Role of Desorption Kinetics in the Rhamnolipid ... - ACS Publications

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Role of Desorption Kinetics in the Rhamnolipid-Enhanced Biodegradation of Polycyclic Aromatic Hydrocarbons Eleonora Congiu and José-Julio Ortega-Calvo* Instituto de Recursos Naturales y Agrobiología de Sevilla (IRNAS-CSIC), Apartado 1052, E-41080-Seville, Spain S Supporting Information *

ABSTRACT: The main aim of this study was to investigate the effect of a rhamnolipid biosurfactant on biodegradation of 14Clabeled phenanthrene and pyrene under desorption-limiting conditions. The rhamnolipid caused a significant solubilization and enhanced biodegradation of PAHs sorbed to soils. The enhancement was, however, negatively influenced by experimental conditions that caused an enrichment of slow desorption fractions. These conditions included aging, a higher organic matter content in soil, and previous extraction with Tenax to remove the labile-desorbing chemical. The decline in bioavailability caused by aging on sorbed 14C-pyrene was partially reversed by rhamnolipids, which enhanced mineralization of the aged compound, although not so efficiently like with the unaged chemical. This loss in biosurfactant efficiency in promoting biodegradation can be explained by intra-aggregate diffusion of the pollutant during aging. We suggest that rhamnolipid can enhance biodegradation of soil-sorbed PAHs by micellar solubilization, which increase the cell exposure to the chemicals in the aqueous phase, and partitioning into soil organic matter, thus enhancing the kinetics of slow desorption. Our study show that rhamnolipid can constitute a valid alternative to chemical surfactants in promoting the biodegradation of slow desorption PAHs, which constitutes a major bottleneck in bioremediation.

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INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are considered as priority environmental pollutants because of their significant toxicity, carcinogenicity, and ubiquity. These compounds have a low aqueous solubility and a strong sorption tendency with solids present in soils, sediments, and waters. In particular, natural organic matter and other geochemical components, such as black carbon and clay minerals, are responsible for this strong sorption, what results in recalcitrance, especially in the case of aged contaminants.1−3 Desorption kinetics is a critical factor that, by limiting the bioavailability of these hydrophobic pollutants, often controls their biodegradation rates. Desorption of PAHs from soils is often biphasic, with a fast desorption phase (rate constant >0.1 h−1), followed of a much slower phase. Once the association between PAHs and the soil matrix occurs, microorganisms can, by active biodegradation, even enhance the desorption rate of the pollutant fractions desorbing rapidly, but not of the slow fractions.4 When connected with bioremediation, such a limitation often results in limited success because the residual concentrations after the treatment periods may still remain above legal requirements. Thus, more extensive studies are needed to enhance the bioavailability of these hydrophobic pollutants but still keep the associated risks of the enhancement at reasonable levels.5,6 The use of biosurfactants is a promising and environmentally benign alternative to chemically synthesized surfactants7,8 and dissolved organic matter9 for enhancing bioavailability of soil© 2014 American Chemical Society

sorbed PAHs. Rhamnolipid, the glycolipid produced by the bacterium Pseudomonas aeruginosa, is one of the biosurfactants most often considered for environmental applications.10 The aerobic and anaerobic biodegradability of this biosurfactant has been confirmed,11 as well as its solubilizing potential for PAHs in soil12,13 and its capacity to enhance bacterial transport.14 Previous studies have also shown that rhamnolipid shows a relatively low sorption tendency, as compared with synthetic surfactants.15 This is relevant because the sorption of surfactants to soils represents a major constraint for their application in bioremediation, since it leads to a significantly high surfactant dose and promotes the sorption of PAHs back to soil because of the modification of the soil surface and the increase in soil organic carbon content caused by the adsorbed surfactants. In spite of these advancements, the effect of rhamnolipid on the bioavailability of slowly desorbing hydrophobic compounds, such as PAHs, has not yet been elucidated. To date, no studies exist about this and other biosurfactants that integrate the processes of slow desorption and biodegradation of PAHs, which is highly necessary to understand the complex mechanisms involved in bioavailability. Therefore, we invesReceived: Revised: Accepted: Published: 10869

March 6, 2014 July 14, 2014 August 14, 2014 August 14, 2014 dx.doi.org/10.1021/es5011253 | Environ. Sci. Technol. 2014, 48, 10869−10877

Environmental Science & Technology

Article

Soils. Two sandy-loam soils, originated from a cork oak (Quercus suber L.) forest in Los Alcornocales Natural Park, Cádiz, Spain, which surrounds the Campo de Gibraltar industrial area, were selected because of their similar characteristics and class texture but their different content of organic matter. Soil I contained 6.1% organic matter and 5.5% clay, whereas soil II had 10% organic matter and 6.2% clay. Other soil properties can be found elsewhere.2 Sorption, Desorption, and Solubilization. The main objective of this work was to understand the effect of rhamnolipid biosurfactant on biodegradation of PAHs initially sorbed to soil. Therefore, the experimental conditions for sorption of this study were designed with mechanistic purposes, to take into account the possible influences of the desorption kinetics on biosurfactant action. This was achieved by changing the solid-to-water (S/W) ratio (800 and 3200 mg L−1), the contact period (from 1 d up to 60 d), the OC content of the soils, and by enrichment of the slow desorption fraction through Tenax extraction. All these conditions allowed precise measurements of the concentration of 14 C-PAHs and biosurfactant in desorption and solubilization experiments. Sorption of 14C-labeled PAHs onto soil was achieved using the batch method following a procedure described previously.3 Briefly, dry soil (16 or 64 mg) was introduced into 50 mL-glass bottles (Schott), together with 10 mL of MM containing 7.9 ± 0.3 ng mL−1 14C-phenanthrene (giving 5000 dpm mL−1) or 8.1 ± 0.4 ng mL−1 of dissolved 14C-pyrene (5400 dpm mL−1). The radiolabeled compound had been added to the aqueous solution dissolved in acetone. The bottles were closed with glass stoppers and, unless otherwise stated, they were shaken for 1 d. In preliminary biodegradation experiments, we determined that this relatively short contact period of 1 d used was sufficient to cause restrictions for biodegradation of soil-sorbed pyrene. It is possible that longer contact times would have caused different kinetics of desorption, as has been observed in other studies with PAHs using contact times ranging from 1 week to several months.2,3 However, because our main objective was to characterize desorption and biodegradation under comparable conditions, we considered it unnecessary to further extend this equilibration period, unless the specific effect of aging was included in the design of the experiments. In this case, the contact period was extended up to 60 days. After the contact period, a homogeneous sample (2 mL) was taken from the bottles, transferred to a glass tube (Pyrex, 15 mL) and centrifuged (7949 g, 1 h). The concentration of the 14C-PAH in the aqueous solution was determined by measurement of radioactivity using a liquid scintillation counter (Model LS6500, Beckman) and the sorbate concentration was calculated by difference. The solid−water distribution coefficient, Kd (in L kg−1), was calculated from the equation

tigated the effect of rhamnolipid to enhance the biodegradation of sorbed PAHs in soil−aqueous systems by comparing different experimental situations leading to dissimilar desorption kinetics.

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MATERIALS AND METHODS Chemicals. 14C-phenanthrene (52.0 mCi mmol−1, radiochemical purity >98%) and 14C pyrene (58.8 mCi mmol−1, radiochemical purity >98%), were purchased from Campro Scientific GmbH (Veenendaal, The Netherlands). Nonlabeled phenanthrene was obtained from Sigma-Aldrich, Madrid, Spain. Hexane, dichloromethane and acetone, with analysis quality, were supplied by Panreac (Barcelona, Spain). Tenax (60−80 mesh, 177−250 μm) was purchased from Teknokroma (Barcelona, Spain). Bacteria, Media, and Cultivation. Mycobacterium gilvum. The strain VM552, originated from a PAHcontaminated soil and able to degrade phenanthrene and pyrene was used as inoculum for mineralization experiments. The bacterium was grown with phenanthrene as the sole source of carbon and energy and prepared for mineralization experiments following a procedure previously described.7 The inorganic aqueous solution used in all the experiments, and called mineralization medium (MM), contained KH2PO4 (0.9 g L−1), K2HPO4 (0.1 g L−1), NH4NO3 (0.1 g L−1), MgSO4·7H2O (0.1 g L−1), CaCl2 (0.080 g L−1), FeCl3·6H2O (0.01 g L−1), and 1 mL L−1of a microelements stock solution to obtain the final concentrations of 0.0014 g L−1 for Na2MoO4·2H2O and 0.002 g L−1 for each of the following: Na2B4O7·10H2O, ZnSO4·H2O, MnSO4·H2O, and CuSO4·5H2O. The pH of this solution was adjusted to 6.7 by adding 0.05 M sodium bicarbonate. Pseudomonas aeruginosa. The strain 19SJ was selected as biosurfactant producer.13 This strain was originally isolated from a petroleum-contaminated soil. The bacterium was stored at −80 °C in glycerol and subcultured on tryptic soy agar (TSA) plates before use for rhamnolipid production. The bacterium was routinely maintained in liquid SWF medium supplied with 2% (w/v) mannitol as the sole source of carbon and energy.13 Rhamnolipid. The rhamnolipid biosurfactant from P. aeruginosa 19SJ is composed mainly of L-rhamnosyl-3hydroxydecanoyl-3-hydroxydecanoate and L-rhamnosyl-L-rhamnosyl-3-hydroxy-decanoyl-3-hydroxydecanoate, besides a small proportion of other congeners with variable length-hydrocarbon chains (C10−C18).13 The biosurfactant was extracted from the culture medium and the total rhamnolipid concentration in the sample was quantified as rhamnose equivalents (RE) with a modified colorimetric orcinol method13 and as total organic carbon (TOC) in solution, which was determined with a Shimadzu TOC-V sch analyzer. The concentration of TOC of a rhamnolipid solution with 200 mg L−1 of rhamnose equivalents was 253 mg L−1. The difference between these two values may correspond to the carbon content of the hydrocarbon chains. For operative reasons, the subsequent quantification of rhamnolipid in sorption and biodegradation experiments was performed with determinations of TOC. The rhamnolipid solutions were kept frozen until further use. Surface tension of the rhamnolipid solutions (in MM) was estimated at 23 °C with a TD1 Lauda ring tensiometer. In our conditions, the CMC of the biosurfactant was 40 mg L−1. This value falls within the commonly determined CMC range for the rhamnolipid produced by P. aeruginosa in liquid culture.

fw = 1/1 + rswKd

(1)

where f w is the compound fraction remaining dissolved in water at equilibrium and rsw is the solid−liquid phase ratio in kg L−1. A theoretical, organic-carbon based Kd value was estimated according to

Kd = foc Koc

(2)

where foc is the weight fraction of organic carbon in the solid phase and Koc is the organic carbon-normalized distribution 10870

dx.doi.org/10.1021/es5011253 | Environ. Sci. Technol. 2014, 48, 10869−10877


Article

approximately 5 × 108 cells mL−1 and a biosurfactant concentration of 400 mg L−1. The suspensions were introduced in 250 mL-Erlenmeyer flasks and the flasks were incubated at 23 ± 2 °C in an orbital shaker operating at 150 rpm. At certain time intervals, duplicate homogeneous samples of 15 mL were collected for the determination of TOC. The background concentration of TOC in MM was 8.9 ± 1.2 mg L−1, whereas biosurfactant-free bacterial suspensions had 45.2 ± 5.4 mg L−1 TOC. In these conditions, the biosurfactant TOC decreased to a final extent of 69.3% after 360 h, with a half-life of 219 ± 59 h (what indicates a reduction in 14.0% after 48 h), therefore excluding any interference from biodegradation of the biosurfactant during mineralization assays. Mineralization Experiments. Mineralization experiments with sorbed 14C-PAHs were performed as described above for sorption−desorption experiments. After equilibration, the soil suspension containing 14C-PAH was transferred to 100 mLErlenmeyer flasks. Sodium bicarbonate and rhamnolipid were added to the soil suspensions together with the bacterial inoculum (in MM, final density of approximately 5 × 108 cells mL−1) to achieve a final volume of 20 mL. The flasks were closed using Teflon−lined stoppers with a suspended 2 mL vial that contained 1 mL of NaOH 0.5 M. The flasks were shaken at 23 ± 2 °C in an orbital shaker operating at 150 rpm. Mineralization of 14 C-PAHs was monitored by 14 CO 2 production appearing in the alkali trap, following the procedure described elsewhere.13 Mineralization experiments with soil suspensions enriched for a slow desorption fraction were performed using a similar procedure. Once dissolved 14C-pyrene (15.6 ng mL−1) was left to sorb in soil suspensions during 1 d as described above, a single-step Tenax extraction was performed subsequently during 30 h. This time period was considered as sufficient to extract more than 95% of the rapid desorption fraction.4 The soil suspension was centrifuged (11 923 g, 20 min) to separate the soil from the aqueous phase. The soil pellet was resuspended with MM and then transferred to an Erlenmeyer flask, were the bacteria and biosurfactant were added to achieve a final solid concentration of 800 mg L−1. The rest of the procedure was the same as described above for mineralization experiments. A control was run to confirm that the pH used in this study was not causing an accumulation of CO2 in the aqueous phase during radiorespirometry measurements. The test was performed in duplicate in 100 mL-Erlenmeyer flasks, where 1500 dpm mL−1 of 14C-NaHCO3 were added in 20 mL of MM, and the flasks then incubated as explained above. The results showed negligible accumulation in the aqueous phase, because 98.3% of 14C-NaHCO3 was recovered as 14CO2 within 48 h. Data Analysis. Desorption and mineralization data were analyzed with the following two-compartment kinetic model:

coefficient for the compound. The value of Koc was determined by the equation log Koc = 0.98 log Kow − 0.32

(3)

where log Kow is the octanol−water partition coefficient (4.57 for phenanthrene and 5.13 for pyrene).16 The value of foc was estimated from the content of organic matter in the soil by using the van Bemmelen factor (1.724).17 Desorption experiments were performed under comparable conditions to those used for mineralization experiments. Therefore, the desorption kinetics was determined at 23 ± 2 °C by the Tenax solid-phase extraction method similar to one described previously using 14C-PAHs.4 For the sorption process, 32 mg of soil were equilibrated as described above during 1 or 60 d with 20 mL of MM that contained dissolved 14 C-pyrene or 14C-phenanthrene. After the sorption period, the suspension was transferred to a 250 mL-separation funnel, and the volume brought to 40 mL with MM (final solid-to-water ratio of 800 mg L−1). Then, 50 mg of Tenax was added, and the funnel was closed with a glass stopper and sealed with Teflon. The funnel was continuously shaken at 150 rpm in a rotary shaker at 25 °C and, at selected intervals, the Tenax was separated from the soil suspension and extracted by shaking 50 mL of acetone-hexane (1:1) during 8 h for the subsequent radioactivity determination. Fresh Tenax was added to the soil suspension, and the cycle was repeated. Solubilization in the presence of biosurfactant was measured in duplicate. Equilibrated 10 mL-soil suspensions were transferred to 100 mL-Erlenmeyer flasks, were the biosurfactant was added and the volume brought to 20 mL with MM, to achieve the desired solid-to-water ratio. At 48 h, a homogeneous sample was analyzed using the same procedure described for the sorption experiment. Control assays without biosurfactant had the same pH and concentration of sodium bicarbonate as suspensions with biosurfactant. The kinetic analysis of desorption in the presence of biosurfactant was not possible due to interferences caused by the foam on the separation of the Tenax beads. Sorption and Biodegradation of Biosurfactant. The sorption of rhamnolipid to soil was determined in 100 mLErlenmeyer flasks. Twenty-milliliter suspensions of soil (800 or 3200 mg L−1) that contained the desired concentration of biosurfactant were introduced into the flasks. The flasks were closed with Teflon-lined stoppers, and then maintained at 23 ± 2 °C in an orbital shaker operating at 150 rpm. At certain time intervals (24, 48, and 72 h), duplicate 10 mL samples were extracted from the flasks and centrifuged at 7949 g during 1 h. The concentration of rhamnolipids in the supernatant was determined as TOC. The solid−water distribution coefficient for rhamnolipids, Kd (in L kg−1), was calculated, using this TOC value, from eq 1. To exclude any interference on these measurements by dissolved organic carbon (DOC) released by the soils, we determined TOC in soil suspensions (800 mg L−1) equilibrated under comparable conditions. The values determined, 2.7 and 4.1 mg L−1 for suspensions of soils I and II, respectively, were significantly low and therefore not taken into account for Kd calculations. Biodegradation of the biosurfactant was tested in M. gilvum VM552 cell suspensions to confirm that rhamnolipid biodegradation was not fast enough to lead to significant losses during the period in which mineralization was determined. The bacterial suspensions (80 mL), which were prepared in the same way as in mineralization experiments, had a cell density of

St /So = Ffast exp( −k fast t) + Fslow exp( −kslowt )

(4)

where St and So (g) are the initial amounts of PAHs at time t (h) and at the start of the experiment, respectively. When applied to desorption, Ffast and Fslow are the fast and slow desorbing fractions, and kfast and kslow (h−1) are the rate constants of fast and slow desorption. Ffast, Fslow, kfast, and kslow were obtained by minimizing the cumulative squared residuals between experimental and calculated values of ln (St/So). Some desorption studies performed with Tenax extraction also consider a third, kinetic fraction desorbing very slowly, or Fvslow.4 However, we considered it unnecessary to use a three10871



10872

± 1.9 A ± 0.6

Errors are standard deviations of duplicate or triplicate experiments. C-PHE, C-phenanthrene; C-PYR, C-pyrene; S/W ratio and St, soil-to-water ratio and initial concentration of sorbate, respectively, in desorption and solubilization experiments; kfast, kslow, and Ffast, kinetic parameters for desorption as defined in eq 4; ND, not determined. c+ or − denotes whether the sorption period was extended (+) or not (−) from 1 to 60 d before desorption and solubilization tests. dKd, solid−water distribution coefficient. The values of log Kd predicted by Koc (log Koc = 0.98 log Kow − 0.32) are indicated in parentheses. eStandard deviation lower than 0.1. fControl and rhamnolipid, percent of PAH in the aqueous phase after 48 h in the absence and presence of 200 or 400 mg L−1 rhamnolipid, respectively. The initial PAH concentration was 4.2 ng mL−1 14C-pyrene or 3.9 ng mL−1 14C-phenanthrene. The statistical comparisons include Ffast.

14 14 14

± 0.0e C-PHE 14

a

b14

A A A B D 1.6 8.3 1.9 1.7 1.1 ± ± ± ± ± 61.7 53.3 52.5 45.1 46.7 C A C B C 3.2 7.8 6.2 6.9 2.5 ± ± ± ± ± ± 0.0 ± 0.0e ± 0.0e C-PYR 14

I I II I I

800 800 800 3200 800

− + − − −

3.7 3.8 3.9 3.7 3.3

± ± ± ± ±

0.1 (3.3) 0.0e (3.3) 0.1 (3.5) 0.0e (3.3) 0.1 (2.8)

4.7 4.8 4.8 1.3 3.5

± ± ± ± ±

0.1 0.1 1.1 0.0e 0.1

0.3 0.1 0.2 ND 0.4

e

1.0 0.5 0.4 ND 2.9

± 0.1 ± 0.0e ± 0.1

70.2 44.6 66.6 ND 78.6

± 1.8 A ± 1.9 A ± 3.7 A

10.9 8.0 9.5 7.0 18.6

± ± ± ± ±

1.2 1.3 0.2 1.8 3.4

B B B A B

44.9 37.6 29.9 27.2 29.7

400 mg L−1

rhamnolipid

200 mg L−1 control Ffast (%)

desorption

kslow (10−2 h−1) kfast (h−1) St (mg kg−1) log Kdd (L kg−1) agingc S/W ratio (mg L−1)

RESULTS Sorption and Desorption of PAHs. Sorption experiments with soils I (6.1% OM) and II (10% OM) showed significant reductions in the concentration of the dissolved PAHs after 1 d (the time course of sorption of 14C-pyrene to the soils is shown in Supporting Information Figure S1). Aging of the chemical, by extending the sorption period up to 60 d, led with soil I to a concentration of dissolved 14C-pyrene (0.76 ± 0.06 ng mL−1) that was not significantly different (P = 0.05) to that observed after 1 d of sorption (0.85 ± 0.06 ng mL−1). A similar observation was made with 14C-pyrene sorbing to soil II during 30 d. Therefore, log Kd was estimated at the end of the selected contact periods (Table 1). Not surprisingly, the values of log Kd were higher with 14C-pyrene than with 14C-phenanthrene, what is attributable to the higher hydrophobicity of 14C-pyrene. Experimental log Kd values were, in the five experimental conditions examined, in good agreement with Koc-based predictions. This agrees with previous findings indicating that organic carbon (OC) was the main factor responsible for sorption of PAHs in these soils.2 The desorption kinetics were determined, under abiotic conditions, by Tenax extraction. The solid line in Supporting Information Figure S2 represents the model results with the mean value of the parameters, obtained from the best fit to each of the duplicate experiments. The figure indicates a good fit of desorption results to the biphasic model (r2 > 0.99). The results with desorption experiments were successfully fitted, and the parameter values are shown in Table 1. The kinetic analysis of desorption in soil I showed that aging caused pyrene to desorb significantly at a slower rate. This was reflected by the high fraction of 14C-pyrene (approximately 25%) that changed into a slowly desorbing fraction. Presumably due to the higher content of OC in soil II, the kinetics of desorption of 14Cpyrene changed, as compared with soil I, by exhibiting a lower kslow value. The analysis also showed that the Ffast value was, under comparable conditions, lower with 14C-pyrene than with 14 C-phenanthrene, whereas kfast and kslow values were 2-fold lower with 14C-pyrene. These results indicate, therefore, that the choice of experimental conditions was appropriate to generate a set of diverse desorption kinetics. Therefore, our main working hypothesis, based on the dependence of biosurfactant action on desorption kinetics, could be validated. Solubilization of Sorbed PAHs by Rhamnolipid. Our working hypothesis was based on the possible dependencies of biosurfactant action on desorption kinetics. Therefore, sufficient rhamnolipid was needed in biodegradation experiments that guaranteed the presence of biosurfactant micelles in solution during incubations, accounting for possible losses of aqueous-phase rhamnolipid because of sorption onto soil and to the biodegradation of the biosurfactant itself. We chose 200 mg L−1 and 400 mg L−1 as the biosurfactant concentrations for the experiments, which were approximately 5 and 10 times the

soil

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PAH

Table 1. Distribution Coefficients, Kinetic Parameters for Desorption As Obtained with Tenax Extraction, and Solubilization in the Presence of Rhamnolipid of 14CPhenanthrene and 14C-Pyrene Sorbed to Soilsa,b

phase model to analyze desorption data, because the biphasic model detected well differences among the desorption conditions studied. The results of mineralization experiments were analyzed with the same eq 4, by substituting desorption parameters by those expressing biodegradation fractions (F′fast and F′slow) and rate constants (k′fast and k′slow). The maximum mineralization rate (ng mL−1 h−1) was calculated as kfast ′ ·Ffast ′ ·S where S is the total concentration of PAHs (ng mL−1). The software used for the minimization was Microsoft Excel 97 (Solver option).

solubilizationf (%)

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Table 2. Sorption of rhamnolipid to soilsa,b rhamnolipid 200 mg L−1

400 mg L−1

soil

S/W ratio (mg L−1)

Cw0 (mg L−1)

Cw48h (mg L−1)

Cads (mg g−1)

focc

log Kdd (L kg−1)

Cw0 (mg L−1)

Cw48h (mg L−1)

Cads (mg g−1)

focc

log Kdd (L kg−1)

I I II

800 3200 800

257 ± 1 294 ± 3 265 ± 2

242 ± 14 208 ± 4 236 ± 9

19 27 36

0.06 0.07 0.10

1.9 2.1 2.2

514 ± 2 543 ± 2 517 ± 2

477 ± 4 428 ± 5 459 ± 11

46 36 73

0.09 0.08 0.13

2.0 1.9 2.2

a Errors are standard deviations of duplicate experiments. bS/W ratio, soil-to-water ratio; Cw0 and Cw48h, initial concentration of rhamnolipid carbon (measured as TOC) in the aqueous phase and after 48 h, respectively; Cads is the rhamnolipid carbon sorbed to soil after 48 h. cfoc, fractional organic carbon of soil resulting from rhamnolipid sorption. Original foc was 0.04 for soil I and 0.06 for soil II. dKd, solid−water distribution coefficient.

other parameter estimates, k′fast and k′slow, are included in Supporting Information Table S1). Sorption reduced the mineralization rate six times, as compared with soil-free controls which contained, in the dissolved state, the same amounts of PAH. This was expected in accordance with the dissolved pyrene half-saturation constant (Km) for the bacterial strain used in this study, 1.2 ng mL−1, reported previously,3 which is well above the aqueous phase concentrations found in the biosurfactant-free controls of solubilization experiments (Table 1). The data also show that mineralization in the absence of soil particles was not affected by the presence of biosurfactant micelles. Initially, a simpler linear regression was applied to the data from the initial phase of mineralization to calculate maximum rates. However, the use of this calculation method was complicated by the limited number of time points that could be used for an optimal regression (i.e., few points were close to the regression lines), especially in the presence of soil particles. In the absence of soil particles, the occurrence of a second, slower phase of mineralization (governed by Fslow ′ and ′ ) is consistent with the production of 14CO2 from 14Ckslow labeled biomass and metabolites, what can be differentiated from that generated directly from 14C-PAHs. In spite of this limitation, which is inherent to radiorespirometry determinations,4,18 and given the good fit (as can be observed in Supporting Information Figure S3), eq 4 was used in this study to analyze the biodegradation results. Figure 1A−C shows the effect of rhamnolipid on mineralization of 14C-pyrene sorbed to soil I at two different contact periods (1 and 60 d) and with two different S/W ratios (800 and 3200 mg L−1). The results from biodegradation of 14 C-pyrene with soil II at 800 mg L−1 are included in Supporting Information Figure S4. The calculation of the maximum rates of mineralization and Ffast ′ values with eq 4 (Table 3) evidenced that, in all these situations, the biosurfactant enhanced significantly the transformation. These two parameters were highly complementary and described well the changes in mineralization caused by the biosurfactant. The enhancement was, however, influenced by aging, an increased S/W ratio, and the higher OM content of soil II. The diminishing effect of aging on the enhancement was apparent in the differences between Figure 1A and 1B, and it was quantitatively translated into smaller differences in Ffast ′ values with the control (Table 3). Doubling the concentration of biosurfactant resulted in statistically higher rates of mineraliza′ values with aged 14C-pyrene, but not with the tion and Ffast nonaged chemical. A similar impact on biosurfactant action, characterized by smaller differences in F′fast values, was observed when the S/W ratio was increased (Figure 1A and C, Table 3). ′ ) did not The kinetic constant for slow biodegradation (kslow change for the two PAHs in the presence of biosurfactants with

CMC. Independent estimations of sorption of the rhamnolipids to the soils indicated that, under the same conditions as in biodegradation assays, 70−94% of the initial TOC was still present in the aqueous phase after 48 h, thus indicating the presence of micelles in solution (Table 2). The values of log Kd for rhamnolipid sorption were, like with PAHs, higher for soil II than for soil I. The log Kd value for rhamnolipid remained unchanged with both soils at the two rhamnolipid concentrations tested, and with soil I at soil-to-water ratios of 800 and 3200 mg L−1. Sorbed rhamnolipid contributed with increments in soil foc between 0.02 and 0.05 for soil I and between 0.04 and 0.07 for soil II. Rhamnolipid caused a significant solubilization of sorbed PAHs (Table 1). Doubling the concentration of biosurfactant caused in all cases a higher aqueous concentration of PAHs. The relative enhancements of aqueous concentration caused by the biosurfactant were higher with 14C-pyrene than with 14Cphenanthrene, what was very likely related to the hydrophobicity of the two PAHs and the tendency to be accommodated into biosurfactant micelles. Without aging, rhamnolipid brought into solution a significant fraction of 14Cpyrene (at 400 mg L−1, 61.7% of the initial pollutant concentration, 4.2 ng mL−1), what led to an aqueous concentration well above that caused by spontaneous desorption (10.9% or 0.5 ng mL −1 ). This extent of solubilization by rhamnolipid was not statistically different to the Ffast value (as revealed by Tenax extraction, Table 1). This means that, in these suspensions, slow desorption 14C-pyrene was still associated with soil particles. Aging caused lower aqueous-phase 14C-pyrene concentrations in these solubilization assays, with and without rhamnolipid in solution. The fraction of 14C-pyrene solubilized in aged suspensions was, at both biosurfactant concentrations, not different to Ffast. Solubilization of 14C-pyrene sorbed to soil II resulted, at both biosurfactant concentrations, in lower aqueous phase 14Cpyrene concentrations than with soil I, possibly as a result of the higher sorption capacity of soil II for 14C-pyrene. Like with soil I, the maximum solubilization achieved by rhamnolipids in soil II did not account for the slow desorbing fraction. The 4fold increase in S/W ratio with soil I (3,200 mg L−1) had a negative effect on biosurfactant effectiveness, what was evidenced by an approximate decrease in 18% of substrate solubilized at both biosurfactant concentrations. Biodegradation of Sorbed PAHs in the Presence of Rhamnolipid. The impact of sorption to soil I on mineralization of pyrene is shown in Supporting Information Figure S3. The same sorption−desorption conditions as those used in experiments from Table 1 (S/W ratio 800 mg L−1, contact time 1 d) were applied to this experiment. Maximum ′ values are shown in Table 3 (the mineralization rates and Ffast 10873


C-PHE 14

soil I, but it was 5-fold lower with pyrene sorbed to soil II (Supporting Information Table S1). The difference is a reflection of the different shape in the mineralization curves observed after 60 h (Supporting Information Figure S4), which lead to progressively lower rates of mineralization with biosurfactant, as compared with the biosurfactant-free controls. The exact reason for the slightly higher mineralization rates of pyrene in the controls with soil II, as compared with soil I (Table 3), remains unknown. It may be related to the slightly higher concentration of DOC in the suspensions (see Materials and Methods section), and the possible enhancement of mineralization of PAHs because of DOC.19 We conducted further solubilization and mineralization experiments to determine how biosurfactant action was affected by a more drastic change in desorption kinetics. For this aim, we eliminated most of the labile fraction of 14C-pyrene sorbed to soil after 1 d by Tenax extraction before inoculation. These experiments were performed using the same procedure as described for Tenax desorption experiments. After 30 h of extraction, the Tenax was removed, the aqueous phase and the soil were transferred to an Erlenmeyer flask, and the flask was immediately inoculated. From that point the procedure was the same as described previously. Figure 1D shows the mineralization results, calculated with respect to the contaminant that remained in the suspensions after Tenax extraction (1.8 ng mL−1, approximately 30% of the initial concentration). We assumed that Tenax acted as a perfect sink for desorption and that the immediate inoculation further maintained this situation through biodegradation. These two facts would minimize redistribution between the fast and slow domains, because sorption back to the soil would be neglected. The results are compared in Figure 1D with the mineralization of 14C-pyrene determined simultaneously under the same sorption conditions, but without the previous extraction step with Tenax. The effect of rhamnolipid on biodegradation was, again, more evident when most of 14C-pyrene was present as a fast desorption compound. However, the biosurfactant also enhanced the transformation in the Tenax-extracted soil. In the latter case, the maximum mineralization rate in the presence of rhamnolipid doubled the rate to that in the biosurfactant-free control, but the difference in F′fast between these two treatments resembled that observed after aging (Table 3). The results from solubilization tests (shown in Supporting Information Table S2) followed a similar pattern as mineralization, i.e., the biosurfactant efficiency decreased significantly. However, the solubilization extent achieved (40% of the total concentration, 1.8 ng mL−1) indicates that the prior removal of most of the fast desorption fraction allowed, at least partially, the solubilization of the slow desorption pyrene.

Errors are standard deviations of duplicate experiments. Statistical comparisons were performed with ANOVA at P = 0.05. b14C-PHE, 14C-phenanthrene; 14C- PYR, 14C-pyrene; S/W ratio and St, soil-to′ in the absence and presence of 200 water ratio and initial concentrations of sorbate, respectively, in biodegradation experiments; ND, not determined. Control and rhamnolipid, mineralization rate or Ffast or 400 mg L−1 rhamnolipid, respectively. c+/− denotes whether the sorption period was extended (+) or not (−) from 1 to 60 d before mineralization tests. d+/− denotes whether equilibrated soil was extracted (+) or not (−) with Tenax for 30 h to remove the most-labile-desorbing fraction before mineralization tests. The concentration of pyrene initially present was 1.8 ± 0.5 and 7.8 ± 0.0 ng mL−1, respectively, for the equilibrated soil extracted (+) or not (−) with Tenax. eStatistical comparisons performed with ANOVA at P = 0.1. fStandard deviation lower than 0.1. gBecause of the relative high ′ in this control, the value was calculated as rate = [(Ffast ′ × kfast ′ ) + (Fslow ′ × kslow ′ )] × S value of kslow

± 1.9 B ± 3.6e B ± 0.9 B ± 9.5e B ± 34.7 A ± 6.3 B

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■

DISCUSSION The coexisting processes of desorption from soil and biodegradation by an active microbial population provided us with a new scenario to investigate the influence of rhamnolipid on the bioavailability of PAHs. A previous study reported the enhanced biodegradation of crystalline phenanthrene in the presence of externally added rhamnolipid.20 The role of this biosurfactant in promoting the dissolution, partitioning and biodegradation of PAHs has also been examined previously by our group.13,21 However, the aim of our study was not to confirm previous findings, but to determine whether biosurfactant-enhanced biodegradation was influenced by the kinetics of desorption of PAHs from soil. Our results not

a

5.8e B 0.5e B 0.2 B 0.4 C 1.5e B 2.7 B ± ± ± ± ± ± ND 36.7 15.7 16.6 44.5 29.4 34.3 ± 3.9 A ± 3.4e B ± 1.5e A

51.8 36.2 15.0 ND 39.9 29.9 35.8 6.3 A 4.8e A 0.4e A 1.8 A 0.5 A 0.3e A 1.1 A ± ± ± ± ± ± ± 51.1 20.8 12.7 11.5 26.1 20.7 21.1 26.3 B 1.2 C 0.0f B 16.1e B 14.0 B 3.5 B ± ± ± ± ± ± ND 175.1 58.3 11.5 173.4 147.4 164.0 ± 52.6 A ± 3.7 B ± 1.6 B

463.2 131.9 34.7 ND 126.7 101.0 155.3 24.5 A 15.7 A 3.4 A 0.1 A 9.0e,g A 0.9 A 8.3 A ± ± ± ± ± ± ± − − − + − − −

421.2 65.1 25.5 4.6 100.3 67.2 83.1

400 mg L−1 200 mg L−1

− − + − − − − 0.1 0.1 0.6 1.1 0.0f 0.1 ± ± ± ± ± ± NA 4.7 4.8 2.4 4.8 1.3 3.5 NA 800 800 800 800 3200 800 none I I I II I I C-PYR 14

rhamnolipid

200 mg L−1 control Tenaxd agingc St (mg kg−1) S/W ratio (mg L−1) soil PAH

mineralization rate (ng L−1 h−1)

Table 3. Enhanced Biodegradation by Rhamnolipid of 14C-Phenanthrene and 14C-Pyrene Sorbed to Soilsa,b

400 mg L−1

control

F′fast (%)

rhamnolipid


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Figure 1. Biodegradation of 14C-pyrene sorbed to soil I (6% organic matter, OM) at 800 (A) and 3200 mg L−1 (C) in the absence (squares) and presence of 200 (triangles) and 400 mg L−1 (circles) of rhamnolipid biosurfactant. Panel B represents mineralization under the same conditions as (A) but extending the previous sorption period to 60 d to induce aging. The sorption period in panels A and C was 1 d. Panel D: Biodegradation of slow desorption 14C-pyrene in soil I at 800 mg L−1 in the absence (squares) and presence (circles) of rhamnolipid at 400 mg L−1. White and black symbols indicate, respectively, whether equilibrated soil was extracted or not with Tenax for 30 h to remove the most-labile-desorbing fraction before biodegradation experiments. The concentration of 14C-pyrene initially present was 7.8 ng mL−1 (white symbols) and 1.8 ng mL−1 (black symbols). In the four panels, the solid and dashed lines represent fitting mineralization results to eq 4

only show that rhamnolipid enhanced biodegradation of PAHs under desorption-limiting conditions, but also that the enhancement was affected by changes in the relative abundance of the fast and slow desorption fractions. The biosurfactant caused an enhancement in biodegradation of sorbed PAHs even when they were mainly present as fast desorption chemicals. This can be understood if the aqueous concentration of PAHs generated as a result of fast desorption did not exceed the bacterial demand. Therefore, biodegradation rates would be controlled by the rates of desorption. This explanation is supported, in the case of pyrene, by the dissolved pyrene concentrations detected in solubilization experiments, which, in the absence of rhamnolipid, were lower than Km. This is significant because, at this concentration range, biodegradation rates are directly related to the aqueous-phase concentrations of substrate. Under the bioavailability restrictions imposed by desorption, active populations of degrading bacteria drive the aqueous concentration of the compound down to a steady-state level caused by desorption inputs and biodegradation removals.3 Therefore, the aqueous concentration during biodegradation would have been even lower than those independent estimations. Given that the acquisition by bacterial cells of pyrene present in biosurfactant micelles was as fast as that of the freely dissolved compound (Table 3), micellar solubilization, and the enhanced desorption resulting from it, caused by rhamnolipid led to higher biodegradation rates due to the bacterial exposure to a higher steady-state concentration of total pyrene in the aqueous phase.

The results showed that pyrene became more resistant to biodegradation as the contact time of the compound with the soil increased. This is in agreement with previous research.1−3 Our results extend those findings by showing that such a decline in bioavailability because of aging can be reversed in the presence of a biosurfactant. The maximum mineralization rate of the aged compound doubled its value with rhamnolipid, and the Ffast ′ values increased significantly. The efficiency of biosurfactant action was, however, lower than with the unaged chemical, as it was evidenced by the relative differences in the F′fast values. This loss in biosurfactant effectiveness in enhancing the biodegradation of pyrene is in apparent contradiction with the relatively high solubilization extents still observed after aging (Table 1). These extents accounted for the fastdesorption compound only, as occurred with the other sorption situations studied. The results can be explained by postulating that slow desorption fractions formed after aging, with a significantly lower kslow, constituted a rate-limiting process for solubilization of pyrene during the active phase of biodegradation. This could be caused by intra-aggregate diffusion of the sorbed chemical during aging. The occurrence of this mechanism is supported in our study by the enhancing effect of doubling the concentration of biosurfactant on biodegradation of aged pyrene, which did not occur without aging. At a higher biosurfactant concentration, the content in sorbed rhamnolipid obviously increased in soil (Table 2). We also observed that sorption of rhamnolipid increased with soil organic carbon content, what suggests that partitioning into soil organic matter was involved in the sorption process. In another study, sorption 10875



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Notes

of rhamnolipid was, however, proposed to occur through an adsorption process at the soil-water interface.12 The reason for this discrepancy may rely on the difference with the OC contents of the soils from that study (0.03−0.36%), which were significantly lower than the OC content in the soils studied here. Therefore, sorbed biosurfactant molecules may have penetrated into the soil aggregates, and caused swelling of the sorbent. This mechanism has already been proposed to explain the enhanced kinetics of desorption of PAHs by synthetic, nonionic polyoxyethylene surfactants in a tar-contaminated soil from a manufactured gas plant (MGP), with an organic carbon content of 62%.22 This, however, should be the subject of further investigation. Nevertheless, in our study, sorption of rhamnolipids may have also occurred through adsorption onto mineral sites, a possibility that was not excluded by our results. Our study extends the limited knowledge of the mechanisms for increasing the bioavailability of PAHs under sorption restrictions, which is one of the bottle-necks in bioremediation and limit its competitiveness. The positive effect on biodegradation of aged pyrene, and the enhancement caused by doubling the biosurfactant dose, are promising. However, the reduction of the biodegradation enhancement by the biosurfactant after the aging period considered (60 days) raises the question on its efficiency with pollutants aged in fieldcontaminated soils for longer periods, of up to several years. To our knowledge, there are no studies focused on the effects of biosurfactants on desorption and biodegradation of contaminants aged under these conditions. Nevertheless, recent studies with synthetic surfactants and recalcitrant PAH fractions present in field contaminated soils indicate that the enhancement is possible. For example, a nonionic surfactant (Brij 30) caused the enhanced desorption and biodegradation of residual PAHs in a soil from a MGP site that had first treated in an aerobic bioreactor.8 In another study, performed with a bioremediated soil from a MGP site that was enriched in slow desorption PAHs, the nonionic surfactant Brij 35 succeeded, through solubilization, in reducing further the residual levels of high-molecular-weight PAHs after slurry-phase bioremediation.7 According to sustainability aspects involving the environmental use of chemical surfactants, such as toxicity and soil quality, the enhancement of biodegradation of slow desorption PAH fractions by biosurfactants is therefore of considerable interest. However, at present the application of rhamnolipid is limited by their high cost of production. Once this limitation is overcome, this microbial biosurfactant could be a good alternative for bioremediation in PAH-contaminated soils. Its action on slow-desorption PAHs would minimize the risk associated with increased concentrations of solubilized PAHs and metabolites up to levels that may be toxic to microorganisms or exceed their metabolic potential.

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The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Support for this research was provided by the Spanish Ministry of Science and Innovation (CGL2010-22068-C02-01 and CGL2013-44554-R), the Andalusian Government (RNM 2337) and the FPI Programme (E.C.).

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ASSOCIATED CONTENT

S Supporting Information *

Tables showing enhanced biodegradation and effect of Tenx extraction and figures showing time course of sorption of 14Cpyrene, kinetics of abiotic desorption of 14C-pyrene and 14Cphenanthrene, and biodegration of 14C-pyrene. This material is available free of charge via the Internet at http://pubs.acs.org/.

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REFERENCES

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*Phone: +34 95 4624711. Fax: +34 95 4624002. E-mail: [email protected]. 10876



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