Environmental Assessment of Triclosan

Consideration of Exposure and Species Sensitivity ofTriclosan in the Freshwater EnvironmentMarie Capdevielle,* Roger Van Egmond,` Mick Whelan, Donald Versteeg, //Matthias Hofmann-Kamensky,# Josef Inauen,# Virginia Cunningham, and Daniel Woltering``

Colgate-Palmolive, 909 River Road, Piscataway, New Jersey 08855, USA`Unilever, Safety and Environmental Assurance Centre, Sharnbrook, Bedford MK44 1LQ, United KingdomCranfield University, Department of Natural Resources, School of Applied Sciences, Building 88, Cranfield, Bedfordshire, Cranfield,Bedfordshire, MK43 0AL, United Kingdom//Procter and Gamble, Central Product Safety, 11810 East Miami River Road, Cincinnati, Ohio 45252, USA#Ciba Specialty Chemicals, Environmental Safety & Compliance, Regulatory Services, Klybeck 424.214, CH-4002, Basel.214, CH-4002,Basel, SwitzerlandGlaxoSmithKline, Corporate Environment, Health and Safety, 200 N. 16th Street, Philadelphia, Pennsylvania 19102, USA``Alexandria, Virginia, USA

(Received 25 March 2007; Accepted 11 September 2007)

ABSTRACTTriclosan (TCS) is a broad-spectrum antimicrobial used in consumer products including toothpaste and hand soap. After

being used, TCS is washed or rinsed off and residuals that are not biodegraded or otherwise removed during wastewater

treatment can enter the aquatic environment in wastewater effluents and sludges. The environmental exposure and toxicity

of TCS has been the subject of various scientific and regulatory discussions in recent years. There have been a number of

publications in the past 5 y reporting toxicity, fate and transport, and in-stream monitoring data as well as predictions from

aquatic risk assessments. State-of-the-science probabilistic exposure models, including Geography-referenced Regional

Exposure Assessment Tool for European Rivers (GREAT-ER) for European surface waters and Pharmaceutical Assessment and

Transport Evalutation (PhATEe) for US surface waters, have been used to predict in-stream concentrations (PECs). Thesemodels take into account spatial and temporal variability in river flows and wastewater emissions based on empirically

derived estimates of chemical removal in wastewater treatment and in receiving waters. These model simulations (based on

realistic use levels of TCS) have been validated with river monitoring data in areas known to be receiving high wastewater

loads. The results suggest that 90th percentile (low flow) TCS concentrations are less than 200 ng/L for the AireCalder

catchment in the United Kingdom and between 250 ng/L (with in-stream removal) and 850 ng/L (without in-stream removal)

for a range of US surface waters. To better identify the aquatic risk of TCS, a species sensitivity distribution (SSD) was

constructed based on chronic toxicity values, either no observed effect concentrations (NOECs) or various percentile adverse

effect concentrations (EC1025 values) for 14 aquatic species including fish, invertebrates, macrophytes, and algae. The SSD

approach is believed to represent a more realistic threshold of effect than a predicted no effect concentration (PNEC) based

on the data from the single most sensitive species tested. The log-logistic SSD was used to estimate a PNEC, based on an

HC5,50 (the concentration estimated to affect the survival, reproduction and/or growth of 5% of species with a 50%

confidence interval). The PNEC for TCS was 1,550 ng/L. Comparing the SSD-based PNEC with the PECs derived from GREAT-

ER and PhATE modeling to simulate in-river conditions in Europe and the United States, the PEC to PNEC ratios are less than

unity suggesting risks to pelagic species are low even under the highest likely exposures which would occur immediately

downstream of wastewater treatment plant (WWTP) discharge points. In-stream sorption, biodegradation, and photo-

degradation will further reduce pelagic exposures of TCS. Monitoring data in Europe and the United States corroborate the

modeled PEC estimates and reductions in TCS concentrations with distance downstream of WWTP discharges. Environ-

mental metabolites, bioaccumulation, biochemical responses including endocrine-related effects, and community level

effects are far less well studied for this chemical but are addressed in the discussion. The aquatic risk assessment for TCS

should be refined as additional information becomes available.

Keywords: Triclosan Aquatic risk assessment Species sensitivity distribution Probabilistic exposure assessment

INTRODUCTIONThe environmental exposure and toxicity of triclosan

(TCS) has been the subject of various scientific and regulatorydiscussions in recent years. These discussions have beenmotivated by the proximity of the predicted environmentalconcentration (PEC) to the predicted no effect concentration(PNEC) based on most sensitive species conventional aquaticrisk assessment methods (for example, as described in the EU

Technical Guidance Documents; EC 2003). Although the

conventional approach provides a reasonable (screening-level)

1st approximation of environmental risk, this risk is derived

using conservative assumptions and assessment factors and

may not reflect actual environmental situations.

Triclosan is a broad-spectrum antimicrobial used in a

variety of consumer products including toothpaste, shampoos,

deodorants, skin lotions, and hand soaps. Its concentration in

these products is typically in the range of 0.1% to 0.3%. After

being used, TCS is washed or rinsed off and may enter the

environment via local wastewater treatment plants (WWTPs)

* To whom correspondence may be addressed: [email protected]

Published on the Web 9/13/2007.

Integrated Environmental Assessment and Management Volume 4, Number 1pp. 1523 2008 SETAC 15

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where typically 90% to 98% is removed as a result ofbiodegradation and sorption (Singer et al. 2002; Bester 2003,2005; Sabaliunas et al. 2003; Thompson et al. 2005; Heidlerand Halden 2006; Waltman et al. 2006). Between 50% and70% of the removal is attributed to biodegradation (Singer etal. 2002; Heidler and Halden 2006). Triclosan has been shownin numerous studies to be biodegradable and photo-instablecausing it to continue to breakdown following its release intothe aquatic environment. Federle and Schwab (2003)reported a mineralization half-life of 2.5 to 3.5 d in alaboratory river water die-away study with radiolabeled TCS.Measured in-stream removal half-lives for TCS have beenreported to range between 2.1 h (Sabaliunas et al. 2003) and11.6 h (Morrall et al. 2004) but are likely to be considerablyhigher (of the order of days) in lakes (e.g., Singer et al. 2002;Tixier et al. 2002). A TCS half life of 11 d for the Ruhr Riverin Germany was suggested by Bester (2005). However, thisestimate was based on concentrations in single grab samplescollected on the same day from stations separated by anestimated time-of-travel of about 6 d and thus is difficult toconsider as part of this evaluation.There is some experimental evidence on TCS environ-

mental degradates. A series of studies (e.g., Latch et al. 2005)show that photodegradation of TCS produced 2,4-dichlor-ophenol and 2,8-dichlorodibenzo-p-dioxin (DCDD). The2,4-dichlorophenol itself is known to be biodegradable aswell as photodegradable (EC 2000). For DCDD, a conversionrate of 1% has been reported and estimated half-lives suggestthat it is photolabile as well (Aranami and Readman 2007).The formationdecay kinetics of DCDD are also reported bySanchez-Prado et al. (2006). Several papers report theformation of the potential biotransformation product meth-yl-triclosan. Observed concentrations in surface waters aregenerally at or below the limit of quantitation (Singer et al.2002). These concentrations suggest that methyl-triclosan isformed, but also degraded with a slower kinetic than TCS.Further discussion of TCS metabolites is beyond the scope ofthis paper, although additional information on formation anddegradation half-lives would be helpful.Heretofore, the most recent aquatic risk assessment for

TCS was published by Reiss et al. (2002). This was based on aprobabilistic exposure estimation that combined measuredand modeled fate and transport data along with the screeninglevel PNEC (that is, using the lowest reported chronicconcentration). According to Reiss et al. (2002), measuredTCS concentrations in US wastewater effluents range from0.2 to 2.7 lg/L, and the lowest no-observed-effect concen-tration (NOEC) for algae, the most sensitive species, is lessthan 1 lg/L. Thus Reiss et al. (2002) concluded that effluentswith TCS concentrations at the high end of the range aredischarged into rivers with low dilution, TCS concentrationsmay exceed the algal NOEC.The conventional PNEC for TCS is 70 ng/L, which was

derived from the lowest reported NOEC for most sensitiveaquatic species, Scenedesmus subspicatus (i.e., 700 ng/Lreported by Orvos et al. [2002]). This effect is believed tobe algistatic rather than algicidal since algae exposed to TCSat concentrations up to 13 lg/L do recover after dilution ofthe test medium to noninhibitory concentrations (Orvos et al.2002). Other species tend to be less sensitive to TCS. Orvoset al. (2002) report acute and chronic toxicity data foractivated-sludge microorganisms, green and blue green algae,marine and freshwater diatoms, duckweed, Daphnia magna

and Ceriodaphnia, fathead minnows, bluegill sunfish, andrainbow trout.Based on probabilistic exposure estimates, Reiss et al.

(2002) concluded that, the risk to fish, invertebrates, andvascular aquatic plants appear to be of no concern. However,for some rivers with small dilution volumes during low flowconditions, it is possible that some algae species will beaffected immediately downstream of WWTP discharges, butbecause of dissipation in the water column environmentalconcentrations are reduced downstream, leading to reducedrisk to aquatic organisms. Reiss et al. (2002) did not includea probabilistic effects assessment endpoint. The speciessensitivity distribution (SSD) approach is believed to providea more robust and realistic indication of the threshold ofeffect in the natural environment, compared with theconventional (single species) approach (Versteeg et al.1999). Guidance provided in the EU Technical GuidanceDocuments (Part II, Section 3.3) suggests that a reliablePNEC can be derived by the SSD method if the databasecontains at least 10 NOECs for different species covering atleast 8 taxonomic groups. The EU Technical GuidanceDocuments further recommends that a PNEC derived usingthis method should be the concentration below which 5% ofspecies are affected, taking into account a 50% confidenceinterval, termed the HC5,50. The US EnvironmentalProtection Agencys (USEPAs) Guidelines for Ecological RiskAssessment (USEPA 1998, section 5.1) describe a comparablemethod to derive effect distributions from single-speciestoxicity data. This is an accepted and often used approach forrisk estimation (e.g., van Straalen 2002; USEPA 2003; Hoseand Van den Brink 2004; EUFRAM 2005; OECD 2005;Straub and Stewart 2007). In fact, Versteeg et al. (1999) haveshown that the HC5,50 with at least 5 species in thedistribution is protective of communities exposed in meso-cosms (i.e., the HC5,50 , NOECmesocosm).In this paper, an aquatic risk assessment for TCS is

presented which combines a probabilistic exposure assess-ment (validated using monitoring data) with a PNEC derivedfrom an SSD. The SSD analysis was based on publishedchronic toxicity data for 14 aquatic species. The exposureassessment uses realistic estimates of wastewater loads,removal during treatment, dilution, and fate of TCS (notthe breakdown products) in receiving surface waters forvarious catchments in Europe and North America. While thefocus of this paper is on those studies that are used in aquatic,pelagic risk assessment as routinely practiced in the UnitedStates and Europe, other studies which evaluate biochemicalresponses, including endocrine-related effects, are consideredin the discussion.

METHODS

Toxicity assessment

A single SSD was constructed from published chronicNOEC and EC1025 values for 14 aquatic species usingmethods described in Versteeg et al. (1999). The data werenot adjusted for ionization although, with a pKa of 8.1, someof the TCS in the aquatic environment is expected to beionized. Biological membranes are generally permeable to un-ionized molecules and relatively impermeable to the ionizedspecies of molecules (Cohn 1979), thus failure to take theionized fraction into account in this toxicity assessment isbelieved to be conservative. Consistent with this, Orvos et al.

16 Integr Environ Assess Manag 4, 2008M Capdevielle et al.

(2002) demonstrated that the ionized form of TCS is lesstoxic to algae than the neutral form. An acute and chronicPNEC for TCS was estimated using the HC5,50 value in anSSD that was generated using published toxicity data (notadjusted for ionization). The method for compiling distribu-tions of single-species toxicity test data is detailed in Versteeget al. (1999). In those situations where more than 1 study wasavailable for a given species, the geometric mean value wasused. Where required, NOEC and EC1025 values werecombined into an estimate of the low effect concentrationfor a species. A log-logistic curve was fitted to the data usingnonlinear regression.

Exposure assessment

A modeling approach was also used to predict TCSexposure in surface waters in Europe and in the UnitedStates. Predicted environmental concentrations were derivedand compared with measured TCS data.

Europe

Geography-referenced Regional Exposure Assessment Toolfor European Rivers (GREAT-ER) was used to predictstatistical distributions of river water TCS concentrations inthe AireCalder catchment in the United Kingdom. TheUnited Kingdom is considered to be a high-consumptioncountry for TCS (typically 1 g TCS per capita per y). TheGREAT-ER is a GIS-based model which was developed tocalculate the PECs (also referred to in GREAT-ER as Csims, orsimulated concentrations) for down-the-drain chemicals inEuropean surface waters (Feijtel et al. 1997; Koormann et al.2006). The model uses spatially referenced data on thelocations and sizes of WWTPs, chemical removal duringwastewater treatment, a digital river network and associatedflow statistics and in-stream removal processes to calculate in-stream exposures. Key input parameters such as removal inWWTPs and river flow are described using probability densityfunctions, which are combined using Monte Carlo simulation(Warn and Brew 1980). Results are expressed as probabilitydistributions for chemical concentrations in individual riverreaches, which can be visualized as color-coded maps or asconcentration profiles. Model output has been compared withmonitoring data for a number of catchments across Europeand has been shown to generate reasonably accuratepredictions (e.g., Gandolfi et al. 2000; Schroder et al. 2002).The AireCalder catchment (approximately 1,960 km2 at

its closing section) is situated in West Yorkshire, in the northof England. It is characterized by a high population density(approximately 925 inhabitants/km2; Keller et al. 2007) andincludes the cities of Leeds and Bradford as well as smallertowns and villages. A large number of WWTPs with varioussizes, discharge effluent into receiving waters with dilutionfactors as low as 2. The catchment was one of the main case-study areas for the GREAT-ER project (Holt et al. 2003). TheAire below its confluence with the Calder carries treatedsewage effluent from a population of 1,900,000 andapproximately 80% of Yorkshires industrial effluent (Holtet al. 2003). In dry weather, approximately 65% of the flow inthe Aire above the confluence with the Calder is derived fromtreated effluent (Holt et al. 2003). A limited number of grabsamples (6 samples from 4 different stations) have beencollected from the AireCalder system and analyzed for TCSby the Environment Agency of England and Wales. Allsampling stations were located downstream of WWTPs.

Samples were analyzed at the Environment Agency Notting-ham laboratory by GC-MS and following the methoddescribed by McAvoy et al. (2002) but substituting acylationin place of silylation for TCS derivatization. The samplingdates were 10, 11, and 18 February; 3 and 17 March; and 5April 2005. Standard laboratory procedures were followed(e.g., glass bottles, 4 8C storage in the dark, less than 10-dholding times, blanks and blank spiked analytical qualitycontrol samples). The recoveries were greater than 90% andthe reported limit of quantification was 5 ng/L.For the application of GREAT-ER to the AireCalder, the

following assumptions were made: a use rate of 1 g TCS percapita per year, a uniform distribution of WWTP removalrates with a range of 90% to 98%, and 1st order kinetics for in-stream removal of TCS with a rate constant of 0.21/hcorresponding to a half-life of 3.3 h. The in-stream removalrate was derived from a monitoring study employing dye-tracing conducted by Sabaliunas et al. (2003) in the MagBrook, a tributary of the river Calder. Sorption, biodegrada-tion, and photodegradation all contributed to the losses ofTCS in the field. This value (i.e., half-life of 3.3 h) isconsidered to be characteristic of relatively small and shallowstreams of Northern Europe and its use here is reasonablesince it was estimated in the same catchment as the onediscussed. However, the actual rate constant for TCS willprobably vary significantly in space (particularly with riverdepth; Boeije et al. 2000) and seasonally (e.g., with temper-ature). It should be noted that the rate constant assumed hererefers to primary degradation (i.e., disappearance of theparent molecule) rather than to complete mineralization.

North America

For modeling river water exposure to TCS in NorthAmerica, the Pharmaceutical Assessment and TransportEvalutation (PhATEe) model was applied. PhATE uses amass balance approach to estimate PECs in 11 catchmentsselected to be representative of the range of hydrologicregions in the United States (Anderson et al. 2004). Thesecatchments range widely in size and include between 19 andmore than 1,100 WWTPs as point sources of TCS. The recentpaper by Anderson et al. (2004) includes the results of TCSsimulations using PhATE. The key input parameters for TCSwere 600,000 kg/y annual use; no loss prior to enteringwastewater treatment; WWTP removal of 32% for primarytreatment only, 90% for primary plus secondary treatment,and 95% for tertiary treatment. In-stream losses of TCS in theevaluations were assumed to take place with a rate constant of0.23/d (corresponding to a half-life of approximately 3 d), orzero. The 0.23/d value was taken from a laboratorymineralization study reported by Federle and Schwab(2003) which is likely to be conservative with respect toTCS degradation in the field. Biodegradation in vitro is oftenlimited by poor contact with fixed-film microbial biomass.Furthermore, complete mineralization in the laboratory willnecessarily take longer than primary conversion of parentTCS to resulting metabolites (Struijs and van den Berg 1995).Better agreement between modeled and measured data wasobtained with an in-stream loss of zero. This may be due tothe fact that the measured results were from samples takenclose to sewage outfalls, while the PhATE model calculatesconcentrations at the end of river segments.In addition to their modeling efforts, Anderson and

colleagues (2004) reported some model corroboration using

Triclosan Aquatic Risk AssessmentIntegr Environ Assess Manag 4, 2008 17

comparisons of field data for TCS reported by Kolpin et al.(2002) with PhATE model PECs for the same locations on apoint-by-point basis. The field data were derived from anationwide US Geological Survey stream survey projectwhich involved 139 streams across the country (Kolpin etal. 2002) chosen from areas known to be receiving relativelyhigh wastewater loads.

Risk assessment

Ecological risk assessments make comparisons between aPEC and a PNEC. In conventional assessments a PEC toPNEC ratio greater than unity indicates a risk (i.e., exposuresexceed the adverse effect threshold for sensitive species). Inthis study the PECs derived from the PhATE and GREAT-ERmodels were compared to PNEC from SSD.

RESULTS AND DISCUSSION

Predicted no effect concentration

Chronic toxicity values for the 14 species range from 0.53lg/L for Scenedesmus to 290 lg/L for Oryzias (Table 1). Thealgae are consistently more sensitive than either invertebratesor fish. The lowest chronic NOEC is 0.69 lg/L, derived froma laboratory toxicity test in which TCS was tested with a pureculture of Scenedesmus subspicatus exposed for 4 d. Forcomparative purposes, acute toxicity values for 13 species areincluded in Table 2. The acute values range from 1.4 lg/L forScenedesmus to 3,000 lg/L for Chironomus. Again, algae arethe most sensitive species to TCS exposure. There is overlapbetween the acute and chronic toxicity values but generallythe acute value is twice (or more) of the chronic value for thesame species.The available chronic toxicity data was used to develop a

log-logistic distribution (Figure 1). The goodness of fit of theSSD to the test data met statistical criteria using theKolmogorov Smirnov test. The resulting chronic PNEC(HC5,50 value) is 1.55 lg/L TCS (Figure 1). The speciesused to construct the SSD are diverse with broad taxonomicdistribution; the chronic toxicity values came from 14 speciesrepresenting fish, invertebrates, and aquatic plant taxa,exceeding the minimum number of species specified in theEU Technical Guidance Documents; the most sensitive taxon,aquatic plants, is over-represented (50% of the data); the mostsensitive biomass endpoint was used to derive chronic valuesfor algae (growth rate was less sensitive); and, the effect onalgae is believed to be algistatic rather than algicidal. Based onthis information, we expect that chronic, continuous ex-posure to the PNEC value would be expected to have someeffect on approximately 5% of the species exposed but wouldnot be expected to have extensive effects on the exposedcommunity. This PNEC is expected to be protective of thecommunity in general as studies have demonstrated theHC5,50 is conservative relative to mesocosm and field-derived NOECs (Emans et al. 1993; Versteeg et al. 1999).Thus, the HC5,50 is used as the PNEC. While not included inthe PNEC derivation, the result of a published Microtox assay(Tatarazako et al. 2004) indicates the sensitivity of thesebacteria (inhibition concentration [IC], IC25 70 lg/L) is inthe same range as that for the other aquatic organisms.

Predicted environmental concentration

EuropeThe modeled PEC distributions in the AireCalder catchment, a densely populated area in the United

Kingdom, provide an example of a representative worst caseTCS exposure scenario. The model generates a PECprobability distributions for each river reach, which reflectvariability in key variables such as removal in WWTPs andriver flow. Figure 2 shows the spatial distribution of the 90thpercentile TCS concentrations in the rivers Aire and Calderpredicted by GREAT-ER. The highest 90th percentile PECsresult from a combination of high emission and low riverflow and are shown in red. A profile of changes in measuredand modeled TCS concentrations along the main channel ofthe river Aire is also shown in Figure 2. Comparison of themodel prediction with the measured TCS concentrationdata, collated by the Environment Agency of England andWales (mean 6 1 standard deviation), suggests that GREAT-ER simulations are reasonable for this system, although themeasured data set is rather limited (4 locations, 6 samplingsper location). This confirms an earlier unpublished compar-ison of GREAT-ER predictions with a more extensive waterquality data set (Holt et al. 2003) that showed that GREAT-ER could match measured concentrations of down-the-drainchemicals for this catchment. The highest PEC wasgenerated for the river reach downstream of Bradford. Atthis point, the 90th percentile PEC is approximately 180 ng/L (see insert in Figure 2) and the mean PEC is approximately100 ng/L.Estimated TCS use in the United Kingdom is high

compared to other EU countries and therefore aquatic risksbased on UK exposure estimates are reasonably conservativefor the European situation. The stream characteristicssimulated in the AireCalder catchment study may bedirectly applied to other relatively small and shallow streamsof Northern Europe. Predicted environmental concentrationscould be different for larger and deeper streams where flowrate, depth, and turbidity would affect in-stream dilution,sorption, biodegradation, and photodegradation.North AmericaFigure 3 provides a comparison of PhATE-

modeled and US Geological Surveymeasured TCS levels forUS surface waters. The cumulative probability distribution ofmeasured TCS concentrations, as reported in Kolpin et al.(2002), is shown as solid triangles. Nondetects are repre-sented by open triangles. Modeled PECs generated by PhATEare also shown as a cumulative distribution of concentrationsgenerated in all river reaches in all catchments considered inthe model. Two model curves are shown in Figure 3: 1 for lowflow conditions (90th percentile PEC 850 ng/L) and 1 formean flow conditions (90th percentile PEC 60 ng/L). ThePhATE simulation depicted in Figure 3 was generatedassuming zero in-stream removal of TCS. The same simu-lation run with an in-stream removal rate of 0.23/d predictsconsiderably lower low-flow and mean-flow TCS 90thpercentile PECs (250 ng/L and 25 ng/L respectively). ThesePECs are between one-third and one-half the concentrationspredicted using zero in-stream removal.However, there is better correlation between the measured

data and PhATE predictions when zero in-stream removal isassumed, especially for the low flow condition. According tothe US Geological Survey authors (Kolpin et al. 2002), theselection of sampling sites in their nationwide survey wasintentionally focused on streams considered susceptible tocontamination from human, industrial, and agriculturalwastewater loading. The higher TCS concentrations generallyrepresent locations close to WWTPs where travel times aretoo short to allow significant in-stream removal.


Risk assessment

The chronic PNEC derived from the SSD and based on

toxicity data for 14 aquatic species is 1,550 ng/L TCS. In

comparison, predicted 90th percentile concentrations based

on a state-of-the science exposure estimation model (GREAT-

ER) for a worst-case catchment in the United Kingdom,

suggest that concentrations rarely exceed 200 ng/L TCS and

this might occur in a limited number of locations (e.g.,

immediately downstream of significant WWTP effluents). For

North America, PhATE model predictions suggest that TCS

concentrations as high as 850 ng/L may occur (without in-

stream removal and 250 ng/L with in-stream removal). Both

of these PECs represent 90th percentile (low-flow) simu-

lations. However, in all cases, the PEC to PNEC ratios are less

than unity suggesting risks to sensitive aquatic species are low

even under the highest likely exposures which would occur

immediately downstream of WWTP discharge points. Hence,at current TCS usage, TCS is not expected to adversely impactaquatic communities immediately belowWWTPeffluents. In-stream sorption, biodegradation, and photodegradation ofTCS will tend to reduce pelagic exposures further down-stream (e.g., Morrall et al. 2004). Monitoring data in Europeand the United States corroborate the modeled PEC estimatesas well as expected reductions in TCS concentrations withincreasing distance downstream of WWTP discharges.

Additional considerations for TCS

This paper presents a risk assessment for pelagic speciescontinuously exposed to TCS released through WWTPeffluents. The authors recognize that based on the physicaland chemical characteristics of TCS, (e.g., log Kow) it will sorbto solids and it may be found in sediments and soils but adiscussion of risk in these compartments is beyond the scope

Table 1. Chronic aquatic toxicity data for triclosan

Taxa Species Durationa Value(lg/L)

Endpoint Referenceb Comment

Fish Oryzias latipes 21 d 156 NOEC 1 Hatchability and fry survival

Fish O. latipes 9 d ph 290 IC25 2 Hatchability and fry survival

Fish Danio rerio 10 d ph 200 NOEC 3 Not reported

Fish D. rerio 14 d ph 160 IC25 2 Hatchability and fry survival

Fish Oncorhynchus mykiss 61 d ph 34.1 NOEC 4 Hatchability and fry survival

Invertebrate Brachionus calyciflorus 2 d 50 NOEC 3 Not reported

Invertebrate Ceriodaphnia dubia 7 d 182 NOEC 4 pH 8.5, survival andreproduction

Invertebrate C. dubia 7 d 6 NOEC 4 pH 7.0, survival andreproduction

Invertebrate C. dubia 7 d 4 NOEC 3 Not reported

Invertebrate C. dubia 7 d 170 IC25 2 Survival and reproduction

Invertebrate Daphnia magna 21 d 40 NOEC 4 Survival and reproduction

Invertebrate Chironomus tentans 10 d 80 NOEC 5 Survival and growth

Invertebrate Hyalella azteca 10 d 50 EC10 5 Survival and growth

Algae Selenastrumcapricornutum

4 d 2.44 EC25 4 Biomass

Algae S. capricornutum 4 d 2.6 NOEC 3 Not reported

Algae S. capricornutum 4 d 3.4 IC25 2 growth

Algae Scenedesmussubspicatus

4 d 0.69 NOEC 4 Biomass

Algae S. subspicatus 3 d 0.53 NOEC 4 Nominal, growth

Algae Skeletonema 4 d 66 EC20 4 Biomass

Algae Anabaena 4 d 0.67 EC25 4 Biomass

Algae Anabaena 4 d 0.97 EC10 4 Biomass

Algae Navicula 4 d 10.7 EC20 4 Biomass

Macrophyte Lemna 10 d 62.5 EC20 4 Biomassa ph posthatch.b 1 Ishibashi et al. 2004; 2 Tatarazako et al. 2004; 3 Ferrari et al. 2002; 4 Orvos et al. 2002; 5 Dussault et al. 2004.


of this paper. This pelagic assessment has focused on the useof predicted and measured water concentrations and data

from laboratory toxicity tests. However, there are a variety ofother fate and effect-related studies that may suggest futureresearch and may be relevant to better understand TCS fate

and effects in the environment. For example, TCS has beendetected in human breast milk (Adolfsson-Erici et al. 2002),

although in its 2002 report on TCS and its use in consumerproducts, the European Commissions Scientific SteeringCommittee concluded that any triclosan absorbed by the

human body is rapidly excreted and no long-term or large-

scale accumulation occurs (EC 2002). In addition, TCSbioconcentrates in fish (Orvos et al. 2002) and other

organisms. Coogan et al. (2007) analyzed filamentous algae(Cladophora spp.) sampled in the vicinity of a WWTP foundincreasing TCS concentrations downstream of the outfall

despite decreasing TCS levels in water. Triclosan has beenreported in the bile of wild fish exposed to a WWTP effluent

(Adolfsson-Erici et al. 2002). While no data are currentlyavailable that correlate tissue concentrations with effect levels(e.g. Environmental Effects Residue Database; USACE 2006),

the contribution of bioconcentration is factored into this

Figure 1. Species sensitivity distribution. The log-logistic nonlinear regression of chronic aquatic toxicity data. Upper and lower confidence limits are shown withdashed lines. The resulting HC5,50 value, belowwhich 5% (a 0.05 cumulative probability) of the species responses (no observed effect concentrations [NOECs] orvarious percentile adverse effect concentrations [ECx]) occur taking into account a 50% confidence interval, is 1.55 lg/L as shown in red on the x-axis.

Table 2. Acute aquatic toxicity data for triclosan

Taxa Species Duration (d) Value (lg/L) Endpoint Referencea

Fish Oryzias latipes 4 399 LC50 1

Fish Pimephales promelas 4 260 LC50 2

Fish Lepomis macrochirus 4 440 LC50 2

Invertebrate Daphnia magna 2 343.8 EC50 2

Invertebrate Ceriodaphnia dubia 2 184.7 EC50 2

Invertebrate Chironomus tentans 10 3,000 LC50 3

Invertebrate Hyalella azteca 10 1,000 LC50 3

Algae Selenastrum capricornutum 4 4.46 EC50 2

Algae Scenedesmus subspicatus 4 1.4 EC50 2

Algae Skeletonema 4 66 EC50 2

Algae Anabaena 4 1.6 EC50 2

Algae Navicula 4 19.1 EC50 2

Macrophyte Lemna 10 62.5 EC50 2a 1 Ishibashi et al. 2004; 2 Orvos et al. 2002; 3 Dussault et al. 2004.


assessment as chronic exposure toxicity tests are sufficient in

duration to include effects of bioconcentration on growth,

survival, and reproduction of test species.

In this assessment, ecologically relevant endpoints of

growth, survival, and reproduction are used to evaluate risk

to multiple species and the ecosystem. Studies that evaluate

biochemical responses, including endocrine-related effects,

are difficult to build into a risk assessment unless effects on

survival and reproduction also occur. However, biochemical

endpoints can be used to understand mode of action, suggest

biomarkers of exposure, and rapidly identify species that may

be uniquely sensitive. For example, Canesi et al. (2007) have

shown that mussel immune cells, hemocytes, when exposed

to TCS at 290 to 29,000 lg/L in vitro and to nominalconcentrations of 0.29, 2.9, and 29 ng/g dry weight by

injection showed effects on various enzymes. Endocrine-

Figure 3. The cumulative probability distribution of all triclosan concentrations in US surface waters reported by the US Geological Survey and predictedenvironmental concentration (PECs) generated by Pharmaceutical Assessment and Transport Evalutation (PhATETM) for all model segments. Detected triclosanshown as solid triangles m and nondetects as open triangles D. Model results are shown based on zero in-stream removal. Reproduced with permission,Copyright 2004, American Chemical Society (Anderson et al. 2004).

Figure 2. Simulated 90th percentile triclosan (TCS) concentrations in the UK AireCalder catchment using Geography-referenced Regional ExposureAssessment Tool for European Rivers (GREAT-ER) simulation model along with direct comparison of modeled and measured (in blue) TCS levels in the same riverreach (upper right insert).


related effects have been investigated by Foran et al. (2000),Ishibashi et al. (2004), and Veldhoen et al. (2006). Veldhoenexposed premetamorphic North American bullfrogs, Ranacatesbeiana, to TCS at 0.15, 1.5, and 22 lg/L and followedtadpole developmental stages for 18 d in 3,5,3-triiodothyr-onine (T3) injected and uninjected individuals. 3,5,3-Triio-dothyronine injection was used to induce metamorphosis inthe bullfrog, although this species was not ready to undergothis process naturally at this life stage. Coadministration of T3and TCS interfered with thyroid-related gene expression andcaused tadpoles to develop faster than controls. Uninjectedtadpoles exposed to TCS showed no difference from controlsin development but did demonstrate a small, transient changein thyroid receptor alpha in the brain. Ishibashi et al. (2004)exposed medaka (Oryzias latipes) to TCS at concentrations upto 137 lg/L for 21 d. Gonadosomatic index was increased inadults exposed to TCS and hepatic vitellogenin levels wereincreased in males exposed to 13 and 61 lg/L TCS. Despitethese changes in biomarkers, there were no adverse effects onreproduction, hatchability, cumulative mortality, growth, orsex ratio of offspring at doses up to 137 lg/L. The authorsconcluded that a metabolite of TCS was probably weaklyestrogenic. Foran et al. (2000) measured sex ratio and anal finlength in medaka and concluded that TCS was not estrogenicbut could be weakly androgenic due to the presence of longerdorsal and anal fins on fish exposed to 100 lg/L.These biomarker-related effects are difficult to build into a

risk assessment as they involve unrealistic exposure modes(Veldhoen et al. 2006; Canesi et al. 2007), demonstrateeffects only at elevated concentrations (Foran et al. 2000), orshow a biochemical effect with no effect on growth, survival,or reproduction of the species (Ishibashi et al. 2004). A recentpaper by Dayan (2007) showed no indication of endocrinedisruption in mammals treated with TCS in relevant highdose tests of fertility and fetal and neonatal development or inchronic toxicity and carcinogenicity experiments. There isgood conservation in endocrine systems (e.g., receptorstructure and function) across wide taxonomic distances andthus an evaluation of endocrine effects on mammals are likelyrelevant to fish and amphibians. Endocrine-related effects inmammals appear to be unlikely and it is reasonable toquestion if endocrine-related effects in fish and amphibianswill be supported by additional studies.The focus of this paper is the application of a recognized

SSD approach to derive a PNEC for pelagic aquatic speciesbased on standardized single species toxicity tests. The singlespecies toxicity data (see Figure 1) demonstrate that algae(including diatoms, green and bluegreen) are the mostsensitive standard test species by a considerable margin overinvertebrates and fish. Investigations of toxicity to algalassemblages or communities (naturally occurring or artifi-cially derived) can provide a point of comparison to thepredictions from the laboratory tests. Several such algalassemblage studies have been reported (Wilson et al. 2003;White et al. 2005; R. Hunt, unpublished data, 2006) in whichexposed natural algal or periphyton communities have beenexposed to TCS. Effects were observed at concentrationsranging from 0.015 to 1,100 lg/L. In these studies, the lack ofanalytical confirmation of exposures, the lack of doseresponse, lack of concurrence across studies and issues withthe level of taxonomy makes interpretation difficult. Inaddition, the use of natural algal and bacterial communitiesin static exposure systems may lead to TCS degradation andrelease of the algal nutrient CO2 which may itself have altered

algal communities in these studies. At this point in time, valid,ecologically interpretable algal assemblage/communityNOEC values do not exist for TCS.

CONCLUSIONSThis pelagic aquatic risk assessment for TCS combined a

more sophisticated toxicity (PNEC) methodology with aprobabilistic exposure (PEC) assessment. Based on the PNECderived using a species sensitivity distribution and the PECsderived from GREAT-ER and PhATE modeling to simulatein-river conditions in Europe and the United States, the PECto PNEC ratios are less than unity suggesting risks to sensitiveaquatic species are low even under the highest likelyexposures that would occur immediately downstream ofWWTP discharge points. In reality, in-stream sorption,biodegradation, and photodegradation of TCS will tend toreduce pelagic exposures further. Monitoring data in Europeand the United States corroborate the modeled PECestimates as well as expected reductions in TCS concen-trations with increasing distance downstream of WWTPdischarges. Environmental metabolites, bioaccumulation,biochemical responses including endocrine-related effects,and community level effects are far less well studied and theaquatic risk assessment for TCS should be continuouslyrefined as additional information becomes available.

REFERENCESAdolfsson-Erici M, Pettersson M, Parkkonen J, Sturve J. 2002. Triclosan, a

commonly used bactericide found in human milk and in the aquatic

environment in Sweden. Chemosphere 46:14851489.

Anderson PD, DAco VJ, Shanahan P, Chapra SC, Buzby ME, Cunningham VL,

Duplessie BM, Hayes EP, Mastrocco FJ, Parke NJ, Rader JC, Samuelian JH,

Schwab BW. 2004. Screening analysis of human pharmaceutical compounds

in U.S. surface waters. Environ Sci Technol 38:838849.

Aranami K, Readman J. 2007. Photolytic degradation of triclosan in freshwater

and seawater. Chemosphere 66:10521056.

Bester K. 2003. Triclosan in a sewage treatment processbalances and

monitoring data. Water Res 37:38913896.

Bester K. 2005. Fate of triclosan and triclosan-methyl in sewage treatment plants

and surface waters. Arch Environ Contam Toxicol 49:917.

Boeije GM, Wagner J-O, Koormann F, Vanrolleghem PA, Schowanek DR, Feijtel TCJ.

2000. New PEC definitions for river basins applicable to GIS-based environ-

mental exposure assessment. Chemosphere 40:255265.

Canesi L, Ciacci C, Lorusso LC, Betti M, Gallo G, Pojana G, Marcomini A. 2007.

Effects of triclosan on Mytilus galloprovincialis hemocyte function and

digestive gland enzyme activities: Possible modes of action on non target

organisms. Comp Biochem Physiol Chem Toxicol Pharmacol 145:464472.

Cohn VH. 1979. Transmembrane movement of drug molecules. In: La Du BN,

Mandel HG, Way EL, editors. Fundamentals of drug metabolism and drug

disposition. New York (NY): Robert E. Krieger. p 316.

Coogan MA, Edziyie RE, LaPoint TW, Venables BJ. 2007. Algal bioaccumulation of

triclocarban, triclosan and methyl-triclosan in a North Texas wastewater

treatment plant receiving stream. Chemosphere 67:19111918.

Dayan A. 2007. Risk assessment of triclosan (Irgasant) in human breast milk. Food

Chem Toxicol 45:125129.

Dussault E, Solomon K, Sibley P. 2004. Aquatic toxicity of carbamazepine,

atorvastatin and triclosan to benthic invertebrates [poster]. In: Society of

Environmental Toxicology and Chemistry (SETAC) 25th Annual Meeting; 2004

Nov 1418; Portland, OR. Pensacola (FL): SETAC. Poster IP006.

[EC] European Commission. 2000. IUCLID Dataset on 2,4-dichlorophenol: CAS No.

120-83-2, February 18, European Chemicals Bureau. European Commission.

58 p.

[EC] European Commission. 2003. Technical guidance document on risk assess-

ment in support of Directive 93/67/EEC (new notified substances) and

Regulation 1488/94 (existing substances). European Chemicals Bureau

Institute for Health and Consumer Protection. Part II. 328 p.


[EC] European Commission. 2002. Health and Consumer Protection Directorate

General, Opinion on Triclosan Resistance. June 2002. Brussels (BE): EC. Section

V1.1.

Emans HJB, van de Plassche EJ, Canton JH, Okkerman PC, Sparenburg PM. 1993.

Validation of some extrapolation methods used for effect assessment. Environ

Toxicol Chem 12:21392154.

[EUFRAM] European Framework for Probabilistic Risk Assessment of the Environ-

mental Impacts of Pesticides. 2005. Introducing probabilistic methods into

the ecological risk assessment of pesticides. Draft EUFRAM Report, Vol 1, Ver

6. York (UK): Central Science Laboratory.

Federle TW, Schwab EL. 2003. Triclosan biodegradation in wastewater treatment

plant effluent diluted in various river waters. 103rd American Society for

Microbiology General Meeting; 2003; Washington, DC.

Feijtel TCJ, Boeije G, Matthies M, Young A, Morris G, Gandolfi C, Hansen B, Fox K,

Holt M, Koch V, Schroeder R, Cassani G, Schowanek D, Rosenblom J, Niessen

H. 1997. Development of a geographically-referenced regional exposure

assessment tool for European riverGREAT-ER. Chemosphere 34:23512374.

Ferrari B, Paxeus N, Pinto G, Pollio A, Andreozzi R, Marotta R, Vogna D, Lyberatos

G, Garric J. 2002. Fate and effect of triclosan in the aquatic ecosystems: data

for a risk assessment [poster]. In: Society of Environmental Toxicology and

Chemistry Annual Meeting in Europe; 2002 May 1216; Vienna, Austria.

Brussels (BE): SETAC Europe. Poster 73-04.

Foran CM, Bennett ER, Benson WH. 2000. Developmental evaluation of a

potential non-steroidal estrogen: Triclosan. Marine Environ Res 50:153156.

Gandolfi C, Facchi A, Whelan MJ, Cassani G, Tartari G, Marcomini A. 2000.

Validation of the GREAT-ER model in the Lambro catchment. CLER Review

6:8291.

Heidler J, Halden RU. 2006. Mass balance assessment of triclosan removal during

conventional sewage treatment. Chemosphere 66:362369.

Holt MS, Fox KK, Daniel M, Buckland H. 2003. Linear alkylbenzene sulfonate and

boron monitoring in four catchments in the UK. Contribution to GREAT-ER

#11. Sci Total Environ 314316:271288.

Hose GC, Van den Brink PJ. 2004. Confirming the species sensitivity concept for

endosulfan using laboratory, mesocosm, and field data. Arch Environ Contam

Toxicol 47:511520.

Ishibashi H, Matsumura N, Hirano M, Matsuoka M, Shiratsuchi H, Ishibashi Y,

Takao Y, Arizono K. 2004. Effects of triclosan on the early life stages and

reproduction of medaka Oryzias latipes and induction of hepatic vitellogenin.

Aquat Toxicol 67:167179.

Keller V, Rees HG, Fox KK, Whelan M.J. 2007. A practical approach for estimating

the concentrations of down-the-drain chemicals locally and regionally: An

application in England and Wales. Environ Pollut 148:334342.

Kolpin DW, Furlong ET, Meyer MT, Thurman EM, Zaugg SD, Barber LB, Buxton HT.

2002. Pharmaceuticals, hormones and other organic wastewater contam-

inants in U.S. streams, 19992000: A national reconnaissance. Environ Sci

Technol 36:12021211.

Koormann F, Rominger J, Schowanek D, Wagner J-O, Schroder R, Wind T, Silvani

M, Whelan MJ. 2006. Modelling the fate of down-the-drain chemicals in

rivers: An improved software for GREAT-ER. Environmental Modelling &

Software 21:925936.

Latch DE, Packer JL, Stender BL, VanOverbek J, Arnold WA, McNeill K. 2005.

Aqueous photochemistry of triclosan: Formation of 2,4-dichlorophenol, 2,8-

dichlorodibenzo-p-dioxin, and oligomerization products. Environ Toxicol

Chem 24:517525.

McAvoy DC, Schatowitz B, Jacob M, Hauk A, Eckoff WS. 2002. Measurement of

triclosan in wastewater treatment systems. Environ Toxicol Chem 21:1323

1329.

Morrall D, McAvoy D, Schatowitz B, Inauen J, Jacob M, Hauk A, Eckhoff W. 2004.

A field study of triclosan loss rates in river water (Cibolo Creek, TX).

Chemosphere 54:653660.

[OECD] Organization for Economic Cooperation and Development. 2005. OECD

SIDS Initial Assessment Report of LAS. In: 20th SIAM meeting; 1921 April

2005; Paris, France. Washington, DC: Council for LAB/LAS Environmental

Research. The CLER Review, 10, December 2005. 104 p.

Orvos DR, Versteeg DJ, Inauen J, Capdevielle M, Rothenstein A, Cunningham V.

2002. Aquatic toxicity of triclosan. Environ Toxicol Chem 21:13381349.

Reiss R, Mackay N, Habig C, Griffin J. 2002. An ecological risk assessment for

triclosan in lotic systems following discharge from wastewater treatment

plants in the United States. Environ Toxicol Chem 21:24832492.

Sabaliunas D, Webb SF, Hauk A, Jacob M, Eckhoff WS. 2003. Environmental fate

of triclosan in the River Aire basin, UK. Water Res 37:31453154.

Sanchez-Prado L, Llompart M, Lores M, Garcia-Jares C, Bayona JM, Cela R. 2006.

Monitoring the photochemical degradation of triclosan in wastewater by UV

light and sunlight using solid-phase microextraction. Chemosphere 65:1338

1347.

Schroder R, Schulze C, Matthies M. 2002. Concentration of LAS and boron in the

Itter. Comparison of measured data with results obtained by simulation with

the GREAT-ER software. Environ Sci Pollut Res 9:130135.

Singer H, Muller S, Tixier C, Pillonel L. 2002. Triclosan: Occurrence and fate of a

widely used biocide in the aquatic environment: Field measurements in

wastewater treatment plants, surface waters, and lake sediments. Environ Sci

Technol 36:49985004.

Straub J, Stewart K. 2007. Deterministic and probabilistic acute-based environ-

mental risk assessment for naproxen for Western Europe. Environ Toxicol

Chem 26:795806.

Struijs J, van den Berg R. 1995. Standardized biodegradability tests: Extrapolation

to aerobic environments. Water Res 29:255262.

Tatarazako N, Ishibashi H, Teshima K, Kishi K, Arizono K. 2004. Effects of triclosan

on various aquatic organisms. Environ Sci 11:130133.

Thompson A, Griffen P, Stuetz R, Cartmell E. 2005. The fate and removal of

triclosan during wastewater treatment. Water Environ Res 77:6367.

Tixier C, Singer HP, Canonica S, Muller SR. 2002. Phototransformation of triclosan

in surface waters: A relevant elimination process for this widely used biocide

Laboratory studies, field measurements, and modeling. Environ Sci Technol

36:34823489.

[USACE] US Army Corps of Engineers. 2006. Environmental Effects Residue

Database. http://el.erdc.usace.army.mil/ered/index.html#misc. Accessed 24

August 2007.

[USEPA] US Environmental Protection Agency. 1998. Guidelines for ecological risk

assessment. Risk assessment forum. Washington DC: Office of Research and

Development. EPA/630/R-95/002F.

[USEPA] US Environmental Protection Agency. 2003. Ambient aquatic life water

quality criteria for atrazine. Washington DC: Office of Water. EPA-822-R-03-

023.

van Straalen NM. 2002. Theory of ecological risk assessment based on species

sensitivity distributions. In: Posthuma L, Suter GW II, Traas TP, editors. Species

sensitivity distributions in ecotoxicology. Boca Raton (FL): Lewis. p 3748.

Veldhoen N, Skirrow RC, Osachoff H, Wigmore H, Clapson DJ, Gunderson MP, Van

Aggelen G, Helbing CC. 2006. The bacterial agent triclosan modulates thyroid

hormone-associated gene expression and disrupts postembryonic anuran

development. Aquat Toxicol 80:217227.

Versteeg DJ, Belanger S, Carr CJ. 1999. Understanding single species and model

ecosystem sensitivity: A data based comparison. Environ Toxicol Chem

18:13291346.

Waltman E, Venables B, Waller W. 2006. Triclosan in a North Texas wastewater

treatment plant and the influent and effluent of an experimental constructed

wetland. Environ Toxicol Chem 25:367372.

Warn AE, Brew JS. 1980. Mass balance. Water Res 14:14271434.

White KL, Haggard BE, Matlock MD, Kim JW. 2005. Periphytic chlorophyll-a

response to triclosan exposure application of a passive diffusion periphy-

tometer. Applied Engineering in Agriculture 21:307311.

Wilson BA, Smith VH, Denoyelles F, Larive CK. 2003. Effects of three

pharmaceutical and personal care products on natural freshwater algal

assemblages. Environ Sci Technol 37:17131719.


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