CRITICAL REVIEW www.rsc.org/jem | Journal of Environmental Monitoring
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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
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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
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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.
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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.
This journal is ª The Royal Society of Chemistry 2009
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,
J. Environ. Monit., 2009, 11, 923–936 | 931
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
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
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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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