Carbon nanotubes (CNTs) have attracted extensive attention since their discovery because of their unique physicochemical properties 1. They have been widely used as catalyst supports and protein biosensors and have potential applications as electrodes in batteries and supercapacitors 2, 3. With sharply increasing applications and production, more and more CNTs will be released into the environment, and their strong adsorption capability may affect the transfer, transformation, and fate of hydrophobic organic contaminants (HOCs) in the environment 4–6. In addition, CNTs have been proposed as a new potential sorbent to remove HOCs from waste water or reduce the bioavailability of HOCs in sediments or soils 7–12. A better understanding of CNT–HOC interaction mechanisms is thus essential for environmental applications of CNTs and health risk assessment of both HOCs and CNTs in the environment.
So far, many studies have been carried out to examine the adsorption of different organic compounds—for example, polycyclic aromatic hydrocarbons and hydroxyl-, amino-, or chloro-substituted aromatics—on CNTs 7, 13, 14. It has been reported that CNTs had much higher efficiency to remove some organic pollutants than activated carbon or black carbon (BC) 15 and that the sorption of HOCs by CNTs was regulated by multiple mechanisms 16. Yang and Xing 13 and Yang et al. 17 reported that the Polanyi model was most appropriate to depict sorption of aromatics by CNTs among several models used for data fitting. These studies are helpful for elucidating sorption mechanisms of HOCs by CNTs.
Biodegradation is a key process for the removal of HOCs from the environment, but until now, only a few studies have been carried out to analyze the biodegradation and mineralization of HOCs sorbed on CNTs. For example, Xia et al. 18 studied the bioavailability of phenanthrene in three artificial sediments amended with two kinds of BC and one kind of multiwalled CNTs (MWCNTs), respectively. They found that microbes can utilize a fraction of phenanthrene sorbed on CNTs after aging for 21 to 40 d, but the biodegradation and mineralization efficiencies of phenanthrene associated with CNTs were significantly lower than with BC. Towell et al. 19 suggested that the addition of CNTs to soil can reduce PAH extractability and bioaccessibility. In addition, some researchers have suggested that CNTs could be toxic to and exert negative influences on bacteria 20–22. Chung et al. 22 investigated the short-term effect of MWCNTs on the activity and biomass of microorganisms inhabiting two different soil types in an incubation study and found that all enzymatic activities as well as microbial biomass C and N were significantly lowered under 5,000 µg MWCNT g−1 soil. However, very few studies have been carried out on the effect of CNT characteristics on the bioavailability of HOCs sorbed on CNTs to bacteria, and the relationship between desorption and biodegradation or mineralization of HOCs sorbed on CNTs is not well understood.
In the present study, 14C-labeled and unlabeled phenanthrene was spiked onto four kinds of MWCNTs with different physiochemical properties and aged for 60 d. The mineralization of sorbed phenanthrene on MWCNTs was examined, and the effects of physiochemical properties of MWCNTs were studied. The effects of MWCNTs on the cell morphology and the growth of degrading bacteria were investigated. Furthermore, the mineralization of phenanthrene on MWCNTs was compared with the abiotic desorption of phenanthrene conducted simultaneously with a porous polymer resin Tenax TA sorbent, which we described previously 23, and the relationship between mineralization and desorption of phenanthrene sorbed on MWCNTs was analyzed.
MATERIALS AND METHODS
The [9-14C] phenanthrene (52 mCi/mmol, 0.05 mCi/ml) and NaH14CO3 (52 mCi/mmol, 0.25 mCi/ml) were obtained from Moravek Biochemicals, and the liquid scintillation cocktail (Ultima Gold) was obtained from PerkinElmer Life and Analytical Sciences. Liquid phenanthrene certified reference materials (CRMs) were from the National Research Center for Certified Reference Materials of China and solid phenanthrene (98 + % pure) from Sigma-Aldrich. For the preparation of phenanthrene aqueous solution, a total of 1.5 mg unlabeled phenanthrene was added to 2 L of sterilized mineral solution (NaN3, 200 mg/L). After sonication for 2 h, the solution was shaken under 24°C and 0.18 g for 12 h, it was then filtered through a 0.45-µm glass fiber 0.18 g membrane to remove the undissolved phenanthrene. The phenanthrene concentration in the solution was determined by high-performance liquid chromatograph (HPLC). M-terphenyl (AccuStandard) was chosen as the internal standard for instrumental determination of phenanthrene, it was prepared as a stock solution in hexane (10 mg/L) and stored in a sealed brown flask with Teflon cap at −4°C until use. NaN3, used to inhibit microbial metabolism, was from Sigma. Methanol (HPLC grade) was from Fisher Scientific, and n-hexane (HPLC grade) was from Beijing Beihua Jingxi Chemical Company of China.
Characterization of MWCNTs
Four kinds of MWCNTs (MWCNT1, MWCNT2, MWCNT3, and MWCNT4) were obtained from Beijing Nachen S.T., with different specific surface area and pore size to represent a wide range of MWCNTs with different characteristics. The MWCNTs were ground and sieved through 100 mesh, then rinsed with Milli-Q water to wipe off suspended parts, dried at 80°C, and extracted with acetone and n-hexane three times. Nitrogen adsorption–desorption isotherms of MWCNTs were measured volumetrically at 77K using a Micromeritics ASAP-2020. Prior to adsorption, the samples were outgassed at 105°C for at least 16 h. Specific surface area (SSA) was calculated by multipoint Brunauer–Emmett–Teller. Mesopore volume (Vmeso) and average diameter (Lpore) were calculated from desorption isotherms by the Barrett–Joyner–Halenda method. Micropore volume (Vmicro) was calculated by t-plot methods. The physicochemical properties of the MWCNTs are shown in Table 1.
Table 1. Select properties of multiwalled carbon nanotubes (MWCNTs) and desorption and mineralization efficiencies of phenanthrene
Specific surface area (SSABET) was calculated from the desorption isotherm of N2 at 77 K by the multipoint Brunauer–Emmett–Teller method.
Micropore volume (Vmicro) was calculated by the t-plot method.
Mesopore volume (Vmeso) and average diameter (Lpore) were calculated from desorption isotherms by the Barrett–Joyner–Halenda method.
Isolation and enrichment of phenanthrene-degrading strains
Phenanthrene-degrading bacteria were isolated from the soil sample collected from a coking plant by continual acclimation at 28°C. Purification was performed by adding an aliquot of the solution containing bacteria to autoclaved minimum basal salts (MBS) containing 1 mg/L phenanthrene, and the pH value was adjusted to 7.4 to 7.6 with NaOH solution. The MBS was composed of 5 ml of phosphate buffer (8.5 g of KH2PO4, 21.75 g of K2HPO4 · H2O, 33.4 g of NaHPO4 · 12H2O, 5 g of NH4Cl, 1000 ml of sterilized water), 3 ml of magnesium sulfate solution (22.5 g of MgSO4, 1000 ml of sterilized water), 1 ml of calcium chloride solution (36.4 g of CaCl2, 1000 ml of sterilized water), 1 ml of ferric chloride solution (0.25 g of FeCl3, 1000 ml of sterilized water), 1 ml of trace-element solution (39.9 mg of MnSO4 · H2O, 42.8 mg of ZnSO4 · H2O, 34.7 mg of [NH4]6Mo7O24 · 4H2O, 1,000 ml of sterilized water), and 990 ml of Milli-Q water. After enrichment in MBS solution with phenanthrene, the cells were resuspended and washed twice with phosphate buffer (2.185 g of Na2HPO4 · 12H2O, 0.609 g of NaH2PO4 · 2H2O, 100 ml of sterilized water, pH 7.4–7.6) prior to the biodegradation experiment. The isolated strain was identified by the 16 rDNA analysis as Methylophilus methylotrophus sp. When the initial concentration of phenanthrene was 0.378 mg/L in the water phase, more than 90% of phenanthrene could be biodegraded by Methylophilus methylotrophus at 24°C in 2 d, indicating that this kind of bacteria has a high biodegrading ability for phenanthrene.
Mineralization of 14C-labeled phenanthrene on MWCNTs
For the spiking of 14C-labeled phenanthrene in MWCNTs, a total of 200 µl 14C-labeled phenanthrene (0.015 mCi/ml) dissolved in methanol was added to each 50-ml glass screw-cap centrifuge tube. The methanol was subject to evaporation before the addition of 50 ml of phenanthrene aqueous solution and 0.05 g of MWCNTs, with an initial 14C radioactivity of 2.928 µCi (0.200 mg/L labeled phenanthrene), unlabeled phenanthrene concentration of 0.378 mg/L, and total phenanthrene concentration of 0.578 mg/L. These tubes were shaken (0.19 g) at 24°C in darkness for 60 d. After standing for 12 h, the supernatant was removed and the radioactivity of the supernatant measured. Then, the MWCNTs were transferred to specially designed bottles with 95 ml fresh mineral solutions for mineralization experiments, and 5 ml cell suspension containing approximately 108 cells was added to each bottle, which included a Teflon-lined screw cap and a CO2 trap containing 2 ml NaOH (1 M) in a suspended 5-ml beaker. All of the bottles were sealed with Teflon-wrapped silicon septa to minimize phenanthrene loss through volatilization and inoculated on a rotary shaker (0.18 g) operating at 24°C under dark conditions. Triplicate CO2 trapping solutions were removed at appropriate time intervals and analyzed for 14C by liquid scintillation counting (Packard Tri-Carb 2900TR), and new CO2 trapping solutions were placed again. Control experiments with 200 mg/L NaN3 and without microbes were conducted in triplicate. At the end of incubation (i.e., after incubation for 70 d), 4 ml H2SO4 (2 M) was added into each bottle to outgas all 14CO2 in the solution. Minimum basal salt solution and each bottle were sterilized before the incubation, so the depletion of oxygen was caused only by the phenanthrene-degrading bacteria, and it was very low (<1 mg/L). In addition, samples were collected at intervals to replenish oxygen in the system. Therefore, the incubations were under fully aerobic conditions. To compare the mineralization and desorption of phenanthrene on MWCNTs, the desorption data obtained simultaneously 23 were used to analyze the relationship between the mineralization and the desorption of phenanthrene. The desorption experiments are described in the Supplemental Data, and the desorption results are shown in Supplemental Data, Figure S1 and Table S1.
To measure the bacteria growth in the mineralization experiment, an experiment was carried out parallel with the mineralization experiments mentioned above but without 14C-labeled phenanthrene (total phenanthrene concentration was 0.578 mg/L). At predetermined time intervals, the samples were mixed in a vortex for homogeneity; 1 ml samples containing MWCNTs and water were collected for microbiological analyses.
Chemical and microbiological analyses
Phenanthrene analysis. Phenanthrene in aqueous phase was analyzed by using HPLC (Waters model 1525) with a Shimadzu ODS2 column (5 µm, 4.0 × 250 mm) and a fluorescence detector (Waters model 474). The instrumental conditions were consistent with our previous research 18.
14C quantification. A total of 200 µl 14CO2 trapping solution in the mineralization experiments was mixed with 10 ml Ultima Gold, shaken vigorously for homogeneity, and placed in the dark for 24 h to dissipate chemiluminescence before counting in the liquid scintillation counter. The test of 14CO2 trapping efficiency was carried out with NaH14CO3 according to Xia et al. 18.
Microbiological analyses. For the counting of bacteria during incubation, cells were enumerated by measurement of colony-forming units on plate-counter agar, following standard microbiological techniques. The detailed processes for the determination of PAH-degrading bacteria have been described by Xia et al. 24.
Quality assurance and quality control
The correlation coefficient for the calibration curve of phenanthrene with HPLC was higher than 0.999. The limit of detection was 0.10 µg/L for HPLC. The radioactivities of unlabeled phenanthrene aqueous solution and Na2CO3/NaHCO3 and NaOH solutions were similar to background levels (1.4 × 10−4 µCi/ml). The trapping efficiency of 14CO2 ranged from 98.0% to 102.3%. The mean relative standard deviations for the mineralization experiments were less than 10%.
RESULTS AND DISCUSSION
Mineralization rate of phenanthrene on MWCNTs
After aging for 60 d, the concentration of 14C-labled phenanthrene was less than 0.05 µg/L in the water phase; because the initial concentration of labeled phenanthrene was 0.200 mg/L, the content of 14C-labled phenanthrene sorbed on each MWCNT was 200 µg/g, and the radioactive concentration was 58.56 µCi/g. The definition of mineralization in the present study is the conversion of phenanthrene to an inorganic substance such as CO2 by the action of bacteria. Thus, the mineralization rate could be expressed as the production rate of CO2 from phenanthrene. As shown in Figure 1, there were fast, slow, and very slow mineralization stages for phenanthrene sorbed on MWCNTs. The production of 14CO2 increased quickly during the first 5 d, then increased slowly thereafter. With the subtraction of 14C release in the control experiments, the mineralization efficiencies of phenanthrene after incubation for 35 d were 2.38, 8.84, 23.88, and 31.47% for MWCNT1, MWCNT2, MWCNT3, and MWCNT4, respectively. This was much lower than the results of our previous study 18, in which the mineralization efficiency of phenanthrene sorbed on MWCNT with a SSA of 88.0 m2/g (aging for 28 d) was 54.2% after incubation for 21 d.
As shown in Table 2, the maximum mineralization rate of phenanthrene on MWCNT4, calculated from the slope of the two consecutive points (accumulated mineralization) showing the largest increase, was approximately 10 times for MWCNT1, 4 times for MWCNT2, and 1.5 times for MWCNT3, like the average mineralization rate calculated from the ratio of the overall mineralization efficiency to the incubation period (35 d). In addition, the first-order kinetics were used to study the mineralization process of phenanthrene during the first 5 d
The integrated form of the above equation is
where C and Co (mg/L) are the phenanthrene concentrations at time t (d) and time zero, respectively, and k (1/d) is the first-order rate constant. For the mineralization of phenanthrene, the evolved 14CO2 molecules were assumed to have a 1:1 stoichiometric relationship with the 14C-phenanthrene molecules, given that the latter were singly labeled (at only the C-9 position). The first-order rate constant of phenanthrene mineralization could be obtained by fitting the mineralization efficiency with the equation
where x (%) is the accumulative mineralization efficiency, corresponding to the ratio of evolved 14CO2 to the injected 14C-phenanthrene. According to the results shown in Figure 1, the first-order kinetic constants (k) were obtained for each system for the first 5 d of mineralization, and the sequence of k values among the four systems was MWCNT1 < MWCNT2 < MWCNT3 < MWCNT4 (Table 2).
Table 2. Maximum and average rates and first-order rate constants for phenanthrene mineralization in the microcosms with different kinds of multiwalled carbon nanotubes (MWCNTs)
Maximum mineralization rates (%/d)
Average mineralization rates (%/d)
First-order rate constants for the first 5 d (1/d)
Growth of bacteria during mineralization process
As shown in Figure 2, for the four kinds of MWCNT systems, the density of bacteria increased significantly during the first 10 d; thereafter, it decreased significantly. During the initial 10 d, the higher desorption rates of phenanthrene from MWCNTs (Supplemental Data, Fig. S1) provided enough carbon source for the growth of bacteria; after that, the lower desorption rate could not provide enough carbon for bacteria, so the population of bacteria decreased quickly. During incubation, the sequence of bacteria population in the four systems was MWCNT1 < MWCNT2 < MWCNT3 < MWCNT4. For example, after incubation for 35 d, the bacteria density of the MWCNT4 system was approximately seven times that of MWCNT1, three times that of MWCNT2, and 1.6 times that of MWCNT3 systems. This was caused by the difference in phenanthrene desorption among the four systems. After incubation for 35 d, the density of bacteria had a significant positive correlation with both the desorption and the mineralization efficiencies (p < 0.05).
To study the effect of MWCNTs on the growth of bacteria during incubation, the morphology of MWCNTs, bacteria, and bacteria mixed with MWCNTs for 5 d was evaluated from scanning electron microscope (SEM) images. As shown in Figure 3 and Supplemental Data, Figure S2, compared with the initial state of bacteria, the integrity of bacterial morphology was modified after mixing with MWCNTs for 5d, and some bacteria changed from round to flat rods. This suggested that the biodegrading bacteria might undergo membrane damage of different degrees resulting from the presence of MWCNTs. Kang et al. 25 and Chipot and Tarek 26 also suggested that CNTs could penetrate through the cell membrane because of their cylindrical shape and high aspect ratio and lead to cell death.
Relationship between desorption and mineralization of phenanthrene
According to the desorption 23 and mineralization efficiencies of phenanthrene at different time points, a positive correlation was observed between Tenax TA-assisted desorption and mineralization efficiencies for the four kinds of MWCNTs (p < 0.01; Fig. 4). This indicated that the mineralization efficiency of phenanthrene associated with MWCNTs was controlled mainly by the desorption process, and its bioavailability to bacteria could be directly predicted from Tenax TA-assisted desorption. As shown in Figure 1, after incubation for 30 d, the mineralization percentage increased very slowly. To determine the reason for this, after incubation for 35 d, a 5-ml cell suspension containing approximately 108 cells was added to each incubation system, and the pH value was adjusted to approximately 7.4. The results showed that during the following 30 d, there was no significant increase in mineralization efficiencies of the four kinds of MWCNT systems (Fig. 1, dashed boxes). This implies that the low mineralization rate after incubation for 30 d was caused not by the bacteria but by the bioavailability of phenanthrene; that is, the bioavailability of sorbed phenanthrene was the key factor that determined its mineralization efficiency in the present study.
As shown in Figure 4, the mineralization rates were lower than the desorption rates. This was mainly because the phenanthrene desorbed from the MWCNTs could not be fully transformed to CO2. As shown in Supplemental Data, Figure S3, after incubation for 35 d, the pH values decreased from 7.6 to 6.64, 6.53, 6.42, and 6.37 for MWCNT1, MWCNT2, MWCNT3, and MWCNT4 systems, respectively, because of the accumulation of acidic metabolites of phenanthrene. Some studies have also indicated that acidic metabolites were produced during phenanthrene biodegradation and that the dynamics of pH changes in cultures were consistent with the concentration change of PAHs 27–30. In addition, the slopes of mineralization to desorption percentages were over 0.5 for MWCNT2, MWCNT3, and MWCNT4, whereas the slope was only 0.21 for MWCNT1 (Fig. 4). According to the results shown in Table 1, the desorption efficiency of phenanthrene on MWCNT4 was approximately four times that on MWCNT1 after desorption for 35 d, and the mineralization efficiency was approximately 12 times that of MWCNT1 after incubation for 35 d. Because the SSA of MWCNT1 was the largest, the intermediate products of phenanthrene biodegradation might be sorbed by MWCNT1, and its bioavailability would be decreased most. Kang et al. 25, 31 pointed out that the toxic effects of CNTs on bacteria would become stronger with decreasing diameter and increasing SSA of CNTs. In the present study, the diameter was the smallest and the SSA was the highest for MWCNT1, so its toxic effect might be stronger than that of the other three kinds of MWCNTs, making the activity of biodegrading bacteria decrease the most. Although the MWCNTs might have toxic effects on the bacteria, according to the results of bacteria growth and phenanthrene mineralization, the presence of MWCNT2, MWCNT3, and MWCNT4 with relatively lower SSAs did not exert a significant influence on the biodegrading ability of bacteria for phenanthrene.
As shown in Figure 5, after incubation for 35 d, the mineralization efficiencies of phenanthrene were positively correlated with the sum of fast and slow fractions of Tenax TA-assisted desorption (p < 0.01). In addition, the mineralization efficiency of phenanthrene was much higher than the fast desorption fraction; for instance, the mineralization efficiency was 37.99%, whereas the fast fraction was only 3% for MWCNT4 system. These results suggested that the slow desorption fraction on MWCNTs was bioavailable and should not be ignored in toxicity assessment. The sum of fast and slow desorption fractions could serve as an indicator for the bioavailability of phenanthrene sorbed on MWCNTs to bacteria. Cui et al. 32 studied the bioavailability of phenanthrene (without aging) in sediment, they found that the slow desorption fraction (using Tenax TA) was also bioavailable to Chironomus tentans. In addition, as shown in Figure 5, the mineralization efficiencies of phenanthrene significantly correlated with the initial 12 h desorption fraction (R = 0.998, p < 0.01), suggesting that the initial 12 h desorbing fraction could be a fast indicator for the bioavailability of phenanthrene on MWCNTs to bacteria.
Effects of MWCNT physicochemical properties on the mineralization and desorption of phenanthrene
The effect of MWCNT characteristics on the desorption and mineralization processes of phenanthrene was significant. As shown in Figure 6, the increasing SSA and mesopore and micropore volume of MWCNTs caused a significant decrease in the mineralization efficiency of phenanthrene (p < 0.01). As shown in Supplemental Data, Figure S4, the desorption efficiencies of phenanthrene had significant negative correlations with Vmeso (p < 0.01) and SSA (p < 0.05) and a positive correlation with Lpore (p < 0.05) of MWCNTs. Also, the desorption efficiency of phenanthrene was negatively correlated with Vmicro.
According to the results shown in Table 1 and Supplemental Data, Table S1, the sum of the fast and slow fractions of desorption was significantly negatively correlated with the sum of Vmeso and Vmicro (r = −0.996, p < 0.01), whereas the very slow fraction of desorption was significantly positively correlated with the sum of Vmeso and Vmicro (r = 0.996, p < 0.01). Based on these results, the influencing mechanism of MWCNT on the desorption and mineralization of phenanthrene could be as follows. The phenanthrene sorbed on the external surface and macropores of MWCNTs could be desorbed and biodegraded relative quickly; the fast and slow desorption and mineralization stages corresponded to this fraction of phenanthrene. The phenanthrene sorbed in mesopores and micropores would diffuse out of the pores very slowly. The very slow desorption and mineralization stages corresponded to this fraction of phenanthrene, leading to the residual phenanthrene having a positive correlation with mesopore and micropore volume. As shown in Figure 1 and Supplemental Data, Figure S1, the desorption and mineralization efficiencies of phenanthrene sorbed by MWCNT1 were the lowest among the four kinds of MWCNTs. This probably was due to the fact that MWCNT1 had the highest mesopore and micropore volume. After aging for 60 d, phenanthrene could enter the mesopores and micropores. In addition, it was suggested that formation of deformable closed spaces in CNT aggregates was responsible for irreversible hysteresis 13. As shown in Figure 3, aggregation of MWCNT1 occurred; the deformation of closed space in the aggregates would lead to the irreversible sorption and low bioavailability of phenanthrene sorbed on MWCNT1.
The present study shows that the mineralization efficiency of phenanthrene sorbed on MWCNTs was positively correlated with the desorption efficiency and that characteristics of MWCNTs exerted great influence on the desorption and mineralization processes. Although the cell morphology of bacteria changed during incubation, indicating that MWCNTs might have toxic effect on the bacteria, the biodegrading ability of bacteria might not decrease significantly under the influence of MWCNTs with a low SSA. Further research is needed to study the interaction mechanism among HOCs, MWCNTs, and bacteria. The present study suggests that the presence of CNTs in sediments and soils will decrease the bioavailability of HOCs to bacteria. However, because we used only MWCNTs in a water system in this research, which might be different from the real environment, further research is needed to study the effects of CNTs on the bioavailability of HOCs and their toxic effect on bacteria in soils and sediments. In addition, other environmental conditions such as the dissolved organic carbon are also important factors influencing the bioavailability of HOCs and might also influence the toxicity of CNTs. Therefore, further research should be carried out to study the effects of the toxicity and specific properties of CNTs on the bioavailability of HOCs under different environmental conditions.
Tenax TA aided abiotic desorption of phenanthrene from MWCNTs.
Figures S1–S4. (3 MB DOC).
This study was supported by the National Science Foundation of China (grants 51121003 and 51079003), the Program for New Century Excellent Talents in University (grant NCET-09-0233), and the Fundamental Research Funds for the Central Universities (grant 2009SD-8).