Waterborne and sediment toxicity of fluoxetine
to select organisms
Bryan W. Brooksa,b,*, Philip K. Turnera, Jacob K. Stanleya, James J. Westonc,
Elizabeth A. Glidewella,b, Christy M. Forand, Marc Slatteryc,
Thomas W. La Pointa, Duane B. Huggettc
aDepartment of Biological Sciences, Institute of Applied Sciences, University of North Texas, P.O. Box 310559, Denton, TX 76203, USA
bDepartment of Environmental Studies, Baylor University, P.O. Box 97266, Waco, TX 76798-7266, USA
cSchool of Pharmacy, Environmental Toxicology Research Program, University of Mississippi, University, MS 38677, USA
dDepartment of Biology, 316 Brooks Hall, West Virginia University, Morgantown, WV 26506-6057, USA
Received 25 July 2002; received in revised form 3 January 2003; accepted 21 January 2003
Ecological risk assessments of pharmaceuticals are currently difficult because little-to-no aquatic hazard and ex-
posure information exists in the peer-reviewed literature for most therapeutics. Recently several studies have identified
fluoxetine, a widely prescribed antidepressant, in municipal effluents. To evaluate the potential aquatic toxicity of
fluoxetine, single species laboratory toxicity tests were performed to assess hazard to aquatic biota. Average LC50values
for Ceriodaphnia dubia, Daphnia magna, and Pimephales promelas were 0.756 (234 lg/l), 2.65 (820 lg/l), and 2.28 lM
(705 lg/l), respectively. Pseudokirchneriella subcapitata growth and C. dubia fecundity were decreased by 0.044 (14 lg/l)
and 0.72 lM (223 lg/l) fluoxetine treatments, respectively. Oryias latipes survival was not affected by fluoxteine ex-
posure up to a concentration of 28.9 lM (8.9 mg/l). An LC50of 15.2 mg/kg was estimated for Chironomus tentans.
Hyalella azteca survival was not affected up to 43 mg/kg fluoxetine sediment exposure. Growth lowest observed effect
concentrations for C. tentans and H. azteca were 1.3 and 5.6 mg/kg, respectively. Our findings indicate that lowest
measured fluoxetine effect levels are an order of magnitude higher than highest reported municipal effluent concen-
? 2003 Elsevier Science Ltd. All rights reserved.
Keywords: Environmental pharmaceuticals; Fluoxetine; Ecological risk assessment; Sediment toxicity; SSRI
A goal of the US Clean Water Act is to protect
aquatic life from deleterious anthropogenic activities.
For example, Section 303 provides statutory support
for water quality standards and the National Pollutant
Discharge Elimination System (NPDES) regulates point
source pollutant discharges to surface waters (Grothe
et al., 1995). Prior to developing water quality criteria,
minimum toxicity data are necessary for pollutants
(USEPA, 1985). Whereas criteria exist for many con-
taminants and NPDES permits regulate discharges,
recent research indicates that multiple classes of phar-
maceutical chemicals are present in municipal waste-
water effluents (Ternes, 1998; Kolpin et al., 2002;
Huggett et al., 2003) and in sediments downstream from
Chemosphere 52 (2003) 135–142
*Corresponding author. Present address: Department of
Environmental Studies, Baylor University, P.O. Box 97266,
Waco, TX 76798-7266, USA. Tel.: +1-254-710-6553; fax: +1-
E-mail address: firstname.lastname@example.org (B.W. Brooks).
0045-6535/03/$ - see front matter ? 2003 Elsevier Science Ltd. All rights reserved.
municipal discharges (Furlong etal., 2002). Currently,
US EPA water quality criteria do not exist for steroid
or non-steroid pharmaceuticals.
Webb (2001) recently identified several pharmaceu-
ticals that have the potential to adversely affect aquatic
ecosystems. Included among these compounds was flu-
oxetine, a selective serotonin reuptake inhibitor (SSRI).
One of the most prescribed antidepressants (RxList,
2000), fluoxetine blocks serotonin reuptake transport-
ers (Ranganathan et al., 2001). Fluoxetine and other
SSRIs also act at a number of sites including norepi-
nephrine uptake and Sigma receptors (Sanchez and
Meier, 1997) and neuronal and muscle nicotinic acety-
lcholine receptors (Garcia-Colunga et al., 1997). Many
SSRIs including fluoxetine bind with high affinity to
other receptors like Sigma receptors, with unknown
consequences. Fluoxetine exposure has also been shown
to increase extracellular norepinephrine and dopamine
in rats (Bymaster et al., 2002) and to potentially influ-
ence neuroendocrine function in Japanese medaka
(Brooks et al., 2003). Fluoxetine, a racemic mixture of
two lipophilic enantiomers, is metabolized by cyto-
chrome P450 isoenzymes to norfluoxetine, its active
metabolite. Fluoxetine is primarily excreted in the urine
with less than 10% as unchanged parent compound
(Hiemke and H€ a artter, 2000).
Whereas municipal effluents laden with steroids
and other estrogenic compounds have received con-
siderable attention, few peer-reviewed environmental
hazard and exposure data are available for non-ster-
oidal pharmaceuticals (Huggett et al., 2002, 2003;
Brooks et al., 2003). Fewer studies have evaluated en-
vironmental effects of antidepressants (Fong, 2001).
Daphnia magna and Daphnia pulex immobilization
EC50 values for amitriptyline, a tricyclic antidepres-
sant, were reported at 4.15 and 3.73 nM, respectively
(Lilius et al., 1995), while Flaherty et al. (2001) found
nominal treatment of 116 nM fluoxetine to significantly
affect D. magna reproduction. Nominal fluoxetine
treatment also affected embryonic Physa elliptica rota-
tional behavior at 0.1–1 lM (Uhler et al., 2000) and
potentiated male mussel spawning at 50 nM (Fong
et al., 1998). However, responses of benthic organ-
isms to sediment fluoxetine exposure have not been
Recent findings indicate that fluoxetine is discharged
in municipal effluents (Weston et al., 2001). Therefore,
the primary objective of this study was to evaluate the
environmental hazard of fluoxetine to select benthic and
pelagic toxicity test organisms. This study provides
novel waterborne fluoxetine toxicity data for Pseudo-
kirchneriella subcapitata, Ceriodaphnia dubia, Daphnia
magna, Pimephales promelas and Oryias latipes, and
sediment toxicity information for Hyalella azteca and
2. Materials and methods
2.1. Aqueous toxicity test methods
P. subcapitata stock cultures were maintained in 250-
ml Erlenmeyer flasks containing 125 ml Gorham?s media
(ATCC, 1984). New stock cultures were inoculated from
existing cultures every 7–10 days. Algal tests followed
general procedures recommended by the USEPA (1989).
Briefly, fluoxetine was added to three replicate 125-ml
flasks at 0, 43.6, 87.3, and 174.4 nM. Twenty-five mil-
liliter sterile AAP media (ATCC, 1984) was inoculated
with 4 to 7-day-old stock culture algae such that flasks
contained approximately 1 ? 104cells/ml at test initia-
tion. Flasks were placed under cool white fluorescent
lighting (2:15 ? 103lmm?2) with a light cycle of 16 h
light, 8 h dark. During a 120 h exposure period, tem-
perature was maintained at 25 ? 1 ?C and flasks were
hand-swirled twice daily.
Algal growth was evaluated by enumeration and
turbidity measurements (USEPA, 1989). Cell counts
were made using a hemocytometer and a Nikon?com-
pound microscope. Turbidity was determined as ab-
sorbance with a Beckman DU 64 spectrophotometer
at 750 nm.
C. dubia and D. magna were mass cultured as previ-
ously described (Turner et al., 2001; Hemming et al.,
2002). Static C. dubia and D. magna 48-h acute toxicity
tests were performed with less than 24-h old neonates
following standard methods (USEPA, 1991). A 7-day
study was conducted to evaluate C. dubia fecundity re-
sponses to fluoxetine treatments at nominal levels of 0,
45.5, 91, 181, 362 and 724 nM (USEPA, 1989). This
toxicity test received daily static renewals of test solu-
tions that were stored in dark, 4 ?C environmental
chambers between renewals. Risley and Bopp (1990)
previously demonstrated the persistence of fluoxetine for
six weeks in water stored at temperatures up to 65 ?C or
exposed to ultraviolet light. Organisms were fed a 0.5 ml
algae-Cerophyll?suspension following daily renewals of
test solutions (Knight and Waller, 1992; Hemming et al.,
2002). All studies were performed at 25 ? 1 ?C with 16 h
light, 8 h dark cycles.
P. promelas eggs were acquired from breeding cul-
tures maintained at the Institute of Applied Sciences,
Denton, TX. Prior to toxicity testing, juvenile fish were
fed 24–48-h old Artemia nauplii twice daily. Two acute
48 h toxicity tests, hereafter tests 1 and 2, were con-
ducted with 11- and 14-day old fathead minnows, re-
spectively (USEPA, 1991). Organisms were not fed
B.W. Brooks et al. / Chemosphere 52 (2003) 135–142
during 48-h tests. A light cycle of 16 h light, 8 h dark was
maintained at 25 ? 1 ?C. Oryias latipes were cultured
according to methods of Yamamoto (1975). Ten juvenile
organisms each were placed in three replicate beakers at
nominal concentrations of 0, 1.8, 3.6, 7.2, 14.5, and 28.9
lM for a 48 h acute study. Temperature was maintained
at 25 ? 1 ?C with a 16 h light, 8 h dark cycle.
2.2. Sediment toxicity test methods
Sediment toxicity tests were performed using a Zum-
walt testing system (USEPA, 2000). Dechlorinated, ac-
tivated carbon treated tap water served as overlying
water in sediment tests (USEPA, 2000). Two volume
additions per day were performed on overlying water.
Reference sediments were obtained from pond meso-
cosms at the University of North Texas Water Research
Field Station. Physical sediment parameters were eval-
uated including organic carbon, particle characteriza-
tion, and percent water. Sediments were spiked with
fluoxetine according to methods of Suedel and Rodgers
(1996). Ten-day C. tentans and H. azteca, and 42-day
H. azteca toxicity tests followed standard methods
(USEPA, 2000). Following preliminary range finding
toxicity tests, 10-day C. tentans and H. azteca treatment
levels were selected at 0, 1.4, 2.8, 5.6, 11.2 and 22.4 mg/
kg, and 0, 5.4, 10.8, 21.6 and 43.2 mg/kg, respectively.
Treatment levels for the 42-day H. azteca reproduction
study were selected at 1.4, 2.8, 5.6, 11.2, and 22.4 mg/kg.
2.3. Analytical measures
Fluoxetine ([?]-N-methyl-c-[4- trifluoromethyl)phen-
oxy]benzenepropanamine) hydrochloride was obtained
from Sigma Chemical (St. Louis, MO). Waterborne flu-
oxetine treatment concentrations were verified according
to methods of Weston et al. (2001). Briefly, 100 ml sub-
samples of lowest, mid, and highest treatment level sol-
utions were adjusted to pH¼9, pre-filtered with 1.2 lm
GFF, and extracted with hexane and methanol condi-
tioned Empore?C18 solid phase extraction disks. Disks
were eluted with methanol, brought to dryness, resolub-
lized in 100 ll methanol, and analyzed by a Waters Al-
liance model2690 liquid chromatograph with Waters 996
photo diode array and Waters Micromass ZQ mass se-
lective detectors. Method detection limits were 0.2 ng/l.
Sediment fluoxetine concentrations were not verified.
C. dubia, D. magna, and P. promelas bioassays were
performed in reconstituted hard water (RHW) (APHA
et al., 1995). O. latipes tests were performed in a bal-
anced saline solution (BSS) (Yamamoto, 1975). For each
study, temperature, pH, dissolved oxygen, conductivity
or salinity, alkalinity and hardness measures followed
standard methods (APHA et al., 1995).
2.4. Statistical analyses
P. subcapitata EC50 levels were estimated using
nonlinear regression with the SAS system (Version
8.0 for Windows; Bruce and Versteeg, 1992). LC50val-
ues for C. dubia, D. magna, P. promelas and C. tentans
tests were estimated using Trimmed Spearman Karber
(Hamilton et al., 1977). P. subcapitata, H. azteca and C.
tentans growth, and C. dubia and H. azteca fecundity
responses were evaluated using a one-way ANOVA with
Dunnett?s test for multiple comparisons (Zar, 1984).
3. Results and discussion
3.1. Analytical measures
Reconstituted hard water quality parameters were
within guidelines recommended by USEPA (1991) and
are listed in Table 1. Percent fluoxetine recoveries
from aquatic treatment levels ranged from 95% to
113%; therefore, nominal concentrations are presented
in Figs. 1 and 2, Table 2, and throughout this manu-
script. Sediment total organic carbon was 22340 mg/kg;
percent moisture was 60%. Sediment grain size was
distributed as 41.2% sand, 39.2% silt and 19.6% clay.
3.2. Acute fluoxetine toxicity
O. latipes and H. azteca survival was unaffected
across treatment levels. An LC50of 15.2 mg/kg was es-
timated for C. tentans. Average LC50values for C. dubia,
D. magna, and P. promelas toxicity tests were 0.756,
2.65, and 2.28 lM, respectively. This average D. magna
LC50is similar to a nominal value of 2.72 lM previously
reported for a Daphnia spp. (FDA-CDER, 1996).
P. promelas LC50?s for tests 1 and 2 were 2.22 and 2.88
lM, respectively. Such a toxicity decrease from test 1 to
test 2 may be related to age of test organisms; test 1
organisms were 11 d old, whereas organisms were 14 d
old in test 2. Fluoxetine, similar to other xenobiotics, is
metabolized from the parent compound by cytochrome
P-450 isoenzymes to an active metabolite, norfluoxetine.
Water quality characteristics of aqueous toxicity tests
Dissolved oxygen (mg/l)
Hardness (mg/l CaCO3)
Alkalinity (as mg/l
B.W. Brooks et al. / Chemosphere 52 (2003) 135–142
For example, Hamm et al. (2001) found that bioacti-
vation of diazinon to its potent metabolite increased
with development in juvenile Medaka. Observed differ-
ences in fluoxetine toxicity to fathead minnows may
be associated with an age dependent developmental
increase in P-450 metabolism. In this study, older
P. promelas were less sensitive to fluoxetine treatment.
3.3. Sublethal fluoxetine toxicity
Lowest observed and no observed fluoxetine effect
levels on C. dubia reproduction were 360 and 180 nM,
respectively (Fig. 1a; P ¼ 0:05, F ¼ 5:093, df ¼ 40;4).
However, such an observed statistically significant de-
crease in fecundity may not be of ecological relevance
because the mean difference between 360 nM and 0
treatments was just 2.1 neonates. Each fluoxetine treat-
ment level significantly reduced C. tentans growth such
that a lowest observed effect level (LOEC) of 1.3 mg/kg
(P ¼ 0:05, F ¼ 10:632, df ¼ 15;2) was observed. H.
azteca growth was also significantly reduced by all
treatment levels such that a LOEC was determined at 5.6
mg/kg (P ¼ 0:05, F ¼ 6:566, df ¼ 15;4). Therefore, C.
tentans was more sensitive to fluoxetine than H. azteca.
A plausible explanation for greater C. tentans sensitivity
to fluoxetine is increased exposure via ingestion of sed-
iments. Because H. azteca is epibenthic, it is likely that
dietary exposure to fluoxetine is less than that C. tentans
would experience during sediment burrowing.
Following a 42-day study period, H. azteca fecundity
was not significantly affected by fluoxetine treatments
(Fig. 1b). However, all treatment levels appeared to
stimulate H. azteca reproduction (Fig. 1b). We also
0 1.42.8 5.611.2 22.4
H. azteca Fecundity (Young Female-1)
0 45 90 180 360 720
C. dubia Fecundity (Neonates Female-1)
Fig. 1. Effects of fluoxetine on: (a) C. dubia reproduction (?SE;
N ¼ 10; F ¼ 5:093; df ¼ 40;4; *: P < 0:05) and (b) H. azteca
reproduction (? SE; N ¼ 8, P ¼ 0:05).
Growth (absorbance cm-1103)
Growth (cells ml-1104)
0 43.687.3 174.4
Fig. 2. P. subcapitata growth responses, measured as turbidity
(?SD; N ¼ 3; F ¼ 101:7, df ¼ 8, 3; *: P < 0:05) and enumer-
ation (?SD; N ¼ 3; F ¼ 28:9, df ¼ 8, 3; P < 0:05), to fluoxetine
P. subcapitata growth (EC50) and C. dubia, D. magna, P.
promelas and C. tentans survival (LC50) responses to fluoxetine
0.126 lM (39 lg/l)
0.077 lM (24 lg/l)
0.756 lM (234 lg/l)a
2.65 lM (820 lg/l)a
2.28 lM (705 lg/l)a
aLC50values are averaged for duplicate experiments.
B.W. Brooks et al. / Chemosphere 52 (2003) 135–142
observed an increase in C. dubia fecundity at the 180 nM
treatment level (Fig. 1a). Flaherty et al. (2001) observed
a similar reproductive stimulation when D. magna were
nominally exposed for thirty days to 116 nM fluoxetine.
In fish, serotonin stimulates release of gonadotropin
in some species, which stimulates sex steroid synthesis
and controls the development of oogenesis, including
vitellogenesis (Arcand-Hoy and Benson, 2001). In in-
vertebrates, serotonin has been shown to stimulate ec-
dysteroids, ecdysone, and juvenile hormone which
control oogenesis and vitellogenesis (Nation, 2002).
Whereas serotonergic effects on ecdysteroids, ecdysone,
and juvenile hormone are less understood (LeBlanc et al.,
1999), observed fecundity stimulation may result from
increased synaptic serotonin levels. Such an increase in
C. dubia and H. azteca reproduction need not be asso-
ciated with maintenance of offspring fitness. Numerous
studies identified that low-level contaminant exposure
may result in higher fecundity; however, increased egg
and neonate production is often associated with reduced
egg and neonate body size (Bodar et al., 1988; Ebert,
1993). Future research is warranted to identify the po-
tential mechanism(s) of crustacean reproductive stimu-
lation by fluoxetine and potential changes in the timing
of invertebrate reproduction (Hoonkoop et al., 1999)
following fluoxetine exposure.
P. subcapitata growth responses to fluoxetine treat-
ments were evaluated by two methods: cell density and
turbidity, the later reported previously as a sensitive
measure of algal growth (USEPA, 1989). P. subcapitata
EC50 values were estimated at 126 nM by cell density
and 77 nM by turbidity (Table 2). These values bracket a
previously reported nominal value of 89 nM for an
unknown green alga (FDA-CDER, 1996). P. subcapitata
growth as turbidity was significantly reduced at 43.6 nM
(P ¼ 0:05, F ¼ 101:7, df ¼ 8;3) and at 174.4 nM (P ¼
0:05, F ¼28:9, df ¼8;3) as cell number (Fig. 2). Whereas
cell density was not significantly affected by fluoxetine
treatment at 87.3 and 174.4 nM, cell deformities were
observed and cell sizes were reduced at both treatment
levels. Cells appeared shriveled and frequently were not
crescent shaped, a normal characteristic of P. subcapi-
tata. The mechanism of toxicity for observed cell de-
formities is not known; however, fluoxetine is known to
have bacteriostatic effects, perhaps exerted by efflux
pump inhibition (Munoz-Bellido et al., 2000). Although
cell deformities and biovolumes were not quantified in
this study, such an effect of fluoxetine on algal cells
warrants further investigation.
3.4. Preliminary ecological risk characterization for
Presently, the US Food and Drug Administration
requires environmental assessments of new pharma-
ceuticals only if predicted environmental introduction
concentrations (EIC) are greater than 1 lg/l (3.32 nM
(1999) identified that this approach: does not consider
additive effects of therapeutics with similar mechanisms
of action, does not consider interactions of compounds
with different mechanisms of action, and relies on tra-
ditional ecotoxicology bioassay response variables. In
addition, a default dilution factor of 10 is applied to an
EIC in an attempt to approximate realistic environ-
mental exposures. Such an approach is generally ap-
propriate because 77% of permitted effluent dischargers
receive greater than 10 fold dilution at annual mean
flow (Dorn, 1996). This exercise becomes problematic in
areas where effluent discharges do not benefit from up-
stream dilution. For example, perennial municipal ef-
fluents often influence historically ephemeral streams in
the southwestern US; 90% of the flow of the Trinity
River south of Dallas/Fort Worth, TX is dominated by
reclaimed waters from treatment plants (Dickson et al.,
1989). Environmental pharmaceuticals in these regions
may present the greatest risk to aquatic organisms.
Ecological risk is often simplistically characterized by
evaluating the relationship between an ambient expo-
sure concentration (AEC) and a toxicologically effective
concentration (TEC; Suter et al., 2000). Conservative
safety factors are also applied to overcome the uncer-
tainties inherent in risk estimation procedures (Warren-
Hicks and Moore, 1998). If a hazard quotient (HQ)
derived from a ratio of AEC/TEC is >1, then risk to the
environment may become a concern. Weston et al.
(2001) found fluoxetine concentrations in municipal ef-
fluents to range from 0.3 to 1.6 nM. This range is con-
sistent with an ambient exposure concentration of 1.2
nM predicted for United Kingdom effluents (Webb,
2001). The lowest fluoxetine effect level reported in this
study was 43.6 nM for algal growth. Therefore, the
lowest observed adverse effect level of fluoxetine on
aquatic biota is an order of magnitude higher than the
highest reported environmental fluoxetine concentra-
tions. Based on data presented in this study an HQ
is calculated at ?1.
Previous investigators discussed reliance on aquatic
toxicity test responses for regulatory decisions (Cairns,
1983; Mount, 1994). Laboratory bioassays may not be
representative of the most sensitive species (Cairns,
1986) and are not intended to predict structural or
functional community responses to stressors (La Point
and Waller, 2000). Others suggested that longer term
studies with low concentrations are necessary to assess
pharmaceuticals effects on higher levels of biological
organization (Daughton and Ternes, 1999; Brooks et al.,
2003). Further, Daughton and Ternes (1999) identified
that bioassays focused on specific mechanisms of action
are necessary to assess therapeutic effects on non-target
biota. Identifying such mechanistic toxicity will be chal-
lenging because the responses of aquatic organisms to
B.W. Brooks et al. / Chemosphere 52 (2003) 135–142
fluoxetine are largely unknown. For example, the beta-
adrenergic receptor blocking pharmaceuticals propra-
nolol and metaprolol significantly reduce cladoceran
heart rate and metabolic activity at concentrations an
order of magnitude lower than those that affect repro-
duction (Dzialowski et al., 2002).
Sensitivities of organisms tested in this study to flu-
oxetine were P. subcapitata>C. dubia>P. promelas>D.
magna>C. tentans>H. azteca. Our results indicate that
fluoxetine adversely affects aquatic organisms at levels
at least an order of magnitude higher than that reported
in municipal effluent concentrations. When pesticide
exposure and effect levels are within an order of mag-
nitude, additional data are often necessary to adequately
characterize risk. Unlike pesticides, environmental phar-
maceutical concentrations that we measured are not
acutely toxic to aquatic organisms. Chronic responses of
non-target biota, many of which are unidentified, may
result from nanomolar or potentially even picomolar
exposure (Fong, 1998). Aquatic systems in regions
where municipal effluents do not benefit from upstream
dilution may present maximal risk to aquatic organisms.
This research was supported in part by a US Con-
gressional Environmental Sensors and Signals grant to
M. Slattery and C.M. Foran, a Texas Water Resources
Institute/United States Geological Survey grant to B.W.
Brooks and T.W. La Point, and the Environmental
Toxicology Research Program at the University of
Mississippi. The authors thank Erica March and Beth-
any Peterson for technical support.
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Bryan Brooks is an assistant professor in the Department of
Environmental Studies at Baylor University. He received a
Ph.D. in environmental science from the University of North
Texas in 2002 and a MS in Limnology from The University of
Mississippi in 1998. His research focuses on environmental
pharmaceuticals, chemical mixture ecotoxicology, aquatic food
webs, metal bioavailability, TIE and TMDL techniques, and
Philip Turner is a doctoral student, research assistant and
teaching fellow with the Institute of Applied Sciences, Univer-
sity of North Texas. His dissertation research focuses on epi-
sodic exposure of metals. He has also been involved with
projects concerning the presence of aquatic pollutants at the
watershed scale, hormone disruption in constructed wetlands,
and TMDL development for aquatic systems.
Jacob Stanley is a graduate student and research assistant at the
University of North Texas, Institute of Applied Sciences. He
completed a B.S. in biology at the University of Mississippi in
1999. His research focuses on metal bioavailability and the use
of stream mesocosms in ecotoxicological research. His other
research interests include pharmaceuticals in the aquatic envi-
ronment and TIEs and TMDLs.
Jim Weston is a research and development marine biologist for
the Environmental Toxicology Research Program at the Uni-
versity of Mississippi. Currently, he is evaluating the fate and
effects of prescription drugs found in aquatic environments in
the Mississippi area. His general interest is in environmental
stressors. His future plan is to pursue a doctoral degree in bio-
Beth Glidewell is currently a graduate student in the Depart-
ment of Environmental Studies at Baylor University. She
completed a B.S. degree in Biology at the University of North
Texas in 2002. Her research interests include pharmaceuticals in
the environment, wetlands, ecological restoration and aquatic
Christy Foran is an assistant professor in the Department of
Biology at West Virginia University. She held a position as a
research assistant professor at The University of Mississippi in
the Environmental Toxicology Research Program until 2001.
Her research focuses on physiological and developmental
responses to environmental signals, focusing in endocrine
Marc Slattery is an associate professor in the Department of
Pharmacognosy at the University of Mississippi. He is a
chemical ecologist with a joint position in the Environmental
Toxicology Research Program at UM. His research interests
focus on developmental responses and phenotypic plasticity to
environmental signals and stressors in aquatic systems.
Tom La Point is Director of the Institute of Applied Sciences,
University of North Texas. His research interests include con-
taminant effects on freshwater aquatic communities, specifically
how anthropogenic perturbations affect fisheries and benthic
population dynamics. He has published on ecosystem measures,
contaminant bioaccumulation, and sub-lethal effects on popu-
Duane Huggett is currently a Research Scientist with Pfizer Inc.
Prior to this position, he was a Post Doctoral Research Asso-
ciate at The University of Mississippi in the Environmental
Toxicology Research Program. He received his M.S. in Biology
and his Ph.D. in Pharmacology and Toxicology from The
University of Mississippi. His research interests include phar-
maceuticals in the environment, physiological toxicology and
ecological risk assessment.
B.W. Brooks et al. / Chemosphere 52 (2003) 135–142