organic micropollutants such as pesticides 
 will only cause detrimental effects to organisms if they are taken 
 up by the organism and can reach a target site where they can do harm .
the processes of uptake , biotransformation , and elimination , also 
 termed bioaccumulation or toxicokinetics , modify the concentration 
 of organic chemicals in organisms , and kinetic rate constant models 
 of these processes quantify and yield the time course of internal 
 concentrations .
bioaccumulation and biotransformation are key factors 
 modifying toxicity , and bioaccumulation itself is one of the assessment 
 end points in risk assessment of chemicals .
bioaccumulation based on total radioactivity measurements 
 of c - labeled compounds varies greatly among species and compounds , but 
 the causes remain partially unresolved because the contribution of 
 biotransformation can not be quantified with these methods .
biotransformation 
 in freshwater arthropods has been shown to greatly modify internal 
 concentrations of organic chemicals for various biological species 
 and chemical compounds . however , there are no studies that investigate the uptake , biotransformation , 
 and elimination kinetics of larger sets of chemicals ( n > 4 ) with diverse properties in freshwater invertebrates . in this study , we investigated 
 the uptake , biotransformation , and elimination kinetics of 15 organic 
 chemicals from diverse chemical classes and physicochemical properties 
 ( octanol  water partition coefficient log kow from 0.33 to 5.18 , including acids and bases ) that 
 exhibit a wide range of modes of toxic action in the freshwater amphipod gammarus pulex
. g. pulex are of ecological 
 importance due to their role as shredders in the detritus processing 
 in streams , are used extensively in ecotoxicological research , and are suitable for bioaccumulation studies due to their slow growth 
 and relatively large size . in a previous study 
 of bioaccumulation kinetics in g. pulex , several 
 organic compounds exhibited unexpectedly slow apparent elimination 
 kinetics , but the cause of these could not be elucidated because only 
 total radioactivity of c - labeled compounds was measured 
 and biotransformation was not investigated . in order to understand the role of biotransformation for the 
 kinetics of bioaccumulation , we measured and modeled not only the 
 uptake and elimination of the parent compounds but also biotransformation 
 and , if applicable , tried to identify the biotransformation products 
 ( metabolites ) .
we carried out metabolite screening and identification 
 tests with a 24 h exposure period as well as biotransformation kinetics 
 experiments consisting of a 24 h uptake phase followed by a depuration 
 phase of variable length ( adapted from refs ( 11 , 17 , 21 , and 22 ) ) .
further , toxicokinetic rate constant models 
 were parametrized for each compound to facilitate process - based modeling 
 of toxicity .
kinetic rate constants were derived for all relevant processes : 
 uptake , biotransformation , and elimination of parent compound as well 
 as formation and elimination of the metabolites .
the uptake and elimination 
 rate constants comprise uptake via food , respiratory surfaces , and 
 dermal absorption and elimination via excretion and through respiratory 
 or dermal surfaces ; i.e. , we measured and modeled bioaccumulation sensu mackay and fraser .
the 
 terms metabolism , xenobiotic metabolism , or biotransformation have 
 been used to describe the biochemical modification and transformation 
 of chemicals in an organism . in the following , we will use the term 
 biotransformation for the process , because it emphasizes the transformation 
 aspect and can not be confused with energy metabolism in organisms . 
 for the biotransformation products ,
adult g. pulex were collected 
 during 2008 to 2010 from a small headwater stream in the itziker ried , 
 ca .
20 km southeast of zrich , switzerland ( e 702150 , n 2360850 ) , 
 and acclimatized to the test conditions ( 13 c , 12 h:12 h light / dark ) 
 for at least 3 days prior to experiments .
experiments generally followed 
 previous studies conditions and were carried 
 out in beakers with 500 ml of preaerated artificial pond water , where organisms were fed ad libitum with horse - chestnut 
 leaf discs ( 20 mm diameter , 3 to 5 discs per beaker , replaced when 
 necessary ) that were inoculated with cladosporium herbarum .
experiments were carried out with c - labeled compounds ( see table 1 for compounds 
 and log kow values , figures 14 for positions of the c - label , and the supporting information , text si-1 , for further details).table 1parent compounds ( p ) and metabolites 
 ( m1 , m2 and m3 ) as well as metabolite enrichment factors ( mefs ) for 
 metabolites and bioaccumulation factors ( bafs ) for parent compoundsparent compound(cas number)log  kowa  and  pka values 
 [ -]molecules ( in order of increasing 
 retentiontime for each compound)method of detection and identificationmef or baf[l / kgwet  weight]imidacloprid(cas  138261 - 41 - 3)0.33 ( pka  =  11.12)no metabolites detectedradio - hplcbafp = 7fethylacrylate(cas  140 - 88 - 5)1.32no metabolites detecteddradio - hplcbafp = 87f4,6-dinitro - o - cresol(cas  534 - 52 - 1)2.22 ( pka  =  4.46)no metabolites detectedradio - hplcbafp = 37f1,2,3-trichlorobenzene(cas  87 - 61 - 6)4.05no metabolites 
 detectedradio - hplcbafp = 191fmalathion(121 - 75 - 5)2.36m1 : not identifiedradio - hplcmefm1 = 5p : malathion + 
 malaoxonehplc - radio / uvbafp = 3chlorpyrifos(cas  2921 - 88 - 2)4.96m1 : not 
 identifiedradio - hplcmefm1 = 492m2 : chlorpyrifos - oxonhplc - radio / uvmefm2 = 183p : chlorpyrifoshplc - radio / uvbafp = 412aldicarb(cas  116 - 06 - 3)1.13m1 : aldicarb - sulfonehplc - radio / uvmefm1 = 0.3p : aldicarbhplc - radio / uvbafp = 0.9carbofuran(cas  1563 - 66 - 2)2.32m1 : hydroxy - carbofuranhplc - radio / uvmefm1 = 69p : carbofuranhplc - radio / uvbafp = 0.7carbaryl(cas  63 - 25 - 2)2.36m1 : naphtol - sulfateradio - hplc , orbitrapmefm1 = 153p : carbaryl 
 + naphtolchplc - radio / uvbafp = 42,4-dichlorophenol(cas  120 - 83 - 2)3.17 ( pka  =  7.97)m1 : 
 2,4-dichlorophenol - sulfatebradio - hplc , orbitrapmefm1 = 195m2 : not identifiedradio - hplcmefm2 = 14p : 2,4-dichlorophenolhplc - radio / uvbafp = 42,4,5-trichlorophenol(cas  95 - 95 - 4)3.72 ( pka  =  6.94)m1 : 2,4,5-trichlorophenol - sulfateradio - hplc , orbitrapmefm1 = 1454m2 : not identifiedradio - hplcmefm2 = 109p : 2,4,5-trichlorophenolhplc - radio / uvbafp = 82pentachlorophenol(cas  87 - 86 - 5)5.18 ( pka  =  4.72)m1 : not identifiedradio - hplcmefm1 = 318p : pentachlorophenolradio - hplcbafp = 2024-nitrobenzyl - chloride(cas  100 - 14 - 1)2.61m1 : not identifiedradio - hplcmefm1 = 15m2 : 4-nitrobenzyl - alcoholhplc - radio / uvmefm2 = 50m3 : not identifiedradio - hplcmefm3 = 87p : 4-nitrobenzyl - chloridehplc - radio / uvbafp = 52,4-dichloroaniline(cas  554 - 00 - 7)2.78m1 : not identifiedradio - hplcmefm1 = 58m2 : not identifiedradio - hplcmefm2 = 22p : 2,4-dichloroanilinehplc - radio / uvbafp = 29sea - nine(cas  64359 - 81 - 5)2.8m1 : not identifiedradio - hplcmefm1 = 35m2 : not identifiedradio - hplcmefm2 = 316p : sea - nineradio - hplcbafp = 272m3 : not identifiedradio - hplcmefm3 = 37aexperimental log kow values from sangster s database and from jacobson and williams for sea - nine . for 4-nitrobenzyl - chloride , no experimental value 
 was available ; thus , we calculated it with the epi suite kowwin software 
 ( http://epa.gov/oppt/exposure/pubs/episuite.htm ) .
the pka values were also taken from sangster s 
 database ( http://logkow.cisti.nrc.ca/logkow/).bsulfate must be metabolite 1 because 
 orbitrap likely detected the largest peak only.cpeaks of carbaryl and naphtol can not 
 be separated ; naphtol was likely formed during sample preparation.dethylacrylate reacts with methanol 
 during sample preparation ( spiked samples ) .
we did not observe any 
 peaks of potential biotransformation products that were distinguishable 
 from the artifact peaks of this adduct.ethe peaks of malathion and malaoxon 
 could not be clearly separated in the biotransformation kinetics experiment . 

both peaks also come closely after each other in the hplc - uv chromatogram.fbioaccumulation factors for 
 compounds 
 that were not biotransformed are taken from ref ( 11).figure 1molecular structures and label positions of compounds that are 
 not biotransformed in gammarus pulex ( a ) or where 
 biotransformation could not be quantified ( b).figure 2molecular structures , label positions , exposure concentration 
 ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for malathion , chlorpyrifos , aldicarb , carbofuran , and carbaryl .
experimental log kow values from sangster s database and from jacobson and williams for sea - nine . for 4-nitrobenzyl - chloride ,
no experimental value 
 was available ; thus , we calculated it with the epi suite kowwin software 
 ( http://epa.gov/oppt/exposure/pubs/episuite.htm ) .
peaks of carbaryl and naphtol can not 
 be separated ; naphtol was likely formed during sample preparation .
we did not observe any 
 peaks of potential biotransformation products that were distinguishable 
 from the artifact peaks of this adduct .
the peaks of malathion and malaoxon 
 could not be clearly separated in the biotransformation kinetics experiment . 

bioaccumulation factors for 
 compounds 
 that were not biotransformed are taken from ref ( 11 ) .
molecular structures and label positions of compounds that are 
 not biotransformed in gammarus pulex ( a ) or where 
 biotransformation could not be quantified ( b ) .
molecular structures , label positions , exposure concentration 
 ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for malathion , chlorpyrifos , aldicarb , carbofuran , and carbaryl .
adult g. pulex were exposed to c radio - labeled compounds 
 for 24 h and then sampled , blotted dry , weighed , and frozen in glass 
 vials at 20 c until analysis . in each experiment , four 
 organisms composed one sample ( in one vial ) and four samples per experiment 
 were taken ( 16 organisms total ) .
two of those four samples were later 
 pooled for carbofuran and malathion to raise the limit of detection , 
 because exposure to those two compounds had to be at very low concentrations 
 due to their toxicity .
experimental details can be found in the text 
 si-2 and table s1 , supporting information .
for those compounds , 
 which showed biotransformation in the metabolite screening experiments 
 ( table 1 ) , we carried out biotransformation 
 kinetics experiments ( design adapted from ref ( 22 ) , details in text si-3 , supporting information ) .
organisms were exposed 
 to c radio - labeled compounds for 24 h and then removed 
 from the test solution , rinsed , and transferred to fresh media .
eight 
 replicate beakers of 500 ml of preaerated artificial pond water and 
 one solvent control beaker each contained 10 , 15 , or 20 adult g. pulex ( see table s2 , supporting information ) , depending on the length of the depuration phase .
the length of 
 the depuration phase depended on the previously measured elimination 
 times for total radioactivity . during 
 the experiment , the concentration of the parent compound and its metabolite(s ) 
 in the organisms
one organism was sampled from each 
 beaker at each time and blotted dry , and four organisms from different 
 beakers were combined in one glass vial , weighed , and frozen at 20 
 c until analysis .
compounds with slow elimination required more sampling time points , hence , the 
 different numbers of organisms per beaker .
each beaker also contained 
 five leaf discs as food and shelter , which were renewed during the 
 depuration phase if needed .
g. pulex were sampled 
 frequently throughout the experiment ( see figures 2 to 4 and data file in supporting information for timing ) .
immediately after dosing 
 ( time 0 h ) and at each sampling time , we also sampled 1 ml of test 
 solution to quantify exposure concentrations .
initial exposure concentrations 
 were chosen so that organisms would not be strongly affected by toxic 
 effects , while still being close enough to toxic levels to be meaningful 
 for toxicity modeling purposes and maximizing 
 the limit of quantification for internal concentrations of parent 
 compounds and biotransformation products ( see text si-3 , supporting information , for more details ) .
concentrations 
 of the test compounds in the exposure medium were quantified by liquid 
 scintillation counting ( lsc , see text si-4 , supporting 
 information , for details ) .
samples were ground with a glass rod , while methanol was added 
 to extract the compounds .
then , filtered samples were concentrated 
 to about 1 ml using a genevac ( ez-2 plus , genevac , uk ) and under nitrogen 
 flow to 90 l .
subsequently , 210 l of distilled water 
 were added to obtain a ratio of 30/70 ( v / v ) methanol to water . subsequently , 

samples were split : 100 l were analyzed by lsc and another 
 100 l were analyzed by high - performance liquid chromatography 
 ( hplc , hp 1100 , agilent ) with a radio - detector ( 500 tr , packard ) to 
 quantify amounts of parent compound and metabolites ( details in text 
 si-5 , supporting information ) .
blank 
 organism samples were spiked with a known amount of parent compound 
 at the beginning of the extraction method .
comparison of the spiked 
 radioactivity with the radioactivity measured by lsc yielded recovery 
 of the extraction steps and comparison with the radioactivity measured 
 on the hplc yields overall recovery ( table s1 , supporting information ) .
known metabolites 
 as well as the parent compound were identified by spiking unlabeled 
 standard material of these to samples of control organisms during 
 the grinding step and identification of these peaks via uv - detection . 

peaks with corresponding retention times in chromatograms of the uv - detector 
 and the radio - detector were assumed to originate from identical compounds . 
 in case
no reference standards were available or biotransformation 
 products were unknown , extracts were analyzed after hplc separation 
 and electrospray ionization in either positive or negative mode with 
 high resolution mass spectrometry using a ltq - orbitrap ( thermo , waltham , 
 ma ) .
the analytical method was adopted from a method for ultratrace 
 level screening of organic micropollutants .
data - dependent ms / ms acquisition was triggered at masses of possible 
 transformation products and the most intensive masses .
screening for 
 possible metabolites was carried out either by extracting the high - resolution , 
 full - scan ms data at the exact mass of known metabolites from literature 
 or by postacquisition data processing using the metworks software 
 ( thermo , waltham , ma ) . with this software
, we predicted the masses 
 of possible metabolites produced by phase i and phase ii reactions 
 and compared high - resolution , full - scan ms data from the g. 
 pulex control and the treatment extract to identify the metabolite 
 masses that occur only in the treatment ( or only to a much higher 
 extent ) .
candidate metabolites identified by this procedure were confirmed 
 or rejected following manual extraction of the respective c - masses from full - scan chromatograms as well as interpretation of 
 ms and ms / ms spectra if available ( comparison of peak shape and relative 
 intensities of c and c peaks of parent and 
 metabolite ) .
extracts of 1,2,3-trichlorobenzene , imidacloprid , and 
 4,6-dinitro - o - cresol , for which no or minor biotransformation was 
 found , as well as malathion and chlorpyrifos , for which the known 
 oxon - metabolites could be purchased , were not analyzed with hplc - ms . 

compounds that are not subject to biotransformation 
 ( figure 1 ) can be modeled with a simple one - compartment 
 model .
the structures of the toxicokinetic 
 models for compounds that are subject to biotransformation are shown 
 in figures 24 , and metabolites are labeled m1 , m2 , and m3 .
we do not differentiate 
 between passive and active uptake and elimination mechanisms , but assume , as an approximation , 
 first - order kinetics for all toxicokinetic processes and that biotransformation 
 products were formed from parent compounds in separate , independent 
 pathways .
the model structure differs depending on the number of metabolites:1234where cinternal ,  parent(t ) is the time course of the concentration of parent 
 compound in g. pulex [ nmol kg wet  weight ] , cwater(t ) is the time course of parent compound in the medium [ nmol 
 l ] , kin_parent is 
 the uptake rate constant for parent compound [ l kg wet  weight d ] , kout_parent is the elimination rate constant for parent compound 
 [ d ] , kmet1 , kmet2 , and kmet3 are 
 the biotransformation rate constants for biotransformation of parent 
 compound to metabolites m1 , m2 , and m3 , respectively , [ d ] , cinternal ,  m1(t ) , cinternal ,  m2(t ) , and cinternal ,  m3(t ) are the time courses of the concentrations of metabolites m1 , m2 , 
 and m3 in g. pulex , respectively , [ nmol kg wet  weight ] , and kloss_m1 , kloss_m2 , and kloss_m3 are the elimination rate constants of the metabolites m1 , m2 , and 
 m3 , respectively , [ d ] .
the elimination rate constants 
 of the metabolites represent direct elimination or elimination after 
 further biotransformation . for compounds with only one or two metabolites
, 
 the model reduces to eqs 1 and 2 or eqs 1 to 3 , respectively .
model parameter values , constrained 
 to positive values , were found by least - squares minimization ( levenberg  marquardt 
 algorithm ) without weighting of data .
asymptotic standard errors of 
 parameters were derived from the covariance matrix followed by calculation 
 of confidence intervals .
all parameters 
 of a given compound were fitted simultaneously , except for carbaryl 
 and 4-nitro - benzyl - chloride where the uptake rate constant needed 
 to be kept fixed at the value from a previous study in order to achieve a good fit .
simulations and parameter 
 estimation were carried out in modelmaker ( v4.0 , cherwell scientific 
 ltd . ,
metabolite enrichment 
 factors ( mefs ) were calculated by fixing the aqueous concentration 
 of the parent at a constant concentration of 1 nmol / l and running 
 the model for 730 days , the maximum life - span of g. pulex , which is beyond the time to approach steady - state for all compounds 
 here .
thus , in steady - state , the mef 
 of metabolite i is defined as the ratio of the internal 
 concentration of the metabolite cinternal ,  mi [ nmol kg wet  weight ] to the 
 external concentration of the parent cwater ,  parent [ nmol l]:5where mefmi is the metabolite enrichment 
 factor of metabolite i , [ l kg wet  weight ] . the same method was used to calculate the 
 corresponding factors for the parent compounds , which in this case 
 are bioaccumulation factors ( bafparent ) because the organisms 
 were fed during the experiments .
the total bioaccumulation of parent 
 and metabolites is described by6where baftotal , [ l kg wet  weight ] , is equal to the apparent bioaccumulation factor 
 that is measured by quantifying internal concentrations as total c - radioactivity .
the parent compounds 4-nitrobenzyl - chloride , 
 sea - nine , ethylacrylate , and aldicarb were not detected in the lc - ms 
 analysis even after purification of the extracts by solid phase extraction . 

this can be explained by insufficient ionization , low sensitivity , 
 or interference with the g. pulex matrix and , in 
 case of 4-nitrobenzylchloride , sea - nine and ethylacrylate also by 
 rapid hydrolysis of these reactive electrophiles . since metabolites 
 usually show similar behavior during electrospray ionization and ms 
 detection as their parent compounds , we did not screen for metabolites 
 in those extracts .
furthermore , some metabolites were likely 
 not detected , because of their low molecular weight , because they 
 are not easily ionized , and because the matrix of the g. pulex extract prevented detection in the lc - ms analysis . in some cases , 

new masses were detected in g. pulex samples , but 
 these masses could not be assigned with certainty to a possible metabolite 
 of phase i or phase ii reactions .
we measured the concentrations 
 of the chemicals adsorbed to the leaf discs that were used for feeding 
 the test organisms and approximated the dietary uptake of the chemicals 
 by calculating the intake via food using the feeding rate and assimilation 
 efficiency ( see text - si 7 , supporting information , for details ) .
the assumptions needed for modeling dietary uptake , 
 such as the assimilation efficiency , are highly uncertain ; thus , the 
 dietary uptake data needs to be interpreted with caution .
the percentage 
 of dietary uptake of total uptake was less than 1% for ten out of 
 twelve compounds ( table s7 , supporting information ) .
the exceptions , carbaryl and malathion ( 8% and 9% , respectively ) , 
 can be explained with the fixed uptake rate constant for carbaryl 
 and the zero elimination rate ( leading to a relatively small uptake 
 rate ) for malathion .
overall , it appears that dietary uptake contributed 
 much less than uptake via water in our bioaccumulation experiments . 

thus , the bioaccumulation factors measured here could also be viewed 
 as bioconcentration factors .
no biotransformation 
 products were found in g. pulex for 1,2,3-trichlorobenzene 
 and imidacloprid using hplc with radiodetector ( figure 1 , table 1 ) . in one out of four samples 
 of 4,6-dinitro - o - cresol
, we detected a small peak ( 4% of total radioactivity ) , 
 which we considered negligible . in case of ethylacrylate ,
parent compound 
 and possible metabolites could not be distinguished or identified 
 because these compounds are not stable during extraction and analysis 
 ( text si-8 , supporting information ) , also 
 indicated by very low recovery ( table s1 , supporting 
 information ) .
bioaccumulation kinetics of the three nonmetabolized 
 compounds ( 1,2,3-trichlorobenzene , imidacloprid , 4,6-dinitro - o - cresol ) 
 and ethylacrylate are assumed to be well captured by toxicokinetic 
 studies based on total radioactivity .
thus , previously measured rate - constants and the corresponding one - compartment model 
 characterize toxicokinetics of these compounds .
note , that for the 
 insecticide imidacloprid this means that the very slow elimination 
 measured previously ( time for 95% elimination 
 > 11 d ) must be interpreted as accumulation of the parent and could 
 lead to carry - over toxicity from repeated pulsed exposure .
we observed oxidative desulfuration 
 of chlorpyrifos to chlorpyrifos - oxon , a common pathway in aquatic 
 organisms .
previously , we found that g. pulex biotransformed diazinon to 2-isopropyl-6-methyl-4-pyrimidinol and 
 a major metabolite of chlorpyrifos in fish and other species is 3,5,6-trichloro-2-pyridinol ; however , this can not be our m1 , because the 
 label positions of chlorpyrifos ( figure 2 ) 
 are such that 3,5,6-trichloro-2-pyridinol would not be radioactive . 

thus , it is likely that the m1 of chlorpyrifos is the hydrolyzed ester 
 of chlorpyrifos , but we could not confirm that with standards or accurate 
 mass measurements .
also , if the kinetics of malaoxon formation 
 and elimination are similar to those of chlorpyrifos - oxon ( figure 2 ) and diazoxon , then 
 the malaoxon concentrations would be close to the limit of quantification . 

hydrolysis of malathion would result in the metabolite m1 being one 
 or both of the malathion monoacids , the dicarboxylic acid or the carboxylic 
 acid esters , although further possible metabolites have been found 
 in insects and fish .
we could identify the metabolite of each 
 carbamate ( table 1 , figure 2 ) . for carbaryl , the sulfate conjugate of the transformation 
 product naphtol
was tentatively identified by the exact mass of the 
 molecular ion ( m : 0.9 ppm ) , the c - isotope ( m : 3.1 ppm ) , naphthol as specific fragment in the 
 ms / ms spectrum , and the retention time ( 1.3 min less than carbaryl ) . 

the oxidation product of carbofuran , hydroxy - carbofuran was identified 
 with radio - hplc .
oxidation of carbofuran to hydroxy - carbofuran has 
 also been found in various species , and hydrolysis 
 of carbaryl to 1-naphtol and subsequent conjugation to napthol - sulfate 
 is also a very common pathway in a range of organisms .
like hydroxy - carbofuran , 
 this is the product of a phase i reaction , whereas naphtol - sulfate 
 is the product of phase i and subsequent phase ii reaction of carbaryl .
biotransformation of the chlorinated 
 phenols resulted in sulfate conjugates after phase ii reactions directly 
 from the parent molecule ( table 1 , figure 3 ) . in case of 2,4-dichlorophenol 
 and 2,4,5-trichlorophenol but not for pcp , the sulfate conjugate could 
 be tentatively identified as metabolites in g. pulex .
identification was based on detecting the exact mass of the molecular 
 ion ( m : 0.8 ppm ( dcp ) , 1.8 ppm ( tcp ) ) , the 
 pattern of the mass spectrum with c and cl isotopes , and the occurrence of di- and trichlorophenol as a specific 
 fragment of the sulfate conjugate in the ms / ms spectrum ( m : 3 ppm ( dcp ) , 1.5 ppm ( tcp ) ) as well as a shorter retention 
 time compared to the parent compound ( 2 min ( dcp ) and 3.5 min ( tcp ) 
 less ) .
m1 is likely the pcp - sulfate , based on the retention time of 
 m1 and the fact that the two other chlorinated phenols were sulfate 
 conjugated , although it could not be identified , presumably because 
 the concentration was below the limit of detection of the orbitrap.figure 3molecular structures , 
 label positions , exposure concentration ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for 2,4-dichlorophenol , 2,4,5-trichlorophenol , and pentachlorophenol .
molecular structures , 
 label positions , exposure concentration ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for 2,4-dichlorophenol , 2,4,5-trichlorophenol , and pentachlorophenol .
sea - nine and 4-nitrobenzyl - chloride 
 had three biotransformation products each , and 2,4-dichloroaniline 
 had one ( table 1 , figure 4 ) .
we identified m2 of 
 4-nitrobenzyl - chloride as the alcohol , and on the basis of retention 
 times and possible pathways , m3 very likely is the conjugate even 
 though we could not confirm that with standards or accurate mass measurements . 

we could not identify the metabolites of sea - nine and 2,4-dichloroaniline.figure 4molecular structures , 
 label positions , exposure concentration ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for 4-nitrobenzyl - chloride , sea - nine , and 2,4-dichloroaniline .
molecular structures , 
 label positions , exposure concentration ( left ) , 
 bioaccumulation , biotransformation , and elimination kinetics ( right , 
 transfer to fresh media indicated by dashed line ) and model structure 
 ( middle ) for 4-nitrobenzyl - chloride , sea - nine , and 2,4-dichloroaniline .
we observed products of 
 phase i reactions ( aldicarb - sulfone , chlorpyrifos - oxon , hydroxy - carbofuran ) , 
 the products of phase i and subsequent phase ii reactions ( naphtol - sulfate ) 
 and products of phase ii reactions directly from the parent molecule 
 ( 2,4-dichlorophenol - sulfate , 2,4,5-trichlorophenol - sulfate , 4-nitrobenzyl - alcohol , 
 4-nitrobenzyl - conjugate , and pentachlorophenol - conjugate ) .
generally , 
 biotransformation in g. pulex appears to follow similar 
 routes as in other aquatic organisms .
conjugation 
 with sulfate was observed for three compounds ( naphtol - sulfate , 2,4-dichlorophenol - sulfate , 
 2,4,5-trichlorophenol - sulfate ) and seems to be a dominant phase ii 
 pathway in g. pulex .
quantification of aqueous concentrations yields a combined measure 
 of parent compound and any metabolite in the water .
we assumed that 
 degradation of the parent compounds in the water during the one day 
 exposure phase was negligible and that reuptake of metabolites excreted 
 by the organisms during the exposure phase is negligible because the 
 ratio of water to organisms is very large .
radioactivity in the water 
 during the depuration phase can consist of parent compound and metabolites 
 that were excreted by the organisms .
we simplify and assume that the 
 radioactivity in the water represents the parent compound and only 
 model uptake of the parent .
this assumption is justified because ( a ) 
 the radioactivity levels were very small compared to the exposure 
 phase and ( b ) we did not observe raised levels of radioactivity during 
 the depuration phase for aldicarb , carbofuran , carbaryl , and 2,4-dichloroaniline 
 ruling out the possibility for reuptake of metabolites ( figures 2 and 4 ) . slightly increasing 
 levels of radioactivity during the depuration phase
were observed 
 for malathion , chlorpyrifos , 4-nitrobenzyl - chloride , and sea - nine 
 ( figures 2 and 4 ) .
however , 
 reuptake of metabolites from water can also be neglected for these 
 compounds because the radioactivity levels were very small compared 
 to the exposure phase ( < 10% ) .
furthermore , some proportion of that 
 radioactivity can also be attributed to the parent compound because 
 measured internal concentrations indicate that a mix of parent and 
 metabolites would be excreted into the water ( figures 2 and 4 ) .
the chlorinated phenols ( figure 3 ) exhibited raised levels of radioactivity during 
 the depuration phase . for pentachlorophenol
, approximately half of 
 that radioactivity can be attributed to the parent assuming a similar 
 ratio between parent and metabolite as for the internal concentrations , 
 thus reducing the potential error by not including reuptake of metabolites 
 into model equations . for 2,4-dichlorophenol and 2,4,5-trichlorophenol , 

our method is likely to overestimate parent concentrations in the 
 water phase and thus underestimate uptake rates .
therefore , these 
 uptake rates must be considered uncertain , even more so as other factors 
 cause large confidence intervals of the uptake rate constants for 
 these two compounds ( explanation below ) . some compounds , in particular carbofuran , 
 2,4-dichlorophenol , and 2,4,5-trichlorophenol , exhibit large confidence 
 intervals for their uptake and elimination rate constants .
these large 
 confidence intervals can be explained by very fast uptake so that 
 steady - state concentrations were reached fast and experimental values 
 did not capture well the shape of the internal concentration curve . 

thus , the data contained insufficient information to narrow down the 
 confidence intervals for the corresponding model parameters .
further , 
 there is systematic uncertainty in the model because we had to assume 
 that all metabolites were directly formed from their parents .
of course , 
 alternative pathways are possible but can not be inferred from our 
 data .
the time course of internal concentrations of the parent compounds 
 and their biotransformation in g. pulex can be simulated 
 for any exposure pattern , including pulsed or fluctuating exposures 
 using the models and parameter values derived in this study ( table 2 ) .
toxicokinetics of compounds without biotransformation 
 could be simulated with the model and parameter values from our previous 
 study , because in those cases all the 
 radioactivity in the organism can be attributed to the parent compound . 

in addition , the parameter values can be used to compare among chemicals 
 to find those that are more or less susceptible to biotransformation . 

however , the parameter values must be interpreted with care , especially 
 those of carbaryl and 4-nitrobenzyl - chloride , where the uptake rate 
 constants were kept fixed to previously measured values as well as 
 the parameters of malathion , pentachlorophenol , chlorpyrifos , 2,4,5-trichlorophenol , 
 and sea - nine , where the elimination rate constant of the parent compound 
 converged to its lower boundary zero ( table 2 ) .
nevertheless , we conclude that biotransformation of xenobiotics 
 in g. pulex occurs for a wide range of compounds 
 and results in a variety of metabolites and the kinetics differ widely 
 among compounds ( first - order biotransformation rate constants range 
 from 0.13 to 44 d , table 2).table 2uptake , elimination , and biotransformation 
 rate constants with 95% confidence intervals ( ci)a  parentm1m2m3compound ( n = # of internalconcentration datapoints , 
 df = degreesof freedom)bkin_parent[95%  ci]l  kg dkout_parent[95% ci]dkmet1[95% 
 ci]dkloss_m1[95% 
 ci]dkmet2[95% 
 ci]dkloss_m2[95% 
 ci]dkmet3[95% 
 ci]dkloss_m3[95% 
 ci]daldicarb(n  =  28 ,  df  =  24)9.86[6.8 ;  12.9]9.31[5.1 ;  13.6]2.05[0.8 ;  3.3]6.19[2.3 ;  10.0 ]  carbofuran(n  =  36 ,  df  =  32)4930[0 ;  595904]7063[0 ;  878858]16.22[0 ;  32.6]0.16[0.12 ;  0.20 ]  carbaryl(n  =  40 ,  df  =  37)23.4[n.a.]c2.258[1.4 ;  3.1]3.36[2.4 ;  4.4]0.09[0.05 ;  0.13 ]  malathion(n  =  36 ,  df  =  32)8.86[4.4 ;  13.3]0[0 ;  3.1]2.99[1.1 ;  4.9]1.91[0.5 ;  3.3 ]  pentachlorophenol(n  =  40 ,  df  =  36)646[143 ;  1149]0[0 ;  5.5]3.19[0 ;  6.7]2.03[0 ;  4.6 ]  chlorpyrifos(n  =  45 ,  df  =  39)1499[1072 ;  1926]0[0 ;  2.4]3.50[1.9 ;  5.1]2.93[1.5 ;  4.4]0.132[0 ;  0.4]0.298[0 ;  1.8 ]  2,4-dichloroaniline(n  =  42 ,  df  =  36)293[82 ;  504]8.23[0.4 ;  16.0]0.949[0.6 ;  1.3]0.478[0.05 ;  0.90]0.799[0.2 ;  1.4]1.07[0 ;  2.2 ]  2,4-dichlorophenol(n  =  60 ,  df  =  54)723[0 ;  6343]153[0 ;  1478]19.62[0 ;  43]0.413[0.34 ;  0.49]3.17  [ 0 ;  8.2]0.911[0 ;  2.3 ]  2,4,5-trichlorophenol(n  =  54 ,  df  =  48)1389[0 ;  4073]0[0 ;  36]14.52[1.1 ;  28.0]0.822[0.42 ;  1.23]2.35[0 ;  9.7]1.78[0 ;  9 ]  4-nitrobenzyl - chloride(n  =  56 ,  df  =  49)576[n.a.]c31.23[0 ;  90]0.86[0 ;  4.2]0.278[0 ;  3.4]42.5[0 ;  111]4.13[2.2 ;  6.1]44.3[0 ;  115]2.46[1.8 ;  3.2]sea - nine(n  =  80 ,  df  =  72)345[295 ;  395]0[0 ;  1.1]0.349[0 ;  1.2]2.70[0 ;  9.8]0.623[0.5 ;  0.8]0.536[0.34 ;  0.73]0.299[0 ;  0.9]2.18[0 ;  7.4]aconfidence intervals extending below 
 zero were truncated at zero.bthe raw data used for modeling is 
 available as supporting information.cthe uptake rate constants kin for carbaryl and 4-nitrobenzyl - chloride were 
 kept fixed to previously measured values .
the uptake rate constants kin for carbaryl and 4-nitrobenzyl - chloride were 
 kept fixed to previously measured values .
comparison of rate constants and baftotal derived in this 
 study ( eq 6 ) with those from our previous study based on total c measurements was 
 undertaken for ten compounds where biotransformation occurred ( carbofuran 
 and 2,4-dichlorophenol were excluded because of large uncertainty ; 
 diazinon was added from another study ) .
the uptake rate is independent of biotransformation , which is 
 confirmed by good agreement between the uptake rate constants derived 
 with both experimental methods ( figure 5a , 
 excluding data for carbaryl and 4-nitrobenzyl - chloride , see supporting information ) .
the variability of 1 
 order of magnitude is due to interexperimental variability , different 
 amounts of data in the two studies , and possible influences of elimination 
 and biotransformation on the fitted uptake rates in this study because 
 all parameters were fitted simultaneously.figure 5correlation of rate constants ( top ) and 
 bioaccumulation factors 
 ( bottom ) from studies based on total c internal concentrations 
 ( y - axis ) compared to this study where metabolites 
 were measured and modeled explicitly ( x - axis ) .
correlation of rate constants ( top ) and 
 bioaccumulation factors 
 ( bottom ) from studies based on total c internal concentrations 
 ( y - axis ) compared to this study where metabolites 
 were measured and modeled explicitly ( x - axis ) .
the elimination rates ( parent compounds from 
 this study ) 
 fell into two groups ( figure 5b ) .
one group 
 contains compounds , which were biotransformed very fast and had a 
 negligible elimination rate in this study , but had very different 
 elimination rates in the total c - based study , because there the elimination rate reflects 
 elimination of the biotransformation products .
thus , total c - studies overestimate parent compound elimination , unless elimination 
 is explicitly defined as the sum of parent and biotransformation products . 

the second group contains those compounds where the elimination rate 
 differs from zero . here , parameter values from the total c - based study are consistently lower . 
 that means elimination in the total c - based study is 
 slower than in this study , because there metabolites are also counted 
 toward internal concentrations and the sum of parent and metabolites 
 in the organism declines slower than the parent alone .
when 
 apparent bioaccumulation factors from a study based on total c internal concentrations ( i.e. , baftotal values ) 
 are compared to the baftotal values from this study , i.e. , 
 the sum of bafparent and mefs for each compound , they correlate 
 and scatter around the line of unity ( figure 5c , text si-9 , supporting information ) . 
 in the study based on total c internal concentrations , 

approximating 
 bafparent values with apparent bioaccumulation factors 
 from a total c - study ( i.e. , 
 baftotal values ) consistently overestimates bafparent values , with the exception of pentachlorophenol ( figure 5d , see also table s8 and text si-9 , supporting information ) .
we calculated mefs ( eq 5 ) for the metabolites and compared them to the bafs 
 of their parent compounds ( table 1 ) . for all 
 compounds with biotransformation , except aldicarb ,
the baf of the 
 parent was smaller than the mef of one or more of its metabolites , 
 illustrating that biotransformation products generally achieve larger 
 internal concentrations in g. pulex than the parent 
 compounds .
note , however , that further metabolites may have been formed 
 with smaller bioaccumulation factors , but we would not have detected 
 those .
other studies with crustaceans also found higher internal concentrations 
 of some metabolites than the respective parent compounds , for example , 
 in daphnia magna , hyalella azteca , gammarus fossarum , and gammarus pulex . together with our result that 14 out of 19 mefs are larger 
 than the bafs of their parent compounds ,
there is enough evidence 
 to suggest that field biomonitoring studies using crustaceans may 
 also need to look for metabolites and confirms that assessment of 
 toxic effects may need to include metabolites . out of the 19 metabolites that we observed in the radio - chromatograms , 
 we could only identify seven with standards or accurate mass measurements . 

the mefs of 14 metabolites were larger than the bafs of their parents , 
 indicating that metabolites could be found more often in field biomonitoring 
 studies than their parent compounds , if they are being looked for . 

we found that biotransformation dominated toxicokinetics and strongly 
 affected internal concentrations of parent compounds and metabolites . 

many metabolites reached higher internal concentrations than their 
 parents , characterized by large metabolite enrichment factors .
thus , 
 comparing toxicokinetics across organisms and chemicals without explicit consideration 
 of biotransformation has limited explanatory power .
extensive biotransformation , 
 as observed in this study , also implies that the development of quantitative - structure 
 activity relationships can not rely on descriptors for partitioning 
 alone but must also include descriptors for susceptibility to biotransformation . 

we conclude that mechanistic understanding of differences in species 
 sensitivity or toxic potency of chemicals is more likely to succeed if biotransformation kinetics are considered .
 bioaccumulation and biotransformation are key toxicokinetic 
 processes 
 that modify toxicity of chemicals and sensitivity of organisms . 
 bioaccumulation 
 kinetics vary greatly among organisms and chemicals ; thus , we investigated 
 the influence of biotransformation kinetics on bioaccumulation in 
 a model aquatic invertebrate using fifteen 14c - labeled 
 organic xenobiotics from diverse chemical classes and physicochemical 
 properties ( 1,2,3-trichlorobenzene , imidacloprid , 4,6-dinitro - o - cresol , 
 ethylacrylate , malathion , chlorpyrifos , aldicarb , carbofuran , carbaryl , 
 2,4-dichlorophenol , 2,4,5-trichlorophenol , pentachlorophenol , 4-nitrobenzyl - chloride , 
 2,4-dichloroaniline , and sea - nine ( 4,5-dichloro-2-octyl-3-isothiazolone ) ) . 
 
 we detected and identified metabolites using hplc with uv and radio - detection 
 as well as high resolution mass spectrometry ( ltq - orbitrap ) . 
 kinetics 
 of uptake , biotransformation , and elimination of parent compounds 
 and metabolites were modeled with a first - order one - compartment model . 
 
 bioaccumulation factors were calculated for parent compounds and metabolite 
 enrichment factors for metabolites . out of 19 detected metabolites 
 , 
 we identified seven by standards or accurate mass measurements and 
 two via pathway analysis and analogies to other compounds . 
 1,2,3-trichlorobenzene , 
 imidacloprid , and 4,6-dinitro - o - cresol were not biotransformed . 
 dietary 
 uptake contributed little to overall uptake . 
 differentiation between 
 parent and metabolites increased accuracy of bioaccumulation parameters 
 compared to total 14c measurements . 
 biotransformation dominated 
 toxicokinetics and strongly affected internal concentrations of parent 
 compounds and metabolites . 
 many metabolites reached higher internal 
 concentrations than their parents , characterized by large metabolite 
 enrichment factors . 
Introduction
Experimental Methods
Data Analysis and Modeling
Results and Discussion