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Evaluation of dissipation gradients of polycyclic aromatic hydrocarbons in rice rhizosphere utilizing a sequential extraction procedure
Environmental Pollution 162 (2012) 413e421
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Evaluation of dissipation gradients of polycyclic aromatic hydrocarbons in rice rhizosphere utilizing a sequential extraction procedure Bin Ma, Jiaojiao Wang, Minmin Xu, Yan He, Haizhen Wang*, Laosheng Wu, Jianming Xu* College of Environmental and Natural Resource Sciences, Zhejiang Provincial Key Laboratory of Subtropical Soil and Plant Nutrition, Zhejiang University, Hangzhou 310029, China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 26 July 2011 Received in revised form 13 October 2011 Accepted 26 October 2011
The aim of this study was to evaluate the spatial dissipation gradient of PAHs, including phenanthrene, pyrene, and benzo[a]pyrene, with various bioavailability represented with sequential extraction. Dissipation rates of PAHs in the rhizosphere were greater than those in the bulk soil. The n-butanol extracted fraction showed a general trend of dissipation during phytoremediation. Moreover, the formation of bound PAH residues was inhibited in the rhizosphere. While concerning the PAH toxicity, the reduction rates of PAH toxicity were significantly greater than total soil PAH concentrations. Microbial biomass was the highest at four mm away from the root surface. However, the PAH dissipation rates were the highest at one mm and two mm away from the root surface in high and low PAH treatments, respectively. These results suggest that rhizoremediation with rice is a useful approach to reduce the toxicity of PAHs in soil. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Benzo[a]pyrene Paddy ecosystem Phenanthrene Pyrene Toxicity equivalency concentration
1. Introduction Paddy ecosystems have received close attention recently, not only because of their role in food production but also as an ecological server, performing functions such as phytoremediation of polycyclic aromatic hydrocarbons (PAHs) (Gao et al., 2010b; Li et al., 2009; Su and Zhu, 2008; Tao et al., 2006). Soil is a sink and source for PAHs since hydrophobic compounds readily partition into humic substances in soils (Bozlaker et al., 2008). Phytoremediation is considered as one of the best potential approaches for PAH-polluted soils (Mackova et al., 2006). Given that paddy fields are one of the most widely distributed agronomic ecosystems, the investigation of PAH remediation in the rice rhizosphere is of special significance in terms of its ecological functions. There are two redox gradients in paddy soils under flooded conditions. One redox gradient is at the interface between water and soil (Ludemann et al., 2000). Another is in the rhizosphere, caused by radial oxygen loss from rice roots (Ludemann et al., 2000; Van Bodegom et al., 2001). Microbial communities and electron acceptors, which are essential for PAH degradation, are vary spatially along the redox gradients (Grishanin et al., 1991; Ludemann et al., 2000). Phytoremediation is mainly based on rhizosphere effects, particularly the influence of root activities,
* Corresponding authors. E-mail addresses: [email protected] (H. Wang), [email protected] (J. Xu). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.10.034
such the release of root exudates and absorption of nutrients, on microbial activity and increased bioavailability of pollutants (Mackova et al., 2006). It is generally agreed that mineral nutrients and root exudates in the rhizosphere change along a linear gradient along the root surface (Baudoin et al., 2003; Dieffenbach and Matzner, 2000). Dissipation of pentachlorophenol in the ryegrass rhizosphere was reported as a nonlinear gradient because microbial degraders were stimulated at the middle distance from the root, but not close to the root surface (He et al., 2009). However, in previous studies, the dissipation of PAHs in the rhizosphere was measured in the entire rhizosphere and the heterogeneous nature of remediation in the rhizosphere was not recognized (Gao et al., 2010b; Li et al., 2008; Su and Zhu, 2008). Maintaining the bioavailability of PAHs in the soil is vital for bioremediation. Such bioavailability can be evaluated from bioaccumulation by organisms (Talley et al., 2002), chemical extraction (Johnson et al., 2002), and passive sampling devices (ter Laak et al., 2006). Given the small amount of available rhizosphere soil in this study, biological methods and passive sampling devices were impractical for evaluating bioavailability. Instead we used chemical extraction of PAHs, using solvents of different hydrophobicity, to represent different bioavailability of discreet PAH fractions (Macleod and Semple, 2003). The initial object of remediation is to reduce the environmental toxicity of pollutants and the risk to human health. Our previous study indicated that toxicity equivalency concentrations of PAHs were meaningful for assessing PAH remediation in the rhizosphere
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(Ma et al., 2010a). The toxicity equivalency factors for PAHs with higher hydrophobicity were greater (Nisbet and LaGoy, 1992). The effects of rhizosphere activities on more hydrophobic PAHs may be greater than on lower hydrophobic PAHs. For example, with increasing dissolved organic matter concentrations, the enhancement factors for diffusion of thirteen PAHs increased with their increasing hydrophobicity (Mayer et al., 2007). The proportion of PAHs dissipated in the rhizosphere also increased with increasing hydrophobicity (Binet et al., 2000a). As a result, it is important to determine toxicity equivalency factors of PAHs in order to fully understand the role of rhizosphere in remediation when multiple PAHs with different toxicity factors are considered in soil. In the present study, the dissipation of PAHs was evaluated from changes in bioavailability in the rhizosphere. Since PAHs stimulate root growth at low concentrations but have the opposite effect at high concentrations (Smreczak and Maliszewska-Kordybach, 2003), we spiked soils with both high and low PAH concentrations. We demonstrated the dissipation gradient of phenanthrene, pyrene, and benzo[a]pyrene in the rhizosphere region. The bioavailability of PAHs in soil was evaluated by a sequential extraction with various organic solvents. The toxicity of phenanthrene and pyrene was transformed into toxic equivalency concentration of benzo[a]pyrene based on reported toxic equivalency factors of PAHs. 2. Materials and methods 2.1. Chemicals The n-butanol, dichloromethane (DCM), and methanol were HPLC grade (Scharlau Chemie, S.A.). HCl, NaOH, and anhydrous Na2SO4 were p.a. grade (Sinopharm Chemical Reagent Co. Ltd, China). EPA 610 polycyclic aromatic hydrocarbons (16 PAHs), acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12 were obtained from Supelco. All water used in the experiment was Milli-Q water.
2.2. Soil treatments The soil used in this study was collected from the 0 to 20 cm top layer of a rice field at Jiaxing County in Zhejiang Province of China. The pH of the soil was 5.7 and the total organic carbon was 3.13%. The total nitrogen, phosphorus, and potassium concentrations were 1.72, 0.82, and 15.6 g kg1, respectively. The particle size distribution was 22.3% sand, 53.1% silt, and 27.6% clay. The soil was air-dried and passed through a 2 mm sieve. Soil was divided into two equal portions and one portion was spiked with a mixture of phenanthrene, pyrene, and benzo[a]pyrene dissolved in acetone to give spiked phenanthrene, pyrene, and benzo[a]pyrene concentrations of 200, 100, and 20 mg kg1 as high PAH treatment, and 20, 10, and 2 mg kg1 as low PAH treatment, respectively. Nutrients were added to the soils at concentrations of 120 mg N ((NH4)2SO4) kg1, 40 mg P (KH2PO4) kg1 and 50 mg K (KCl) kg1. Totally 4.5 kg spiked soils were packed into rhizoboxes modified from He et al. (2005) which divided the rhizosphere into five 1-mm-thick layers with nylon mesh with 40 mm pore size. Soils in the rhizoboxes were flooded and pre-incubated at 25 C for 4 weeks. After this pre-incubation time, rice seedlings were transplanted into the rhizoboxes and the concentrations of phenanthrene, pyrene, and benzo[a] pyrene in the soil were determined. 2.3. Preparation of seedlings China rice seeds, variety Zhenong-71 (Oryzae sativa L.), were disinfected in 30% ethanol (v:v) solution for 10 min, followed by thorough washing with Milli-Q water. The seeds were germinated in a growth cabinet at 30 C in the dark. One week after germination, rice seedlings were transferred to nutrient solutions to grow for two more weeks. After that, seedlings of uniform size were transplanted to the plant chambers of rhizobox. Control soil samples with PAHs but without seedlings were prepared at the same time. Three replicates were prepared for each treatment. Distilled water was supplied every day to ensure that soils were submerged in water throughout the experiment. The experiment was carried out in a greenhouse at 25 C with 14 h light/16 h darkness and rhizoboxes were randomly arranged. 2.4. Sequential extraction of PAHs in soil Rhizosphere soil samples and control soil samples without rice seedlings were taken at 15, 30 and 45 days after transplanting. Soil samples were lyophilized immediately after collection from the rhizoboxes.
Fig. 1. Sequential extraction profiles of phenanthrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting. The dashed lines indicate the sequential extraction profiles of PAHs at the day of planting. The CK treatment means control without plant.
B. Ma et al. / Environmental Pollution 162 (2012) 413e421 Methanol/water (6 ml, v/v ¼ 1:1) was added to 0.5 g of high PAH soils and 2 g of low PAH soils in 50 ml centrifuge tubes. The tubes were shaken for 24 h at 200 rpm at 25 C and then centrifuged at 2600 g for 20 min. Each supernatant was carefully removed, and the volume was determined. A 3 ml aliquot spiked with standards was added to a conditioned 6 ml C-18 solid-phase extraction cartridge. The solid-phase extraction cartridges were treated with 5 ml DCM, followed by 5 ml methanol. The PAHs were then eluted from the sorbents with 8 ml of DCM. The eluted solution was dried under a gentle stream of nitrogen, and then the solvent was exchanged with n-hexane. PAH analysis was performed using GCeMS (Agilent 6890N/5975B). Separation was achieved using a HP-5MS column (30 m 250 mm 0.25 mm). The temperature was kept at 60 C for 2 min, then 30 C min1 to 120 C and 5 C min1 to 270 C, where it was held for 15 min. The mass spectrometry operated in selective ion monitoring mode. The centrifuged soil pellet was loosened using a clean stainless steel spatula and 6 ml of n-butanol added. The tubes were shaken for 24 h on a tumble shaker and then centrifuged at 2600 g for 20 min. The supernatant was carefully removed, filtered, and PAHs determined with GC/MS. The soils remaining in the centrifuge tubes were extracted twice each for 0.5 h in an ultrasonic bath with dichloromethane, the extracts were combined, filtered, evaporated to 1 ml and PAHs determined with GC/MS. After exhaustive extraction, NaOH (2 M) was added to each vial and the vials were sealed under nitrogen gas. The vials were closed with Teflon-lined caps and heat-treated at 100 C for 2 h. The aqueous fraction was then obtained by centrifugation at 2600 rpm for 25 min, acidified with 6 M HCl to pH < 2, and subjected to liquideliquid extraction three times with 10 mL of dichloromethane. The combined organic phases were dehydrated with anhydrous Na2SO4, and concentrated by rotary evaporation. The concentrated solution was filtered and the PAHs determined with GC/MS. 2.5. Dissolved organic carbon concentrations and DNA concentrations Dissolved organic carbon was extracted from 1 g soil with milli-Q water at a water/soil ratio of 1:5. After centrifuging at 2600 g for 20 min, the supernatant was carefully removed and filtered through a 0.45-mm membrane. Dissolved organic carbon concentrations were determined with a Jena Mulit N/C 3100 (Analytic Jena). The extraction of DNA was done using the FastDNA SPIN Kit for soil sample combined with three replicates (Qbiogene).
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2.6. Quality control Procedural blanks were determined by going through the same extraction and cleanup procedures for each series of samples. Average surrogates recoveries (n ¼ 3) of acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12 were 76.6 3.5%, 85.6 3.2%, 91.4 2.7%, and 93.7 2.1%, respectively. The determine limits of phenanthrene, pyrene, and benzo[a]pyrene were 0.71 mg kg1, 0.68 mg kg1, and 0.66 mg kg1, respectively. 2.7. Data analysis Results were obtained from three replicates and compared at each time point. Solvent and soil controls were included. Statistical analyses were performed using the statistical package R, with p < 0.05 taken to indicate statistical significance.
3. Results 3.1. Available fractions of phenanthrene in rice rhizosphere soils The concentration of available phenanthrene, including water/ methanol extracted, n-butanol extracted, and dichloromethane extracted fractions, decreased to 92.9 mg kg1 in the high PAH treatment at planting (incubation time 0 days) (Fig. 1a). The Phenanthrene concentrations in the rhizosphere soils declined more rapidly than those in the control soil at 15 days after planting. However, the differences between rhizosphere soils and control samples were decreased at 30 days after planting. At 45 days after planting, the sum of available phenanthrene concentrations in all soils declined to about 25 percent of the initial concentrations. Methanol extracted phenanthrene fractions and dichloromethane extracted fractions were maintained at about 20 and 18 mg kg1, respectively, during the experiment. Phenanthrene concentrations extracted with n-butanol, however, decreased with time from 28.6
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Fig. 2. Sequential extraction profiles of pyrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30 and 45 days after planting.
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to 10 mg kg1. At 15 days after planting, n-butanol extracted phenanthrene concentrations in control samples were about twice those in rhizosphere soils. But 45 days after planting, phenanthrene concentrations in both control and rhizosphere samples decreased to about 10 mg kg1. The amounts of available phenanthrene increased with increasing distance from plant roots at 45 days after planting. In the low PAH soil, the sum concentrations of available phenanthrene decreased to 10.1 mg kg1 by time 0 (Fig. 1b). Phenanthrene concentrations in control soils were slightly higher than those in the rhizosphere of planted soils. However, this difference was undetectable at 30 days after planting. The amount of dissipated phenanthrene in the low PAH soil could be neglected, but the profile of fractions changed during the experiment. The proportions of water/methanol extracted and n-butanol extracted fractions decreased, whereas that of the dichloromethane extracted fraction increased. At about 15 days after planting, the n-butanol extracted phenanthrene concentration in the control samples was about twice that in the rhizosphere. Available phenanthrene concentrations in the two and three mm layers tended to be lower than in the other layers at 45 days after planting. 3.2. Available fractions of pyrene in rice rhizosphere soils The initial concentrations of pyrene in the high and low PAH soils were 100 and 10 mg kg1, respectively. Seventy percent of spiked pyrene was detected in the high pollutant treatment at zero days (Fig. 2a). The sum of pyrene concentrations in the control samples was much greater than those in rhizosphere soils at all three samplings. The concentration and proportion of pyrene in the dichloromethane extracted fraction increased in the control samples but the n-butanol extracted fraction of pyrene decreased during the experiment. The water/methanol extracted pyrene
fraction was maintained at about 10 mg kg1. After transplanting rice into the soils (time 0), the sum of available pyrene concentrations in the rhizosphere soils decreased to around 35 percent of the initial pyrene concentration at 15 days after planting. However, 45 days after planting, available pyrene in the rhizosphere was reduced to about 30 percent. Furthermore, the n-butanol extracted fractions in the rhizosphere soils decreased with time, but the pyrene concentrations in the water/methanol extracted and dichloromethane extracted fractions were stable at about 10 and 6 mg kg1, respectively. The sum of available pyrene concentrations in the three and four mm layers was less than those in other layers at 15 and 30 days after planting. At 45 days after planting, however, the available pyrene concentrations decreased nearer to the root. At zero days, the total concentration of available pyrene was 65 percent of spiked pyrene in the low PAH soil (Fig. 2b). During 45 days after planting, available pyrene decreased from 6.5 to 5.9 mg kg1. Pyrene concentrations in rhizosphere soils were significantly lower than those in control soils at 15 days after planting. The amounts of pyrene dissipated in rhizosphere soils were less than 1 mg kg1 during the following 30 days. The concentrations of water/methanol extracted and n-butanol extracted pyrene fractions decreased during the 45 days after planting. The sum concentrations of available pyrene in the two mm layer tended to be less than in other layers in the rhizosphere at the end of the experiment. 3.3. Available fractions of benzo[a]pyrene in rice rhizosphere soils The sum of available benzo[a]pyrene concentration at time 0 in the high PAH soils decreased to 8.4 mg kg1 (Fig. 3a). At 15 after planting, the sum of available benzo[a]pyrene was 7.9 mg kg1 in the control soil, and about 6 mg kg1 in rhizosphere soils. At 45 days after planting, available benzo[a]pyrene in the control soils was close
Fig. 3. Sequential extraction profiles of benzo[a]pyrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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Fig. 4. Bound residual phenanthrene (a, d), pyrene (b, e), benzo[a]pyrene (c, f), in rice rhizosphere of high PAH soils (a, b, c) and low PAH soils (d, e, f) at 15, 30, and 45 days after planting.
to that in the fifth mm layer of the rhizosphere soils. The proportions of water/methanol extracted fractions of benzo[a]pyrene were considerably less than those of phenanthrene and pyrene. Dichloromethane extracted benzo[a]pyrene concentrations decreased during 45 days after planting, but the differences between control and different layers of rhizosphere soils was small. At 45 days
after planting, available benzo[a]pyrene concentration in the first mm layer of rhizosphere soil was significantly lower than in other layers because of the decrease in the n-butanol extracted fraction. Similarly, the higher available benzo[a]pyrene concentrations in the control soils at 15 and 30 days after planting was also due to the higher concentrations in the n-butanol extracted fraction.
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Fig. 5. Sequential extraction profiles of total PAH concentrations in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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In low PAH soils, the sum concentration of extractable benzo[a] pyrene at time 0 was 1.1 mg kg1 (Fig. 3b). Benzo[a]pyrene concentrations in soils without plants then decreased, but were significantly greater than those in rhizosphere soils. In the rhizosphere soil layers at increasing distance away from the root surface, benzo[a]pyrene concentrations were not significantly different. Benzo[a]pyrene concentrations in the rhizosphere soils decreased to approximately 0.4 mg kg1 at 15 days after planting but no further decreases were observed. The methanol extracted fraction in the rhizosphere soils decreased, while the n-butanol extracted fraction in the rhizosphere soils increased with increasing incubation time after planting.
(Fig. 5). At 45 days after planting, total PAH concentrations increased with distance from the root surface in high PAH rhizosphere soils, and were the lowest in the 2 mm layer of low PAH soils. Phenanthrene and pyrene concentrations were converting into toxicity equivalency concentrations of benzo[a]pyrene based on toxicity equivalency factors. The differences in toxicity equivalency concentrations between soils with and without plants were greater than those of total PAH concentrations (Fig. 6). Compared with total PAH concentrations, the toxicity of methanol extracted fraction reduced and n-butanol extracted fraction increased. 3.6. Extracted DNA concentrations and dissolved organic carbon in rice rhizosphere soils
3.4. Bound residual PAHs in rice rhizosphere soils Although the proportions of bound residual PAHs in soils were less than one percent of the total PAHs in the present study, there was considerable variation in their concentrations in the rhizosphere. The bound residual phenanthrene and pyrene concentrations in soils near the root surface were markedly lower in high PAHs soils (Fig. 4). In low PAH soils the bound residual concentrations of phenanthrene and pyrene were near zero during 30 days after planting, but increased to similar levels for the soils without plants at 45 days after planting. The concentrations of bound residual phenanthrene and pyrene decreased with increasing distance away from the root surface. The difference of bound residual benzo[a]pyrene concentrations among soils were not significant.
At 15 days after planting, extracted DNA concentrations in high PAH soils were all lower than 10 ng g1, but increased to similar concentrations as in low PAH soils at 30 days after planting (Fig. 7). In low PAH soils, the extracted DNA concentrations increased with time. The extracted DNA concentrations in the rhizosphere soils increased up to four mm away from the root surface. The concentrations of DOC in high PAHs soils were significantly greater than in corresponding low PAH soils (Fig. 8). There was a general tendency for DOC concentrations to decrease with plant growth. The DOC concentrations in soil at three mm were lower than in soils nearer or further away from the root. 4. Discussion 4.1. Sequential extraction fractions of PAHs in rhizosphere gradient
3.5. Toxicity equivalency concentrations of PAHs in rice rhizosphere soils Total PAH concentrations in soils without plants were greater than in the rhizosphere between day 0 and day 45 after planting
Phytoremediation, particularly rhizoremediation, of pollutants in soils is based upon removal, immobilization, and transformation of pollutants to less harmful metabolites. For rhizoremediation of PAHs with rice, the contribution of removal pathways by which
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Fig. 6. Sequential extraction profiles of equivalency toxicity concentrations in the rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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Extracted DNA concentrations(ng g−1)
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Fig. 7. Extracted DNA concentrations in the rice rhizosphere. The dash line indicates low PAH soil and solid line indicated the high PAH soil.
TOC in low PAHs soil(mg g−1) TOC in high PAHs soil(mg g−1)
PAHs are translocated to shoots is very low (Binet et al., 2000b; Gao et al., 2006; Su and Zhu, 2008). Some studies reported high concentrations of PAHs in stem tissues of lettuce (Khan et al., 2008), cabbage, ryegrass, and amaranth (Zhu and Gao, 2004). This might be due to contamination from PAHs in the ambient air (Li et al.,
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2009). The immobilization of PAHs in soils by rhizoremediation can be achieved by decreasing their bioavailability. The widely used analytical methods for PAHs, however, entail vigorous extraction of soils with organic solvents. Such methods do not reveal bioavailability (Alexander, 2000). The scheme of sequential extraction with different solvents from soils in our study was designed to operationally remove PAHs with different availabilities in soils (Macleod and Semple, 2003). Methanol removes labile nonpolar parent compounds and polar metabolites, n-butanol extracts weakly associated PAHs, and the DCM-extracted fraction represents chemically and physically associated PAHs. The decrease in the proportions of fractions extracted with methanol with increasing hydrophobicity of PAHs reveals that the methanol extracted fraction represented the dissolved PAH fraction, determined by partition characteristics of PAHs between soil organic matters and water (Koc). The Koc values of PAHs can be estimated from their related physical parameters, such as octanolewater partition coefficients (Kow) or water solubility (Sw) (Chiou et al., 1998). The Kow of benzo [a]pyrene was ten times and 36.3 times that of pyrene and phenanthrene Kow, respectively. The solubility of benzo[a]pyrene was, however, only approximately 0.01 times and 0.001 times that of pyrene and phenanthrene Kow, respectively. The difference among different proportions of methanol extracted fractions of the PAHs corresponds better to Kow than Sw. There were decreases in methanol extracted fractions with time for pyrene and benzo[a]pyrene. This tendency indicates reduced availability of PAHs with increasing soil contact time during the experiment. The n-butanol-extracted PAHs fraction from soil was very similar to concentrations accumulated by earthworms (Liste and Alexander, 2002) and therefore indicates the bioavailability of PAHs in soils. In the present study, the n-butanol fraction mainly contributed to the variation of PAHs in the rhizosphere. The proportions of methanol and DCM-extracted fractions were relatively constant, and the differences among different soils were mainly attributed to the changes in the n-butanol-extracted fraction. The ratio between methanol- and n-butanol-extracted fractions increased in the soils where more PAHs were dissipated. A lower ratio indicates higher availability because more labile PAHs were present. This result indicated that one of the mechanisms for rhizoremediation in the rice rhizosphere increased availability of PAHs, rather than immobilization. This result is approved by the previously reported result that desorbing fraction accounted for the dominant contribution to dissipated PAHs, but immobilized bound residues contributed little to the dissipation of PAHs in soils (Gao et al., 2010b). 4.2. The response of microbial activities in rhizosphere gradient
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Distance (mm) Fig. 8. Dissolved organic carbon concentrations in the rice rhizosphere of high PAH soils (a) and low PAH soils (b).
The destruction of PAHs in soils can be achieved through intraand extracellular enzyme activities of microbes (Johnsen et al., 2005). It is widely reported that microbial community and activity are affected by the rhizosphere. Our previous study based on metaanalysis indicated that PAH-degrading bacteria are significantly enhanced in the rhizosphere (Ma et al., 2010b). The extracted DNA concentrations were greatest at 4 mm away from the root surface. Bacterial biomass could be evaluated by quantifying the level of extracted DNA concentration (Aoshima et al., 2006). Therefore, the biggest stimulation of microbial activity by the rhizosphere occurred at 4 mm away from the rice root surface. The greatest PAH dissipation in the rhizosphere was, however, determined at one and two mm away from the root surface in both low- and high PAH soils, respectively. The separation between maximum microbial biomass concentration (Fig. 5) and maximum dissipation rates (Fig. 7) indicates that the growth of PAH degraders was not consistent with other microbes in the rhizosphere. The suppression of microbial activity near the root surface might be due to the depletion of
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mineral nutrients in rhizosphere. Such depletion of nutrient elements favors the growth of PAH-degrading bacteria which are well-adopted to such conditions (Johnsen et al., 2005). The bioavailability of PAHs is influenced by the rhizosphere effect as well. Root exudates have been widely reported to enhance PAH degradation through modification of microbial community structure, the promotion of microbial activity (Gao et al., 2011; Sun et al., 2010; Yi and Crowley, 2007) and the induction of PAH-cometabolic degrading enzymes (Muratova et al., 2009). Moreover, root exudates could increase desorption of PAHs from soils, thus increasing their bioavailability (Gao et al., 2010a). The difference in dissipation gradients in the rhizosphere of low- and high-PAH soils might indicate that the toxicity of PAHs to rice root activity differs. Exposure to low doses of PAHs can stimulate plant growth, but high doses of PAHs inhibit root growth and root hair initiation (Zhang et al., 2010). Consequently, the stress of high PAH concentrations in soil to rice root might be stronger than in low PAH concentration soils. 4.3. The response of dissolved organic carbons in rhizosphere gradient The high concentration of DOC in high PAH soils might be due to the metabolic products of degraded PAHs. Although the solubility of PAHs is low, the solubility of their derivatives is much higher (Yi and Crowley, 2007). The low DOC concentrations in the 3 mm soil layer indicated that oxygen released from the root could promote DOC utilization. However, the higher TOC concentrations in the one to 2 mm soil layers may also arise from degraded PAH derivatives. The low proportion of bound residual PAHs in our study indicated that only small amounts of PAHs were sequestrated into humic substance during the 45 days. The half-lives for the apparent loss of phenanthrene, pyrene, and benzo[a]pyrene in spiked soils by sequestration were 1789, 254 and 254 days, respectively (Northcott and Jones, 2001). Furthermore, the availability of PAHs was enhanced in rhizosphere. The effects of the rhizosphere on the toxicity of PAHs, rather than the total concentrations of PAHs dissipated. The toxicity of benzo[a]pyrene was one thousand times greater than phenanthrene and pyrene. The more effective remediation of PAH toxicity in the rhizosphere indicated that highly hydrophobic PAH compounds, e.g. benzo[a]pyrene, were more effectively degraded there than phenanthrene and pyrene. Soils collected from industrially polluted sites also showed that the rhizosphere contributed more to the bioremediation of highly PAH-polluted soils due to their high hydrophobicity (Ma et al., 2010a,b). 5. Conclusion The rhizoremediation effect of rice on PAH dissipation was efficient and the role of the rhizoremediation approach should be recognized when the toxicity risk of PAHs is evaluated. Concentrations of PAHs influence PAH dissipation in the rice rhizosphere. Extraction with mild extracting solvents such as n-butanol provided a useful model of the rhizosphere effect on PAH remediation. Furthermore, the greatest dissipation rate did not coincide with the greatest microbial biomass in rhizosphere. Acknowledgments This work was jointly supported by the National Natural Science Foundation of China (41090284, 20977077, 40971136), the Project Supported by Zhejiang Provincial Natural Science Foundation of China (R5110079), and the Key Scientific and Technological Program of Zhejiang Province (2008C13024-3).
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Phytoremediation and Rhizoremediation: Theoretical Background. Springer-Verlag GmbH. Macleod, C.J.A., Semple, K.T., 2003. Sequential extraction of low concentrations of pyrene and formation of non-extractable residues in sterile and non-sterile soils. Soil Biology and Biochemistry 35, 1443e1450. Mayer, P., Fernqvist, M.M., Christensen, P.S., Karlson, U., Trapp, S., 2007. Enhanced diffusion of polycyclic aromatic hydrocarbons in artificial and natural aqueous solutions. Environmental Science and Technology 41, 6148e6155. Muratova, A., Pozdnyakova, N., Golubev, S., Wittenmayer, L., Makarov, O., Merbach, W., Turkovskaya, O., 2009. Oxidoreductase activity of sorghum root exudates in a phenanthrene-contaminated environment. Chemosphere 74, 1031e1036. Nisbet, I.C.T., LaGoy, P.K., 1992. Toxic equivalency factors (TEFs) for polycyclic aromatic hydrocarbons (PAHs). Regulatory Toxicology and Pharmacology 16, 290e300.
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Evaluation of dissipation gradients of polycyclic aromatic hydrocarbons in rice rhizosphere utilizing a sequential extraction procedure Bin Ma, Jiaojiao Wang, Minmin Xu, Yan He, Haizhen Wang*, Laosheng Wu, Jianming Xu* College of Environmental and Natural Resource Sciences, Zhejiang Provincial Key Laboratory of Subtropical Soil and Plant Nutrition, Zhejiang University, Hangzhou 310029, China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 26 July 2011 Received in revised form 13 October 2011 Accepted 26 October 2011
The aim of this study was to evaluate the spatial dissipation gradient of PAHs, including phenanthrene, pyrene, and benzo[a]pyrene, with various bioavailability represented with sequential extraction. Dissipation rates of PAHs in the rhizosphere were greater than those in the bulk soil. The n-butanol extracted fraction showed a general trend of dissipation during phytoremediation. Moreover, the formation of bound PAH residues was inhibited in the rhizosphere. While concerning the PAH toxicity, the reduction rates of PAH toxicity were significantly greater than total soil PAH concentrations. Microbial biomass was the highest at four mm away from the root surface. However, the PAH dissipation rates were the highest at one mm and two mm away from the root surface in high and low PAH treatments, respectively. These results suggest that rhizoremediation with rice is a useful approach to reduce the toxicity of PAHs in soil. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Benzo[a]pyrene Paddy ecosystem Phenanthrene Pyrene Toxicity equivalency concentration
1. Introduction Paddy ecosystems have received close attention recently, not only because of their role in food production but also as an ecological server, performing functions such as phytoremediation of polycyclic aromatic hydrocarbons (PAHs) (Gao et al., 2010b; Li et al., 2009; Su and Zhu, 2008; Tao et al., 2006). Soil is a sink and source for PAHs since hydrophobic compounds readily partition into humic substances in soils (Bozlaker et al., 2008). Phytoremediation is considered as one of the best potential approaches for PAH-polluted soils (Mackova et al., 2006). Given that paddy fields are one of the most widely distributed agronomic ecosystems, the investigation of PAH remediation in the rice rhizosphere is of special significance in terms of its ecological functions. There are two redox gradients in paddy soils under flooded conditions. One redox gradient is at the interface between water and soil (Ludemann et al., 2000). Another is in the rhizosphere, caused by radial oxygen loss from rice roots (Ludemann et al., 2000; Van Bodegom et al., 2001). Microbial communities and electron acceptors, which are essential for PAH degradation, are vary spatially along the redox gradients (Grishanin et al., 1991; Ludemann et al., 2000). Phytoremediation is mainly based on rhizosphere effects, particularly the influence of root activities,
* Corresponding authors. E-mail addresses: [email protected] (H. Wang), [email protected] (J. Xu). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.10.034
such the release of root exudates and absorption of nutrients, on microbial activity and increased bioavailability of pollutants (Mackova et al., 2006). It is generally agreed that mineral nutrients and root exudates in the rhizosphere change along a linear gradient along the root surface (Baudoin et al., 2003; Dieffenbach and Matzner, 2000). Dissipation of pentachlorophenol in the ryegrass rhizosphere was reported as a nonlinear gradient because microbial degraders were stimulated at the middle distance from the root, but not close to the root surface (He et al., 2009). However, in previous studies, the dissipation of PAHs in the rhizosphere was measured in the entire rhizosphere and the heterogeneous nature of remediation in the rhizosphere was not recognized (Gao et al., 2010b; Li et al., 2008; Su and Zhu, 2008). Maintaining the bioavailability of PAHs in the soil is vital for bioremediation. Such bioavailability can be evaluated from bioaccumulation by organisms (Talley et al., 2002), chemical extraction (Johnson et al., 2002), and passive sampling devices (ter Laak et al., 2006). Given the small amount of available rhizosphere soil in this study, biological methods and passive sampling devices were impractical for evaluating bioavailability. Instead we used chemical extraction of PAHs, using solvents of different hydrophobicity, to represent different bioavailability of discreet PAH fractions (Macleod and Semple, 2003). The initial object of remediation is to reduce the environmental toxicity of pollutants and the risk to human health. Our previous study indicated that toxicity equivalency concentrations of PAHs were meaningful for assessing PAH remediation in the rhizosphere
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(Ma et al., 2010a). The toxicity equivalency factors for PAHs with higher hydrophobicity were greater (Nisbet and LaGoy, 1992). The effects of rhizosphere activities on more hydrophobic PAHs may be greater than on lower hydrophobic PAHs. For example, with increasing dissolved organic matter concentrations, the enhancement factors for diffusion of thirteen PAHs increased with their increasing hydrophobicity (Mayer et al., 2007). The proportion of PAHs dissipated in the rhizosphere also increased with increasing hydrophobicity (Binet et al., 2000a). As a result, it is important to determine toxicity equivalency factors of PAHs in order to fully understand the role of rhizosphere in remediation when multiple PAHs with different toxicity factors are considered in soil. In the present study, the dissipation of PAHs was evaluated from changes in bioavailability in the rhizosphere. Since PAHs stimulate root growth at low concentrations but have the opposite effect at high concentrations (Smreczak and Maliszewska-Kordybach, 2003), we spiked soils with both high and low PAH concentrations. We demonstrated the dissipation gradient of phenanthrene, pyrene, and benzo[a]pyrene in the rhizosphere region. The bioavailability of PAHs in soil was evaluated by a sequential extraction with various organic solvents. The toxicity of phenanthrene and pyrene was transformed into toxic equivalency concentration of benzo[a]pyrene based on reported toxic equivalency factors of PAHs. 2. Materials and methods 2.1. Chemicals The n-butanol, dichloromethane (DCM), and methanol were HPLC grade (Scharlau Chemie, S.A.). HCl, NaOH, and anhydrous Na2SO4 were p.a. grade (Sinopharm Chemical Reagent Co. Ltd, China). EPA 610 polycyclic aromatic hydrocarbons (16 PAHs), acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12 were obtained from Supelco. All water used in the experiment was Milli-Q water.
2.2. Soil treatments The soil used in this study was collected from the 0 to 20 cm top layer of a rice field at Jiaxing County in Zhejiang Province of China. The pH of the soil was 5.7 and the total organic carbon was 3.13%. The total nitrogen, phosphorus, and potassium concentrations were 1.72, 0.82, and 15.6 g kg1, respectively. The particle size distribution was 22.3% sand, 53.1% silt, and 27.6% clay. The soil was air-dried and passed through a 2 mm sieve. Soil was divided into two equal portions and one portion was spiked with a mixture of phenanthrene, pyrene, and benzo[a]pyrene dissolved in acetone to give spiked phenanthrene, pyrene, and benzo[a]pyrene concentrations of 200, 100, and 20 mg kg1 as high PAH treatment, and 20, 10, and 2 mg kg1 as low PAH treatment, respectively. Nutrients were added to the soils at concentrations of 120 mg N ((NH4)2SO4) kg1, 40 mg P (KH2PO4) kg1 and 50 mg K (KCl) kg1. Totally 4.5 kg spiked soils were packed into rhizoboxes modified from He et al. (2005) which divided the rhizosphere into five 1-mm-thick layers with nylon mesh with 40 mm pore size. Soils in the rhizoboxes were flooded and pre-incubated at 25 C for 4 weeks. After this pre-incubation time, rice seedlings were transplanted into the rhizoboxes and the concentrations of phenanthrene, pyrene, and benzo[a] pyrene in the soil were determined. 2.3. Preparation of seedlings China rice seeds, variety Zhenong-71 (Oryzae sativa L.), were disinfected in 30% ethanol (v:v) solution for 10 min, followed by thorough washing with Milli-Q water. The seeds were germinated in a growth cabinet at 30 C in the dark. One week after germination, rice seedlings were transferred to nutrient solutions to grow for two more weeks. After that, seedlings of uniform size were transplanted to the plant chambers of rhizobox. Control soil samples with PAHs but without seedlings were prepared at the same time. Three replicates were prepared for each treatment. Distilled water was supplied every day to ensure that soils were submerged in water throughout the experiment. The experiment was carried out in a greenhouse at 25 C with 14 h light/16 h darkness and rhizoboxes were randomly arranged. 2.4. Sequential extraction of PAHs in soil Rhizosphere soil samples and control soil samples without rice seedlings were taken at 15, 30 and 45 days after transplanting. Soil samples were lyophilized immediately after collection from the rhizoboxes.
Fig. 1. Sequential extraction profiles of phenanthrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting. The dashed lines indicate the sequential extraction profiles of PAHs at the day of planting. The CK treatment means control without plant.
B. Ma et al. / Environmental Pollution 162 (2012) 413e421 Methanol/water (6 ml, v/v ¼ 1:1) was added to 0.5 g of high PAH soils and 2 g of low PAH soils in 50 ml centrifuge tubes. The tubes were shaken for 24 h at 200 rpm at 25 C and then centrifuged at 2600 g for 20 min. Each supernatant was carefully removed, and the volume was determined. A 3 ml aliquot spiked with standards was added to a conditioned 6 ml C-18 solid-phase extraction cartridge. The solid-phase extraction cartridges were treated with 5 ml DCM, followed by 5 ml methanol. The PAHs were then eluted from the sorbents with 8 ml of DCM. The eluted solution was dried under a gentle stream of nitrogen, and then the solvent was exchanged with n-hexane. PAH analysis was performed using GCeMS (Agilent 6890N/5975B). Separation was achieved using a HP-5MS column (30 m 250 mm 0.25 mm). The temperature was kept at 60 C for 2 min, then 30 C min1 to 120 C and 5 C min1 to 270 C, where it was held for 15 min. The mass spectrometry operated in selective ion monitoring mode. The centrifuged soil pellet was loosened using a clean stainless steel spatula and 6 ml of n-butanol added. The tubes were shaken for 24 h on a tumble shaker and then centrifuged at 2600 g for 20 min. The supernatant was carefully removed, filtered, and PAHs determined with GC/MS. The soils remaining in the centrifuge tubes were extracted twice each for 0.5 h in an ultrasonic bath with dichloromethane, the extracts were combined, filtered, evaporated to 1 ml and PAHs determined with GC/MS. After exhaustive extraction, NaOH (2 M) was added to each vial and the vials were sealed under nitrogen gas. The vials were closed with Teflon-lined caps and heat-treated at 100 C for 2 h. The aqueous fraction was then obtained by centrifugation at 2600 rpm for 25 min, acidified with 6 M HCl to pH < 2, and subjected to liquideliquid extraction three times with 10 mL of dichloromethane. The combined organic phases were dehydrated with anhydrous Na2SO4, and concentrated by rotary evaporation. The concentrated solution was filtered and the PAHs determined with GC/MS. 2.5. Dissolved organic carbon concentrations and DNA concentrations Dissolved organic carbon was extracted from 1 g soil with milli-Q water at a water/soil ratio of 1:5. After centrifuging at 2600 g for 20 min, the supernatant was carefully removed and filtered through a 0.45-mm membrane. Dissolved organic carbon concentrations were determined with a Jena Mulit N/C 3100 (Analytic Jena). The extraction of DNA was done using the FastDNA SPIN Kit for soil sample combined with three replicates (Qbiogene).
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2.6. Quality control Procedural blanks were determined by going through the same extraction and cleanup procedures for each series of samples. Average surrogates recoveries (n ¼ 3) of acenaphthene-d10, phenanthrene-d10, chrysene-d12, and perylene-d12 were 76.6 3.5%, 85.6 3.2%, 91.4 2.7%, and 93.7 2.1%, respectively. The determine limits of phenanthrene, pyrene, and benzo[a]pyrene were 0.71 mg kg1, 0.68 mg kg1, and 0.66 mg kg1, respectively. 2.7. Data analysis Results were obtained from three replicates and compared at each time point. Solvent and soil controls were included. Statistical analyses were performed using the statistical package R, with p < 0.05 taken to indicate statistical significance.
3. Results 3.1. Available fractions of phenanthrene in rice rhizosphere soils The concentration of available phenanthrene, including water/ methanol extracted, n-butanol extracted, and dichloromethane extracted fractions, decreased to 92.9 mg kg1 in the high PAH treatment at planting (incubation time 0 days) (Fig. 1a). The Phenanthrene concentrations in the rhizosphere soils declined more rapidly than those in the control soil at 15 days after planting. However, the differences between rhizosphere soils and control samples were decreased at 30 days after planting. At 45 days after planting, the sum of available phenanthrene concentrations in all soils declined to about 25 percent of the initial concentrations. Methanol extracted phenanthrene fractions and dichloromethane extracted fractions were maintained at about 20 and 18 mg kg1, respectively, during the experiment. Phenanthrene concentrations extracted with n-butanol, however, decreased with time from 28.6
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Fig. 2. Sequential extraction profiles of pyrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30 and 45 days after planting.
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to 10 mg kg1. At 15 days after planting, n-butanol extracted phenanthrene concentrations in control samples were about twice those in rhizosphere soils. But 45 days after planting, phenanthrene concentrations in both control and rhizosphere samples decreased to about 10 mg kg1. The amounts of available phenanthrene increased with increasing distance from plant roots at 45 days after planting. In the low PAH soil, the sum concentrations of available phenanthrene decreased to 10.1 mg kg1 by time 0 (Fig. 1b). Phenanthrene concentrations in control soils were slightly higher than those in the rhizosphere of planted soils. However, this difference was undetectable at 30 days after planting. The amount of dissipated phenanthrene in the low PAH soil could be neglected, but the profile of fractions changed during the experiment. The proportions of water/methanol extracted and n-butanol extracted fractions decreased, whereas that of the dichloromethane extracted fraction increased. At about 15 days after planting, the n-butanol extracted phenanthrene concentration in the control samples was about twice that in the rhizosphere. Available phenanthrene concentrations in the two and three mm layers tended to be lower than in the other layers at 45 days after planting. 3.2. Available fractions of pyrene in rice rhizosphere soils The initial concentrations of pyrene in the high and low PAH soils were 100 and 10 mg kg1, respectively. Seventy percent of spiked pyrene was detected in the high pollutant treatment at zero days (Fig. 2a). The sum of pyrene concentrations in the control samples was much greater than those in rhizosphere soils at all three samplings. The concentration and proportion of pyrene in the dichloromethane extracted fraction increased in the control samples but the n-butanol extracted fraction of pyrene decreased during the experiment. The water/methanol extracted pyrene
fraction was maintained at about 10 mg kg1. After transplanting rice into the soils (time 0), the sum of available pyrene concentrations in the rhizosphere soils decreased to around 35 percent of the initial pyrene concentration at 15 days after planting. However, 45 days after planting, available pyrene in the rhizosphere was reduced to about 30 percent. Furthermore, the n-butanol extracted fractions in the rhizosphere soils decreased with time, but the pyrene concentrations in the water/methanol extracted and dichloromethane extracted fractions were stable at about 10 and 6 mg kg1, respectively. The sum of available pyrene concentrations in the three and four mm layers was less than those in other layers at 15 and 30 days after planting. At 45 days after planting, however, the available pyrene concentrations decreased nearer to the root. At zero days, the total concentration of available pyrene was 65 percent of spiked pyrene in the low PAH soil (Fig. 2b). During 45 days after planting, available pyrene decreased from 6.5 to 5.9 mg kg1. Pyrene concentrations in rhizosphere soils were significantly lower than those in control soils at 15 days after planting. The amounts of pyrene dissipated in rhizosphere soils were less than 1 mg kg1 during the following 30 days. The concentrations of water/methanol extracted and n-butanol extracted pyrene fractions decreased during the 45 days after planting. The sum concentrations of available pyrene in the two mm layer tended to be less than in other layers in the rhizosphere at the end of the experiment. 3.3. Available fractions of benzo[a]pyrene in rice rhizosphere soils The sum of available benzo[a]pyrene concentration at time 0 in the high PAH soils decreased to 8.4 mg kg1 (Fig. 3a). At 15 after planting, the sum of available benzo[a]pyrene was 7.9 mg kg1 in the control soil, and about 6 mg kg1 in rhizosphere soils. At 45 days after planting, available benzo[a]pyrene in the control soils was close
Fig. 3. Sequential extraction profiles of benzo[a]pyrene in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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Fig. 4. Bound residual phenanthrene (a, d), pyrene (b, e), benzo[a]pyrene (c, f), in rice rhizosphere of high PAH soils (a, b, c) and low PAH soils (d, e, f) at 15, 30, and 45 days after planting.
to that in the fifth mm layer of the rhizosphere soils. The proportions of water/methanol extracted fractions of benzo[a]pyrene were considerably less than those of phenanthrene and pyrene. Dichloromethane extracted benzo[a]pyrene concentrations decreased during 45 days after planting, but the differences between control and different layers of rhizosphere soils was small. At 45 days
after planting, available benzo[a]pyrene concentration in the first mm layer of rhizosphere soil was significantly lower than in other layers because of the decrease in the n-butanol extracted fraction. Similarly, the higher available benzo[a]pyrene concentrations in the control soils at 15 and 30 days after planting was also due to the higher concentrations in the n-butanol extracted fraction.
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Fig. 5. Sequential extraction profiles of total PAH concentrations in rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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In low PAH soils, the sum concentration of extractable benzo[a] pyrene at time 0 was 1.1 mg kg1 (Fig. 3b). Benzo[a]pyrene concentrations in soils without plants then decreased, but were significantly greater than those in rhizosphere soils. In the rhizosphere soil layers at increasing distance away from the root surface, benzo[a]pyrene concentrations were not significantly different. Benzo[a]pyrene concentrations in the rhizosphere soils decreased to approximately 0.4 mg kg1 at 15 days after planting but no further decreases were observed. The methanol extracted fraction in the rhizosphere soils decreased, while the n-butanol extracted fraction in the rhizosphere soils increased with increasing incubation time after planting.
(Fig. 5). At 45 days after planting, total PAH concentrations increased with distance from the root surface in high PAH rhizosphere soils, and were the lowest in the 2 mm layer of low PAH soils. Phenanthrene and pyrene concentrations were converting into toxicity equivalency concentrations of benzo[a]pyrene based on toxicity equivalency factors. The differences in toxicity equivalency concentrations between soils with and without plants were greater than those of total PAH concentrations (Fig. 6). Compared with total PAH concentrations, the toxicity of methanol extracted fraction reduced and n-butanol extracted fraction increased. 3.6. Extracted DNA concentrations and dissolved organic carbon in rice rhizosphere soils
3.4. Bound residual PAHs in rice rhizosphere soils Although the proportions of bound residual PAHs in soils were less than one percent of the total PAHs in the present study, there was considerable variation in their concentrations in the rhizosphere. The bound residual phenanthrene and pyrene concentrations in soils near the root surface were markedly lower in high PAHs soils (Fig. 4). In low PAH soils the bound residual concentrations of phenanthrene and pyrene were near zero during 30 days after planting, but increased to similar levels for the soils without plants at 45 days after planting. The concentrations of bound residual phenanthrene and pyrene decreased with increasing distance away from the root surface. The difference of bound residual benzo[a]pyrene concentrations among soils were not significant.
At 15 days after planting, extracted DNA concentrations in high PAH soils were all lower than 10 ng g1, but increased to similar concentrations as in low PAH soils at 30 days after planting (Fig. 7). In low PAH soils, the extracted DNA concentrations increased with time. The extracted DNA concentrations in the rhizosphere soils increased up to four mm away from the root surface. The concentrations of DOC in high PAHs soils were significantly greater than in corresponding low PAH soils (Fig. 8). There was a general tendency for DOC concentrations to decrease with plant growth. The DOC concentrations in soil at three mm were lower than in soils nearer or further away from the root. 4. Discussion 4.1. Sequential extraction fractions of PAHs in rhizosphere gradient
3.5. Toxicity equivalency concentrations of PAHs in rice rhizosphere soils Total PAH concentrations in soils without plants were greater than in the rhizosphere between day 0 and day 45 after planting
Phytoremediation, particularly rhizoremediation, of pollutants in soils is based upon removal, immobilization, and transformation of pollutants to less harmful metabolites. For rhizoremediation of PAHs with rice, the contribution of removal pathways by which
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Fig. 6. Sequential extraction profiles of equivalency toxicity concentrations in the rice rhizosphere of high PAH soils (a) and low PAH soils (b) at 15, 30, and 45 days after planting.
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PAHs are translocated to shoots is very low (Binet et al., 2000b; Gao et al., 2006; Su and Zhu, 2008). Some studies reported high concentrations of PAHs in stem tissues of lettuce (Khan et al., 2008), cabbage, ryegrass, and amaranth (Zhu and Gao, 2004). This might be due to contamination from PAHs in the ambient air (Li et al.,
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2009). The immobilization of PAHs in soils by rhizoremediation can be achieved by decreasing their bioavailability. The widely used analytical methods for PAHs, however, entail vigorous extraction of soils with organic solvents. Such methods do not reveal bioavailability (Alexander, 2000). The scheme of sequential extraction with different solvents from soils in our study was designed to operationally remove PAHs with different availabilities in soils (Macleod and Semple, 2003). Methanol removes labile nonpolar parent compounds and polar metabolites, n-butanol extracts weakly associated PAHs, and the DCM-extracted fraction represents chemically and physically associated PAHs. The decrease in the proportions of fractions extracted with methanol with increasing hydrophobicity of PAHs reveals that the methanol extracted fraction represented the dissolved PAH fraction, determined by partition characteristics of PAHs between soil organic matters and water (Koc). The Koc values of PAHs can be estimated from their related physical parameters, such as octanolewater partition coefficients (Kow) or water solubility (Sw) (Chiou et al., 1998). The Kow of benzo [a]pyrene was ten times and 36.3 times that of pyrene and phenanthrene Kow, respectively. The solubility of benzo[a]pyrene was, however, only approximately 0.01 times and 0.001 times that of pyrene and phenanthrene Kow, respectively. The difference among different proportions of methanol extracted fractions of the PAHs corresponds better to Kow than Sw. There were decreases in methanol extracted fractions with time for pyrene and benzo[a]pyrene. This tendency indicates reduced availability of PAHs with increasing soil contact time during the experiment. The n-butanol-extracted PAHs fraction from soil was very similar to concentrations accumulated by earthworms (Liste and Alexander, 2002) and therefore indicates the bioavailability of PAHs in soils. In the present study, the n-butanol fraction mainly contributed to the variation of PAHs in the rhizosphere. The proportions of methanol and DCM-extracted fractions were relatively constant, and the differences among different soils were mainly attributed to the changes in the n-butanol-extracted fraction. The ratio between methanol- and n-butanol-extracted fractions increased in the soils where more PAHs were dissipated. A lower ratio indicates higher availability because more labile PAHs were present. This result indicated that one of the mechanisms for rhizoremediation in the rice rhizosphere increased availability of PAHs, rather than immobilization. This result is approved by the previously reported result that desorbing fraction accounted for the dominant contribution to dissipated PAHs, but immobilized bound residues contributed little to the dissipation of PAHs in soils (Gao et al., 2010b). 4.2. The response of microbial activities in rhizosphere gradient
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Distance (mm) Fig. 8. Dissolved organic carbon concentrations in the rice rhizosphere of high PAH soils (a) and low PAH soils (b).
The destruction of PAHs in soils can be achieved through intraand extracellular enzyme activities of microbes (Johnsen et al., 2005). It is widely reported that microbial community and activity are affected by the rhizosphere. Our previous study based on metaanalysis indicated that PAH-degrading bacteria are significantly enhanced in the rhizosphere (Ma et al., 2010b). The extracted DNA concentrations were greatest at 4 mm away from the root surface. Bacterial biomass could be evaluated by quantifying the level of extracted DNA concentration (Aoshima et al., 2006). Therefore, the biggest stimulation of microbial activity by the rhizosphere occurred at 4 mm away from the rice root surface. The greatest PAH dissipation in the rhizosphere was, however, determined at one and two mm away from the root surface in both low- and high PAH soils, respectively. The separation between maximum microbial biomass concentration (Fig. 5) and maximum dissipation rates (Fig. 7) indicates that the growth of PAH degraders was not consistent with other microbes in the rhizosphere. The suppression of microbial activity near the root surface might be due to the depletion of
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mineral nutrients in rhizosphere. Such depletion of nutrient elements favors the growth of PAH-degrading bacteria which are well-adopted to such conditions (Johnsen et al., 2005). The bioavailability of PAHs is influenced by the rhizosphere effect as well. Root exudates have been widely reported to enhance PAH degradation through modification of microbial community structure, the promotion of microbial activity (Gao et al., 2011; Sun et al., 2010; Yi and Crowley, 2007) and the induction of PAH-cometabolic degrading enzymes (Muratova et al., 2009). Moreover, root exudates could increase desorption of PAHs from soils, thus increasing their bioavailability (Gao et al., 2010a). The difference in dissipation gradients in the rhizosphere of low- and high-PAH soils might indicate that the toxicity of PAHs to rice root activity differs. Exposure to low doses of PAHs can stimulate plant growth, but high doses of PAHs inhibit root growth and root hair initiation (Zhang et al., 2010). Consequently, the stress of high PAH concentrations in soil to rice root might be stronger than in low PAH concentration soils. 4.3. The response of dissolved organic carbons in rhizosphere gradient The high concentration of DOC in high PAH soils might be due to the metabolic products of degraded PAHs. Although the solubility of PAHs is low, the solubility of their derivatives is much higher (Yi and Crowley, 2007). The low DOC concentrations in the 3 mm soil layer indicated that oxygen released from the root could promote DOC utilization. However, the higher TOC concentrations in the one to 2 mm soil layers may also arise from degraded PAH derivatives. The low proportion of bound residual PAHs in our study indicated that only small amounts of PAHs were sequestrated into humic substance during the 45 days. The half-lives for the apparent loss of phenanthrene, pyrene, and benzo[a]pyrene in spiked soils by sequestration were 1789, 254 and 254 days, respectively (Northcott and Jones, 2001). Furthermore, the availability of PAHs was enhanced in rhizosphere. The effects of the rhizosphere on the toxicity of PAHs, rather than the total concentrations of PAHs dissipated. The toxicity of benzo[a]pyrene was one thousand times greater than phenanthrene and pyrene. The more effective remediation of PAH toxicity in the rhizosphere indicated that highly hydrophobic PAH compounds, e.g. benzo[a]pyrene, were more effectively degraded there than phenanthrene and pyrene. Soils collected from industrially polluted sites also showed that the rhizosphere contributed more to the bioremediation of highly PAH-polluted soils due to their high hydrophobicity (Ma et al., 2010a,b). 5. Conclusion The rhizoremediation effect of rice on PAH dissipation was efficient and the role of the rhizoremediation approach should be recognized when the toxicity risk of PAHs is evaluated. Concentrations of PAHs influence PAH dissipation in the rice rhizosphere. Extraction with mild extracting solvents such as n-butanol provided a useful model of the rhizosphere effect on PAH remediation. Furthermore, the greatest dissipation rate did not coincide with the greatest microbial biomass in rhizosphere. Acknowledgments This work was jointly supported by the National Natural Science Foundation of China (41090284, 20977077, 40971136), the Project Supported by Zhejiang Provincial Natural Science Foundation of China (R5110079), and the Key Scientific and Technological Program of Zhejiang Province (2008C13024-3).
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