Environ. Sci. Technol. 1997, 31, 3426-3431
Bioconcentration of LAS: Experimental Determination and Extrapolation to Environmental Mixtures JOHANNES TOLLS,* MANUELA HALLER, ILJA DE GRAAF, MAARTJE A. T. C. THIJSSEN, AND DICK T. H. M. SIJM Environmental Chemistry, Research Institute of Toxicology, Utrecht University, Padualaan 8, 3584 CH Utrecht, The Netherlands
Linear alkylbenzenesulfonate (LAS) is the most widely used synthetic surfactant. Environmental risk assessment of LAS requires quantitative information on the bioconcentration properties which, as yet, are unavailable. Here, we present compound- and isomer-specific bioconcentration data for n-(p-sulfophenyl)alkanes, the constituents of LAS, determined in flow-through experiments. BCF values ranged between 2 and 1000 L kg-1. We derived bioconcentrationhydrophobicity relationships and observed that bioconcentration factors as well as uptake rate constants increase with increasing log KOW (estimated) while the elimination rate constants do not vary with log KOW. In an attempt to account for the variability of the composition of LAS mixtures, we found the average length of the alkyl chain of the LAS mixture unsuitable for estimation of the bioconcentration potential. In contrast, an estimation based on the composition of the mixture of LAS and the bioconcentration factors of the individual constituents overpredicts measured values by less than a factor of 1.5. Using the composition of two LAS mixtures, one typical for LAS in detergents and one representing LAS in rivers, we calculated the respective bioconcentration potentials to be 91 and 22 L kg-1. This indicates that environmental processes decrease the bioconcentration potential of LAS mixtures.
Introduction LAS (linear alkylbenzenesulfonate) is the most important synthetic surfactant with an annual production of 1.8 Mt worldwide in 1987 (1), being used in industrial as well as domestic, primarily laundry detergents. Chemically, LAS is a mixture of n-(p-sulfophenyl)alkanes (n-p-SPA), with alkyl chain lengths ranging typically between 10 and 13 carbon atoms (Figure 1) and the alkyl chain being substituted with the benzenesulfonate moiety in all but the terminal carbon atom. Due to the use pattern, it is discharged into the wastewater (2). During wastewater treatment, a large portion of the LAS load is removed by biodegradation as well as by sorption to the biomass (3). Hence, the amounts discharged into surface waters are largely determined by the efficiency of wastewater treatment (4). Aquatic life will be exposed to LAS in the environment since even highly effective wastewater treatment cannot * Corresponding author telephone: 0031-30-2532578; fax: 003130-2532837; e-mail:
[email protected].
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FIGURE 1. Structures of 2- and 5-n-(p-sulfophenyl)dodecane, C12-2and C12-5-LAS, respectively, two constituents of LAS used as model compounds in this study. prevent LAS from being introduced into surface waters via the effluents of wastewater treatment facilities (3-6). Therefore, the question arises whether LAS poses a risk to aquatic organisms. One aspect of that question is whether LAS has the potential to bioaccumulate in organisms. It is generally believed that a compound must have a significant potential to be bioconcentrated from the water in order to be bioaccumulated via the food web (7, 8). Therefore, the investigation of the bioconcentration potential is part of the assessment of the bioaccumulation potential of a compound. While the intrinsic toxicity of LAS has been intensively investigated in the past (9), we found in a recent review that the information on bioconcentration is scarce (10). In addition, the reported bioconcentration studies were performed with radiolabeled LAS. Since LAS is biotransformed in fish (11-13) and since biotransformation products were not separated from the intact LAS, the hitherto available data do not allow for a quantitative description of the bioconcentration of LAS. In this study, we investigated the bioconcentration behavior of individual n-p-SPA in flow-through bioconcentration experiments. In the series of test compounds, alkyl chain length as well as the substitution position of the p-sulfophenyl moiety were systematically varied to allow for studying the influence of structure on the bioconcentration behavior. LAS measurements were performed by HPLC after extraction of the test compounds from fish and water. LAS bioconcentration was made comparable to other compounds by deriving relationships between the bioconcentration factor and the rate constants for uptake and elimination of LAS on one hand and estimated values of log KOW (14) on the other. The composition of LAS mixtures is altered by the processes they undergo from discharge into wastewater until entering surface water (5, 6). As a result, the bioconcentration potential of the mixture is also likely to change. In order to account for the variability of LAS mixtures in risk assessment, we attempted to develop a method to estimate the bioconcentration potential of LAS mixtures from the composition of the mixture and the bioconcentration factors of its constituents.
Materials and Methods Chemicals. Salts and sorbents were of analytical grade. The solvents were of HPLC grade or residue analyzed and used without purification except for hexane, which was redistilled
S0013-936X(97)00100-4 CCC: $14.00
1997 American Chemical Society
prior to use. Syringes (20 mL, Becton and Dickinson) used as columns in matrix solid-phase dispersion (MSPD) extraction of the fish were rinsed with ethyl acetate (EtOAc) and methanol (MeOH) prior to use. Quartz wool, also used in assembling the MSPD columns, was purified by Soxhlet extraction with MeOH for 6 h. The general abbreviation of the n-p-SPA is Cn-m, with the subscript n and m specifying the length of the alkyl chain and the position at which the sulfophenyl moiety is substituted to the alkyl chain, respectively. The individual n-p-SPA compounds tested (C10-2, C11-2, C12-2, C13-2, C11-5, C12-5, C135, C12-3, and C12-6) were synthesized in our laboratory with a minimum purity of 97.4%. The use of selected individual compounds allows for complete analytical separation of the test compounds during HPLC analysis and thus for compound-specific determination of bioconcentration behavior. C10-C13 LAS materials consisting of at least 88% of one alkylbenzene homolog and with an isomer composition corresponding to that of LAS in European laundry detergents (15) were kindly provided by Petresa, Spain. Variations in water chemistry were prevented from influencing our experiments by employing an artificial freshwater (H2Oreco), which we reconstituted from distilled water by the addition of salts (CaCl2‚2H2O 0.75 mM, MgSO4‚7H2O 0.46 mM, NaHCO3 1.51 mM, KH2PO4 0.04 mM, NaNO3 1.01 mM, Na2SiO3 0.10 mM) with a resulting hardness of 1.21 mM. Stock solutions were prepared by dissolving the test compounds in H2Oreco and kept under N2 atmosphere to prevent aerobic biodegradation. The exposure solution was continuously prepared by dilution of the stock solution of the test compounds in H2Oreco with O2-saturated H2Oreco. Two peristaltic pumps were employed to deliver the desired volumes of the stock solution as well as H2Oreco to a mixing vessel from where the exposure solution entered the exposure aquarium. Bioconcentration Experiments. The fish used in our study were fathead minnows (Pimephales promelas) reared in the hatchery of Utrecht University ranging in weight between 0.5 and 1 g. They were acclimated for at least 1 week before being exposed to LAS. Only individuals free from observable diseases and abnormalities were selected. A bioconcentration experiment consisted of an uptake and an elimination phase. During the uptake phase, the fish were kept in a flow-through system receiving the exposure solution. The exposure experiment was terminated by transferring fish to clean water, in which fish were allowed to depurate. The water in the elimination aquarium was renewed continuously. The exposure of the fish was performed according to OECD guideline 305E (16). An exception is experiment A, which was considered an exploratory experiment for which we chose a lower flow rate (0.5 L d-1 g-1 fish), chose a shorter exposure period, and decided to starve the fish during the experiment. In the experiments B-D, the specific water renewal rate during exposure was 1 L d-1 g-1 fish. The flow rate during elimination ranged between 0.5 and 1 L d-1 g-1. Exposure lasted for 48 and 168 h to 192 h in experiment A and experiments B-D, respectively. Four fish were sampled at five time points during the steady-state period. Prior to and during experiments B-D, fish were fed at a rate of 1% of their body weight per day. In the elimination phase fish were fed in the elimination aquarium. In an attempt to minimize the concentration of suspended solids in the exposure solution, the fish were allowed to feed for 30 min in a separate aquarium receiving the effluent of the exposure aquarium. Transfer to and back from the feeding aquarium was done with a net. LAS residues were below the limit of detection in the feed (300 nmol/kg). Sampling. Water (ca. 40 mL, at least once per day) and fish samples were taken after predetermined times. Water samples were either extracted immediately or conserved by
addition of 10 vol % of MeOH and stored in the refrigerator under N2 atmosphere for no longer than 2 days. No decrease of the water concentration was observed within a storage period of days (data not shown). Fish were netted out of the aquarium, carefully blotted with a paper tissue, and subsequently killed by immersion into liquid N2 and stored at -20 °C until analysis. Single or duplicate water samples were taken in experiment A and experiments B-D, respectively. During the initial and the steady-state phase of experiment A, fish were sampled in triplicate and in duplicate, respectively. In experiments B, D, and C, the number of replicates was 4, 4, and 3 during the whole experiment. Chemical Analysis. n-p-SPA not tested in the particular experiment was added as an internal standard to the samples, and the results obtained were corrected for the recovery of the internal standard. Water samples were extracted using octadecyl solid-phase extraction cartridges (Baker or J&W) according to the method of de Henau and Matthijs (17). Fish homogenates were extracted employing a MSPD method (18), the details of which will be published elsewhere. In short, fish were homogenized by grinding with a pestle in a mortar. Homogenization was continued after addition of octadecylsilica. The obtained mass is allowed to dry; transferred to a column; and eluted with hexane, ethyl acetate, and a mixture of CH3OH:ethyl acetate (1:1, v:v). The latter fraction containing the LAS analytes was evaporated to dryness, resuspended, and partitioned between 1 M NaOH and CH2Cl2:ethyl acetate (3:1) in the presence of tetrabutylammonium hydrogensulfate as the phase transfer catalyst. The organic phase containing LAS was evaporated to dryness and resuspended in MeOH prior to HPLC analysis. The recovery of the analytes as well as the internal standard exceeded 80% in the water as well as in the fish samples. The method limit of quantitation in fish and water was 500 nmol/kg (0.2 mg/kg) and 15 nM (5 µg/L) per individual test compound. The relative standard deviation of the recovery of the internal standard in water and fish samples ranged around 10% and 20%, respectively. Extracts and standards were analyzed by HPLC with fluorescence detection. The analytes were eluted from the octadecyl column (100 mm × 3 mm, 5 µm, Chrompack) by a gradient elution of acetonitrile and water with NaClO4 as ion-pair reagent (17). The gradient was delivered by a M480 Gynkotek HPLC pump. Samples were injected (20 µL) by a Spark Holland Basic Marathon autosampler. The fluorescence detector (Jasco 920) was set at excitation and emission wavelengths of 225 and 295 nm, respectively. Data were collected by a personal computer equipped with Chromcard (Fisons) data collection software. In experiment D, we used HPLC separation to define the unresolved isomers as ‘inner isomers’ in contrast to the 2-isomer. The inner isomers were quantitated as the sum of the 3-, 4-, 5-, 6-, and 7-sulfophenylalkanes and are abbreviated as Cn-in. In each of the experiments, C12-2 was tested as the reference compound. Concentrations of total organic carbon (TOC) were determined by the Department of Geochemistry of Utrecht University with a Dohrmann DOC analyzer. Data Analysis. The first-order one-compartment model of bioconcentration (19, 20) describes the time course of the concentration of the chemical in fish, Cf (eq I) as function of the concentration in water (Cw) and time (t). It was applied to the measured data to obtain rate constants of uptake (k1) and elimination (k2). Since a minor decrease (typically 10%) in the water concentration occurred during the initial phase of the experiment, we did not assume Cf to be constant but determined k1 from the initial linear part of the uptake curve by linear regression of the ratio Cf/Cw(t) against time. Values of k2 were determined by a nonlinear curve fit (21) of (Cf(t)/ Cf(t)0)) to eq Ib. The bioconcentration factor was determined on the one hand as BCFss, which we define as the ratio of Cf/Cw at steady state (dCf/dt ) 0), and on the other hand as
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TABLE 1. Values of Steady-State Bioconcentration Factor (BCFss); Uptake and Elimination Rate Constants, k1 and k2, Respectively; Fraction of Individual Compounds in Tested Mixture (Oi,w); Average Length of Alkyl Chain (nC,Av, eq III); and Relative Bioconcentration Potential of Mixture ((∑Cf,i/ ∑Cw,i)rel, eq IV)a expt comp
BCFss (L kg-1)
k1 (L (kg d)-1)
k2
(d-1)
Oi,w
A C12-2 C10-2 C11-2 C13-2
47.6 (26) 115.2 (6) 1.7 (29) 21.9 (6) 5.8 (27) 43.7 (2) 353.8 (nc) 419.5 (2)
1.1 (20) 0.9 (73) 1.4 (27) 0.6 (24)
0.133 0.501 0.276 0.090
99.1 (19) 134.4 (1) 1.5 (12) 6.1 (49) 4.3 (16) 1.2 (17) 10.0 (44) 11.1 (24) 1.2 (22) 34.0 (34) 45.5 (14) 1.2 (18)
0.163 0.420 0.247 0.170
B C12-2 C11-5 C12-5 C13-5
FIGURE 2. Uptake curve of C12-2-LAS as obtained in experiment B. The solid line in panel a is calculated according to the integrated form of eq 1 using independently determined values of the uptake and the elimination rate constant.
C
the ratio of k1/k2, which will be referred to in the following as BCFkin.
D
dCf/dt ) + k1Cw - k2Cf
(I)
(Cf(t)/Cf(t)0)) ) e(-k2t)
(Ib)
In analogy to the bioconcentration factor of individual chemicals, we define the bioconcentration potential for mixtures (eq II):
Cf,tot/Cw,tot )
∑C /∑C f,i
w,i
Results and Discussion Average temperature and fish weight ranged between 20.7 and 22.5 °C and between 0.63 and 0.82 g, respectively. The sum of the concentrations of the individual test compounds (∑Cw,i) in the exposure water during steady state ranges between 2.7 and 4.1 µM, and the values of the body burden, defined as ∑Cf,i of the individual LAS constituents in the fish during the steady-state phase, range between 76 and 116 µmol/kg. This demonstrates that the experimental conditions were very similar to each other. A large difference exists between the values of ∑Cf,i measured in our experiment and the LAS lethal body burdens (LBBs) of 1300-3500 µmol/kg (23). This as well as the fact that no mortality occurred during any of the experiments indicate that a sufficient margin between ∑Cf,i and the LBB existed during the experiments. During the steady-state phase of the different experiments, the relative standard deviation of Cw,i did not exceed 20% for any of the compounds tested (data not shown). In the elimination experiments, the concentrations in the water were below the limit of detection. The total organic carbon (TOC) concentrations ranged between 1.6 and 2.8 mg/L. These TOC levels are lower than those at which Traina et al. (24) observed reduction of bioavailability of LAS due to association of LAS to dissolved organic matter. During exposure, the ratio Cf,i/Cw,i increases steadily during the first 24 h (Figure 2). An apparently stable ratio of Cf,i/Cw,i has been reached after 96 h. During elimination, Cf,i decreases until it is below the limit of detection of our method (not shown). The solid line in Figure 2 represents the calculated time course of Cf/Cw during the uptake experiment employing
9
168.4 (37) 251.4 (10) 9.8 (53) 12.4 (8) 31.9 (48) 47.0 (24) 42.1 (42) 128.4 (11)
0.7 (39) 0.8 (34) 0.5 (38) 1.5 (29)
10.8
0.51
11.7
0.29
11.4
0.18
10.6
0.10
0.075 0.592 0.214 0.118
C12-2 211.5 (27) 260.1 (21) 0.7 (16) 0.005 6.0 (46) 10.9 (42) 0.9 (25) 0.065 C10-2 31.9 (29) 61.7 (23) 0.8 (17) 0.017 C11-2 C13-2 987.2 (22) 642.2 (8) 0.6 (16) 0.008 3.0 (50) 6.4 (19) 0.9 (25) 0.623 C10-in 9.1 (41) 26.8 (12) 1.4 (23) 0.165 C11-in C12-in 29.9 (27) 98.9 (18) 1.1 (25) 0.047 C13-in 112.4 (28) 187.8 (9) 0.6 (12) 0.071 a The values in parentheses specify the relative standard deviation (BCFss) or error (k1 and k2) in %.
(II)
The subscript i indicates that the data refer to an individual chemical. If least squares regression between two variables yielded a slope not significantly different from zero (P < 0.05, t-test (22)), we concluded that the variables are independent from each other.
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C12-2 C11-5 C12-6 C12-3
nC,Av
(∑Cf,i/ ∑Cw,i)rel (L kg-1)
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the integrated form of eq I and values of k1 and k2 determined independently from the initial uptake phase and the elimination phase, respectively. Figure 2 demonstrates that the firstorder one-compartment model is suitable to describe the data in terms of first-order rate constants for uptake and elimination. Bioconcentration Parameters. Visual inspection of the uptake curves indicated that steady state was reached between 74 and 96 h. Regression of Cf/Cw for samples taken after 74 h against time indicated that in experiments B-D the slope was not significantly (P < 0.05) different from 0 for all compounds, indicating that steady state had been reached. A shortage of data precluded regression analysis in experiment A. Hence, steady state was assumed to be reached if the increase of Cf/Cw during the last 24 h of the experiment was less than 15%. This was the case for C10-2, C11-2, and C12-2. For C13-2, steady state had not been reached in experiment A. The average of the values of the ratio Cf,i/Cw,i for individual fish at steady state is the steady-state bioconcentration factor (BCFss). The values of BCFss from all experiments are detailed in Table 1. They span a range of about 3 orders of magnitude with C10-2 and C13-2 having the lowest and the highest BCFss values, respectively. The value of C13-2 from experiment A is an estimate of BCFss derived by extrapolation of the uptake curve using a nonlinear curve-fitting procedure. The absolute values of k1 range between 4.3 and 642.2 L kg-1 d-1 with C11-5 and C13-2 having the lowest and the highest k1, respectively (Table 1). The variation in k2 values, between 0.5 and 1.5 d-1 (Table 1) for the different compounds, is about a factor of 3 and therefore small in comparison to the variation observed in the values of k1. From experiments A, B, and D, it can be seen that BCFss increases with increasing length of the alkyl chain length for a given isomer. In addition, the results of experiments B and C demonstrate that the closer the p-sulfophenyl moiety is positioned to the terminal C-atom of the alkyl chain, the higher
is the BCFss. The present results confirm the dependence of bioconcentration parameters on alkyl chain length observed in experiments with fathead minnows employing radiolabeled LAS and nonspecific analysis (25). However, the values of ∑Cf,i/∑Cw,i, 173-245 L kg-1 for C12- and 291-345 L kg-1 for C13-LAS, are based on radioactivity measurements and are higher than values calculated from the results of experiment D, 62 and 182 L kg-1 for C12- and C13-LAS, respectively. This is presumably due to radioactive biotransformation products that were quantitated as LAS. BCFss vs BCFkin. While BCFss is model independent, BCFkin is model dependent because k1 and k2 are determined from the time course of Cf assuming that uptake and elimination follow first-order kinetics. We compared BCFss with BCFkin using linear regression. BCF values of C10-2 determined in experiment A were omitted because of the large standard error (>50%) in the estimate of k2. The slope of the regression line (not shown) is not significantly different from unity (P < 0.05), indicating a linear relationship between BCFss and BCFkin and a regression line being approximately parallel to the line of equality. The intercept of 0.39 can be employed to quantify the degree of overestimation of BCFss by BCFkin. Hence, on the average, the model-dependent BCFkin overestimates BCFss by a factor of approximately 2.5. The average values of the BCFss of the reference compound, C12-2, as obtained in the four experiments, differ from each other by a factor of 4.4 and follow the order A < B (P < 0.05), B < C (P < 0.05), C < D (P < 0.0625) (Table 1) and were significant at the significance levels (ANOVA) specified in parentheses. This indicates that interexperimental differences exist. Since the water chemistry was identical in all experiments, it cannot explain the variation in BCFss of C12-2. Feeding has been suggested to increase elimination of surfactants (13). However, since the values determined in unfed fish (experiment A) are lower than those in fed fish feeding does not explain the observed differences, which are unclear to us at the present time but deserve future attention. Compound Properties Related to Bioconcentration. When comparing the results of different compounds from different experiments, we compensated for these differences by normalization of a result from an experiment to the results for the reference compound as measured in the same experiment according to eq III, with X being either k1, k2, or BCFss and the subscript i indicating the compound to be normalized. When normalizing in this manner, we assume that the reference compound does not interfere with bioconcentration of the indexed compound.
Xi,rel ) Xi/XC12-2-1
(III)
We initially employed the number of carbon atoms of the alkyl chain of the individual LAS constituents (nC,i) as measure of hydrophobicity and found that BCFss,rel increases with increasing alkyl chain length. Regression of BCFss,rel against nC,i yielded strong relationships (r 2 values are 0.96 and 0.88 for the 2- and 5-isomers, respectively) with the slopes being 0.77 and 0.37 log BCFss,rel per nC,i for the 2- and 5-isomers, respectively. The respective increases of BCFss,rel,i per additional C-atom are 5.9 and 2.3, indicating that bioconcentration of n-p-SPA is isomer specific. We employed log KOW estimates (14) in the present investigation to obtain a first impression of bioconcentrationhydrophobicity relationships even though log KOW is not considered a suitable descriptor of the hydrophobicity of dissociated species (26, 27). Figure 3 demonstrates that a linear relationship between log KOW and log BCFss,rel exists. The results of linear least squares regression for log k1,rel and log k2,rel were log k1,rel ) 1.02log KOW - 3.69 (n ) 16, r 2 ) 0.88) and log k2,rel ) 0.05log KOW - 0.32 (n ) 16, r 2 ) 0.05). Hence, it can be concluded that log BCFss,rel and log k1,rel are dependent on hydrophobicity. In contrast to metabolically stable,
FIGURE 3. Plot of the relative values of the steady-state bioconcentration factor BCFss vs estimates of log KOW (according to ref 14) as hydrophobicity parameter. The solid line represents the regression lines. neutral, and rather hydrophobic compounds, such as polychlorinated aromatic compounds (28), k2 does not vary with hydrophobicity. This indicates that the processes involved in bioconcentration might be different for LAS than for the polychlorinated compounds. Application to LAS Mixtures. (a) Mixture Characterization. The constituents of LAS mixtures are affected by environmental processes to different degrees. The average alkyl chain length of an LAS mixture, expressed as average number of C-atoms (nC,Av), is the sum of the number of C-atoms in the alkyl chain of the individual constituents in the water (nC,i,w) weighted with the fraction of the respective constituents in the LAS mixture in water on mole basis (φi,w) (eq IV). In wastewater treatment effluents nC,Av is lower than in the wastewater entering the sewage system (5, 6). In addition, in river waters the relative contribution of the 2-isomers is decreased as compared to the laundry detergent (29). Different production processes yield different isomer compositions, with the proportion of the 2-isomers appearing to be particularly dependent on the production process (30). As a result, LAS composition varies in nC,Av as well as in isomer distribution:
nC,Av )
∑(φ
i,w‚nC,i,w)
(IV)
Since nC,Av is frequently reported in studies that monitor levels of LAS in wastewater treatment effluents (5, 6), it can be regarded an established descriptor of LAS mixtures. Hence, we investigated whether a predictive relationship between nC,Av and (∑Cf,i/∑Cw,i)rel can be derived. In a second approach, we assumed that the individual LAS constituents bioconcentrate additively and that ∑Cf,i/∑Cw,i can be calculated according to
∑C /∑C
(
f,i
w,i)rel
)
∑(φ
i,w‚BCFi,rel)
(V)
Values of (∑Cf,i/∑Cw,i)rel and nC,Av from the experiments are given in Table 1. From Figure 4, it may appear that for experiments B and C a significant relationship exists between nC,Av and log (∑Cf,i/∑Cw,i)rel. However, the data from experiment A confirm the isomer specificity of LAS bioconcentration, since the highest value of (∑Cf,i/∑Cw,i)rel was determined for the homologous series of the 2-isomers, which have a higher bioconcentration potential than the remaining isomers. In experiment D, the 2-isomers were also tested. However, they constitute a minor fraction of the total LAS (approximately 10%), therefore their influence on (∑Cf,i/∑Cw,i)rel is not as pronounced as in experiment A. Furthermore, the results of the different experiments have been made comparable by normalization to C12-2. Therefore, we conclude that the
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TABLE 2. Mixture Compositions Used for Calculation of ∑Cf,i/∑Cw,i As Expressed in Values of Oi,w and Resulting Values of ∑Cf,i/∑Cw,i, Calculated According ∑Cf,i/∑Cw,i ) ∑Oi,w‚BCFss,i Employing Values of BCFss,i As Determined in Experiment D (Table 1) ∑Oi,w mixture
C10-in
C10-2
C11-in
C11-2
C12-in
C12-2
C13-in
C13-2
∑Cf,i/∑Cw,i (L kg-1)
typicala Mississippib expt D
0.10 0.39 0.623
0.02 0.06 0.065
0.24 0.26 0.165
0.05 0.04 0.017
0.28 0.20 0.047
0.06 0.03 0.005
0.20 0.02 0.071
0.04 0.002 0.008
87.2 21.7 22.6
a
European domestic formulation (15).
b
Filtered Mississippi water (37).
FIGURE 4. Plot of log (∑Cf,i/∑Cw,i)rel vs nC,Av. The dots give the values of ∑Cf,i/∑Cw,i as determined for the mixtures in experiments A-D. outlier (experiment A) demonstrates that no relationships exists between nC,Av and log (∑Cf,i/∑Cw,i)rel. The second approach was evaluated by calculating values of log (∑Cf,i/∑Cw,i)rel for the inner isomers and comparing them with experimental values (experiment D). We estimated values for log BCFss,rel of those individual n-p-SPA for which no experimental data were available employing the relationship between log KOW and log BCFss,rel. Estimated and experimental data of log BCFss,rel were used as input into eq V. Since experimental data of BCFss,rel were available for none of the constituents of C10-in, we excluded the C10-homolog from linear regression of experimental and estimated values of log (∑Cf,i/∑Cw,i)rel. The value of r 2 (1.0) indicates a strong relationship between estimated and experimental values. Slope and intercept of the regression line are 1.05 and 0.17, respectively. Even though the slope is significantly different (P < 0.05) from 1, the regression line is approximately parallel to the line of equality. Hence, the intercept represents the average degree of overestimation, which upon transforming from logarithmic to linear scale is approximately 1.5. Given the large safety factors usually employed in environmental risk assessment we think that the above estimation approach yields satisfactory results. (b) Environmental Mixtures. We calculated ∑Cf,i/∑Cw,i as function of the composition of two particular LAS mixtures according to eq V. One is the ‘typical’ mixture of LAS as it is formulated into European domestic formulations (15) and released into wastewater. The composition of the second mixture is based on data measured by GC-MS (after derivatization) in filtered Mississippi River water samples (29) with an estimated value of φC13-2 since the levels of C13-2 were below the limit of detection. ∑Cf,i/∑Cw,i was calculated by taking the inner isomers to be one substance and the 2-sulfophenyl alkanes as the second and employing the corresponding experimentally determined data of BCFss (experiment D). We preferred this approach above calculation of ∑Cf,i/∑Cw,i based on individual n-p-sulfophenyl alkane data because, for the latter, BCFss,i data are available for 9 out 20 LAS constituents, and estimation of the missing values using
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the relationship between log KOW and log BCFss,rel would have introduced additional uncertainty. In Table 2, the mixture compositions and the values of ∑Cf,i/∑Cw,i are specified with ∑Cf,i/∑Cw,i being 22 and 87 L kg-1 for the Mississippi river and the typical mixture, respectively. While both values are below 100 L kg-1, the value considered critical in European risk assessment (31), ∑Cf,i/∑Cw,i in the river water mixture is approximately four times lower than in the typical mixture. This indicates that the modification of the LAS composition in sewer, in wastewater treatment, and in river leads to a reduction of the bioconcentration potential of LAS as a mixture. The above calculation also indicates that ∑Cf,i/∑Cw,i of LAS is not a compound property but a parameter characteristic of a specific mixture with φi,w characterizing the mixture composition. Since (a) the composition of LAS is variable and (b) the bioconcentration properties of its constituents differ widely, the variability of LAS composition has to be taken in into account when assessing the bioconcentration potential of a given mixture.
Acknowledgments The authors gratefully acknowledge the financial support granted by AISE/CESIO as well as the discussions with the members of the AISE/CESIO steering committee Surfactant Bioconcentration.
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Received for review February 5, 1997. Revised manuscript received July 14, 1997. Accepted August 18, 1997.X ES970100D X
Abstract published in Advance ACS Abstracts, October 1, 1997.
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