Short communication

Aquatic ecotoxicology of fluoxetine

Bryan W. Brooks a,1,*, Christy M. Foran b, Sean M. Richards c,2,James Weston d, Philip K. Turner a, Jacob K. Stanley a, Keith R. Solomon c,

Marc Slattery d, Thomas W. La Point a

a Institute of Applied Sciences, University of North Texas, Denton, TX, USAb Department of Biology, West Virginia University, Morgantown, WV, USA

c Centre for Toxicology, University of Guelph, Guelph, Ont., Canadad Environmental Toxicology Research Program, School of Pharmacy, University of Mississippi, University, MS, USA

Received 16 September 2002; received in revised form 15 October 2002; accepted 12 November 2002

Abstract

Recent studies indicate that the pharmaceutical fluoxetine, a selective serotonin reuptake inhibitor, is discharged in

municipal wastewater treatment plant effluents to surface waters. Few data on environmental fluoxetine exposure and

hazard to aquatic life are currently available in the literature. Here, we summarize information on fluoxetine detection

in surface waters and review research on single-species toxicity test, Japanese medaka (Oryzias latipes ) reproduction

and endocrine function, and freshwater mesocosm community responses to fluoxetine exposure. Based on results from

our studies and calculations of expected introduction concentrations, we also provide a preliminary aquatic risk

characterization for fluoxetine. If standard toxicity test responses and a hazard quotient risk characterization approach

are solely considered, little risk of fluoxetine exposure may be expected to aquatic life. However, our findings indicate

that: (1) the magnitude, duration and frequency of fluoxetine exposure in aquatic systems requires further investigation;

(2) mechanistic toxicity of fluoxetine in non-target biota, including behavioral responses, are clearly not understood;

and (3) an assessment of environmentally relevant fluoxetine concentrations is needed to characterize ecological

community responses.

# 2003 Elsevier Science Ireland Ltd. All rights reserved.

Keywords: Fluoxetine; Serotonin reuptake inhibitor; Risk assessment; Environmental pharmaceuticals

1. Introduction

Kolpin et al. (2002) recently identified wide-

spread occurrence of multiple pharmaceuticals in

United States surface waters. Included among

these contaminants was fluoxetine (Fig. 1; Table

1), a selective serotonin reuptake inhibitor (SSRI).

SSRIs are primarily indicated for depression, but

are also prescribed to treat compulsive behavior,

* Corresponding author. Tel.: �/1-254-710-6553; fax: �/1-

254-710-3409.

E-mail address: bryan_brooks@baylor.edu (B.W. Brooks).1 Present Address: Department of Environmental Studies,

Baylor University, Waco, TX, USA.2 Present Address: Department of Biological and

Environmental Sciences, University of Tennessee at

Chattanooga, Chattanooga, TN, USA.

Toxicology Letters 142 (2003) 169�/183

www.elsevier.com/locate/toxlet

0378-4274/03/$ - see front matter # 2003 Elsevier Science Ireland Ltd. All rights reserved.

doi:10.1016/S0378-4274(03)00066-3

and eating and personality disorders. SSRIs are

preferred to monoamine oxidase inhibitors and

tricyclic antidepressants for treatment of affective

disorders due to a lack of receptor antagonism and

few anticholinergic and cardiovascular side effects

(Rang et al., 1995). The prototype SSRI and a

highly prescribed antidepressant (NDC Health,

1999), ProzacTM (fluoxetine HCl) blocks serotonin

reuptake from the pre-synaptic nerve cleft (Ran-

ganathan et al., 2001). A racemic mixture of two

lipophilic enantiomers, fluoxetine is metabolized

by cytochrome P-450 isoenzymes to norfluoxetine,

its active metabolite, and is primarily excreted as

less than 10% unchanged parent compound in

urine (Hiemke and Hartter, 2000).

Whereas the occurrence and estrogenicity of

steroid therapeutics in municipal effluents has

received attention (Harries et al., 1997; Nichols

et al., 1999; Hemming et al., 2001, 2002), environ-

mental hazard and exposure information is sparse

for non-steroid pharmaceuticals (Huggett et al.,

2002, 2003). Such limited environmental pharma-

ceutical data pertains mostly to detection (Stumpf

et al., 1996; Stan and Heberer, 1997; Buser and

Muller, 1998; Ternes, 1998; Buser et al., 1999;

Hirsch et al., 1999; Stumpf et al., 1999; Suter and

Fig. 1. Chemical structures of (a) fluoxetine and (b) norfluoxetine.

Table 1

Physiochemical and environmental fate parameters of fluoxetine and norfluoxetine

Parameter Fluoxetine Norfluoxetine

Physiochemical parameter

Empirical formula C17H18F3NO C16H16F3NO

Molecular weight 309.33 295.3

PKa 10.069/0.10 9.059/0.13

pH pH

Environmental fate parameter 2.0 7.0 11.0 2.0 7.0 11.0

log Kow 1.25 1.57 4.30 0.97 2.05 4.06

BCF �/1 2.00 1071.52 �/1 6.97 716.12

log Koc 0.64 0.97 3.70 0.49 1.57 3.58

Values calculated by ACD/Labs Software Version 5.0 (Toronto, Ontario, Canada).

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183170

Giger, 2000; Golet et al., 2001; Rossknecht et al.,2001; Kolpin et al. 2002; Huggett et al., 2002) and

degradation (Richardson and Bowron, 1985;

Guarino and Lech, 1986; Velagaleti and Robin-

son, 2000) in rivers and lakes. Of the few studies

that have examined pharmaceutical effects, single-

species, acute laboratory toxicity tests were pri-

marily used (Guarino and Lech, 1986; Lanzky and

Halling-Sorensen, 1997; Fong, 1998; Huber andDelago, 1998; Honkoop et al., 1999; Uhler et al.,

2000). Fewer studies have evaluated fish biochem-

ical and reproduction responses to non-steroid

therapeutics (Huggett et al., 2002). Pharmaceutical

effects on higher levels of biological organization

are not reported in the peer-reviewed literature.

Very little information is also available for

fluoxetine exposure (Weston et al., 2001; Kolpinet al., 2002) and effects (Fong, 2001; Brooks et al.,

2003a; Weston et al., 2003; Richards et al., 2003)

in aquatic ecosystems. Therefore, the objectives of

this paper are: (1) to summarize available data on

the occurrence and detection of fluoxetine in

surface waters; (2) to summarize our research

with aquatic biota and community responses to

fluoxetine; and (3) to provide a preliminaryfluoxetine aquatic risk characterization.

2. Fluoxetine exposure and detection in surface

waters

Environmental exposure to norfluoxetine has

not been reported, but several investigators de-

tected fluoxetine in waterbodies and municipaleffluents (Jones-Lepp et al., 2001; Weston et al.,

2001; Kolpin et al., 2002). In surface waters,

Kolpin et al. (2002) estimated maximum fluoxetine

concentrations at 0.012 mg/l. Weston et al. (2001)

indicated that effluent fluoxetine concentrations

may reach as high as 0.540 mg/l. However, the

magnitude, frequency and duration of fluoxetine

exposure have not been fully explored. Further,detection of fluoxetine in sediments has not been

reported. Although reliable detection of fluoxetine

in aqueous and sediment matrices is essential to

assessing environmental exposure, Weston et al.

(2001) identified that extraction and recovery of

fluoxetine and norfluoxetine in aqueous samples

requires further development. Here, we summarizeseveral reported methods for fluoxetine and nor-

fluoxetine detection in water matrices. Methods of

Kolpin et al. (2002), Jones-Lepp et al. (2001), and

Weston et al. (2001) follow a general scheme: pre-

filtration, extraction, concentration, and detection

and quantitation.

Kolpin et al. (2002) pre-filtered 1 l surface water

and extracted analytes from samples using solidphase extraction (SPE) cartridges. Cartridges were

eluted, evaporated to near dryness and then

brought to a final volume of 1 ml, resulting in a

1000:1 concentration ratio. High-performance li-

quid chromatography (HPLC) with a reverse

phase octylsilane (C8) column was used to detect

and quantify analytes. Jones-Lepp et al. (2001)

collected effluent samples from nine wastewatertreatment plants (WWTP); whether samples were

pre-filtered prior to solid phase extraction was not

indicated. Two liter samples were adjusted to a pH

of 2.5 and subsequently extracted using SPE C18

discs. Analytes adsorbed to discs were eluted and

concentrated to 0.3 ml, resulting in a 6667:1

concentration ratio. Liquid-chromatography elec-

tron spray ion trap mass spectrometry (LC-ES/ITMS) was used for detection and quantitation.

Weston et al. (2001) sampled municipal effluent

discharge from two WWTPs. One liter samples

were pre-filtered, adjusted to pH 9, and extracted

with C18 SPE discs. Weston et al. (2001) selected a

pH adjustment to 9.0 because the ionization state

and lipophilicity of fluoxetine changes with in-

creasing pH (Table 1). Adsorbed analytes wereeluted, evaporated to dryness, and reconstituted to

a final volume of 0.1 ml, resulting in a 100 000:1

concentration ratio. LC-ES/MS was utilized to

detect and quantitate effluent fluoxetine and nor-

fluoxetine levels.

Percent recovery of fluoxetine varied between

methods: less than 60% in Kolpin et al. (2002),

88% (triplicate extractions) in Jones-Lepp et al.(2001), and 79�/82% in Weston et al. (2001). In

addition, Weston et al. (2001) observed percent

recoveries of 67�/77% for norfluoxetine matrix

spikes, based on triplicate extractions. Kolpin et

al. (2002) reported that out of 84 streams sampled

fluoxetine concentrations did not exceed an esti-

mated 0.012 mg/l. Although Jones-Lepp et al.

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183 171

(2001) sampled nine WWTP effluents for fluox-etine, detection frequency and specific effluent

fluoxetine concentrations were not reported. Wes-

ton et al. (2001) reported fluoxetine levels up to

0.54 mg/l; however, norfluoxetine was not detected

in two WWTP effluents.

3. Single species toxicity test organism responses tofluoxetine

Standardized single species toxicity tests are

used to screen for potential hazards of aquatic

contaminants, to develop water quality criteria,

and to monitor whole effluent toxicity in the

United States. Environmental assessments of

pharmaceutical compounds also rely on singlespecies responses if an expected environmental

introduction concentration (EIC) exceeds 1 mg/l

(FDA-CDER, 1998). Such laboratory studies are

attractive because they use clean water or sedi-

ments and are less expensive to perform than field

studies. Further, laboratory responses are often

less variable than data collected in bioassessments

or mesocosm studies (Dickson et al., 1996).Whereas fluoxetine exposure may affect pelagic

organisms (Fong, 2001), fluoxetine may also bind

to sediments and affect benthic organisms; benthic

macroinvertebrate responses to sediment fluoxe-

tine exposures have not been reported. Therefore,

standardized aquatic and sediment toxicity tests

were performed to assess potential effects of

fluoxetine on freshwater biota (Brooks et al.,2003a). The green algae, Pseudokirchneriella sub-

capitata , two cladocerans, Ceriodaphnia dubia and

Daphnia magna , and the fathead minnow, Pime-

phales promelas were chosen for aquatic toxicity

tests. For sediment toxicity tests, the midge,

Chironomus tentans , and the amphipod, Hyalella

azteca , were utilized. Waterborne exposure con-

centrations were verified according to Weston etal. (2001).

3.1. Freshwater toxicity tests

P. subcapitata (formerly Selanastrum capricor-

nutum ) toxicity tests followed recommended pro-

cedures (USEPA, 1989, 1991). An EC50 for P.

subcapitata growth was estimated by nonlinearregression (Bruce and Versteeg, 1992) at 24 mg/l

(Table 2). This value is almost identical to a

previously reported EC50 of 28 mg/l for an un-

named green algae (FDA-CDER, 1996). P. sub-

capitata growth was also evaluated for treatment

level effects using ANOVA with a Dunnett’s post

hoc test. A lowest observed effect concentration

(LOEC) was observed at 13.6 mg/l (Table 2), whichwas also the lowest treatment level tested. P.

subcapitata cell deformities were observed at 13.6

and 27.2 mg/l treatment levels. Cells also appeared

smaller than untreated controls at these concen-

trations. The mechanism by which fluoxetine may

induce deformations in algal cells is unknown.

However, fluoxetine has antimicrobial properties

and potentially exerts its toxicity by inhibitingcellular efflux pumps (Munoz-Bellido et al., 2000).

Although cell deformities and biovolumes were

not quantified, a possible bacteriostatic mechan-

ism of algal and microbial fluoxetine toxicity

deserves future investigation.

C. dubia , D. magna and P. promelas 48-h acute

toxicity tests were performed in reconstituted hard

water (USEPA, 1991). Each test was repeated, andLC50s estimated by Trimmed Spearman Karber

(Hamilton et al., 1977). Average LC50s for C.

dubia , D. magna and P. promelas were 234, 820,

and 705 mg/l, respectively. A D. magna LC50 of 820

mg/l is similar to a nominal value of 940 mg/l

reported for a Daphnia spp. (FDA-CDER, 1996).

The LC50 of 705 mg/l for P. promelas is lower than

a previously reported 48-h LC50 of 2 mg/l forrainbow trout (Oncorhynchus mykiss ; FDA-

CDER, 1996).

In addition, a 7-day C. dubia static-renewal

study was performed to evaluate potential fluox-

etine effects on cladoceran reproduction. This test

also followed standard methods (USEPA, 1989);

however, organisms were fed an algae-Cerophyll†

suspension following daily renewals (Hemming etal., 2002). C. dubia no observed effect (NOEC)

and lowest observed effect concentrations were

determined at 56 and 112 mg/l, respectively, using

ANOVA with a Bonferroni’s adjustment.

Although a treatment level of 112 mg/l was

statistically different from control organisms

(a�/0.05), the observed difference may not be of

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183172

ecological relevance because the difference was

only a mean of 2.1 neonates per female.

3.2. Sediment toxicity tests

Ten-day C. tentans and H. azteca sediment

toxicity tests followed standard methods and

were performed using a Zumwalt testing system

(USEPA, 2000). Reference sediments were ob-

tained from pond mesocosms at the University of

North Texas Water Research Field Station. Sedi-

ments were characterized for total organic carbon

(22340 mg/kg), percent moisture (60%), and grainsize distribution (41.2% sand, 39.2% silt, 19.6%

clay). In addition to physical characterization,

sediments were evaluated for 17 metals, 44 volatile

organics, 56 semi-volatile organics, four triazine

herbicides, six organophosphorus insecticides,

three organochlorine herbicides, 20 organochlor-

ine pesticides, two carbamate pesticides, and seven

PCB congeners (La Point et al., unpublished data).Sediments were considered ‘clean’ and were spiked

with fluoxetine according to Suedel and Rodgers

(1996). Following preliminary range finding toxi-

city tests, 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.

C. tentans survival was reduced by fluoxetinetreatments; an LC50 of 15.2 mg/kg was estimated

(Table 2). In addition, each fluoxetine treatment

level significantly reduced C. tentans growth such

that a LOEC of 1.3 mg/kg was observed (Table 2).

H. azteca survival was not affected by the highest

treatment level tested (43 mg/kg; Table 2). How-

ever, H. azteca growth was significantly reduced

by all treatment levels (LOEC�/5.6 mg/kg; Table

2). In addition to 10-day tests with H. azteca , a 42

day study was performed to evaluate potentialfluoxetine effects on H. azteca reproduction

(USEPA, 2000). H. azteca fecundity (young per

female) was not significantly reduced by fluoxetine

treatment levels.

3.3. Fluoxetine effects on invertebrate reproduction

Fluoxetine treatments stimulated H. azteca

reproduction, though not significantly. An in-crease in C. dubia fecundity was also observed

with 56 mg/l fluoxetine treatment. Flaherty et al.

(2001) observed a comparable reproductive stimu-

lation when D. magna were exposed to 36 mg/l

fluoxetine for 30 days. Similarly, Fong et al. (1998)

observed fluoxetine to induce mussel spawning. In

invertebrates, serotonin may stimulate ecdyster-

oids, ecdysone, and juvenile hormone, which areresponsible for controlling oogenesis and vitello-

genesis (Nation, 2002). In some fish species (see

further discussion below) serotonin may stimulate

the release of gonadotropin. Gonadotropin stimu-

lates sex steroid synthesis and controls oogenesis

development, including vitellogenesis (Arcand-

Hoy and Benson, 2001). Although serotonergic

effects on ecdysteroids, ecdysone, and juvenilehormone are less understood (LeBlanc et al.,

1999), observed stimulation in fecundity may

result from increased synaptic serotonin levels.

However, because invertebrate reproduction is

energy intensive, such an increase in C. dubia or

H. azteca reproduction should not necessarily be

Table 2

Standardized toxicity test organism responses to fluoxetine

Organism EC50 NOEC LOEC Matrix

Pseudokirchneriella subcapitata 24 mg/l ND 13.6 mg/l AAP

Ceriodaphnia dubia 234 mg/l 56 mg/l 112 mg/l RHW

Daphnia magna 820 mg/l N/A N/A RHW

Pimephales promelas 705 mg/l N/A N/A RHW

Hyalella azteca �/43 mg/kg ND 5.4 mg/kg Sediment

Chironomus tentans 15.2 mg/kg ND 1.3 mg/kg Sediment

AAP�/AAP media; RHW�/reconstituted hard water; Sediment�/University of North Texas Water Research Field Station

reference sediment; ND�/not determined; N/A�/not available.

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183 173

associated with maintenance of offspring viabilityor fitness.

4. Medaka reproduction and endocrine function

responses to fluoxetine

Among the issues raised by environmental

detection of fluoxetine are concerns over potential

sub-lethal effects on aquatic organisms, includingbehavioral responses. Considering the potential

for environmental SSRIs to act as they do in

humans, to alter or increase serotonin concentra-

tions, the potential disruptive effects of chronic

exposure must be considered. Serotonin is likely to

be one of the most potent and ubiquitous neuro-

modulators in vertebrates (Azmitia, 1999). It is

synthesized in cells lining the gut, in neurons of thehypothalamus that regulates pituitary activity, and

in the brainstem of vertebrates. Many of these

neurons release serotonin into the synaptic cleft

where it acts as a neurotransmitter. In addition,

cerbrospinal-fluid contacting neurons in the hy-

pothalamus and cells in the periphery release

serotonin into general circulation where it acts

on more distant target tissues in the centralnervous system or vascular and gastrointestinal

muscle, T cells and platelets.

Because of the critical nature of the functions

regulated by serotonin, there is a potential for

environmental SSRIs to alter appetite, the immune

system, and reproduction as well as other beha-

vioral functions (Meguid et al., 2000; Mossner and

Lesch, 1998; Fong, 2001). Serotonin acts directlyon the immune system by modulating cellular

function and indirectly through actions on the

central nervous system (Mossner and Lesch, 1998).

In studies with lymphocytes from HIV-positive

patients, treatment with a serotonin receptor

agonist resulted in increased T cell counts (Hof-

mann et al., 1996) whereas serotonin itself was

found to increase their proliferative capacity (Eu-gen-Olsen et al., 1997). A similar relationship

between serotonin and immune function was also

described for fish (Khan and Deschaux, 1997). To

the extent that serotonin alters immune function,

an increase in serotonin may produce beneficial

changes in the immune response but also may

elevate the rate of negative impacts such asautoimmune disease.

Serotonin is an important neuromodulator of

sexual function in vertebrates and invertebrates.

Changes in serotonin metabolism or concentration

are correlated with reproductive phases of human

females (Hindberg and Naesh, 1992) and other

animals, including female fish (Hernandez-Rauda

et al., 1999). Fish studies indicated that serotoninpotentiates effects of gonadotropin-releasing hor-

mone on gonadotropin release from the pituitary

(Khan and Thomas, 1994). In some seasonally

reproductive animals, serotonin concentration var-

ies with reproductive potential and gonadal recru-

descence (Hernandez-Rauda et al., 1999).

Literature relating serotonin and SSRIs to repro-

ductive function has been recently reviewed formany groups of invertebrates and vertebrates,

including fish (Fong, 2001). The role of serotonin

in reproduction, and therefore the potential for

SSRIs to disrupt normal serotonin function, varies

across family groups. Serotonin and SSRIs po-

tentiate spawning and oocyte maturation in some

bivalves and crustaceans (Fong, 2001). Serotonin

also induces oocyte maturation in Japanese me-daka (Oryzias latipes ; Iwamatsu et al., 1993) but

inhibits this process in another teleost, the mum-

michog (Fundulus heteroclitus ; Cerda et al., 1998).

Vertebrate studies evaluating linkages between

reproductive and endocrine function changes and

aquatic SSRI exposure are limited.

To assess potential fish endocrine function and

reproduction responses to environmental SSRIs,Weston et al. (2003) exposed Japanese medaka for

4 weeks to fluoxetine treatments of 0, 0.1, 0.5, 1.0

and 5.0 mg/l. Japanese medaka was chosen because

this species is a widely used model organism for

the study of contaminant-induced developmental

effects (Metcalfe et al., 1999) and reproductive

impairment (Arcand-Hoy et al., 1998). Reproduc-

tion, including fecundity, rate of fertilization, egghatching success and abnormal development, and

endocrine function, including vitellogenin and

circulating plasma steroids, were assessed follow-

ing the exposure period. Methods for reproduction

endpoint assessment and vitellogenin, plasma

steroids, and ex vivo gonadal steroid release

followed those reported elsewhere (Foran et al.,

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183174

2002; Zhang et al., 2003). Nominal exposure

concentrations were verified following Weston et

al. (2001).

Fluoxetine exposure for 4 weeks resulted in few

changes in medaka reproductive success. Japanese

medaka pairs produced an average of 1589/41

eggs over the 2-week reproductive assessment

period. Fecundity was unaffected by fluoxetine

treatments (ANOVA, P�/0.55); however, statisti-

cal power was limited by treatment level replica-

tion. Greater than 87% of all eggs were fertilized in

each group. Percentage of fertilized eggs hatched

30 days post-fertilization ranged from 84 to 94% in

all treatment levels. During observations of devel-

oping embryos, several abnormalities were noted.

These included edema, curved spine, incomplete

development (no pectoral fins, reduced eyes), and

non-responsiveness. Whereas few abnormalities

were noted in untreated organisms (4 of 820, or

0.49%), developmental abnormalities were ob-

served more frequently at all fluoxetine treatment

levels. The number and percent of developmental

alterations for each treatment level were: 0.1 mg/l,

21/863 or 2.43%; 0.5 mg/l, 17/637 or 2.53%; 1.0 mg/l,

18/913 or 1.97%; 5.0 mg/l, 17/758 or 2.24%. These

observations indicated that developmental ab-

normalities were 4�/5 times more frequent in

fluoxetine treatments.

Most adult physiological measurements were

also unaffected by fluoxetine exposure. However,

female circulating estradiol levels were increased

by exposure to two fluoxetine treatment levels.

Condition factor [weight (g)/length (mm)3] did not

differ for animals across treatment levels (females,

P�/0.84; males, P�/0.50). Gonadal somatic in-

dex, measured as gonad profile area normalized by

somatic weight, was also unchanged with exposure

(females, P�/0.32; males, P�/0.23). Hepatic VTG

content and circulating T concentrations were not

affected by fluoxetine treatment levels (females,

P�/0.72; males, P�/0.98). Release of E2 and T

from ex vivo gonadal tissue incubated with 25-

hyroxycholesterol did not change with treatment

for either ovarian tissue (E2, P�/0.59; T, P�/0.75)

or testes (E2, P�/0.80; T, P�/0.50). Although

plasma E2 concentrations were unaffected in

males, female circulating E2 was significantly

increased by 0.1 mg/l (P�/0.01) and 0.5 mg/l (P�/

0.054) fluoxetine treatments.

These results provide some early information to

associate physiological change with environmental

fluoxetine exposure and demonstrate the limita-

tion of assessing reproductive impacts with only

one model organism. A 4-week exposure of

environmentally relevant fluoxetine concentra-

tions did not affect Japanese medaka fecundity,

egg fertilization or hatching success. However,

developmental abnormalities were noted at all

fluoxetine exposure levels. Further, a complex

response was noted among the endocrine end-

points; female circulating steroid concentrations

were elevated at 0.1 and 0.5 mg/l exposure levels.

Because of the small plasma volumes collected

from medaka, blood from two animals of the same

sex was pooled, leaving three tissue samples for

plasma steroid analysis in each treatment group.

However, statistically significant responses with

such sample numbers may indicate a dramatic

effect of fluoxetine at 0.1 and 0.5 mg/l exposure

levels. Absence of a concentration-response rela-

tionship in the change of circulating E2 highlights

the potential for different factors to affect re-

sponses to SSRIs, including basal circulating

steroid levels, sexual dimorphisms in cytochrome

P450 enzyme activity, and potential sexual di-

morphisms in serotonin systems (Hernandez-

Rauda et al., 1999). A wide range of impacts and

the potential for regulatory biofeedback to coun-

teract elevations in serotonin raises an issue as to

whether a traditional concentration-response rela-

tionship would be expected with a long term

aquatic exposure to SSRIs.

Although Japanese medaka is a commonly used

model organism in reproductive assessments of

contaminant effects, the response of oocytes to

serotonin is known to vary between species

(Iwamatsu et al., 1993; Cerda et al., 1998; Fong,

2001). Therefore, results of Weston et al. (2003),

which indicate no statistically significant reproduc-

tion changes in response to fluoxetine treatment,

may not be representative of effects in other

teleosts. Clearly, further studies on the sub-lethal

consequences of fluoxetine exposure are necessary,

and these studies should consider study species

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183 175

sensitivity, behavioral responses, and endpointselection to serotonin modulation.

5. Freshwater community responses to fluoxetine,

ibuprofen and ciprofloxacin mixtures

Whereas a direct assessment of fluoxetine effects

on lotic or lentic freshwater communities has notbeen performed, Richards et al. (2003) used lentic

mesocosms to investigate responses to fluoxetine,

ibuprofen, and ciprofloxacin mixtures. Eight

12 000 l mesocosms and their aquatic communities

were established, maintained, and treated with

ibuprofen, ciprofloxacin, and fluoxetine according

to methods described by Richards et al. (2003).

Control mesocosms (n�/3) received no treatment,low treatment mesocosms (LT, n�/1) received 6,

10, and 10 mg/l, medium treatment (MT, n�/1)

received 60, 100, and 100 mg/l, and high treatment

(HT, n�/3) received 600, 1000, and 1000 mg/l of

ibuprofen, ciprofloxacin, and fluoxetine, respec-

tively. These pharmaceuticals were selected based

on mode of action and frequency of prescription in

North America (NDC Health, 1999). Treatmentconcentrations were based on distributional ana-

lyses of upper centiles (e.g. 95th, 99th, and 99.9th)

estimated from actual surface water concentra-

tions for ibuprofen (Buser et al., 1999; Stumpf et

al., 1999; Metcalfe and Koenig, 2001) or on

centiles of distributions of measured environmen-

tal concentrations from similarly prescribed phar-

maceuticals in surface water (Daughton andTernes, 1999; NDC Health, 1999) for ciproflox-

acin and fluoxetine. HT concentrations purposely

exceeded that of individual pharmaceuticals found

in the environment because Richards et al. (2003)

wanted to account for the possibility of additivity

among compounds with the same mode of action

and provide a high-exposure scenario for future

probabilistic risk assessments. Pharmaceutical le-vels within the mesocosms were monitored and

reintroduced as necessary to maintain nominal

concentrations (Richards et al., 2003). Resultant

48-h time-weighted average concentrations were

within 10% of nominal values. Biological samples

(phytoplankton, zooplankton, macrophytes, and

bacteria) were collected every 7 days; fish wereobserved daily.

The initial and most obvious treatment-depen-

dent response was observed in fish. Juvenile

sunfish (Lepomis gibbosus , n�/30 per microcosm)

were contained in mesh cages; naturally occurring

plankton were the primary food source as no

external food was added to the cages or micro-

cosm. Within the first 96 h of HT exposure, allsunfish died (n�/90). The trial was repeated on

day 8; within 4 days, 98.8% mortality was ob-

served. After 35 days of exposure in the MT

microcosm, 46.6% of sunfish had died (n�/14/

30). During the same period, 1.1% mortality was

observed in controls (1/90) and no LT fish (n�/30)

died.

The mechanisms of fish toxicity are unclear;treatment levels were not expected to induce

mortality because MT concentrations were �/

230, 130, and 11-fold lower than those equivalent

to mammalian whole-body therapeutic doses for

ibuprofen, ciprofloxacin, and fluoxetine, respec-

tively (Canadian Pharmacists Association, 2000).

One hypothesis suggested by Richards et al. (2003)

for such a mechanism is that fluoxetine exposureled to increased plasma serotonin levels. Serotonin

constricts the arterio-arterial branchial vasculature

(Nilsson and Sundin, 1998). This would lead to

impaired gas exchange and hypoxia, potentially

leading to death. However, Khan and Thomas

(1992) failed to increase levels or potentiate effects

of serotonin by i.p. injection of 10 mg/g fluoxetine.

Their dose of 10 mg/g was 10-fold greater than theestimated body dose experienced by fish in HT

microcosms (assuming that the concentration in

the fish came to equilibrium with the water 1000

mg/l:/1 mg/g). Other potential factors that could

affect lethality (e.g. dissolved oxygen and pH) were

not significantly different between treatments.

Synergistic interactions, wherein the combination

of the three drugs may have increased the potencyof one or all, may also account for toxicity

observed in sunfish; however, preliminary labora-

tory studies suggest that the observed response was

partially attributable to fluoxetine (Table 2; D.

Johnson, personal communication).

Zooplankton, phytoplankton, and macrophytes

all responded to treatment with the mixture of

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183176

pharmaceuticals (Richards et al., 2003), however,

these changes were not all attributed to fluoxetine.

Zooplankton and phytoplankton communities

were characterized by a decrease in species com-

position but an increase in numbers of some

species. The mixture components responsible for

these responses have not been identified. The

macrophytes, Myriophyllum spicatum , Myriophyl-

lum sibiricum, and Lemna gibba L. all declined at

the HT and L. gibba showed sub-lethal effects

(chlorosis and necrosis) at the MT. These re-

sponses were attributed to ciprofloxacin. Fish

mortality, reduced diversity and community-level

effects among plankton populations, and macro-

phyte mortality observed in this study raises

important questions about the potential for similar

effects in surface waters. However, the causative

agents in the treatment mixture have not all been

identified.

Based on recent data of Kolpin et al. (2002) and

limited data from other sources, the probability of

ibuprofen, ciprofloxacin, and fluoxetine occurring

individually at concentrations high enough to

affect aquatic communities is judged to be low,

however, potential additivity of action must also

be considered. In typical surface waters receiving

wastewater treatment plant effluent, there could be

hundreds of pharmaceuticals; those with similar

modes of action could have additive effects on

indigenous aquatic organisms. In the UK and

Canada alone, over 3000 active pharmaceuticals

are licensed for use (Pfluger and Dietrich, 2001;

Servos et al., 2002); few of which have even been

analyzed for in surface water. Many unaccounted

pharmaceuticals share the same mode of action.

For example, in their review of pharmaceuticals in

the environment, Daughton and Ternes (1999)

discussed 50 pharmaceuticals of concern; of these,

three were SSRIs. Consequently, when surface

water concentrations of pharmaceuticals sharing

a common mechanism of action are further

elucidated, the effective (additive) environmental

concentrations could be more substantial (Daugh-

ton and Ternes, 1999). The types and quantities of

pharmaceuticals present in surface waters will

obviously vary by region; however, it is clear

that there is a high potential for a large number

of pharmaceuticals to simultaneously occur in theenvironment.

6. Ecological risk characterization for fluoxetine

Ecological risk assessment (ERA) procedures

often rely on deterministic hazard quotients to

characterize risk to aquatic organisms. Probabil-

istic risk assessment methods are more attractivethan deterministic ratios because risk can be

expressed as the probability that adverse effects

will occur (Solomon et al., 1996). Further, prob-

abilistic procedures can quantify variability asso-

ciated with exposure and effect measures and

quantify uncertainty inherent to risk assessments

(Hart, 2001). However, lack of environmental

exposure and hazard information for fluoxetinecurrently preclude such probabilistic approaches.

Data presented here provide a foundation for

future probabilistic risk assessments of fluoxetine.

Ecological risk of pharmaceuticals to aquatic

organisms is currently characterized with a hazard

quotient (HQ); however, alternative approaches

have been suggested (Lange and Dietrich, 2002). A

HQ is expressed as the relationship between apredicted environmental concentration (PEC) and

a predicted no effect concentration (PNEC). If a

HQ derived from a PEC/PNEC ratio is B/1, then

risk to the environment is considered low. The US

Food and Drug Administration requires that an

environmental assessment, a modified ERA, be

performed for pharmaceuticals if predicted envir-

onmental introduction concentrations (EIC) aregreater than 1 mg/l (FDA-CDER, 1998). This

approach does not address additive effects of

therapeutics with similar mechanisms of action,

does not consider interaction effects of compounds

with different mechanisms of action, and relies on

acute toxicity test responses. Also, a 10-fold

dilution factor is generally applied to an EIC to

predict expected environmental concentrations(EEC or PEC). This technique may be appropriate

for many lotic systems; Dorn (1996) indicated that

at annual mean flows greater than 75% of per-

mitted, effluent dischargers in the United States

receive 10-fold dilution. However, such an exercise

becomes problematic in regions where effluent

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183 177

discharges do not receive upstream dilution. Per-

ennial municipal effluents influence historically

ephemeral streams in the arid southwestern United

States. For example, flow of the Trinity River

south of Dallas/Fort Worth, TX, is often domi-

nated by greater than 90% municipal effluents

(Dickson et al., 1989). Because the EIC is approxi-

mately equal to the PEC in these streams, fluox-

etine concentrations in effluent dominated systems

may represent maximal hazard to aquatic organ-

isms (Marsh et al., 2003).

Although default EIC calculations do not con-

sider effluent dominated streams, EICs may be

conservative if they are not adjusted for metabo-

lism. Fluoxetine EICs in the United States were

roughly estimated using annual consumption data

for the year 2000 (Table 3). If instream dilution,

degradation, and metabolism are not included in

these estimations, an EIC or PEC for fluoxetine is

approximately 0.439 mg/l (Table 3). Webb (2001)

reported a similar PEC of 0.37 mg/l for fluoxetine

in the UK. However, fluoxetine is normally

excreted as 10% parent compound or fluoxetine

N -glucuronide in urine (Hiemke and Hartter,

2000). When such metabolism is included in EIC

calculations, a value of 0.0439 mg/l was calculated

for systems not receiving dilution. Further, an EIC

of 0.00439 mg/l was generated when a 10-fold

dilution factor and metabolism were considered

(Table 3). This is similar to a PEC of 0.003 mg/l

reported by Webb (2001), which included WWTPbiodegradation and 10-fold dilution factors.

Both Kolpin et al. (2002) and Weston et al.

(2001) measured fluoxetine concentrations in sur-

face waters and municipal effluents at higher levels

than those predicted by lowest PEC calculations in

Webb (2001) and Table 3. Weston et al. (2001)

detected fluoxetine in municipal effluents from

0.32 to 0.54 mg/l and Kolpin et al. (2002) reportedmaximum fluoxetine levels of 0.012 mg/l in surface

waters. The lowest fluoxetine effect level, as

required by USFDA in environmental assessments

of pharmaceuticals, is 13.6 mg/l for P. subcapitata

growth (Table 2). Based on standardized toxicity

test data, a HQ for fluoxetine is calculated at B/1,

suggesting little risk to the aquatic environment.

However, when fish physiological and reproduc-tive responses were evaluated, lower fluoxetine

exposure levels of 0.1 and 0.5 mg/l affected female

Japanese medaka plasma estradiol levels and the

number of developmental abnormalities were

elevated at all exposure levels. When these non-

standard steroid and developmental data are

considered, the lowest observed response level of

fluoxetine on aquatic biota occurs at concentra-tions detected in municipal effluents and at one

order of magnitude higher than highest surface

water concentrations reported (Kolpin et al.,

2002).

7. Conclusions

Herein, we summarized current data on fluox-etine occurrence in surface waters and aquatic

organism and community responses to fluoxetine

exposure, and a preliminary aquatic risk charac-

terization for fluoxetine was provided. Fluoxetine

is reported in effluents and surface waters at low to

mid ng/l concentrations (Weston et al., 2001;

Kolpin et al., 2002). Adverse effects of fluoxetine

are observed in standardized aquatic toxicity testsat mg/l levels. Little risk to aquatic systems is

expected from such fluoxetine exposure levels if a

hazard quotient approach is utilized to character-

ize risk. However, such a deterministic ratio of

exposure and effect levels should not preclude

fluoxetine from further risk consideration.

Table 3

Aquatic predicted environmental concentration (PEC) of

fluoxetine in the United States with and with out corrections

for dilution and metabolism

No metabolism 90% metabolism

No dilution 0.439 mg/l* 0.0439 mg/l

10-fold dilution 0.0439 mg/l 0.00439 mg/l

90% metabolism based on 10% parent compound and

glucoronide conjugate excreted in urine (Hiemke and Hartter,

2000).

* EIC aquatic (mg/l)�/A�/B�/C�/D. A�/kg/year pro-

duced for direct use (active moiety)**; B�/1/l per day entering

POW#; C�/year/365 days; D�/109 mg/kg (conversion fac-

tor).Estimated annual consumption of Prozac (2000): Annual

sales of Prozac�/$2.7 billion USD (CNN, 2001); $2.50 USD

per 20 mg tablet (McLean, 2001); �/19.44 metric tones per year

fluoxetine.# 1.214�/1011 l/day entering public treatment works

on average in the US (FDA-CDER, 1998).

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183178

Sole reliance on aquatic toxicity test endpoints

for regulatory contaminant decisions may not be

sufficient (Cairns, 1983). Standardized toxicity

tests are not intended to predict structural or

functional ecological responses to contaminants

(Dickson et al., 1992; La Point and Waller, 2000)

and may not represent most sensitive species

responses (Cairns, 1986). Further, standardized

test endpoints do not provide information on

biochemical, developmental, behavioural or trans-

generational responses to fluoxetine exposure. For

example, Weston et al. (2003) observed ng/l

treatment levels of fluoxetine to affect Japanese

medaka plasma estradiol levels and increase devel-

opmental abnormalities. Whereas the mechanism

by which fluoxetine induced these responses and

whether such responses may impact population

viability is not clear, potential fluoxetine effects on

fish populations warrants further study.

Daughton and Ternes (1999) suggested that

chronic studies with environmental concentrations

of pharmaceuticals are necessary to assess aquatic

ecosystem responses. Because pharmaceuticals are

continuously released into the environment, it is

relevant to perform such chronic life-cycle-type

tests which encompass sensitive stages of organism

development. Extrapolation from single-species

toxicity tests alone does not assess the potential

for pharmaceuticals to affect aquatic communities.

Aquatic microcosms and mesocosms are well

suited for evaluation of multi-trophic level stressor

responses (Boudou and Ribeyre, 1997; Brooks et

al., 1997; Halling-Sorensen et al., 1998; Kennedy

et al., 2002) and are ideal for assessment of direct

and indirect effects of parent compounds and

metabolites on complex aquatic communities

(Hill et al., 1994). Richards et al. (2003) found

that a mixture of fluoxetine with other pharma-

ceuticals impacted aquatic microcosms over a 35-

day study period. Fluoxetine treatment levels used

by Richards et al. (2003) are higher than reported

environmental concentrations; however, their

study clearly identified the importance of evaluat-

ing contaminant effects on multiple levels of

biological organization. Further, such results sug-

gest that a more definitive assessment of risks

posed by parent compounds and mixtures of

pharmaceuticals in surface waters should be con-ducted.

Daughton and Ternes (1999) also suggested that

bioassays or biomarkers should be developed that

focus on specific mechanisms of pharmaceutical

action on non-target biota. This is decidedly

critical because environmental pharmaceuticals,

unlike pesticides, are not acutely toxic to aquatic

life. For example, the beta-adrenergic receptorblocking therapeutics propranolol and metaprolol

reduce cladoceran heart rate and respiration at

levels lower than those affecting survival, growth

and fecundity (Brooks et al., 2003b). Further,

Brooks et al. (2003a) observed that fluoxetine

adversely reduced growth of a green algae, P.

subcapitata (Table 2); the mechanism(s) by which

fluoxetine exerts its toxicity on algae has not beenreported in the peer-reviewed literature. However,

Munoz-Bellido et al. (2000) identified that fluox-

etine has antibacterial properties, potentially inter-

fering with efflux pumps. Richards et al. (2003)

evaluated effects of fluoxetine, ibuprofen, and

ciprofloxacin mixtures on bacterial communities

in aquatic microcosms. Whereas initial measures

of bacterial cell numbers were not affected bytreatments, phylotypes of bacterial community

samples are currently being evaluated for treat-

ment effects.

A review of existing data clearly indicates a need

for greater understanding of fluoxetine effects on

non-target biota, and of pharmaceutical interac-

tions and effects on multiple levels of biological

organization. Such information is required beforemore definitive assessments of pharmaceuticals in

the environment may be performed. Similarly, lack

of information on temporal and spatial occurrence

of fluoxetine in aquatic systems presently limits

predicted environmental concentration estimates.

Acknowledgements

This research was supported by a US Congres-

sional Environmental Sensors and Signals grant, a

Texas Water Resources Institute/United States

Geological Survey grant, the Institute of Applied

Sciences at the University of North Texas, the

Environmental Toxicology Research Program at

B.W. Brooks et al. / Toxicology Letters 142 (2003) 169�/183 179

the University of Mississippi, and the CanadianNetwork of Toxicology Centres and Canada

Rx&D. Beth Glidewell, Erica March, Bethany

Peterson, Lindsey Odom, John Rimoldi, Richard

Brain, David Johnson, Monica Lam, Scott Ma-

bury and Christian Wilson were vital for sample

collection, analysis, and data processing.

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