Effect of Rhamnolipids on the Dissolution, Bioavailability, and

of MSM with or without rhamnolipid (0.35 or 3.5 mM) was added to the flask. The flask was placed in a gyratory shaker. (200 rpm) maintained at 23 °C...
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Environ. Sci. Technol. 1997, 31, 2211-2217

Effect of Rhamnolipids on the Dissolution, Bioavailability, and Biodegradation of Phenanthrene YIMIN ZHANG,† WALTER J. MAIER,‡ AND R A I N A M . M I L L E R * ,† Department of Soil, Water and Environmental Science, University of Arizona, Tucson, Arizona 85721 and Department of Civil and Mineral Engineering, University of Minnesota, Minneapolis, Minnesota 55455

Biodegradation rates for polycyclic aromatic hydrocarbons (PAH) in the environment are limited by their low solubility and sorption to solid surfaces. The purpose of this study was to quantify the effect of biosurfactants on the dissolution, bioavailability, and biodegradation of a slightly soluble PAH, phenanthrene, in a series of batch solution studies. A mathematical model that describes the combined effects of solubilization and biodegradation, including description of bioavailability within surfactant micelles, was used to analyze the experimental results. Two forms of the biosurfactant, a monorhamnolipid and a dirhamnolipid, were tested; it was found that both surfactants increased the solubility and enhanced the rate of phenanthrene biodegradation. Monorhamnolipid was more effective than dirhamnolipid for solubilization; however, overall rates of mineralization were essentially the same. This seems to result from variable bioavailability of substrate: phenanthrene within monorhamnolipid micelles was less bioavailable than phenanthrene within dirhamnolipid micelles. Therefore, the effect of a surfactant on biodegradation is a combination of the solubilizing power of the surfactant and the bioavailability of the substrate within the surfactant micelles. Model analysis of the solubilization data showed that the overall solubilization rate coefficient, KL, increased with increasing biosurfactant concentration. Analysis of biodegradation data showed that enhanced biodegradation rates depend upon both KL and R, the coefficient for substrate bioavailability from micelles. Model simulations using parameters developed from test data are discussed.

Introduction Polycyclic aromatic hydrocarbons (PAH) are a group of solid phase organic chemicals containing two or more fused benzene rings. Because they are toxic, carcinogenic, and mutagenic (1), remediation of PAH-contaminated sites is desirable. Although most PAH compounds are biodegradable (2), rates of PAH biodegradation in the environment are limited due to their hydrophobicity and low water solubility. These factors cause PAHs to sorb to the soil matrix or to form a separate solid phase. It is generally believed that sorbed and solid phase PAH are not directly available to microorganisms (3), and it is the limited rate of mass transfer of PAH from the sorbed and the solid phase to the aqueous phase that is responsible for slow biodegradation rates. * Corresponding author phone: (520)621-7231; fax: (520)621-1647; e-mail: [email protected]. † University of Arizona. ‡ University of Minnesota.

S0013-936X(96)00687-6 CCC: $14.00

 1997 American Chemical Society

FIGURE 1. Rhamnolipid structure. For monorhamnolipid, R ) H; for dirhamnolipid, R ) rhamnosyl. Surfactants can enhance the rate of mass transfer from the solid and sorbed phases by increasing the rates of dissolution and desorption of PAHs (4). The surfactant concentration at which surfactant monomers begin to assemble in ordered colloidal aggregates is termed critical micelle concentration (cmc). Surfactant solubilization of PAHs generally commences at the cmc and then is a linear function of surfactant concentration (5). Solubilization is attributed to incorporation of the hydrophobic PAH molecules into the hydrophobic core of micelles in solution. Most surfactants can increase the apparent solubility of PAH; however, the rate of biodegradation is not always similarly enhanced (6-9). One possible explanation for this is that PAH solubilized into micelles may not be readily available to microorganisms (8, 9). So clearly, surfactants above the cmc can play two opposing roles in the biodegradation of PAH: they increase the mass transfer of substrate into the aqueous phase, but the PAH within surfactant micelles may have reduced bioavailability as compared to the “truly soluble” aqueous phase PAH. Despite this, no link between solubilization and bioavailability has yet been established, nor has a quantitative understanding of the relationship between the solubilization, bioavailability, and biodegradation been developed. The purpose of this study was to quantify the effect of biosurfactant structure and concentration on the solubilization, bioavailability, and biodegradation of a PAH. Two biosurfactants were used in this study, monorhamnolipid and dirhamnolipid, the most commonly isolated surfactants from Pseudomonas spp. Phenanthrene was chosen as a model PAH because it has relatively low vapor pressure (0.018 pa at 25 °C) (10) and water solubility (1.26 µg mL-1) (11). A mathematical model was formulated to quantify the role of the biosurfactants in phenanthrene biodegradation.

Materials and Methods Microorganism. Pseudomonas putida CRE 7 was obtained from Mike Montgomery, Geo-Center, Inc., Naval Research Lab, Washington, DC. This culture was maintained at 23 °C on mineral salts medium (MSM) (12) agar plates using phenanthrene as sole carbon source and transferred monthly. This strain did not produce or utilize biosurfactants during growth on MSM containing phenanthrene. Biosurfactant. Monorhamnolipid and dirhamnolipid were used in this study (Figure 1). Monorhamnolipid was produced and purified from P. aeruginosa ATCC 9027 as described previously (12, 13). The average molecular weight of this surfactant is 504, and the cmc is 0.1 mM (15). Surface tension and interfacial tension between hexadecane and water were measured at the monorhamnolipid cmc and found to

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be 28 dyn cm-1 and 2 dyn-1, respectively. Dirhamnolipid was a gift from Kyowa Hakko Kogyo Co., Ltd. (Tokyo, Japan) (14). It has an average molecular weight of 650, and the cmc is 0.1 mM. Surface tension and interfacial tension were measured at the dirhamnolipid cmc to be 36 dyn cm-1 and 5 dyn-1, respectively. Chemicals. Phenanthrene (98% pure) was purchased from Aldrich Chemical Company (Milwaukee, WI), and [9-14C]phenanthrene (specific activity, 13.1 mCi mmol-1, >98% pure) was obtained from Sigma Chemical Company (St Louis, MO). Solubilization Rates. The rates of the mass transfer of phenanthrene from the solid to liquid phase were determined in the presence and the absence of biosurfactant. For each experiment, 50 µL of phenanthrene (3.3 mg) dissolved in chloroform was carefully added to the bottom of a tilted (30°) 125-mL micro-Fernbach flask (Wheaton, Millville, NJ). The chloroform was allowed to evaporate. The coated area was determined to be 1.5 cm2, while the total area of the bottom of the flask was 54 cm2. After evaporation of solvent, 20 mL of MSM with or without rhamnolipid (0.35 or 3.5 mM) was added to the flask. The flask was placed in a gyratory shaker (200 rpm) maintained at 23 °C. After 1, 2.5, 5, 10, 15, and 20 min as well as 24 h, triplicate 0.5-mL samples were removed from each flask using a pipet and each sample was filtered through a 1-mL glass syringe packed with glass wool to remove any solid phenanthrene particles. Glass wool was chosen as the filtering material because phenanthrene does not sorb to this material. The amount of phenanthrene in the aqueous solution was determined by high-performance liquid chromatography (HPLC). The HPLC system used was a Waters 600 delivery system equipped with a Waters 490E UV detector operating at 250 nm and an ODS-2 reverse-phase column (4.6 by 25 mm; 5 µm particle size) (Alltech, Deerfield, IL). Isocratic elution was carried out with methanol:water (95:5). The injection volume was 50 µL. The sensitivity of detection was 0.01 µg mL-1 phenanthrene. Biodegradation of Phenanthrene. Biodegradation of phenanthrene was quantified in two ways: by quantification of mineralization by measurement of 14CO2 evolved during growth on [14C]phenanthrene and by quantification of substrate loss as determined by HPLC. For mineralization experiments, a mixture of phenanthrene and [14C]phenanthrene was coated onto the micro-Fernbach flasks as described earlier. The final mass of phenanthrene added was 3.3 mg (specific activity, 2.3 µCi mmol-1). After evaporation of solvent, 20 mL of MSM containing 0, 0.35, or 3.5 mM rhamnolipid was added to each flask. Finally, each flask was inoculated (2.5%) with a culture pregrown on MSM containing phenanthrene (100 µg mL-1) as the sole carbon and energy source at 23 °C for 3 days. The flasks were sealed with specially designed caps, incubated, and purged periodically to collect 14CO and 14C-labeled volatile organic compounds (16). 2 For determination of substrate loss, phenanthrene was coated, inoculated, and incubated as described above except that only nonradioactive phenanthrene was used. Periodically, duplicate flasks were sacrificed to determine the amount of phenanthrene remaining using the following procedure. Five drops of concentrated HCl were added to each flask to precipitate rhamnolipid to reduce emulsification during solvent extraction. The flasks were then extracted with 40 mL of chloroform twice. The solvent was pooled and evaporated. The residual phenanthrene was dissolved in 20 mL of methanol, and the samples were stored at -20 °C until HPLC analysis was performed. Model Description. A model, BIOSURF, was written to describe a postulated sequence of steps involved when solid phase phenanthrene is “solubilized” in the presence of a surfactant and then biodegraded in a well-mixed batch reactor. The term “solubilized” includes both the transformation of the solid phase chemical into a dispersed true solute phase (represented by S) and a micelle associated dispersed

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phase (represented by Sm). This model was developed for analysis of test data from experiments using a range of surfactant concentrations in order to quantify observations that clearly showed that rates of biodegradation do not increase in proportion to the increase in total “solubilization”. These observations led us to hypothesize that micelleassociated chemicals are less readily available for biodegradation than true solutes. The model consists of a series of coupled differential equations that describe the time-dependent changes in concentrations of residual solid phase phenanthrene (SXS), water-soluble phenanthrene (S), and micelle-associated phenanthrene (Sm). The model also describes the cumulative concentration of active cell mass and the evolution of CO2 as a function of time. Several assumptions were made in constructing this model: (1) First, it is assumed that dissolution occurs at the solid-water interface of the separate phase chemical and involves a phase change from solid to solute, which then mixes into the bulk water phase. (2) Attachment of cells to the solid phenanthrene phase, with attendent direct transport of substrate into cells, is not considered a significant biodegradation mechanism. (3) The surfactant concentration remains constant for the duration of the experiment. (4) A stepwise mechanism is postulated that involves formation of molecular solute S, which is available for partitioning into micelles (Sm) and uptake into cells. Both S and Sm are available for microbial degradation; however, uptake of Sm occurs as micelles release the intercalated phenanthrene to the water phase or to the cell surface as they continually break up and reform. (5) Partitioning of solutes into micelles and the breakup of micelles are assumed to be fast reactions (9). (6) It is assumed that at all times the aqueous phase solute (S) is in equilibrium with the micellar solute (Sm) as shown in eq 1:

Sm ) SKp

(1)

where Kp is the partition coefficient for phenanthrene in the micellar phase. The rate of dissolution of solid phase phenanthrene (SXS) is described by eq 2:

dSXS ) -KL[KSA(SXSi)n(SXS)1-n](Cs - S) dt

(2)

where KL is the dissolution rate coefficient that describes the rate of transfer of phenanthrene from the solid phase into the dissolved phase per unit of surface area per unit time and Cs is the intrinsic water solubility of phenanthrene. The second term of this equation, which includes KSA, SXSi, SXS, and n, specifies the value of the surface area remaining as the initial solid phase (SXSi) decreases to SXS. KSA represents the initial ratio of surface area to mass of solid phase material. It is anticipated that surface area will decrease as the solid phase (SXS) dissolves. However, reduction in surface area depends on the physical configuration of the solid phase. For purposes of modeling, the changes in this ratio (surface area to solid mass) for decreasing SXS and for different physical configurations has been formulated using the term KSA(SXSi)n(SXS)1-n. In this term, n represents the change in the magnitude of the loss in surface area as the solid phase shrinks. At time zero, the surface area term simplifies to KSA(SXSi)1. For times greater than 0, the value of n impacts this term. For example, setting n ) 1.0 results in a simplification of the surface area term to KSA(SXSi)1. In this case, the surface area is constant and does not decrease as SXS decreases. This corresponds to the case where the solid phase is present as a flat film so that dissolution of solid phase does not change the surface area. For n < 1, the surface area decreases as SXS decreases resulting in a decreased solubilization rate. For example, Stealf (17) analyzed the dissolution of cubical and

TABLE 1. Solubilization of Phenanthrene by Rhamnolipids surfactant type

surfactant concn (mM)

phenanthrene solubilitya (µg mL-1)

no surfactant monorhamnolipid monorhamnolipid dirhamnolipid dirhamnolipid

0 0.35 3.5 0.35 3.5

0.69 ( 0.04 3.63 ( 0.45 34.8 ( 0.7 1.30 ( 0.09 12.98 ( 1.33

a

Samples were equilibrated for 24 h before determining solubility.

spherical particles and found n ) 0.33, where the decrease in the term [KSA(SXSi)0.33(SXS)0.67] corresponded to the shrinkage of the solid surface area in the form of cubes. Values of n can be decreased to zero or even negative values. For n ) 0, the dissolution term becomes KSA(SXS)1, which corresponds to the case where surface area shrinks in direct proportion to the shrinkage in mass of solids remaining. Physically, this can be represented by the surface area of a unimolecular deposition of the solid phase on a high surface area solid support. For negative values of n, the reduction in surface area is more rapid than the reduction in mass remaining. The driving force for the solubilization reaction shown in eq 2 is determined by the difference between the aqueous solubility (Cs) and the amount of solute in true solution (S). Compared with the partitioning reaction (eq 1), the solubilization reaction is relatively slow because it depends on the extent of the solid phase surface area in contact with water (KSA) as well as physical factors that affect mass transport at the interface (18). Biodegradation of phenanthrene, production of carbon dioxide, and growth of cell mass are described by the following modified Monod equations:

dX ) A - KdX dt

(3)

dS dSXS A ) dt dt Y

(4)

A)

µmaxS(1 + RKp)X Ks + S(1 + RKp)

dCO2 YCO2grA ) + YCO2emKdX dt Y

(5)

(6)

where X is the active cell mass concentration; Kd is the endogenous cell decay rate coefficient; Y is the cell mass yield coefficient; µmax is the Monod growth rate coefficient; Ks is the half-saturation coefficient; CO2 is the cumulative mass of carbon dioxide produced; YCO2gr is the carbon dioxide yield coefficient from growth metabolism; YCO2em is the carbon dioxide yield coefficient from endogenous cell decay metabolism; and R is the microbial uptake effectiveness factor for micelle-associated phenanthrene. R is a fitted parameter that characterizes availability of micellar phenanthrene for transport into actively growing cells. Micelles are postulated to be in a dynamic flux, undergoing continuous breakup and re-formation with attendant release and re-intercalation of phenanthrene. On the basis of measured micellar uptake, it appears that there is 1 molecule of phenanthrene/18 molecules of surfactant. The fate of phenanthrene molecules therefore depends on the partitioning dynamics that exist in the vicinity of the cell surface. As the surfactant concentration is increased, the concentration of phenanthrene per micelle will decrease. This would be expected to cause a decrease in R due to a decrease in the transfer of phenanthrene from the micelle to the microbial cell.

FIGURE 2. Effect of surfactant concentrations on the dissolution of phenanthrene by monorhamnolipid (A) and dirhamnolipid (B). Symbols are experimental data, and dotted lines are model fits. Symbols: (b) no rhamnolipid, (O) 0.35 mM rhamnolipid, and (9) 3.5 mM rhamnolipid. Numerical values for many of the model parameters were obtained experimentally. Namely, Kp, and Cs were determined from measurements of phenanthrene solubility in aqueous and surfactant solutions. SXSi was the measured value of the initial mass of phenanthrene added per unit volume. KSA was calculated from the surface area of the applied solid phase phenanthrene. KL was fitted from dissolution rate experiments in the absence of biodegradation. Y, YCO2gr, and YCO2em were determined from batch biodegradation tests. The Monod parameters µmax, Ks, Kd, and Xi were fitted for the control case (phenanthrene biodegradation in the absence of surfactant), and these fitted values were used for all model fits of biodegradation in the presence of surfactant. The bioavailability coefficient, R, and n, which describes the dissolution behavior of the solid phase, were fitted during the biodegradation model fits. For each model fitting, the best fit value for the unknown coefficient(s) was calculated using a computer-programed search technique to minimize the sum of the squares of the differences between the calculated and measured values of the dependent variable. At least seven measured data points were used for each fitting.

Results Phenanthrene Solubilization. The aqueous solubility of phenanthrene in the absence of surfactant was measured to be 0.69 ( 0.04 mg L-1. This is close to the solubility reported in the literature, which ranges from 0.71 to 2.67 mg L-1 depending on the method of measurement used (19). The solubility of phenanthrene increased considerably, up to 35 mg L-1, in the presence of rhamnolipids (Table 1).

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TABLE 2. Parameters for Model Fits of Solubility Experiments surfactant concentration (mM) monorhamnolipid

dirhamnolipid

model parameters

0

0.35

3.5

0.35

3.5

µmax (µg mL-1) SXSi (µg mL-1)a Cs (µg mL-1)a KSA (cm2 µg-1)a Kp (1)a KL (cm h-1)b

0 165 0.69 4.55 × 10-4 0 45

0 165 0.69 4.55 × 10-4 4.26 103

0 165 0.69 4.55 × 10-4 49.43 846

0 165 0.69 4.55 × 10-4 0.88 65

0 165 0.69 4.55 × 10-4 17.81 500

a

Values were obtained from direct measurement or calculation.

b

Values were fitted.

Monorhamnolipid, which is a more surface active molecule, showed greater solubilizing capacity for phenanthrene than dirhamnolipid. This can be expressed as a solubilization capacity (moles of organic compounds solubilized per mole of surfactant). The solubilization capacity calculated from the data in Table 1 was 0.057 mol of phenanthrene/mol of monorhamnolipid and 0.021 mol of phenanthrene/mol of dirhamnolipid. The kinetics of the phenanthrene dissolution in rhamnolipid solutions is shown in Figure 2. The rate of phenanthrene dissolution was dependent on both the type and concentration of rhamnolipid (see the slope of each curve). The experimental data from Figure 2 were fitted using BIOSURF to estimate the best values for the dissolution rate coefficient (KL). Table 2 lists the parameters used for these model fits. Specifically, µmax was set to zero in the absence of biodegradation; SXSi (165 µg mL-1) and Cs (0.69 µg mL-1) were measured; n was set to 1 because the short duration of the test (20 min) should result in only a very small change in the residual SXS; KSA (4.5 × 10-4 cm2 µg-1) was estimated using the mass and surface area of the substrate; and Kp was calculated using eq 1 and the solubility data from Table 1. As shown in Figure 2, the model fits were close to the experimental data for both rhamnolipids (see dashed lines). A significant enhancement of the dissolution rate coefficient (KL) was found in the presence of surfactant. At 0.35 mM surfactant, the KL was approximately doubled, while at 3.5 mM surfactant, the KL was 10-20-fold greater than the KL in the absence of surfactant (Table 2). As expected, the fitted KL values were greater for monorhamnolipid than for dirhamnolipid. Phenanthrene Biodegradation. Figure 3 shows the effect of monorhamnolipid and dirhamnolipid on the biodegradation of phenanthrene. In the absence of surfactant, mineralization of phenanthrene was linear as would be expected for a substrate with low water solubility and a limited rate of mass transfer (20, 21). In the presence of the low concentration of surfactant (0.35 mM), there was only a slight enhancement of biodegradation even though the rate coefficient for phenanthrene dissolution was doubled (Table 2, see KL values). However, at the higher surfactant concentration (3.5 mM) there was a significant enhancement of the mineralization rate. In this case, the dissolution rate coefficient was increased 10-20-fold. The data shown in Figure 3 were fitted using the BIOSURF model to determine best values for R, the bioavailability coefficient. Parameters used for these simulations are shown in Table 3. These include SXSi, Cs, KSA, and Kp, which were obtained by direct measurement. The values of Y, YCO2gr, and YCO2em were estimated from a carbon balance of the biodegradation data shown in Figure 3. The fitted values of KL were used from the dissolution study (Table 2). The values of Xi (the initial cell mass concentration), Ks, µmax, and Kd were fitted for the control biodegradation test (no surfactant present) and then set for the model fits of biodegradation in the presence of surfactant. The value for n was optimized

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FIGURE 3. Effect of rhamnolipid concentrations on the mineralization of phenanthrene by P. putida CRE 7: (A) monorhamnolipid, (B) dirhamnolipid. Symbols are experimental data, and dotted lines are model fits. Symbols: (b) no rhamnolipid, (O) 0.35 mM rhamnolipid, and (9) 3.5 mM rhamnolipid. during each model fit. During the course of performing the simulations, it was found that the values of Xi and Kd fitted for the control case and values of n and R fitted for the remaining cases were unique sets of values for each simulation. However, for the control case, the values of µmax and Ks were correlated. Thus, it was not possible to determine a unique set of values for these two parameters. So for all model fits, µmax was set to 0.4 h-1 and Ks was fitted. Other combinations of µmax and Ks would probably give equally good fits. As shown by the dashed lines in Figure 3, the model provided good fits to the experimental data. The implication is that the model adequately describes the dominant process variables that control dissolution and biodegradation. As such, the model serves as a tool for more detailed analysis and additional experimental work to gain more complete

TABLE 3. Parameters for Model Fits of Biodegradation Experiments surfactant concentration (mM) monorhamnolipid

dirhamnolipid

model parameters

0

0.35

3.5

0.35

3.5

Si (µg mL-1) SXSi (µg mL-1)a Xi (µg mL-1)b CO2i (µg mL-1) Cs (µg mL-1)a Ks (µg mL-1)b µmax (h-1)b KSA (cm2 µg-1)a Kd (h-1)b Y (1)a YCO2gr (1)a YCO2em (1)a n (1)b KL (cm h-1)c Kp (1)a R (1)b

0 163.2 21 0 0.69 1.95 0.4 4.55 × 10-4 0.0046 1 0.9 1.5 0.58 45 0 nad

0 164.9 21 0 0.69 1.95 0.4 4.55 × 10-4 0.0046 1 0.9 1.5 -0.25 103 4.26 0.1

0 169.5 21 0 0.69 1.95 0.4 4.55 × 10-4 0.0046 1 0.9 1.5 -0.5 846 49.43 0.01

0 165.0 21 0 0.69 1.95 0.4 4.55 × 10-4 0.0046 1 0.9 1.5 0.3 65 0.88 0.3

0 165.0 21 0 0.69 1.95 0.4 4.55 × 10-4 0.0046 1 0.9 1.5 -0.1 500 17.81 0.03

a

Values were obtained from direct measurement or calculation.

b

Values were fitted. c Fitted values are from Table 2.

d

Not applicable.

understanding of mechanisms. For example, the finding that the fitted R values decreased with increasing surfactant concentration has important practical implications; namely, that there may be an optimal surfactant concentration for increasing overall rates of solubilization and biodegradation, and that more surfactant is not necessarily better. The maximum amount of freely available micellar phase phenanthrene in each experiment can be calculated from the solubility information in Table 1 using

freely available micellar phase phenanthrene ) Sm‚R where Sm is the total phenanthrene solubility, - Cs. Thus, when 0.35 mM monorhamnolipid was added, the fitted value of R was 0.1, indicating that 10% of the micellar phase phenanthrene was freely available. Calculation shows that the freely available micellar phase phenanthrene in this case is 0.29 mg L-1 [(3.63-0.69 mg L-1) × 0.1]. This can be compared to 2.9 mg L-1 for R ) 1.0 (micellar phase substrate is completely available). In the presence of 3.5 mM monorhamnolipid, the value of R decreased further to 0.01, indicating that only 1% of the micellar phase substrate (0.34 mg L-1) was freely available. It is noteworthy that the calculated increase in freely available micellar phenanthrene with high rhamnolipid application is only marginally larger than for the lower application. However, as shown in Figure 3, the rate of phenanthrene degradation is significantly higher with high rhamnolipid application, particularly in the early stages of the experiment when solid phase phenanthrene was not limiting. Application of the model, therefore, has identified a possible anomaly that requires explanation. It is postulated that overall phenanthrene degradation is promoted by more efficient transport of phenanthrene from the micelle phase to suspended cells as compared to transport of phenanthrene from the solid phase, because the release of phenanthrene from the micelle occurs in close proximity to the cells and in effect creates a localized region of high concentration. Bioavailability, as indicated by R, was also dependent on the surfactant structure (Table 3). Phenanthrene in monorhamnolipid solutions was less bioavailable than that in dirhamnolipid solutions. Recall that, in terms of solubilization alone, monorhamnolipid was more effective (Table 1), but in terms of bioavailability, the monorhamnolipid was less effective than dirhamnolipid. Interestingly, the biodegradation behavior in the presence of either surfactant type was similar, indicating that the solubilization and bioavailability factors balanced out (Figure 3). A second coefficient that

FIGURE 4. Effect of varying n on model fitting. The constants used in these simulations are from the degradation of phenanthrene in the presence of 3.5 mM monorhamnolipid (see Table 3). Five different model fits are shown varying n from -1 to 1. was fitted was n, which describes the change in the solid phase surface area as a function of time as the residual mass of phenanthrene shrinks. Addition of rhamnolipids resulted in decreasing n values, indicating a more rapid reduction in the surface area per unit mass of phenanthrene during the experiment as compared to the control that had no surfactant added. The decrease in n was greater for monorhamnolipid, indicating that the reduction in surface area of substrate was faster for monorhamnolipid, which caused a higher dissolution rate, than for dirhamnolipid. The range of fitted n values in Table 3 is from -0.25 to 0.58. Although this seems like a large variation in values, these changes in n actually have a relatively small affect on the model fit (Figure 4). In a separate experiment, the effect of monorhamnolipid on the biodegradation of phenanthrene was determined by measurement of phenanthrene loss. As shown in Figure 5, the substrate loss data are well correlated to the mineralization data (Figure 3A). These data were fitted using BIOSURF and the same parameters shown in Table 3 that were used to fit the mineralization data. Model fits are the dashed lines in Figure 5, which show good fits to the experimental data. Prediction of Biodegradation Behavior in the Presence of Surfactants. The BIOSURF model can also be used to predict biodegradation behavior under altered conditions. Such predictions may be useful for optimizing the use of a

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FIGURE 5. Effect of monorhamnolipid concentrations on the utilization of phenanthrene by P. putida CRE 7. Symbols are experimental data, and dotted lines are model fits. Symbols: (b) no rhamnolipid, (O) 0.35 mM rhamnolipid, and (9) 3.5 mM rhamnolipid. surfactant in remediation. For example, simulations were done for three cases: (A) initial cell mass (Xi) was increased 10-fold, (B) initial substrate concentration (SXSi) was increased 10-fold, and (C) both initial cell mass and initial substrate concentration were increased 10-fold. The results of these simulations are shown in Figure 6. The simulation shown in Figure 6A indicates that, in the presence of additional cell mass, the surfactant was even more effective in increasing biodegradation (compare Figure 6A to Figure 5). A similar enhancement of biodegradation rate was seen for case B (Figure 6), where the initial substrate concentration was increased 10-fold. However, in this case, there was an initial lag period that was necessary to build up a sufficient biodegrading population. In case C, both the initial substrate and cell mass were increased, and the resulting simulation was similar to that shown in Figure 5 (compare Figure 6C with Figure 5).

Discussion Bacterial growth on phenanthrene was linear (Figure 3) due to rate-limited dissolution of the substrate from the solid phase to the liquid phase (22). The addition of rhamnolipids markedly increased the rate of dissolution of phenanthrene (Figure 2) but increased the rate of biodegradation more modestly. Similar behavior has been reported in several previous studies concerning synthetic and biological surfactants (7, 9, 14, 23), and it has been suggested that this behavior is caused by a reduction in bioavailability of the phenanthrene that is partitioned into the surfactant micelles. The results reported in this study present further evidence that micellar phenanthrene has variable bioavailability that is dependent upon both the structure and the concentration of surfactant used. This study investigated the effect of two structural types of rhamnolipid on phenanthrene biodegradation: monorhamnolipid, which exhibits relatively high surface activity and high solubilization capacity, and dirhamnolipid, which is less surface active and has less solubilization capacity. Despite this difference in solubilization capacity, both rhamnolipid types were found to stimulate biodegradation rates equally well. Modeling results suggest that this is because phenanthrene within monorhamnolipid micelles was less bioavailable than phenanthrene within dirhamnolipid micelles. These results support prior evidence showing a relationship between

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FIGURE 6. Prediction of the disappearance of phenanthrene in the presence of (A) 10 × Xi (210 mg biomass/L), (B) 10 × SXSi (1695 mg of phenanthrene L-1), (C) 10 × Xi, 10 × SXSi by BIOSURF model. All other parameters are used from Table 4. Symbols: (s) no rhamnolipid, (‚‚‚) 0.35 mM monorhamnolipid, (-‚-) 3.5 mM monorhamnolipid. solubility enhancement and structure of a surfactant (5, 24). In addition, these results point out the importance of the structure of the surfactant on bioavailability of the micellar substrate. The uptake of micellar phenanthrene requires the transfer of substrate from micelle to cell. In the case that this transfer is a passive partitioning process, the rate of transfer should be inversely related to the affinity of the substrate for micelles. For example, phenanthrene has higher affinity for monorhamnolipid than dirhamnolipid, resulting in lower bioavailability in the presence of monorhamnolipid. It should be noted that in some cases, particularly for biosurfactants, the partitioning process may not be entirely passive. Biosurfactant recognition sites on the cell surface may aid in the delivery and uptake of the substrate within the micelle. For example, Zhang and Miller (13) have shown that rhamnolipids can induce development of cell surface hydrophobicity, which is followed by increased rates of degradation of aliphatic compounds. Thus, for biosurfactants, bioavailability of micellar substrate may not always be predictable from structure and solubilization data. In summary, surfactants have potential for use in improving remediation in sites where biodegradation is affected by rate-limited desorption and dissolution of the contaminant. The results of this research show that the effectiveness of a surfactant in improving contaminant biodegradation is a

combination of the solubilizing power of the surfactant and the bioavailability of micellar contaminant. For synthetic surfactants, both of these parameters should depend on the chemical structure of the surfactant. As has been reported in the literature, biodegradation in the presence of synthetic surfactants is inhibited in some cases (25). Recognized factors that cause such inhibition include surfactant toxicity to degrading cells or the surfactant serving as a preferred carbon source (4, 6). Our results suggest that another factor that may inhibit biodegradation in the presence of surfactants is low bioavailability of the substrate within surfactant micelles.

Acknowledgments This research was supported by Grant P42 ES04940 from the National Institute of Environmental Health Sciences, NIH.

Literature Cited (1) Keith, L. H.; Telliard, W. A. Environ. Sci. Technol. 1979, 13, 416. (2) Gibson, D. T.; Subramanian, V. In Microbial Degradation of Organic Compounds; Gibson, D. T., Ed.; Marcel Dekker: New York, 1984; pp 181-252. (3) Mihelcic, J. R.; Lueking, D. R.; Mitzell, R. J.; Stapleton, M. Biodegradation 1993, 4, 141. (4) Miller, R. M. 1995. In BioremediationsScience and Applications; Skipper, H., Ed.; Soil Science Society of America: Madison, WI, 1995; pp 33-54. (5) Edwards, D. A.; Luthy, R. G.; Liu, Z. Environ. Sci. Technol. 1991, 25, 127. (6) Tiehm, A. Appl. Environ. Microbiol. 1994, 60, 258. (7) Liu, Z.; Jacobson, A. M.; Luthy, R. G. Appl. Environ. Microbiol. 1995, 61, 145.

(8) Volkering, F.; Breure, A. M.; van Andel, J. G.; Rulkens, W. H. Appl. Environ. Microbiol. 1995, 61, 1699. (9) Guha, S.; Jaffe´, P. G. Environ. Sci. Technol. 1996, 30, 605. (10) DeKruif, C. G. J. Chem. Thermodyn. 1980, 12, 243. (11) Bruggeman, W. A.; Van Der Steen, J.; Hutzinger, O. J. Chromatogr. 1982, 238, 335. (12) Zhang, M.; Miller, R. M. Appl. Environ. Microbiol. 1992, 58, 3276. (13) Zhang, M.; Miller, R. M. Appl. Environ. Microbiol. 1994, 60, 2101. (14) Zhang, M.; Miller, R. M. Appl. Environ. Microbiol. 1995, 61, 2247. (15) Magaritis, A.; Kennedy, K.; Zajic, J. E.; Gerson, D. F. Dev. Ind. Microbiol. 1979, 20, 623. (16) Marinucci, A. C.; Bartha, R. Appl. Environ. Microbiol. 1979, 38, 1020. (17) Stealf, A. M.S. Plan B Paper, University Minnesota, 1994. (18) Jahan, K. Ph.D. Dissertation, University Minnesota, 1993. (19) Mackay, D.; Shiu, W-Y.; Ma, K-C. Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals, Vol. 2; Lewis Publishers: Chelsea, MI, 1992; pp 133135. (20) Dunn, I. J. Biotechnol. Bioeng. 1968, 10, 891. (21) Miller, R. M.; Bartha, R. Appl. Environ. Microbiol. 1989, 55, 269. (22) Volkering, F.; Breure, A. M.; Sterkenburg, A.; van Andel, J. G. Appl. Microbiol. Biotechnol. 1992, 36, 548. (23) Falatko, D. F.; Novak, J. T. Water Environ. Res. 1992, 54, 163. (24) Kile, D. E.; Chiou, C. T. Environ. Sci. Technol. 1989, 23, 832. (25) Laha, S.; Luthy, R. G. Environ. Sci. Technol. 1991, 25, 1920.

Received for review August 8, 1996. Revised manuscript received April 21, 1997. Accepted April 29, 1997.X ES960687G X

Abstract published in Advance ACS Abstracts, July 1, 1997.

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