sexually undifferentiated juvenile sea bass ( approximately 1 year of age ) were obtained from coelho & castro lda . ( pvoa do varzim , portugal ) and were kept in 2,000-l glass fiber tanks ( density , 24 kg / m ) supplied with natural sea - water ( 20  1 ) until exposures began . fish used for studies had an average body weight of 30  10 g ( study 1 , n = 224 ) and 32  10 g ( study 2 , n = 120 ) . e2 ( 98% purity ) , ee2 ( 98% purity ) , and bpa ( 99% purity ) were purchased from sigma - aldrich , ( st . louis , mo , usa ) , and the marine salt was supplied by sera premium ( heinsberg , germany ) . we followed the principles in the guide for the care and use of laboratory animals ( institute of laboratory animal resources 1996 ) . master stock solutions of chemicals were prepared in methanol ( analytical grade ) and stored at 20c . aqueous stock solutions were regularly prepared from aliquots of master solutions and administrated to the tanks using a multi - channel peristaltic pump equipped with silicon tubing ( watson - marlow , falmouth , cornwall , uk ) . diluted solution flowed into the experimental tanks , which resulted in a theoretical complete water change every 30 hr . dosing started with the simultaneous addition of a required volume of chemical solution directly into the tanks . water flowing from tanks was filtered through activated carbon before being delivered into the municipal sewage system . seawater used in the system was artificially prepared using marine salt diluted in dechlorinated filtered municipal tap water . dissolved oxygen saturation ( > 80% ) and total ammonia concentrations ( < 0.0010.2 test vessels and tanks were constructed of glass , with a minimum of other materials in contact ( e.g. , silicon rubber tubing ) with the test solutions . exposures were carried out at 15c in artificial seawater ( 20 ) under a photoperiod of 12 hr light:12 hr dark . before exposure , animals were allowed to acclimatize for 1 week to experimental test conditions . fish were fed food pellets ( aquasoja , portugal ) 1% of body weight per day throughout the exposure period . concentrations of methanol served as the solvent control , and 100 ng / l ee2 was the positive control . we planned the concentrations of test chemicals based on range - finding studies , such that the entire effect range was covered between low and maximal effects , allowing statistical estimation of concentration  response functions for effects between 1 and 100% . each experiment was performed using at least six different concentrations , ranging from 0.04 to 3.7 nm ( 101,000 ng / l ) for e2 , 0.03 to 6.8 nm ( 102,000 ng / l ) for ee2 , and 0.04 to 7.0 m ( 101,600 g / l ) for bpa . selected concentrations of e2 ( 0.04 , 0.15 , 0.37 , and 3.7 nm ) and bpa ( 0.04 , 0.13 , and 0.44 m ) were repeated after 3 months . stock solutions of each tested chemical were prepared in methanol ( analytical grade ) and dosed to glass mixed vessels by means of a peristaltic pump at a rate of 0.10.5 ml / min ; these solutions were then mixed with the dilution water flowing to mixing vessels at a rate of 32 ml / min , resulting in 1:3211:65 dilutions . to assess whether joint effects of three estrogenic chemicals were additive , we compared the observed effects with the ca predictions for a wide range of different vtg levels . to keep the number of fish to a minimum , the  response - surface instead , we chose the  fixed - ratio  design ( faust et al . , the ratio of concentrations of individual compounds is kept constant , and only the total concentration of the mixture is varied . by using this approach a complete concentration response relationship for the mixture can be generated , which allows a comparative assessment between observed and predicted ec values for a wide range of effects . we prepared a master stock solution containing each of the chemicals at their nominal ec50 ( median effective concentrations ; equipotent ratio ) . dilutions of this mixture were then tested , such that the entire range of vtg responses between 100 and 0% was covered evenly according to the expectation of the ca model . after 2 months , we repeated three selected mixture concentrations in an independent second mixture study . the first set of water samples was collected 1 week after the chemical dosing of the tanks began ( t0 ) . after this first sampling , the fish were placed into the tanks . the second set ( t7 ) was taken after 2 weeks of dosing , and the third set ( t14 ) was taken after the third and final week , on the day that the experiment was terminated . all samples were solid - phase extracted on a dvb speedisk ( baker , deventer , the netherlands ) . for the single chemical exposures as well as the t0 sampling points of the mixture exposures , we used an isocratic hplc method ( adapted from belfroid et al . 1999 ) coupled to diode array detection ( varian model 9065 ; palo alto , ca , usa ) . separation was done using a mobile phase consisting of methanol / water ( 60/40 , vol / vol ) . for the t7 and t14 sampling points of the mixture experiments , the same method was used as for e2 and ee2 , consisting of gas chromatography coupled to ion trap detection ( gc - itd ) using a saturn 2200 ion trap detector ( varian model 9095 ) , which had the advantage of requiring a smaller sample volume because of the higher sensitivity of the method . for the analyses of e2 , ee2 , and bpa using gc - itd , we added a deuterated standard containing e2-d4 , ee2-d4 , or bpa - d6 to the samples prior to solid phase extraction ( adapted from belfroid et al . derivatization was carried out using silyl reagent before analysis . unless stated otherwise , concentration response data given in this article refer to the arithmetic mean of measured concentrations in the three periods of sampling ( t0 , t7 , t14 ) , as proposed by the organisation for economic co - operation and development ( oecd 2000 ) . fish were anesthetized and blood was extracted from the caudal vein using a heparinized syringe . plasma was obtained after blood centrifugation ( 6,000  g for 7 min at 5c ) in heparinized tubes containing phenylmethylsulfonyl fluoride ( 1 mm ) and stored at 20c until required for vtg analysis . ( n = 8/treatment ) were determined using a competitive elisa assay , as described by maans et al . sea bass vtg used for coating elisa plates and as a standard was isolated as described by maans et al . the secondary antibody ( goat anti - rabbit antibody igg ) and the tetramethyl benzidine peroxidase substrate kit were obtained from bio - rad ( hercules , ca , usa ) . the range of the standard vtg curve was 2100 ng / ml , corresponding to 80% and 20% of binding , with 50% of binding around 15 ng / ml . juvenile plasma samples were tested at a dilution of 1:10 or higher in order to place all measurements within the confidence range of standard curve ; intraassay and interassay coefficients of variation ( cvs ) were similar to those described by maans et al . we accounted for slight differences in absolute control effects between studies by standardizing absolute effect scales to relative effects : the mean vtg concentration in fish from negative - control and positive - control tanks were used as the minimum and maximum responses , respectively , in order to standardize individual vtg measurements to values between zero ( i.e. , no vtg induction ) and one ( i.e. , vtg level in positive control ) . scaling to relative effects was carried out after individual vtg effect data were log10-transformed ( i.e. , an ec50 corresponds to the concentration that produces a log10-transformed vtg induction ) , a value that is the median in relation to negative and positive controls . we performed statistical concentration  response analyses in the same way for all compounds and for the mixture by applying a best - fit procedure : various nonlinear regression models were fitted independently to the same data set , and we selected the best fit on the basis of statistical criteria . if data from repeated studies were obtained , the pooled data was used ; to account for intra- and inter - experimental variability associated with this nested data scenario , we adopted the generalized nonlinear mixed modeling approach [ see brian et al . otherwise , the regression models were fitted to the data as described by scholze et al . effect concentrations ( ecx ) were calculated from the functional inverse of the best fit - ting model . statistical uncertainties for estimated effect concentrations were expressed as 95% confidence intervals ( cis ) and approximately determined by applying the bootstrap method ( efron and tibshirani 1993 ) . additionally , we derived no observed effect concentrations ( noecs ) using the likelihood ratio test under total order restriction ( bretz and hothorn 2003 ) . calculations for statistical inference are based on absolute vtg values , and all approaches were implemented using sas statistical software ( sas 2001 ) . based on the best - fit regression functions of single compounds , we calculated expected effect concentrations for the ternary mixture in definite ratios . quantitative relations between effects of single substances and the mixture are described by the concept of ca . for a multi - component mixture of n components , it is defined by where ecxi denotes the effect concentrations of single compounds 1 to n ( i.e. , those concentrations that alone would produce the same quantitative effect x as the mixture ) , ecxmix is the mixture concentration that induces an overall effect x , and pi is the proportion of the ith component in the mixture . individual effect concentrations , ecxi , are derived from individual concentration response functions , fi , by using their inverse functional form . graphs of predicted concentration response curves were obtained by calculating numerous ecxmix values , with x varying from 1% to 99% in steps of  1% . all individual effect concentrations , ecxi , are estimates and are therefore subject to stochastic variability , which means that the predicted mixture effect concentration also has a measure of statistical uncertainty . this was achieved using the bootstrap method ( efron and tibshirani 1993 ) , which enabled 95% cis to be derived for predicted mean effect . predicted and observed effect concentrations were deemed to be statistically significantly different when the 95% ci of predicted and observed effects did not overlap . the analytical determination of each mixture concentration often necessitated adjustments of the nominal mixture ratios , with relative proportions , pi , of the compounds being different for each measured mixture exposure . this still enabled the calculation of expected mixture concentrations because all the information required is available as demanded by equation 1 . in a strict quantitative sense , however , this can be done only for the tested mixture concentrations ; because extrapolation methods are unsuitable , the mixture data can not be figured out using the same concentration scale . however , if the variation of the individual mixture ratios follows a random process , then smoothing techniques can be applied in order to estimate an average  mixture ratio , which also enables the usual concentration  response relationships to be constructed for the mixture . we determined the mean mixture ratio by calculating the average fraction for each compound within the total mixture using data for the six highest tested mixture concentrations . sexually undifferentiated juvenile sea bass ( approximately 1 year of age ) were obtained from coelho & castro lda . ( pvoa do varzim , portugal ) and were kept in 2,000-l glass fiber tanks ( density , 24 kg / m ) supplied with natural sea - water ( 20  1 ) until exposures began . fish used for studies had an average body weight of 30  10 g ( study 1 , n = 224 ) and 32  10 g ( study 2 , n = 120 ) . e2 ( 98% purity ) , ee2 ( 98% purity ) , and bpa ( 99% purity ) were purchased from sigma - aldrich , ( st . louis , mo , usa ) , and the marine salt was supplied by sera premium ( heinsberg , germany ) . we followed the principles in the guide for the care and use of laboratory animals ( institute of laboratory animal resources 1996 ) . master stock solutions of chemicals were prepared in methanol ( analytical grade ) and stored at 20c . aqueous stock solutions were regularly prepared from aliquots of master solutions and administrated to the tanks using a multi - channel peristaltic pump equipped with silicon tubing ( watson - marlow , falmouth , cornwall , uk ) . diluted solution flowed into the experimental tanks , which resulted in a theoretical complete water change every 30 hr . dosing started with the simultaneous addition of a required volume of chemical solution directly into the tanks . water flowing from tanks was filtered through activated carbon before being delivered into the municipal sewage system . seawater used in the system was artificially prepared using marine salt diluted in dechlorinated filtered municipal tap water . test vessels and tanks were constructed of glass , with a minimum of other materials in contact ( e.g. , silicon rubber tubing ) with the test solutions . exposures were carried out at 15c in artificial seawater ( 20 ) under a photoperiod of 12 hr light:12 hr dark . before exposure , animals were allowed to acclimatize for 1 week to experimental test conditions . fish were fed food pellets ( aquasoja , portugal ) 1% of body weight per day throughout the exposure period . concentrations of methanol served as the solvent control , and 100 ng / l ee2 was the positive control . we planned the concentrations of test chemicals based on range - finding studies , such that the entire effect range was covered between low and maximal effects , allowing statistical estimation of concentration  response functions for effects between 1 and 100% . each experiment was performed using at least six different concentrations , ranging from 0.04 to 3.7 nm ( 101,000 ng / l ) for e2 , 0.03 to 6.8 nm ( 102,000 ng / l ) for ee2 , and 0.04 to 7.0 m ( 101,600 g / l ) for bpa . selected concentrations of e2 ( 0.04 , 0.15 , 0.37 , and 3.7 nm ) and bpa ( 0.04 , 0.13 , and 0.44 m ) were repeated after 3 months . stock solutions of each tested chemical were prepared in methanol ( analytical grade ) and dosed to glass mixed vessels by means of a peristaltic pump at a rate of 0.10.5 ml / min ; these solutions were then mixed with the dilution water flowing to mixing vessels at a rate of 32 ml / min , resulting in 1:3211:65 dilutions . to assess whether joint effects of three estrogenic chemicals were additive , we compared the observed effects with the ca predictions for a wide range of different vtg levels . to keep the number of fish to a minimum , the  response - surface instead , we chose the  fixed - ratio  design ( faust et al . the ratio of concentrations of individual compounds is kept constant , and only the total concentration of the mixture is varied . by using this approach a complete concentration response relationship for the mixture can be generated , which allows a comparative assessment between observed and predicted ec values for a wide range of effects . we prepared a master stock solution containing each of the chemicals at their nominal ec50 ( median effective concentrations ; equipotent ratio ) . dilutions of this mixture were then tested , such that the entire range of vtg responses between 100 and 0% was covered evenly according to the expectation of the ca model . after 2 months , we repeated three selected mixture concentrations in an independent second mixture study . we planned the concentrations of test chemicals based on range - finding studies , such that the entire effect range was covered between low and maximal effects , allowing statistical estimation of concentration  response functions for effects between 1 and 100% . each experiment was performed using at least six different concentrations , ranging from 0.04 to 3.7 nm ( 101,000 ng / l ) for e2 , 0.03 to 6.8 nm ( 102,000 ng / l ) for ee2 , and 0.04 to 7.0 m ( 101,600 g / l ) for bpa . selected concentrations of e2 ( 0.04 , 0.15 , 0.37 , and 3.7 nm ) and bpa ( 0.04 , 0.13 , and 0.44 m ) were repeated after 3 months . stock solutions of each tested chemical were prepared in methanol ( analytical grade ) and dosed to glass mixed vessels by means of a peristaltic pump at a rate of 0.10.5 ml / min ; these solutions were then mixed with the dilution water flowing to mixing vessels at a rate of 32 ml / min , resulting in 1:3211:65 dilutions . to assess whether joint effects of three estrogenic chemicals were additive , we compared the observed effects with the ca predictions for a wide range of different vtg levels . to keep the number of fish to a minimum , the  response - surface instead , we chose the  fixed - ratio  design ( faust et al . the ratio of concentrations of individual compounds is kept constant , and only the total concentration of the mixture is varied . by using this approach a complete concentration response relationship for the mixture can be generated , which allows a comparative assessment between observed and predicted ec values for a wide range of effects . we prepared a master stock solution containing each of the chemicals at their nominal ec50 ( median effective concentrations ; equipotent ratio ) . dilutions of this mixture were then tested , such that the entire range of vtg responses between 100 and 0% was covered evenly according to the expectation of the ca model . after 2 months , we repeated three selected mixture concentrations in an independent second mixture study . the first set of water samples was collected 1 week after the chemical dosing of the tanks began ( t0 ) . after this first sampling , the fish were placed into the tanks . the second set ( t7 ) was taken after 2 weeks of dosing , and the third set ( t14 ) was taken after the third and final week , on the day that the experiment was terminated . all samples were solid - phase extracted on a dvb speedisk ( baker , deventer , the netherlands ) . for the single chemical exposures as well as the t0 sampling points of the mixture exposures , we used an isocratic hplc method ( adapted from belfroid et al . 1999 ) coupled to diode array detection ( varian model 9065 ; palo alto , ca , usa ) . separation was done using a mobile phase consisting of methanol / water ( 60/40 , vol / vol ) . for the t7 and t14 sampling points of the mixture experiments , the same method was used as for e2 and ee2 , consisting of gas chromatography coupled to ion trap detection ( gc - itd ) using a saturn 2200 ion trap detector ( varian model 9095 ) , which had the advantage of requiring a smaller sample volume because of the higher sensitivity of the method . for the analyses of e2 , ee2 , and bpa using gc - itd , we added a deuterated standard containing e2-d4 , ee2-d4 , or bpa - d6 to the samples prior to solid phase extraction ( adapted from belfroid et al . 1999 and houtman et al . 2004 ) . following the cleanup of the extracts with c18 cartridges , derivatization was carried out using silyl reagent before analysis . unless stated otherwise , concentration response data given in this article refer to the arithmetic mean of measured concentrations in the three periods of sampling ( t0 , t7 , t14 ) , as proposed by the organisation for economic co - operation and development ( oecd 2000 ) . fish were anesthetized and blood was extracted from the caudal vein using a heparinized syringe . plasma was obtained after blood centrifugation ( 6,000  g for 7 min at 5c ) in heparinized tubes containing phenylmethylsulfonyl fluoride ( 1 mm ) and stored at 20c until required for vtg analysis . ( n = 8/treatment ) were determined using a competitive elisa assay , as described by maans et al . sea bass vtg used for coating elisa plates and as a standard was isolated as described by maans et al . the antibody against sea bass vtg was raised in rabbits . the secondary antibody ( goat anti - rabbit antibody igg ) and the tetramethyl benzidine peroxidase substrate kit were obtained from bio - rad ( hercules , ca , usa ) . the range of the standard vtg curve was 2100 ng / ml , corresponding to 80% and 20% of binding , with 50% of binding around 15 ng / ml . juvenile plasma samples were tested at a dilution of 1:10 or higher in order to place all measurements within the confidence range of standard curve ; intraassay and interassay coefficients of variation ( cvs ) were similar to those described by maans et al . we accounted for slight differences in absolute control effects between studies by standardizing absolute effect scales to relative effects : the mean vtg concentration in fish from negative - control and positive - control tanks were used as the minimum and maximum responses , respectively , in order to standardize individual vtg measurements to values between zero ( i.e. , no vtg induction ) and one ( i.e. , vtg level in positive control ) . scaling to relative effects was carried out after individual vtg effect data were log10-transformed ( i.e. , an ec50 corresponds to the concentration that produces a log10-transformed vtg induction ) , a value that is the median in relation to negative and positive controls . we performed statistical concentration  response analyses in the same way for all compounds and for the mixture by applying a best - fit procedure : various nonlinear regression models were fitted independently to the same data set , and we selected the best fit on the basis of statistical criteria . if data from repeated studies were obtained , the pooled data was used ; to account for intra- and inter - experimental variability associated with this nested data scenario , we adopted the generalized nonlinear mixed modeling approach [ see brian et al . otherwise , the regression models were fitted to the data as described by scholze et al . effect concentrations ( ecx ) were calculated from the functional inverse of the best fit - ting model . statistical uncertainties for estimated effect concentrations were expressed as 95% confidence intervals ( cis ) and approximately determined by applying the bootstrap method ( efron and tibshirani 1993 ) . additionally , we derived no observed effect concentrations ( noecs ) using the likelihood ratio test under total order restriction ( bretz and hothorn 2003 ) . calculations for statistical inference are based on absolute vtg values , and all approaches were implemented using sas statistical software ( sas 2001 ) . based on the best - fit regression functions of single compounds , we calculated expected effect concentrations for the ternary mixture in definite ratios . quantitative relations between effects of single substances and the mixture are described by the concept of ca . for a multi - component mixture of n components , it is defined by where ecxi denotes the effect concentrations of single compounds 1 to n ( i.e. , those concentrations that alone would produce the same quantitative effect x as the mixture ) , ecxmix is the mixture concentration that induces an overall effect x , and pi is the proportion of the ith component in the mixture . individual effect concentrations , ecxi , are derived from individual concentration response functions , fi , by using their inverse functional form . graphs of predicted concentration response curves were obtained by calculating numerous ecxmix values , with x varying from 1% to 99% in steps of  1% . all individual effect concentrations , ecxi , are estimates and are therefore subject to stochastic variability , which means that the predicted mixture effect concentration also has a measure of statistical uncertainty . this was achieved using the bootstrap method ( efron and tibshirani 1993 ) , which enabled 95% cis to be derived for predicted mean effect . predicted and observed effect concentrations were deemed to be statistically significantly different when the 95% ci of predicted and observed effects did not overlap . the analytical determination of each mixture concentration often necessitated adjustments of the nominal mixture ratios , with relative proportions , pi , of the compounds being different for each measured mixture exposure . this still enabled the calculation of expected mixture concentrations because all the information required is available as demanded by equation 1 . in a strict quantitative sense , however , this can be done only for the tested mixture concentrations ; because extrapolation methods are unsuitable , the mixture data can not be figured out using the same concentration scale . however , if the variation of the individual mixture ratios follows a random process , then smoothing techniques can be applied in order to estimate an average  mixture ratio , which also enables the usual concentration  response relationships to be constructed for the mixture . we determined the mean mixture ratio by calculating the average fraction for each compound within the total mixture using data for the six highest tested mixture concentrations . analytical data ascertaining nominal concentrations for single exposures and for the mixture are given in table 1 . for the sake of simplicity , only average values of data from t0 , t7 , and t14 from the first mixture testing are shown ( the second mixture study yielded nearly identical mixture ratios ) . the average variation of the individual bpa measurements between the sampling days was relatively low ( cv = 18% ) , and the exposures were constant over the testing period . we found relatively good agreement between nominal and measured concentrations , with an average recovery rate of around 95% for single exposures and for within the mixture . for single ee2 and e2 exposures , data variation between the sampling days was higher , with an average cv of 70% for e2 and 80% for ee2 , and 27% for all positive controls ( 100 ng / l ee2 ) . this higher variation mainly occurs because recovery rates at t0 ( 90% ) were higher , but they decreased after 1 week of testing to around 2040% , resulting in an overall mean recovery rate of 47.0% for ee2 and 46.7% for e2 these findings were confirmed by repeated studies ; thus , we exclude laboratory accident as the reason for these rates . within the mixture , the recovery rates for ee2 were slightly higher , with an overall average recovery of 55.2% . the mixture ratio was originally conceived to be proportional to nominal ec50 values of the individual compounds . because of the lower recovery rates for both steroids , it is obvious that the mixture ratio differed when based on measured concentrations . as a result , the fraction of bpa in the mixture was higher than planned , with an average 99.76% instead of 99.52% of the total concentration ( table 2 ) . for the effective exposure ranges of the mixture ( 3.8438.4 mg / l , nominal ) , recovery rates for all three compounds were relatively stable and of low variation , indicating that the corresponding mixture ratios for each mixture concentration were nearly identical . thus , the use of a common average mixture ratio for all mixture exposures appears justified , and observed and predicted mixture effects can be compared for untested concentration ranges by interpolation . figure 1 shows concentration  response data for each tested chemical and their best - fit regression curves . the corresponding functions , model parameters , and statistical estimates of estrogenic potencies are given in table 2 . because of the relatively low data variability ( the cv was 2.55% for the negative controls and 2.57% for the positive controls ) , the statistical power was sufficiently large to detect responses down to 5% ; this allowed precise regression estimations of mean vtg responses . the effect concentrations for median and low vtg induction ( ec50 , ec10 ) are given in table 2 ; the corresponding 95% cis are relatively low for all compounds . using the median effect level as reference , we found that ee2 was the most potent steroid tested , with an ec50 of 0.029 g / l , 3.6 times more potent than e2 ( ec50 = 0.104 g / l ) . as expected , bpa was the least potent of the chemicals , with an ec50 of 77.94 all tested bpa concentrations produced statistically significant responses , with 10 g / l being the lowest tested exposure yielding a vtg induction of around 10% ( ec10 the noecs derived for e2 and ee2 were equal to the lowest tested concentrations and were lower than the respective ec10 . a positive control of 100 ng ee2/l was expected to produce the maximal possible vtg induction in our system . this assumption was based mainly on outcomes from previous range - finding studies ; furthermore , this dose is commonly used as positive control in flow - through systems with fish ( e.g. , brian et al . however , our studies demonstrated clearly that higher ee2 concentrations are able to induce stronger effects , with a maximal induction of around 3  10 ng vtg / ml , 6 times higher than observed for 100 ng ee2/l . consequently , our assumed maximal control reference represents a slight underestimate , with a corresponding impact on every effect and effect concentration estimation . however , because we used the log10-transformed vtg scale for the assessment , the differences were negligible and did not change the statistical estimates . for example , the estimated ec50 for ee2 might in reality refer not to a 50% effect , but to 49.3% . to guarantee comparability between studies , the mixture was prepared from a master stock solution containing e2 , ee2 , and bpa at a ratio of their nominal ec50 values ( table 2 ) ; dilutions of 100 , 66 , 33 , 10 , 6.7 , 3.3 , and 1% were tested , corresponding to expected vtg responses between 100% and 0% . this experiment was repeated for three mixture concentrations ( corresponding dilutions of 66 , 33 , and 10% ) , and the vtg data for both studies agreed excellently ( figure 2 ) . figure 3 shows the individual and mean vtg responses from the first mixture study normalized to the negative and positive controls , and based on measured concentrations . there was excellent agreement with the ca - predicted concentration  response curve , which was generated based on the concentration  response functions shown in table 2 . we did not detect a statistically significant deviation from predictions for either of the selected effect levels ( table 3 ) . when we used the exact mixture ratios measured for each mixture concentration , the resulting mixture effects were almost identical to those in figure 2 and did not fall outside the 95% ci of the prediction curve ( figure 3 ) . figure 4 shows a comparison of the observed mixture effects with the expected effects of the individual compounds that would occur if they were present alone at their respective levels in the mixture . all individual curves are based on the regression fits of the compounds tested alone ( table 2 ) . figure 4 shows that a single compound alone can not explain the observed mixture effect . furthermore , figure 4 shows that concentrations of the compounds without statistically significant effects can still produce a detectable mixture effect when they are present together . for example , the comparison between the median vtg response for the nominal 19.2 mg / l mixture and the corresponding individual effects of the compounds demonstrates how strongly low - steroidal exposures can enhance a weak vtg induction by bpa ; for bpa alone , we observed a 20% vtg response , but the presence of 12 ng / l ee2 and 33 ng / l e2 increased the vtg induction to 50% . analytical data ascertaining nominal concentrations for single exposures and for the mixture are given in table 1 . for the sake of simplicity , only average values of data from t0 , t7 , and t14 from the first mixture testing are shown ( the second mixture study yielded nearly identical mixture ratios ) . the average variation of the individual bpa measurements between the sampling days was relatively low ( cv = 18% ) , and the exposures were constant over the testing period . we found relatively good agreement between nominal and measured concentrations , with an average recovery rate of around 95% for single exposures and for within the mixture . for single ee2 and e2 exposures , data variation between the sampling days was higher , with an average cv of 70% for e2 and 80% for ee2 , and 27% for all positive controls ( 100 ng / l ee2 ) . this higher variation mainly occurs because recovery rates at t0 ( 90% ) were higher , but they decreased after 1 week of testing to around 2040% , resulting in an overall mean recovery rate of 47.0% for ee2 and 46.7% for e2 these findings were confirmed by repeated studies ; thus , we exclude laboratory accident as the reason for these rates . within the mixture , the recovery rates for ee2 were slightly higher , with an overall average recovery of 55.2% . the mixture ratio was originally conceived to be proportional to nominal ec50 values of the individual compounds . because of the lower recovery rates for both steroids , it is obvious that the mixture ratio differed when based on measured concentrations . as a result , the fraction of bpa in the mixture was higher than planned , with an average 99.76% instead of 99.52% of the total concentration ( table 2 ) . for the effective exposure ranges of the mixture ( 3.8438.4 mg / l , nominal ) , recovery rates for all three compounds were relatively stable and of low variation , indicating that the corresponding mixture ratios for each mixture concentration were nearly identical . thus , the use of a common average mixture ratio for all mixture exposures appears justified , and observed and predicted mixture effects can be compared for untested concentration ranges by interpolation . figure 1 shows concentration  response data for each tested chemical and their best - fit regression curves . the corresponding functions , model parameters , and statistical estimates of estrogenic potencies are given in table 2 . because of the relatively low data variability ( the cv was 2.55% for the negative controls and 2.57% for the positive controls ) , the statistical power was sufficiently large to detect responses down to 5% ; this allowed precise regression estimations of mean vtg responses . the effect concentrations for median and low vtg induction ( ec50 , ec10 ) are given in table 2 ; the corresponding 95% cis are relatively low for all compounds . using the median effect level as reference , we found that ee2 was the most potent steroid tested , with an ec50 of 0.029 g / l , 3.6 times more potent than e2 ( ec50 = 0.104 g / l ) . as expected , bpa was the least potent of the chemicals , with an ec50 of 77.94 g / l being the lowest tested exposure yielding a vtg induction of around 10% ( ec10 = 9.12 the noecs derived for e2 and ee2 were equal to the lowest tested concentrations and were lower than the respective ec10 . a positive control of 100 ng ee2/l was expected to produce the maximal possible vtg induction in our system . this assumption was based mainly on outcomes from previous range - finding studies ; furthermore , this dose is commonly used as positive control in flow - through systems with fish ( e.g. , brian et al . 2005 ) . however , our studies demonstrated clearly that higher ee2 concentrations are able to induce stronger effects , with a maximal induction of around 3  10 ng vtg / ml , 6 times higher than observed for 100 ng ee2/l . consequently , our assumed maximal control reference represents a slight underestimate , with a corresponding impact on every effect and effect concentration estimation . however , because we used the log10-transformed vtg scale for the assessment , the differences were negligible and did not change the statistical estimates . for example , the estimated ec50 for ee2 might in reality refer not to a 50% effect , but to 49.3% . to guarantee comparability between studies , we did not change the positive control concentration . the mixture was prepared from a master stock solution containing e2 , ee2 , and bpa at a ratio of their nominal ec50 values ( table 2 ) ; dilutions of 100 , 66 , 33 , 10 , 6.7 , 3.3 , and 1% were tested , corresponding to expected vtg responses between 100% and 0% . this experiment was repeated for three mixture concentrations ( corresponding dilutions of 66 , 33 , and 10% ) , and the vtg data for both studies agreed excellently ( figure 2 ) . figure 3 shows the individual and mean vtg responses from the first mixture study normalized to the negative and positive controls , and based on measured concentrations . response curve , which was generated based on the concentration  response functions shown in table 2 . we did not detect a statistically significant deviation from predictions for either of the selected effect levels ( table 3 ) . the predicted ec values were based on a common , average mixture ratio . when we used the exact mixture ratios measured for each mixture concentration , the resulting mixture effects were almost identical to those in figure 2 and did not fall outside the 95% ci of the prediction curve ( figure 3 ) . figure 4 shows a comparison of the observed mixture effects with the expected effects of the individual compounds that would occur if they were present alone at their respective levels in the mixture . all individual curves are based on the regression fits of the compounds tested alone ( table 2 ) . figure 4 shows that a single compound alone can not explain the observed mixture effect . furthermore , figure 4 shows that concentrations of the compounds without statistically significant effects can still produce a detectable mixture effect when they are present together . for example , the comparison between the median vtg response for the nominal 19.2 mg / l mixture and the corresponding individual effects of the compounds demonstrates how strongly low - steroidal exposures can enhance a weak vtg induction by bpa ; for bpa alone , we observed a 20% vtg response , but the presence of 12 ng / l ee2 and 33 ng / l e2 increased the vtg induction to 50% . current knowledge about the sensitivity of marine fish to estrogenic environmental chemicals is still limited , and our study fills these gaps by providing data about vtg induction in sea bass . the vtg responses observed here are comparable to the sensitivity rank orders reported for freshwater species : in our study , ee2 was the most potent inducer of vtg , an order that was also found in rainbow trout and zebra fish , in which the lowest observed effect concentration ( loec ) for ee2 was 4-fold lower than that for e2 ( 20 however , for fathead minnow and rainbow trout , up to 25-fold higher potency differences between ee2 and e2 have been reported ( brian et al . 2005 ; thorpe et al . the estimated ec10 values for vtg induction , 17 ng / l for ee2 and 31 ng / l for e2 , were of the same magnitude as values shown to induce endocrine disruption in freshwater and seawater fish ( e.g. , folmar et al . exposure of 6 ng / l ee2 to sand goby induced impaired male maturation and reproductive behavior ( robinson et al . 2003 ) . ( 2001 ) reported a lack of sexual differentiation in the male fathead minnow after exposure to 4 ng ee2/l . the sensitivity of the sea bass to bpa ( ec10 = 9 g / l ) is comparable to that of the fathead minnow ( ec10 = 50 g / l ) ( sohoni et al . 2001 ) , but it seems to be higher in comparison to other freshwater species ; that is , the ec10 was about two orders of magnitude lower than the loec for juvenile rainbow trout ( loec = 1,000 g / l ) , and the ec50 ( 78 g / l ) was nearly 2-fold lower than for fathead minnow ( ec50 = 158 g / l ) ( van den belt et al . it is unclear whether this is due to the high stability of the compound in our test system or to differences in metabolism . the relevance of exposure to e2 , ee2 , and bpa in estuaries and coastal zones is increasingly recognized . this is not confined only to surface waters , but also occurs in sediments and marine biota . bpa can be found in sea - water at concentrations of up to 2 g / l and in seafood at levels of up to 213 g / g ( basheer et al . 2004 ) . because both steroidal estrogens are poorly removed in coastal treatment plants ( braga et al . 2005a ) , they could be detected in sediments in the proximity of a coastal primary sewage treatment plant in ranges of 0.222.48 ng / g ( e2 ) and up to 0.5 ng / g ( ee2 ) ( braga et al . however , it is more likely that the measured environmental exposures , at which we observed significant vtg responses in juvenile sea bass under controlled experimental conditions , reflect the levels encountered at highly exposed point sources rather than the average diffuse environmental burden , which is likely to be much lower . although a realistic exposure assessment for edcs in the marine environment is still missing , it is conceivable that even very low concentrations of these compounds can contribute to an overall significant mixture effect , as confirmed in this and other mixture studies ( brian et al . 2005 ; silva et al . thus , it can not be excluded a priori that the complex mixture scenario of the edcs in the marine environment contributes significantly to adverse health effects in fish . if a sufficient number of edcs is present , theoretically significant vtg inductions can occur when the individual compounds are below the technical detection limit . because of the difficulties of trace analysis in seawater matrix , the detection limits for estrogenic chemicals are quite high each tank ( 30 l ) was supplied with a relatively low water exchange rate ( 30 l / day ) , mainly to minimize the relatively high costs of artificial seawater . because of these testing limitations , we detected strong differences between nominal and measured concentrations , in particular for e2 and ee2 , where the measured exposures were remarkably lower at the middle and the end of the study ( t7 , t14 ) than shortly before the placement of fish ( t0 ) . this can be explained by active microbial degradation ( methanol solvent influence ) , adsorption to various surfaces such as organic matter ( fish feces ) in the water and glassware , or the uptake of the chemicals by fish . in contrast , bpa was very stable in seawater ( 95.2% average recovery ) , which is in good agreement with the recent findings of kang and kondo ( 2005 ) . in theory , measured concentrations of the test chemicals in the water should provide quite accurate reflections of the actual exposure conditions . however , it is not clear whether analytical data give a valid estimation for exposures that are responsible for the observed vtg induction after 14 days . even when recovery rates are extremely low , we can not always assume that the fish were exposed to small amounts . if we exclude the possibility that technical issues were responsible for a lower chemical in flux ( which was proven not to be the case in our study ) , then low recovery rates could be the result of large uptake by fish , leading to low concentrations . in this case the resulting uncertainties in terms of effective concentrations are difficult to model statistically , and the use of arithmetic means of all measured concentrations may be a poor reflection of biologically active concentrations . however , because of the unknown complex relationship between intake , uptake , technical testing environment , and the observed biological effect at the end of the study , more valid approaches are not available . this general uncertainty explains some of the differences in potency when concentration  response results for the same compound and species are compared from laboratories using flow - through systems with different flow rates . the question arises whether these technical issues influence the outcomes of mixture studies , especially when they are aiming to investigate the predictability of mixture effects . ( 2001 , 2003 ) did not demonstrate relevant differences between nominal and measured exposures over the whole exposure period , both for the single exposures and mixtures . thus , the experimental design of their mixture studies put the assessment of the predictive power of the ca model on a sound footing . the high replacement rates of the test chemicals , together with relatively large tanks and small fish , were largely responsible for the small variations between nominal and actual concentrations in these studies . if the recovery patterns between nominal and measured concentrations are reproducible for each compound over the whole testing period , then the important prerequisite of comparability for the ca model is fulfilled and the observed and predicted mixture effects allow a comparative assessment . moreover , the relative relationship between observed and predicted effects for the mixture remains the same when using nominal instead of measured concentrations , and thus a comparative assessment based on nominal concentrations will come to the same conclusions as when based on measured concentrations . in the present study , we detected higher recovery rates of ee2 in the mixture than we observed in the single studies , which were confirmed by the second mixture study . thus , the reproducibility assumption was violated , and the comparison between observed and predicted mixture effects could only be done on measured concentrations . the detailed information about actual concentrations of the components in the mixture was essential . a mixture analysis based on the fixed - ratio design can be problematic when the mixture composition varies significantly between the tested mixture concentrations ( e.g. , when the recovery rate for one individual component is not constant ) . the observed effect data are not comparable in the sense that they can be analyzed on a common concentration scale , consequently , statistical concentration  effect regression approaches can not be used for the generation of a fitting curve . thus , observed and predicted mixture effects however , it might be possible to model the variable mixture compositions in function of the single measured exposures ( variable mixture ratio ) . the present study highlights the potential hazard of joint exposure to steroidal estrogens ( e2 and ee2 ) and bpa to marine fish . we have shown that the pharmacologic concept of ca very accurately describes the vtg induction of estrogenic chemicals in marine fish . this is in good agreement with the outcomes from previous multicomponent mixture studies in freshwater fish ( brian et al . taken altogether , these studies provide sufficient evidence that edcs act dose - additively in inducing vtg , independent of the fish species . if information is available about each compound present in the mixture and if exposure patterns and the effects are reproducible , then ca can be used as a tool to predict accurately the joint effects of edcs . because the effect assessment for edcs is receiving more and more attention , a proper exposure assessment remains important for assessing the risk . one factor that limits progress in risk assessments for edc mixtures in marine life is a lack of information on relevant exposure levels . monitoring programs for most of the known edcs are not implemented for the marine environment . moreover , it is unknown how many edcs are in the field , although new exposure assessment tools , such as the adaptation of the toxicity identification and evaluation for edcs , might be a promising solution for the future .