, by adding an

additional layer of resistance to analyte mass transfer from water

into the sorbent. Polyethersulfone was adopted as a membrane

for POCIS, as it resisted biofouling more effectively than other

materials.22 Discrepancies between TWA concentrations and

spot sampling for drugs was suspected in POCIS in wastewater

effluent, in which some biofilm growth on devices was

observed.24 Chemcatchers without membranes had a fourfold

reduction in sampling rate after biofilms developed.31 It is also

possible that bacterial activity in biofilms can biotransform

analytes diffusing into passive samplers, but this confounding

effect has not been investigated to date.

Parameter-specific calibrations may be alleviated by the

addition of performance reference compounds (PRCs) to passive

samplers before deployment. The PRC, which is not present in

the environment, desorbs during deployment. If both sorption

and desorption kinetics are similar, then release of the PRC

during deployment would be indicative of Rs for other analytes,

and the PRC essentially serves as an internal standard. Such an

approach has been applied for other passive sampling devices

926 | J. Environ. Monit., 2009, 11, 923–936

such as SPMDs.34 For POCIS in particular, PRCs have not been

widely reported as the sorptive capacity of the polymer adsor-

bents is quite high32 and equilibrium is not reached. Desisopro-

pylatrazine is a possible PRC for POCIS uptake of herbicides, as

its desorption from the sorbent was measurable within a suitable

sampling timeframe.35 However, further research in this area is

required to further the applicability of passive samplers to

monitor pharmaceutical pollution. For example, it is not clear if

Oasis HLB has anisotropic sorption characteristics (i.e., kinetics

for sorption and desorption are similar), given that adsorption of

analytes via strong lone-pair interactions with the sorbent is

a major mechanism of sorption.35 Other approaches, such as

PRC use on a sampler with a different sorbent but the same

surface area and membrane configuration, have been suggested22

but not yet extensively investigated.

Though passive samplers are widely used for air sampling,

particularly for occupational health applications, they have not

yet gained wide acceptance in water sampling for polar pollut-

ants.18 A number of important challenges have been laid out for

passive sampling of pharmaceuticals in the aquatic environ-

ment.19 Some of these challenges have been discussed: calibration

rates for new analytes must be determined, a major limiting

factor; a better understanding is needed to relate analyte and

environmental properties to Rs; and calibration methods should

be standardized. In addition, stringent quality control proce-

dures are necessary, as exemplified by an unsuccessful attempt to

determine POCIS Rs for the widely-used antibacterial agent,

triclosan, given laboratory contamination found in blanks.36 In

addition, new sorbents and devices are needed.19 A major chal-

lenge in the adoption passive sampling devices to a regulatory

framework for water analysis is the limited availability of

rigorous, field-based validation.14 Finally, studies which link

passive sampler extract toxicity with targeted chemical analysis

would be valuable, if specific biological effects of mixtures of

pharmaceuticals and other pollutants are suspected.19,27,37

Solid phase extraction

Sequestration of pharmaceuticals from water samples is most

frequently accomplished using SPE, which serves to concentrate

analytes from trace ng L�1 level concentrations to levels suitable

for instrumental analysis. Samples are nearly always filtered,

typically through 0.45 mm filters, to remove particulate matter

that can clog SPE phases and LC columns, although some

samples with high particulate loads (e.g., wastewater influent) are

also centrifuged.38,39 Depending on the analyte, reagents may be

added or pH adjusted for optimum recovery to waters prior to

SPE, which is typically assessed from spike-and-recovery

experiments.40–42 Chelation agents (e.g., Na2EDTA) are the

most common reagents added, to sequester metal ions

that could otherwise complex some analytes, particularly

tetracyclines.38,43–48 Adjustment of sample water pH may increase

the affinity of target analytes for the SPE sorbent, depending on

the makeup of the sorbent and specific analytes.45,49–53 An

aqueous wash following SPE extraction may be performed,

followed by collection of analytes via elution with an organic

solvent.38,40–42,44,46,47,54–59

There are a number of detailed reviews on optimizing SPE for

extraction of drugs from wastewaters and surface

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waters.8,10,40,60,61 Thus, our focus is to discuss recent alternative

and complementary techniques. The most commonly used SPE

sorbent for pharmaceutical extraction from environmental

waters is Oasis HLB, a hydrophilic and lipophilic polymer that

can simultaneously extract acidic, neutral and basic polar ana-

lytes62 at a wide range of pH values, wets easily, and can be

allowed to run dry without adversely affecting extraction effi-

ciency. These properties make it useful for extraction of multi-

class analytes.40,52,62,63 Other SPE materials have also been used,

such as octadecyl carbon and ISOELUTENV+. Other sorbents

continue to be developed, including sol–gels64 and carbon

nanotubes.65

It is crucial to note that SPE is a concentration procedure, not

a cleanup procedure per se, as the SPE sorbent will sequester all

material, with varying degrees of efficiency, with an affinity to the

phase.66 This fact is problematic in the collection of analyte from

complex environmental samples such as wastewater, as matrix

components which may interfere with analysis (see Ionization

sources and matrix effects section) at much higher concentrations

than the analytes of interest are also sequestered by SPE. Some

investigators have been able to remove more interferences from

water samples via the use of multiple SPE sorbents with different

stationary phases in tandem, such as anion-exchange with HLB

for basic antibiotics such as fluoroquinolones, sulfonamides, and

trimethoprim;63 phenyl SPE extraction followed by elution

through a Bond Elut NH2 SPE phase;42,57 and cleanup of

wastewaters for diclofenac, aceclofenac, and hydroxylated

metabolites through Oasis HLB and ISOELUTENV+.67

Cleanup using a single SPE cartridge with mixed-mode reversed-

phase cation exchange media68 such as Oasis MCX, built upon

HLB copolymer53 and providing multiple modes of sorption of

analytes to the SPE phase, has been successful at reducing

interferences and matrix effects of drugs in waters.49,51,53,69 Size

exclusion chromatography has also been used to remove high

molecular weight material, such as dissolved organic carbon in

waters and wastewaters.70,71

A promising extraction and cleanup procedure involves the use

of molecular imprinted polymers (MIPs). These polymeric

stationary phases are made with a molecular template corre-

sponding to the analyte or analyte class of interest. The phase

polymerizes around the template, which is then washed off,

leaving behind sorbent sites shaped like the template molecule for

selective extraction of the analyte. Ideally, chemicals with

structures not fitting the template site will not be collected. This

approach has been successfully applied for extraction from

environmental waters of estrogenic compounds,72 tetracycline73

and fluoroquinolone74 antibiotics, NSAIDs,75–77 b-blockers,78

and clofibric acid.77 Careful optimization of extraction proce-

dures (e.g., pH, extraction time, composition of extraction

solvents) and assessment of non-specific analyte sorption and

cross-reactivity (i.e., how well the MIP collects analytes struc-

turally similar to the template) is necessary in MIP analysis. Most

such materials must be custom made to date, and the robustness

and reliability of MIPs needs to be evaluated before widespread

commercial adoption of these materials for trace environmental

pharmaceutical analysis can take place.

Most SPE processing is performed offline; i.e., separately from

chromatographic separation and detection via instrumental

analysis. Offline SPE is time-consuming and laborious. Online

This journal is ª The Royal Society of Chemistry 2009

methods have been developed to streamline and automate

extraction, concentration, and instrumental analysis by directing

solvents to elute analytes from the sorbent directly to chro-

matographic columns. This is achieved either through dedicated

online SPE cartridge systems,79,80 or through column switching,

in which a cleanup LC column with large diameter stationary

phase particles (e.g., 5–12 mm) collected analytes which are then

flushed via an elution solvent system into a chromatographic LC

column.39,81–83 The main disadvantages of online SPE are the

capital cost for commercial systems, and limited sample size.

Microextraction

Another method to sample pharmaceuticals in waters is micro-

extraction, wherein analytes in aqueous solution are sorbed to

a stationary phase, then desorbed for instrumental analysis

thermally for GC or via solvents for LC. Microextraction has

a number of advantages over SPE. There is minimal use of both

sample (i.e., several mL at most versus 1 L and more for SPE) and

extraction solvent in the case of LC. Microextraction is also

a rapid way to process samples, as the amount of stationary

phase is small resulting in relatively short equilibration times for

analyte collection. In addition, microextraction simultaneously

extracts, cleans up, and concentrates samples. Both solid phase

microextraction (SPME) and liquid-phase microextraction

(LPME) techniques have been developed for measuring drugs in

environmental waters, and are discussed here.

In SPME, a fiber coated with a stationary phase is exposed to

the sample, typically until equilibrium is reached. These fibers are

reusable, unlike single-use SPE cartridges, resulting in cost-

savings along with the reduction in time, labor, sample, and

supplies. Most SPME environmental applications to date have

focused on extraction of nonpolar analytes,84 as few polar phases

currently exist commercially. Accordingly, much earlier work on

SPME for polar analytes has focused on derivatization to

increase their affinity for nonpolar polydimethylsiloxane

(PDMS) SPME fibers and to increase volatility for GC analysis,

as reviewed elsewhere.84 The use of SPME for LC is natural as

analytes can simply be desorbed into a multiport injector appa-

ratus directly to chromatographic columns. As with all method

development, optimization of multiple parameters is crucial. In

the case of SPME to LC, these parameters include choice of fiber;

sample pH, ionic strength, temperature; and extraction and

desorption time. In the case of drugs in environmental waters,

SPME optimization has been successfully applied to antibiotics

such as tetracyclines,85 sulfonamides,66,86 macrolides and

trimethoprim.86 Only a small amount of matrix components in

the sample is transferred to the SPME fiber during equilibration,

so sample cleanup is efficient.66,86 However, for the same reason,

detection limits and precision also tend to be worse than tech-

niques that collect most to all available analyte in the sample.60,86

Liquid-phase microextraction is essentially miniaturized

liquid–liquid extraction,87 an early method for extracting polar

materials from aqueous samples with the disadvantages of being

time-consuming, tedious, and wasteful of solvent. In LPME,

water samples are extracted typically into a porous hollow fiber

impregnated with an immiscible (e.g., organic) liquid phase col-

lecting analytes of interest; this liquid acceptor phase is then

analyzed. Three-phase LPME is also possible, in which the

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extracted analytes are back-extracted into a separate aqueous

phase to allow for adjustment of pH and analysis by reversed-

phase LC. The various LPME techniques have the advantages of

significant concentration of analytes from samples, efficient

cleanup, and minimal use of solvent (e.g., mL). The use of LPME

in environmental applications has recently been reviewed.84,87,88

As with SPME, much work with LPME has focused on nonpolar

contaminants, or polar chemicals with derivatization for GC-

based analysis. Acidic drugs and NSAIDs have been extracted

using LPME without significant coextraction of matrix compo-

nents interfering with LC/MS/MS analysis.89 Using an LPME

method, the occurrence of SSRIs was investigated in urban and

remote sewage and receiving waters in Norway.90 The efficient

cleanup afforded by LPME makes it possible, at least potentially,

for non-MS-based LC instrumental analysis to be used, as shown

for salbutamol and terbutaline in aqueous samples.91 As with

SPME, LPME precision tends to be lower than with other

techniques,88 in part because all aspects of using LPME to date

are manual, from fiber preparation, to conditioning, to pro-

cessing of very small extract volumes. In addition, variability in

fiber wall thickness and pore size also affect precision.88

Stirbar sorptive extraction

One way to overcome the limited analyte amount transferred

from equilibrium microextraction techniques is to use more

stationary phase. However, doing so has the disadvantage of

slow kinetics and time to equilibrium. Stirring will increase mass

transfer. Stirbar sorptive extraction (SBSE) is based upon this

premise, by building the sorbent onto the stirbar itself, which

provides rapid mass transfer in the solution to be extracted. A

much larger proportion of the analyte is transferred into the

sorbent by SBSE than by microextraction, resulting in greater

sensitivity. The SBSE technique was originally developed for

nonpolar analytes such as persistent organic pollutants.92 These

analytes sorb well to the PDMS sorbent used in Gerstel’s

commercial Twister SBSE, and are efficiently transferred via

thermal desorption in heated injectors for analysis by GC. For

analysis of polar analytes such as pharmaceuticals, two separate

approaches have been taken. The first involves derivatization of

polar analytes to make them more amenable to sorption to

PDMS phases and for GC analysis.93–96 Alternatively, stationary

phases more polar than PDMS, such as sol–gels for estrogens97

and polyurethane98 for NSAIDs and other acidic drugs, have

been developed for extraction followed by LC analysis. More

research is needed into optimizing SBSE extraction for a larger

number of drugs for widespread adoption of this promising

technique to take place.

Analytical separations

Although many of the first reports of environmental pharma-

ceutical residues were based on analysis by GC, most work now

uses LC to avoid the increased time, potential for analyte loss,

and possible safety issues associated with derivatization proce-

dures necessary to analyze polar or thermally labile analytes

including many pharmaceuticals.8–10 For LC, reversed-phase

chromatography with a octadecyl-based stationary phase is most

commonly used, with eluent systems generally consisting of

928 | J. Environ. Monit., 2009, 11, 923–936

combinations of acetonitrile, methanol and water with additives

to improve peak shape, retention, and resolution.8–10 Good

chromatographic separations are desirable, even with sophisti-

cated detectors like mass spectrometers, as water samples may

contain many substances that can interfere with the analytes of

interest.

Recently, ultraperformance liquid chromatography (UPLC)

has been explored for this type of analysis.56 By using columns

with smaller particles (1.7 mm diameter versus regular diameter of

5 mm), UPLC results in higher back-pressures requiring special

pumps, but uses less solvent and provides improved speed,

resolution, and sensitivity from narrower and sharper chro-

matographic peaks.56 Four rapid and sensitive UPLC methods

were recently used to target 48 prescription drugs and 6 metab-

olites in wastewater and surface water. The total combined

runtime was 48 min and the limits of detection (LODs) ranged

from 0.1 to 26 ng/L in real samples.51 Another UPLC method

was used target several drugs of abuse (cannabinoids and

opiates) along with some of their metabolites in wastewater and

surface water with separation in less than 8 min.99 Another

advantage of UPLC is the reduction of matrix effects during MS/

MS detection, a topic discussed further in that section.

Analysis of novel pharmaceuticals often requires novel

analytical separation techniques. Enantiomers of chiral phar-

maceuticals require some means of enantioselective separations

to resolve. Highly polar pharmaceuticals,100 or polar metabolites

of pharmaceuticals also require appropriate consideration of

interactions between the analyte and the chromatographic

stationary phase. These classes of analytes are resolved with LC

using enantioselective chromatography and hydrophilic inter-

action chromatography (HILIC), respectively, which are dis-

cussed in the section on novel analytes.

Ionization sources and matrix effects

The electrospray ionization (ESI) source is the most common

interface to mass spectrometers for trace drug analysis in waters,

given high sensitivity. Electrospray ionization efficiency is

heavily influenced by the composition of the LC mobile phase

and the chemistry among mobile phase molecules, analytes, and

matrix components.8,10,60 As a result, mobile phase eluents and

additives for LC/MS/MS must be selected carefully, to provide

good chromatographic peak shape, retention and resolution and

to have efficient ionization in the MS source.53,101 As an example

of the impact of solution chemistry on ESI performance,

ammonium bicarbonate provided much better sensitivity as

a mobile phase eluent than ammonium acetate, in the analysis of

basic tricyclic antidepressants and SSRIs.102

The dependency of ESI performance on solution chemistry

also has the drawback that ESI is quite susceptible to ion

suppression or enhancement as a result of co-eluted sample

components or ‘‘matrix effects’’,103 particularly with complex

samples like wastewater and surface water8,10,101 (Fig. 2).

Although the exact mechanism is not well understood,103 it is

likely that a major source of this interference in extracts from

aqueous environmental samples is the presence of organic

matter63 such as humic acids.80 These effects must be taken into

account for proper quantitative analysis of drugs in environ-

mental waters.

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Fig. 2 LC/ESI/MS/MS Ion suppression for analytes in water from the river Rhine as well as influent and effluent of a municipal WWTP. Reproduced

with permission from ref. 55. Copyright ª 2006 American Chemical Society.

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To evaluate matrix effects, comparisons can be made between

the signals for analytes spiked into blank water and drinking,

surface, or wastewater. For samples with analyte already present

(i.e., most wastewaters and some surface water), the corre-

sponding signal can be subtracted from that of unspiked samples

(i.e., simple standard addition).40,41,52,56,80,81,83,104 Significant

analyte-dependent40 matrix-related ESI signal modification is

frequently reported in wastewater, with less severe effects in

surface water samples.55,56,81,83 Suppression ranged from 15 to

35% for five sulfonamides, trimethoprim and diclofenac in spiked

wastewater extracts compared to spiked deionized water.41 Three

sets of samples–unspiked, spiked before extraction, spiked after

extraction–were compared to standards for nine basic drugs, but

the spiked samples could not be used to compensate adequately

for matrix effects, as matrix-spiked calibration curves did not

overlap with standard calibration curves.57 Matrix-matched

calibration standards are uncommon, as it is difficult to find an

ideal matrix that does not already contain analytes of interest. As

matrix components may vary by location and time, the use of

matrix-matched calibration is impractical for a multi-site and/or

temporal study.48

Recent reviews of analytical techniques for pharmaceuticals in

environmental samples have discussed matrix effects as an

important area where improvements must be made for accurate

quantitation.10,13,60,101 Some of the advantages and disadvantages

of several such strategies to reduce or circumvent matrix effects

are discussed below.

Changes in sample preparation may reduce the severity of

matrix effects by removing some of the unwanted components

from extracts. For example, MIP cleanup, because of its speci-

ficity in analyte extraction, can provide extracts with little

endogenous material and minimal matrix effects, even for

traditionally dirty samples such as wastewater.77 Likewise,

samples collected by SPME also had minimal matrix effects

This journal is ª The Royal Society of Chemistry 2009

compared to those collected by SPE, as little endogenous mate-

rial was transferred in equilibrium SPME partitioning.66,86

However, some such changes, such as extensive sample

cleanup,57 different SPE cartridges53,57 and online SPE39,79–83 may

result in increased workload and/or cost without being effective

for multi-analyte and/or multi-matrix studies.47 For example, the

use of restricted access materials to remove high molecular

weight matrix interferences from wastewater for drug analysis

was not effective, as the dissolved organic carbon in the tested

wastewater consisted mostly of low molecular weight material

that the restricted access material could not remove.70 Other

researchers, however, have reported reductions in matrix effects

using size exclusion chromatography cleanup procedures, so this

effect is likely to be heavily matrix-dependent.71 Recently,

cleanup of wastewater using micellular sodium dodecyl sulfate

surfactants was reported, with no matrix effects to enable LC

with ultraviolet absorbance detection of ibuprofen and naproxen

at mg L�1 concentrations.105 However, no confirmatory analysis

by LC/MS/MS was performed, and it is unclear if this cleanup

procedure is effective at the ng L�1 concentrations typical of

surface water concentrations, or is applicable for other drugs.

The use of extract dilution may reduce the severity of matrix

effects by reducing the amount of matrix introduced into the

ionization source.106,107 However, care must be taken as not to

dilute samples below analyte detection limits.40,53 For example, 1

: 2 and 1 : 4 dilutions of wastewater extract were effective for

reducing suppression for some drugs, but decreased method

sensitivity was noted as an important consideration.40 Similar

cautions apply to reduced injection volumes and/or splitting of

post-column flow. Although a post-column split (1 : 5) did reduce

matrix effects for the analysis of nine basic drugs, it was not

sufficient to allow for accurate quantification.57

One way to reduce ion suppression is to use UPLC, as co-

elutions of analytes with matrix materials would be reduced

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given narrowed elution bands, better resolution, and increased

peak height compared to conventional LC. As a result, fewer

matrix effects are possible.42 However, some matrix effects were

still noted for many analytes in multiresidue methods for phar-

maceuticals in waters,51,52,56 as co-elution cannot be completely

eliminated. It should be noted, that matrix effects were indeed

reduced when UPLC was applied to the analysis of biofluids, in

combination with improved sample preparation techniques, such

as polymeric mixed-mode SPE.68 Use of UPLC for environ-

mental analysis of pharmaceuticals51–53,56,59,67,99,108–110 may reduce

the incidence of co-elutions, but additional techniques will likely

be needed to compensate fully for interferences.

The use of alternate ionization techniques can also compensate

for matrix effects. Atmospheric pressure chemical ionization

(APCI) has been used less frequently for the analysis of drugs in

water extracts,71,111,112 given generally lower sensitivity compared

to ESI. Although APCI has been reported to be less susceptible

to matrix effects for some analytes,71 signal enhancement has

been noted for some drugs in wastewater extracts.112 Atmo-

spheric photoionization is less susceptible to matrix effects than

ESI. This technique has been applied to the analysis of sulfon-

amides in food products113 and is a promising tool for drug

discovery.114 However, it has so far not been evaluated for

pharmaceutical analysis in waters.

Overall, standard addition and isotope dilution appear to be

the most effective techniques for dealing with matrix effects.

Standard addition quantifies analytes in spite of interferences, by

creating a matrix-influenced calibration curve. Its major disad-

vantage is the significant increase in workload and instrumental

analysis time. However, it is the only reliable way to quantify

analytes in a complex matrix if isotope-labeled standards are not

available. Standard addition was applied to quantification of

antibiotics in wastewater treatment plant effluent.81 It was also

chosen for quantitative analysis of nine basic drugs in wastewater

and surface water associated with a pharmaceutical company,

after several unsuccessful attempts to minimize or eliminate

matrix effects through changes to sample preparation.57,115 Four-

point standard addition curves were used for quantification of 13

drugs from various classes in river water, because the matrix

resulted in variable suppression between 8 and 80%.52 To

compensate for matrix effects on antibiotics and diclofenac in

wastewater samples, standard addition was used for all analytes

except sulfamethoxazole for which an isotope-labeled standard

was available.41

A simpler way to compensate for ionization suppression/

enhancement is the addition of isotope-labeled analogues of each

analyte to samples. The isotope dilution method is common in

environmental analysis, and has been increasingly applied to

environmental pharmaceutical analysis as more isotope-labeled

standards become available. The matrix components affect the

isotope-labeled analog in an identical manner as the unlabeled

analyte, and thus correct for signal modification across different

matrices to allow for direct cross-comparison of results. The

isotope dilution method was successfully applied to negate

matrix effects in five different waters to quantify 15 pharma-

ceuticals along with 4 metabolites, 3 EDCs and one personal care

product.47 Deuterated internal standards were employed to

quantify 17 drugs of abuse and two metabolites in a wastewater

matrix with 47 to 94% signal reduction.80 Surrogate and internal

930 | J. Environ. Monit., 2009, 11, 923–936

standards of 13C and 15N labeled analytes accounted adequately

for matrix effects for psychoactive drugs in wastewater, rivers,

and creeks.55 For quantification of 38 pharmaceutically active

compounds, 10 EDCs and 3 perfluorinated acids, both deuter-

ated and 13C-labeled standards added prior to extraction could

account for recovery and matrix effects in grab samples of

wastewater, surface water and drinking water.48 A multi-class

method for 29 pharmaceuticals in surface and wastewater

involved use of structurally similar labeled internal standards for

quantification after evaluating other options to deal with matrix

effects.40 Because isotope-labeled analogues were not available

for every analyte, this approach involved two labeled internal

standards per ionization mode, and was effective in compen-

sating for suppression of most analytes.40 Isotope dilution was

also successfully applied to the analysis of several controlled and

non-controlled stimulants and some of their metabolites in

surface and wastewater.59 Matrix-induced signal enhancement

with APCI was combated by the use of isotope labeled standards

for six neutral pharmaceuticals in wastewater.112 For the isotope

dilution method, drawbacks include the cost of labeled stan-

dards, and limitations on quantitation for analytes without

available labeled analogues.

Mass analyzers

Mass analyzers are the detector of choice to identify and quantify

drugs in the aquatic environment, given the needs for sensitivity

and selectivity in samples with complex matrices. The mass

analyzer typically interfaced to LC for these analyses is the triple

quadrupole (QqQ), chosen for its sensitivity, availability, and

relatively low cost.9,10,54,60,88,101 Numerous methods for measuring

drugs in the aquatic environment exist using QqQ, for both

multi-class to target a wide range of drugs, including those of

known ecological importance, as well as other analytes of

interest10,40,41,47,51–53,55,59,83,99,102 and single class to target drugs

with similar structures or modes of action or both.5,42,50,57,80,81 In

a typical QqQ analysis, precursor ions generated during source

ionization of the analyte are selected in the first quadrupole for

collision induced dissociation in the second quadrupole,

producing product ions selected in the third quadrupole. While

QqQ analysis can provide accurate and sensitive quantification

of known targets, its power is limited for identification, confir-

mation and screening given low mass resolution and limited

sensitivity in full scan mode. The European Union Commission

Decision 2002/657/EC63 recommended four identification points

for positive confirmation of a target drug in a complex matrix.116

Using QqQ, this is accomplished by comparing retention time,

two multiple reaction monitoring (MRM) precursor-to-product

ion transitions, and the ratios between the two MRMs, between

samples and standards. Fragmentation details and MRM tran-

sitions for environmental drugs have been detailed elsewhere.54

Other mass analyzers, such as time of flight (ToF) or ion trap

(IT) are increasing in popularity for analysis of drugs in the

aquatic environment as complementary or alternative techniques

to QqQ-based analysis.54 Advantages of ToF are increased

selectivity and reduced false positives due to improved mass

resolution compared to QqQ. However, the ToF has a smaller

linear dynamic range than QqQ88,102 and less sensitivity. A ToF

has been used for confirmation, through retrospective analysis,

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of target and non-target drugs and other contaminants in

wastewater extracts, although co-eluted interferences limited the

accuracy in identifying low concentration non-target analytes.58

With an IT, continued fragmentation of ions, or MSn, provides

reliable confirmation.60 For antibiotics in surface and ground-

water, full scan IT experiments provided LOD between 0.03 and

0.19 mg L�1;45 similar LOD (0.03 to 0.07 mg L�1) were achieved

with an IT for antibiotics in surface water44 and wastewater.38

Hybrid mass analyzers, such as quadrupole-time of flight

(QToF) and quadrupole-linear ion trap (QLIT), are also

becoming important.54 The hybrid QToF can provide accurate

mass of precursor ions and high mass accuracy in full-scan

product ion detection.60,88 In a comparison between QToF and

QqQ for determination of NSAIDs in surface water, the selec-

tivity of the QToF was superior because of its higher resolving

power and therefore higher accuracy in product ion selection.117

That was one of the first reports of QToF use for detection of

drugs in the aquatic environment. With fewer matrix interfer-

ences in full scan, the QToF had a greater signal-to-noise ratio

and better certainty in identification of drugs at trace levels

(Fig. 3).

A limitation of QToF is its lower sensitivity in general

compared to QqQ. An early comparison found that for the same

method, a QqQ had limits of quantification <1.2 ng L�1, while

a QToF had worse limits (<3 ng L�1) with similar precision and

linear dynamic range for both.117 However, the method detection

limit for a QToF-based method for 29 pharmaceuticals in

river and wastewater was an order of magnitude higher than the

Fig. 3 Enhanced selectivity of UPLC-TOF analysis corresponding to

reconstructed ion chromatogram of carbamazepine (m/z 237.103) in an

urban wastewater sample with varying mass windows. Reproduced with

permission from ref. 56. Copyright ª 2006 Elsevier.

This journal is ª The Royal Society of Chemistry 2009

QqQ-based analogue, and was thus insufficient to detect analytes

at low ng L�1 levels in river water.56,108 Even with sample volumes

>1 L, high LOD were encountered with the QToF method,39

which was used for confirmation while QqQ was used for

quantification. However, current QToF models offer sensitivity

to drugs in waters approaching that of QqQ,109 albeit at a

significant price premium.

A more recent application of QToF screened for non-target

drugs in surface and wastewaters.118 Using a 500 compound

homemade library and deconvolution software, the authors used

accurate mass spectra to determine elemental composition and

identify paracetamol, caffeine, ofloxacin, and ciprofloxacin in

wastewater. The difficulty associated with identifying non-target

analytes not in a library was highlighted with a caution that such

work may be futile unless the compounds are environmentally

relevant. The authors suggested ensuring relevance by including

as many compounds of environmental concern as possible in the

library, particularly to avoid false negatives.118

Although not frequently employed, the QLIT allows for both

quantification and confirmation of target analytes, in using the

third quadrupole as a linear ion trap. Use of the information-

dependent analysis (IDA) function of a QLIT can give structural

information from fragmentation experiments, which are trig-

gered when criteria suggesting the presence of the analyte in

question are met during analysis. This has been demonstrated in

surface water with agricultural and wastewater inputs of 27

antibiotics and neutral drugs, as the IDA function of a 2000

QTrap was employed to confirm target analytes and identify

unknowns.46 A QLIT was employed with enhanced product ion

scanning and IDA for accurate quantification and unequivocal

identification of 56 pollutants in wastewater, including 38 phar-

maceuticals and 10 of their metabolites.58 The enhanced product

ion method provided accurate structural information for analy-

tes without strong secondary MRM transitions (e.g., ibuprofen)

by allowing for full scans of product ions trapped in the third

quadrupole.58

A newly developed mass analyzer, the hybrid linear ion trap

Orbitrap from Thermo (LTQ FT Orbitrap), has recently been

used for accurate mass screen and identification of drugs in the

aquatic environment.119 The LTQ portion of the system provides

MS/MS capability, while Orbitrap radially traps ions about

a central spindle electrode to determine accurate mass over

a large dynamic range by relating the frequency of rotation to

mass via Fourier transforms. This system offered good sensitivity

in full scan while providing high mass resolution of precursor and

product ions, advantages over QqQ. Its higher ion transmission

and higher mass range provides improved LOD and better

accurate mass data over QToF. This system has been used to

identify several pharmaceuticals in wastewater with linear range

between 0.05 and 1 mg L�1.119

Novel analytes

Chiral drugs

Some drugs are produced and ingested as a racemic mixture of

enantiomers, or non-superimposable mirror image molecules.

However, biological processes in the human body, WWTPs and

the environment may change the enantiomer composition,

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Fig. 4 Enantiomer fractions (EF) for the chiral drugs propranolol and

fluoxetine (chiral centres indicated with asterisk) in racemic standards as

well as raw influent and treated effluent from an urban WWTP. Values of

EF for a drug indicate the proportion of its (+) enantiomer relative to the

sum of both its (+) and (�) enantiomers. Reproduced with permission

from ref. 104. Copyright ª 2007 Elsevier.

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resulting in ecological exposure to non-racemic residues. Drug

enantiomers can have widely varying effects and toxicity, ranging

from innocuous (e.g., R-(�)-ibuprofen is inactive3) to deadly

(e.g., birth defects from thalidomide). For non-target organisms,

recent reports showed that S-(�)-propranolol had greater

chronic toxicity than R-(+)-propranolol to fathead minnows

(Pimephales promelas),120 while S-(+)-fluoxetine had a greater

effect than R-(�)-fluoxetine on sublethal behavioral endpoints,

particularly feeding rate, for this species.121 Thus, enantiomer-

specific analysis is crucial for accurate exposure and risk assess-

ment for chiral drugs. While the relative amounts of enantiomers

in human urine and faeces is generally well understood, there is

little attention to characterizing individual enantiomers of chiral

drugs in environmental waters.122

The few studies that have investigated the environmental fate

of drug enantiomers have relied initially on enantioselective GC/

MS analysis after derivatization, to quantify ibuprofen,3

propranolol,6 and metoprolol123 enantiomers in wastewaters and

surface waters. Recently, enantioselective LC/MS/MS methods

have been developed104,124 for measurement without derivatiza-

tion of individual enantiomers of eight chiral drugs, including

propranolol and fluoxetine, in wastewater with LODs104 between

0.2 and 7.5 ng L�1 comparable to conventional LC/MS/MS

methods. Enantiomer analysis in the complex matrices of

wastewater and natural waters present significant challenges. As

with conventional LC/MS/MS analysis of pharmaceuticals in

environmental waters, analysts must ensure that co-eluted matrix

components do not differentially affect each enantiomer’s signal

and skew results.104 As with conventional LC/MS/MS analysis,

corrections can be performed through standard addition and/or

the use of isotope-labeled analyte analogues as internal stan-

dards.104

Stereoisomer-specific analysis has been used to gain further

insight into the aquatic fate of chiral drugs that would otherwise

remain hidden.122 Abiotic processes in the environment generally

affect both enantiomers identically, but biologically-mediated

processes, which may involve chiral molecules such as enzymes,

could be more important for one enantiomer than the other. In

the first report of individual drug enantiomer measurement in

natural waters,3 untreated wastewater was more enriched in S-

(+)-ibuprofen from biotransformation by aquatic microbes.

Surface water had more R-(�)-ibuprofen than S-(+)-ibuprofen,

but it was unclear whether this was a result of microbially-

mediated enantiomerization, a common conversion for

ibuprofen and similar compounds. Racemic metoprolol was

found in wastewater, but water further downstream from

wastewater inputs contained nonracemic metoprolol, indicating

that biotransformation occurred as metoprolol moved down-

stream.123 These results suggest that chiral pharmaceuticals could

be tracers of untreated wastewater inputs to natural waters,6

a suggestion consistent with the significant differences in enan-

tiomer compositions observed between untreated and treated

wastewater from a biologically-mediated and enantiomer-

specific process during treatment.104 However, the same plug of

water was not followed through the plant, so conclusions

regarding changes during treatment should be carefully drawn.104

In addition, untreated wastewater also had nonracemic propor-

tions of many pharmaceuticals investigated104 (Fig. 4). This

observation confounds the use of enantiomer signatures as raw

932 | J. Environ. Monit., 2009, 11, 923–936

wastewater tracers, as input sources must have different enan-

tiomer compositions than treated and receiving waters for this

approach to be successful. Treated wastewater was enriched in

S-(�)-propranolol, the more acutely toxic enantiomer to

P. promelas, while treated wastewater was enriched in

R-(�)-fluoxetine over its more toxic antipode.104 As many drugs

are chiral, additional data regarding enantiomer-specific chronic

effects of drugs on aquatic organisms are clearly needed, and as

such data appears, methods should also emerge to measure

individual enantiomers accurately.

Metabolites and transformation products

Although most current literature on drugs in the aquatic envi-

ronment is focused on occurrence, fate and toxicity of drugs, the

potential importance of their degradates, by human and micro-

bial transformation, is also important. Metabolites are the

dominant form in which some drugs are excreted from humans,

while un-metabolized parent compounds may undergo biological

transformations during wastewater treatment and in the open

environment.125,126 In batch reactors, bacteria mediated rapid

ester cleavage of aceclofenac to form diclofenac, a potential

source of the latter in biological wastewater treatment67 (Fig. 5).

Metabolites may also be present in significant amounts in waters.

For example, the 10,11-dihydroxy-10,11-dihydro metabolite of

carbamazepine has been detected at concentrations three times

higher than carbamazepine itself in treated wastewater.127

Metabolites may even be a source of the original drug, as with the

case of glucuronide-conjugated metabolites, which can be enzy-

matically deconjugated during wastewater treatment to release

the parent compound,128 but have generally not been studied to

date in environmental waters other than for estrogen conju-

gates.129 Metabolites may also be toxic on their own, but little is

known about their environmental fate and toxicity. Recent

This journal is ª The Royal Society of Chemistry 2009

Page 11: JEM Spotlight: Recent advances in analysis of pharmaceuticals in the aquatic environment

Fig. 5 Concentration-time profile for biodegradation of aceclofenac

(ACF) in an activated sludge batch reactor spiked at 10 mg L�1.

Biodegradation experiments in batch reactors loaded with mixed liquor

demonstrated that ACF underwent rapid ester cleavage to liberate

diclofenac (DCF) and is thus a potential source of diclofenac in biolog-

ically treated sewage. Reproduced with permission from ref. 67. Copy-

right ª 2008 American Chemical Society.

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reviews have compiled advances in analysis for metabolites126

and biotransformation products125 of drugs, particularly anti-

microbials.128

As with parent compounds, accurate metabolite quantification

and identification is best accomplished if standards are

commercially available for comparison. In such cases, trans-

formation products have been included in some targeted analysis

in wastewater, surface water, and drinking water for degradates

of caffeine,58,59,110 carbamazepine,55,127 cocaine55,59,110,130,131 dil-

tiazem,51 dypirone,58 ibuprofen,51 nicotine,59,82,110 lysergic acid

diethylamide80 and verapamil,51 along with a host of metabolites

of cannabinoids,80,99,131 opiates,80,99,131 SSRIs,51,102 and tricyclic

antidepressants.51,102 Benzoylecgonine, a cocaine metabolite, was

detected in screening analysis of wastewater using UPLC/QToF/

MS/MS and comparison to an in-house library created from

available standards of environmental contaminants.118 When

standards are not commercially available, they may be synthe-

sized, as with the hydroxylated metabolites of diclofenac and

aceclofenac produced using cytochrome P-450 2C9-mediated

oxidation of the parent compounds.67 These metabolites were

removed from wastewater by membrane bioreactor treatment

with 56% and 96% efficiency, respectively.67 Alternatively,

metabolites can be isolated. For example, hydroxylated diclofe-

nac was collected by preparative LC from urine of a human

volunteer who took the drug,132 and was used to quantify these

metabolites at <mg L�1 concentrations in WWTP effluent.

However, limited work has been done to date using some func-

tionalities of hybrid tandem MS to metabolite identification. For

example, linear ion traps have been used for MSn experiments to

determine drug metabolites in human plasma.133 Using a QToF,

exact mass neutral loss tandem MS/MS has been used to identify

glutathione conjugates of drugs degraded by liver microsomes, to

distinguish from false negatives from neutral loss using low

resolution QqQ instruments.134

This journal is ª The Royal Society of Chemistry 2009

Aside from the paucity of available standards, another diffi-

culty in analyzing drug metabolites is the fact that these are

highly polar and poorly retrained on the reversed-phase LC

columns typically used for analytical separation of pharmaceu-

ticals. As a result, they are not easily resolved from one another,

or from co-eluting polar matrix components which would

interfere with their ionization and detection. In addition, the use

of more polar mobile phase eluents to elute polar metabolites

during chromatography could adversely impact ionization effi-

ciency.129 While normal phase chromatography with a nonpolar

mobile phase and polar stationary phase can be used for sepa-

rations of polar analytes, analysis of both parent compounds by

reversed-phase LC and metabolites by normal-phase chroma-

tography is difficult and laborious. Finding compatibility of

normal phase eluents and with ESI is also nontrivial. Use of

HILIC allows for greater interaction of polar analytes to the

stationary phase, and the use of reversed-phase mobile phases to

elute and ionize highly polar analytes efficiently. Cocaine and its

metabolites in wastewater and surface water were successfully

analyzed using HILIC and LC/MS/MS.130 A dual column-

switching apparatus was developed to shunt estrogens in river

water extracts into a reversed-phase LC column, and estrogen

conjugates from the same sample into a HILIC column.129 The

two fractions were delivered to the MS/MS at different times in

the same chromatographic run. These column-switching and T-

switching techniques135 increase throughput and efficiency.

Controlled experiments have been carried out to identify

potential biotransformation products of drugs using hybrid mass

spectrometers for structure elucidation. Microbial degradation

products of the antibiotic drug trimethoprim in nitrifying acti-

vated sludge were investigated using both IT and QToF.136

Hydrogen/deuterium (H/D) exchange experiments were carried

out with the IT to locate acidic protons and thus identify

degradates.136 In a similar study, three aerobic microbial

metabolites of diclofenac were identified using UPLC/ESI/

QToF/MS/MS with H/D exchange, which helped identify the

portion of the diclofenac molecule used to produce each

metabolite.136 Microbial degradation products of cyclofosfamide

and ifosfamide were not found in other activated sludge experi-

ments.137 Microbial aerobic biodegradation of verapamil was

investigated using a test that mimicked surface water conditions,

and a test for inherent biodegradability at high bacterial

density.138 In this case, standards of potential degradates were

available for confirmation of microbial metabolite identity with

IT MSn experiments.138

Other studies have focused on drug degradates formed by

abiotic processes such as photolysis.125,139 Photodegradation can

play a significant role in the fate of drugs in surface waters, and

can be used to remove drugs from wastewater.139 Non-photo-

chemical abiotic degradation of cyclophosphamide and ifosfa-

mide was shown experimentally, by determination of photo- and

chemical degradates, to be a potentially important removal

process with a half-life of years, but other abiotic processes such

as photolysis would only be important in shallow, clear nitrate-

rich water.137 This work highlighted the lack of available stan-

dards and occurrence data, and the lack of studies presenting

possible degradates for some drugs. Simulated sunlight was used

to assess photodegradation of enalapril and its metabolite ena-

laprilat, while both QLIT and QToF were used to elucidate

J. Environ. Monit., 2009, 11, 923–936 | 933

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possible structures of breakdown products.125 Photodegradation

of ofloxacin and ciprofloxacin by solar radiation with a TiO2

catalyst was investigated using a LTQ Orbitrap to identify

intermediates.140 That study also tested the toxicity of interme-

diates to assess whether successful treatment would result in less

toxic products.140

Directions for future innovations

In this review, we have discussed recent advances in the collec-

tion, processing, and analysis of pharmaceuticals and personal

care products in natural and engineered waters by LC/MS/MS.

While our understanding of occurrence, fate, and effects of drugs

in waters is far more complete than it was only a few years ago,

there clearly remain significant data gaps. As we have discussed,

the spatial and temporal distribution and loading of drugs to

receiving waters is not well characterized to date. Monitoring

programs could potentially benefit greatly from integration with

geographic information systems.141 There remain considerable

challenges in using LC/MS/MS, the instrumental method of

choice for measuring trace polar analytes in complex matrices, in

terms of accounting for processing and instrumental artifacts.

And the characterization of the impact of chiral drugs and of

metabolites and degradates is currently in its infancy.

In addition, it is also important to link occurrence and fate

studies on drugs in waters with toxicity assessment. While both

lines of research have been followed, little currently exists to tie

both together, as measurements are of limited use if they are not

environmentally relevant. Such work would include evaluation

of toxicity identification for an integrated approach to assessing

the impact of environmental pharmaceuticals,101 and inter-

laboratory studies to quantify toxicity.109 Attention should be

paid to environmental media other than water, such as fate and

effects of pharmaceuticals in biosolids from wastewater treat-

ment and applied to fields as fertilizer.142 Addressing these issues

would provide sound data by which to address risks posed by

pharmaceuticals in the aquatic environment.

Acknowledgements

Funding was provided by the Natural Sciences and Engineering

Research Council of Canada (Discovery Grant) and the Canada

Research Chairs program to Charles Wong. Sherri MacLeod

was also funded by the former in the form of a postgraduate

fellowship, as well as an Alberta Ingenuity Fund Studentship and

an American Chemical Society Division of Analytical Chemistry

2008 summer fellowship, sponsored by DuPont.

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