Environ Sci Pollut Res (2013) 20:3989–3999 DOI 10.1007/s11356-012-1350-x

RESEARCH ARTICLE

Biostimulation of the autochthonous microbial community for the depletion of polychlorinated biphenyls (PCBs) in contaminated sediments Simona Di Gregorio & Hassan Azaizeh & Roberto Lorenzi

Received: 17 August 2012 / Accepted: 19 November 2012 / Published online: 4 December 2012 # Springer-Verlag Berlin Heidelberg 2012

Abstract In this study, the effect of the biostimulation of the autochthonous microbial community on the depletion of polychlorinated biphenyls (PCBs) in historically contaminated sediments (6.260±9.3 10−3 μg PCB/ g dry weight) has been observed. Biostimulation consisted of (1) the amendment of an electron donor to favor the dehalogenation of the high-chlorinated PCBs and (2) the vegetation of sediments with Sparganium sp. plants to promote the oxidation of the low-chlorinated PCBs by rhizodegradation. The effects of the treatments have been analyzed in terms of both PCB depletion and changes of the autochthonous bacterial community structure. The relative abundance of selected bacterial groups with reference to untreated sediments has been evaluated by quantitative real-time PCR. The amendment of acetate determined the enrichment of anaerobic dechlorinators like Dehalococcoides sp. Vegetation with Sparganium sp. plants determined the enrichment of either (3) the dechlorinators, Dehalococcoides and the Chloroflexi o-17/DF-1 strains or (4) the Acidobacteria, β-Proteobacteria, Actinobacteria, α-Proteobacteria, Bacteroidetes, and Firmicutes. The combination of the two biostimulation strategy determined the 91.5 % of abatement of the initial PCB content. Responsible editor: Robert Duran S. Di Gregorio (*) : R. Lorenzi Department of Biology, University of Pisa, Via Luca Ghini 5-56123, Pisa, Italy e-mail: [email protected] H. Azaizeh Institute of Applied Research (affiliated with University of Haifa), The Galilee Society, P.O. Box 437, Shefa-Amr 20200, Israel H. Azaizeh Tel Hai College, Upper Galilee 12210, Israel

Keywords Sparganium sp. . Chloroflexi o-17/DF-1 strains . Dehalococcoides sp. . Acidobacteria . β-Proteobacteria . Actinobacteria . α-Proteobacteria . Bacteroidetes . Firmicutes

Introduction Polychlorinated biphenyls (PCBs) have been utilized for a variety of industrial application. In the 1990s, the already known health effects and recalcitrance of PCBs in the environment led to prohibitions on their manufacturing, processing, and distribution all over the world. However, at that time hundreds of thousands of tons of PCBs have been already introduced into the market. Due to their high chemical stability and high affinity for hydrophobic organic solids (Girvin and Scott 1997), their detection in the environment is still ubiquitous and sediments are among the environmental compartments affected by this contamination. Despite their low bioavailability, microbial PCB biodegradation has been reported (Borja et al. 2005; Pieper and Seeger 2008). PCB congeners with four or more chlorine atoms undergo microbial anaerobic reductive dechlorination, an energy yielding process, where PCBs serve as electron acceptor for the oxidation of organic carbon (Abraham et al. 2002). Lower-chlorinated PCBs congeners, possibly resulting from anaerobic dechlorination, undergo co-metabolic aerobic oxidation mediated by dioxygenase enzymes, resulting in ring opening and potentially in the complete mineralization of the molecule (Furukawa et al. 2004; Vasilyeva and Strijakova 2007). One of the factors determining the recalcitrance of PCBs in sediments is the absence of favorable conditions for biodegradative processes. In the case of high-chlorinated PCBs, the shortage of electron donors, supporting microbial reductive dehalogenation, might be involved. Actually, the amendment

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of electron donors has been described as capable to elicit the transformation of PCBs in contaminated matrices (Lloyd and Paige 2008). On the other hand, the recalcitrance of lowchlorinated PCBs to biodegradation might be related to the dominant reductive conditions associated to anoxic sediments, impeding processes of oxidation of the contaminants. Biostimulation of the biodegradative capacity of the autochthonous microbial communities results promising in enhancing bioremediation in contaminated matrices. Among biostimulation approaches, phytoremediation, favoring rhizodegradation processes, found application for accelerating the depletion of disparate types of contamination, comprising PCBs (Chekol et al. 2004; Li et al. 2011; Mackova et al. 2009; Slater et al. 2011). In relation to sediments, plants, by exploring the matrix with oxygen-transporting roots, offer the possibility of exposing the anoxic sediments to aerobic conditions, favoring the oxidation of low-chlorinated PCBs. The differential effects of diverse plant species on the depletion of the low-chlorinated fraction of PCBs in soil and sediment have actually been reported (Chekol et al. 2004; Perelo 2010). Only a limited number of attempts on the use of plant species to deplete the high-chlorinated fraction in contaminated soil and sediment have been performed (Epuri and Sorensen 1997; Slater et al. 2011; Smith et al. 2007). The application of phyto-based technologies in the decontamination of dredged sediments is frequent because of sustainability and efficacy. The optimization of the exploitation of these technologies for the depletion of priority pollutants is actually desirable. The objective of the current study consisted in validating the biostimulation of the autochthonous microbial community as an exploitable approach to elicit the depletion of PCBs in historically contaminated sediments, collected from a water canal crossing a dismissed industrial site in Italy. During the last 10 years, the site underwent physical–chemical treatments for its reclamation; however, the residual level of PCB contamination is still reason of public concern. In this study, the biostimulation approach for the depletion of the residual PCBs in the sediments has been designed as a direct intervention on the water saturated matrix by (1) the spiking of acetate as electron donor to elicit high-chlorinated PCBs dechlorination and by (2) the vegetation of sediments with a low-transpiring aquatic plant to establish a rhizoeffect on the autochthonous microbial communities, eliciting the oxidation of low-chlorinated PCBs. A lowtranspiring plant has been selected to avoid any excess in the increase of the redox potential in the sediment, possibly inhibiting reductive dehalogenation (Smith et al. 2007). The selected plant species was Sparganium sp., naturally growing in the area and actually described as a low-transpiring plant species (Calhoun and King 1997). The effects of the different strategy of biostimulation were evaluated either in terms of PCB depletion or, with

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reference to untreated sediments, in terms of changes in the autochthonous bacterial community structure with reference to bacterial taxa already detected in PCB-contaminated matrices of different origins.

Material and methods Chemicals, plants, and sediments Chemicals used throughout the experiments were of analytical grade and purchased from Sigma-Aldrich (Milan, Italy). PCB congeners and the industrial PCB mixture Aroclor 1248, 1254, and 1260 were purchased by Ultrascientific (USA). Sparganium sp. plants were collected from a local nursery at the same developmental stage. PCBcontaminated sediments (6.260± 9.3 10−3 μg PCB/g dry weight, Table 1) have been collected from a canal crossing, a dismissed industrial site in Italy. The texture of the sediments was sandy–loam (45 % silt, 55 % sand, and 15 % clay) with total phosphorous, 1.2 %; total organic carbon, 1.1 %; total organic matter, 2.2 %; and total nitrogen, 1.3 %. Any sediment contamination by total petroleum hydrocarbon (TPH), policyclic aromatic hydrocarbons (PAH), and heavy metals have been detected. Experimental conditions A total of 36 experimental replicates containing 5 kg of air-dried dredged sediments were prepared in plastic pots and maintained in a temperature (24±1 °C) and lightening controlled growth chamber (14-h light/10-h dark). Sediment in pots was saturated with a water volume exceeding (5 % v/v) the one corresponding to the maximum water-holding capacity. A total of 30 pots, out of the 36, were spiked (biostimulated) with Na–acetate dissolved, at the nominal concentration of 3 mmol/l, in the water volume used to saturate the sediment. The six pots that were not spiked with acetate were analyzed after 12 months of incubation in the growth chamber as control of the biostimulation process. After 6, 9, and 12 months of incubation in the growth chamber, six acetate-spiked pots for each time point were processed for analysis. In parallel, after 6 and 9 months of anaerobic incubation in the growth chamber, a total of six acetate-spiked pots for each time-points were vegetated with 3 Sparganium sp. plants per pot and analyzed after further 3 months of incubation in the chamber. The redox potential in sediments was monitored every 3 days by using a redox tester (HI 98120, HANNA Instruments), recording the parameter in different points per pots up to the stabilization of the recorded values.

Environ Sci Pollut Res (2013) 20:3989–3999 Table 1 PCB congeners in sediment at the beginning of the experimentation

IUPAC number is the number given to the compound by the International Union of Pure and Applied Chemistry

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PCB congeners

IUPAC number

μg PCB congeners/g dry weight sediment

PCB-128 PCB-138 PCB-153 PCB-156 PCB-169 PCB-180 ∑ PCBs

2,2′,3,3′,4,4′-Hexachlorobiphenyl 2,2′,3,4,4′,5′-Hexachlorobiphenyl 2,2′,4,4′,5,5′-Hexachlorobiphenyl 2,3,3′,4,4′,5-Hexachlorobiphenyl 3,3′,4,4′,5,5′-Hexachlorobiphenyl 2,2′,3,4,4′,5,5′-Heptachlorobiphenyl

0.003±1.1×10−4 0.002±1.1×10−4 0.004±1.1 10−4 2.567±1.3×10−3 0.006±2.3×10−4 3.678±2.4×10−3 6.260±9.3×10−3

PCB extraction and analytical procedures Volatile fatty acids were determined from the water layer of the different pots by HPLC system equipped with an electrical conductivity detector CDD10AVP (Shimadzu). The HPLC system was operated with a Shim-pack SCR column (300 mm L×7.9 mm ID; Shim.pack SCR-102H). A 10-μl aliquot of sample was loaded and eluted isocratically using 5 mM p-toluenesulphonate (pH2.7) at a flow rate of 0.8 ml/ min. The oven temperature was 40 °C. Sediments and plant samples, analyzed for PCB content, derived from the independently destructive dissection of six pots per time-points of analysis in (1) plants that were pulled out from pots and representative sub-samples were prepared from roots and shoots (the total dry mass of the plant was determined by applying the dry/wet factor from the subsampled tissues to the total wet biomass) and in (2) the sediments that, in each pots, were roughly mixed and a total of 10 different samples of 0.5 g of sediments were randomly collected. The different 10 sediment samples per pot were mixed together and the resulting 5 g of sediments were divided into two parts and separately extracted and analyzed. One part was qualitatively and the second part was quantitatively analyzed for PCB content. Sediments and plant samples were dried overnight in a vented oven at 25 °C for approximately 24 h. All samples were extracted using a soxhlet apparatus with dichloromethane as the solvent. Qualitative and quantitative analysis of PCBs was performed with an Agilent 6890 GC-5975B Series MS system in the selective ion monitoring mode. Injection of 1 μL of samples was conducted in the splitless mode with a purge time of 1.0 min. Separation of PCB congeners was carried out with a DB-5 (60 m×0.25 mm× 0.25 μm) capillary column. The oven temperature was programmed from 120 to 180 °C at a rate of 6 °C/min to 240 °C at 1 °C/min and to 290 °C at 6 °C/min and finally hold on for 15 min. The temperature of both the injector and detector was 260 °C. He was used as the carrier gas. Qualitative analysis of PCBs was performed by comparing the retention time of the GC peaks obtained from the analysis of the sediment organic extracts with those of standard Aroclor 1248, 1254, and 1260 and those of the 12 dioxin-

like PCB congeners (PCB-77, PCB-81, PCB-105, PCB-114, PCB-118, PCB-123, PCB-126, PCB-156, PCB-157, PCB167, PCB-169, and PCB-189) analyzed under identical conditions. Mixtures of standards of the PCBs identified in sediments during the qualitative analysis at the different timepoints were used for the determination of the single PCB congener concentrations at the corresponding time of analysis. Linear five-points calibration curve has been obtained for each PCB congeners mixture used for the quantification. The PCB30, PCB65, and PCB204 were used as surrogate standards. A known amount of internal standard (PCB24, PCB82, and PCB198) was added before instrumental analysis. The PCB concentrations were calculated from the calibration curves. Recoveries were calculated from the surrogates and quantification was performed by the internal standard method. Quality assurance and quality control For each of the samples, a procedural blank (solvent), a spiked blank (targeted PCB congeners spiked into the solvent), a spiked matrix sample (targeted PCB congeners spiked into pre-extracted sediment), and a spiked matrix duplicate were processed. The limit of detection (LOD) was defined as three times the ratio of signal to noise for analytes which were not detected in the blanks and three times of the standard deviation of the values in the blanks. The LODs for the analytes were ranged from 0.097 to 0.96 pg/g dry weight (d.w.). The recoveries of the surrogates PCB30, PCB65, and PCB204 were 82–99 %, 94–105 %, and 98–114 %, respectively, and recoveries of the targeted PCB congeners were 82–105.0 % in three spiked blank duplicates, while they ranged from 80 to 109 % in three spiked matrix duplicates. Three sample duplicates were analyzed and their relative standard deviations were <5 %. Bacterial community structure The rhizospheric portion of sediments was collected as the portion of sediment adhering to the root apparatus of the three plants vegetating each pots. A total of 30 different sediment samples per plant was collected. Each sample

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accounted for a total of 250 mg. All the samples collected per pot were deeply mixed before sampling the 250 mg of sediment processed for the extraction of the rhizospheric metagenomic DNA. The bulk portion of sediments in vegetated pots, as well as in not-vegetated pots, has been collected by deeply mixing the sediments in each pot after intact plants have been pulled out from the pots. A total of 10 different samples of 250 mg of sediments per pot were randomly collected and mixed before sampling the 250 mg of sediment processed for the extraction of the bulk metagenomic DNA. The metagenomic DNA has been purified using the UltraClean™ Soil DNA Isolation Kit (MO BIO Laboratories, Carlsbad, CA) following the manufacturer’s instructions. DNA extracts were quantified by the OD260 using a NanoDrop ND-2000 spectrophotometer (NanoDrop, Wilmington, DE). The quality of the DNA was assessed by the ratio OD260/OD280. Quantitative real-time PCR was applied to assess the community structure at level of taxonomic group by quantifying the relative abundance of different bacterial groups in biostimulated sediments in comparison to not-biostimulated sediments. Taxa specific 16S rDNA primers were used for quantification of the Actinobacteria, Acidobacteria, α- and β-Proteobacteria, Bacteroidetes, Firmicutes, Dehalococcoides sp., and the Chloroflexi o-17 and DF-1 strains by quantitative real-time PCR. Primers used are described in Table 2. PCR reactions were carried out on an ABI Prism® 7300 Sequence Detection System using Sybr® Green PCR Master Mix (Applied Biosystem, Monza, Milan, Italy). The presence of PCR inhibitors was estimated by diluting the metagenomic DNA and by mixing a known amount of standard DNA to the metagenomic DNA prior to quantitative real-time PCR. No inhibition was detected in both cases. For each reaction, approximately 2 ng of DNA was used as template, forward and reverse primers were used at a 0.6 mM final concentration, and water to a final volume of 50 μl. Samples were always measured in triplicate. Cycling condition consisted of an initial denaturation at 95 °C for 10 min: 40 cycles—denaturation at 95 °C for 15 min and annealing–extension at 55–62 °C for 1 min. After this step, Table 2 Bacterial groupspecific primers

acquisition took place. A melting curve ranging from 50 to 99 °C was performed with steps of 1 °C and a hold of 5 s to check the specificity of the assays. For each 16S rRNA target, a standard curve was established using serial dilutions of linearized plasmid pGEM-T (102 to 107 copies) containing cloned 16S rDNA. Two no-template controls (NTC) were also included in all the assays. At each run, two standard curves per sample were included and all the samples were related to the corresponding standard curve. Amplification efficiencies were calculated from the slopes of the standard curves (Pfaffl 2001). To confirm the specificity of each of the primer sets, samples from the first run using the primer set were loaded onto a 2 % agarose gel. Primer sets did not yield any side products in lowcopy number dilutions and in NTCs. The amplification levels of group-specific genes were first normalized by the amplification level of 16S rDNA (Blackwood et al. 2005; Fierer et al. 2005). The fractional copy number of each taxa was calculated as the ratio between the normalized amplification levels for each taxa in biostimulated sediments and the normalized amplification levels of the taxa in not-biostimulated sediments. All the results were computed with ABI Prism 7300 software v1.3.1. Statistical analysis All the data were elaborated with the aid of ANOVA and means were separated by the Fisher’s least significant difference (LSD) test (P≤0.05) using the specific software Statgraphics 5.1 (Statistical Graphics Corp., USA).

Results PCB depletion The PCB congeners identified in the sediment at the beginning of the experimentation were shown in Table 1. PCB180 and PCB-156 resembled nearly the total content in PCBs. Actually PCB-128, PCB-138, PCB-153, and PCB-

Target group

Primer

All groups α-Proteobacteria β-Proteobacteria Acidobacteria Actinobacteria Bacteroidetes Firmicutes Dehalococcoides o-17/DF-1 strains

341F/534R Eub338/Alf685 Enb338/Bet680 Acid31/Eub518 Actino235/Eub518 Cfb319/Eub518 Lgc353/Eub518 DHC 1/DHC1377 14 F/Dehal1265

Annealing T (°C)

PCR efficiency (%)

60 60 55 55 60 60 55 55 62

93 100 96 94 90 95 95 98 97

Reference

Muyzer et al. 1993 Lane 1991 Overmann et al. 1999 Barns et al. 1999 Stach et al. 2003 Manz et al. 1996 Meier et al. 1999 Hendrickson et al. 2002 Watts et al. 2005

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169 were present only in trace amounts, resulting to be under the detection limits at the successive time of analysis. Any contamination by TPH, PAH and heavy metal was detected. Volatile fatty acids in the water phase of sediments were measured at the beginning of the experimentation and resulted to be under the detection limits. Thus, acetate was spiked in each pots at the concentration of 3 mmol/l in the water volume used to saturate the sediment (Fig. 1). The spiked-acetate was nearly completely depleted during the first 6 months of not-vegetated incubation (Fig. 1A) and resulted to be mandatory to promote the depletion of the total PCBs since in not-acetate-spiked sediments (nAcSnV)

A

Acetate (mmol/l of water)

3

2

1

0

-1 120

B

80

60

a

a

b

AcS6V3

AcS9

40

AcS6

% total PCB depletion

100

b

nAcSnV

AsS12

BE

0

AcS9V3

20

Fig. 1 a Acetate content in the water phase of sediments at the successive time-points of analysis. b Percentage of abatement of total PCBs at the successive time-points of analysis. The same letters on the bars indicate values not significantly different at the 5 % level by LSD test (P>0.05). Abbreviations: BE, beginning of the experimentation; AcS6, acetate-spiked sediments incubated for 6 months in notvegetated conditions; AcS6V3, acetate-spiked sediments incubated for 6 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS9, acetate-spiked sediments incubated for 9 months in not-vegetated conditions; AcS9V3, acetate-spiked sediments incubated for 9 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS12, acetate-spiked sediments incubated for 12 months in not-vegetated conditions; nAcSnV, notbiostimulated sediments: not spiked with acetate and not vegetated with Sparganium sp.

PCB depletion was absent (Fig. 1B). In acetate-spiked sediments, after 6 months of not-vegetated incubation (AcS6), a percentage of depletion of 19.2 % (P00.01) of the total PCBs has been observed followed by an increment up to 25.5 % during the successive 3 months of not-vegetated incubation (AcS9) (P00.02). The percentage of total PCB depletion did not significantly changed between 9 months (AcS9) and 12 months (AcS12) of not-vegetated incubation (P00.07), reaching the maximum value of 26.9 % after 12 months of not-vegetated incubation (AcS12). Vegetation of acetatespiked sediments with Sparganium sp. determined a significant improvement in total PCB depletion (91.5 %) only when vegetation occurred after 9 months of not-vegetated incubation (AcS9V3). On the contrary, when vegetation occurred after 6 months of not-vegetated incubation (AcS6V3), PCB depletion was lower than the one recorded in not-vegetated sediment incubated for the same interval of time in the growth chamber (AcS9; P00.04; Fig. 1b). The PCB quantification in sediments indicated that the congeners that contributed to the definition of the total PCB content changed with the time of incubation (Fig. 2). As previously stated, at the beginning of the experimentation (BE), the total PCB content in sediments was ascribable to PCB-180 and PCB-156 (Fig. 2A). After 6 months of notvegetated incubation (AcS6), in addition to the highchlorinated PCB-156 and PCB-180 (Fig. 2A), the lowchlorinated PCB-52, PCB-77, and PCB-81 were also detected (Fig. 2B). The latter reached easily measurable concentrations (one order of magnitude lower than PCB180 and PCB-156) only in not-vegetated sediments. Their concentrations increased between 6 (AcS6) and 9 (AcS9) months of not-vegetated incubation (0.02
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c

e

e

2

60

40 ab 20

0

0

B

0.10

m l

lm n

0.06 0.04

p qg

p gq hr i

hr i

nAcSnV

AcS12

AcS9V3

AcS9

AcS6V3

AcS6

0.00 BE

ab

n

0.08

0.02

AcS6

1

PCB-81 PCB-77 PCB-52

cd

cd

nAcSnV

c

a

f

80

AcS12

f

a

0.12 µ g PCB congeners/g dry weight sediment

d

AcS9V3

3

b

AcS9

d

PCB-180 PCB-156 PCB-81 PCB-77 PCB-52

100

AcS6V3

b

4

A

PCB-180 PCB-156

% depletion PCB congeners

µ g PCB congeners/g dry weight sediment

5

Fig. 2 PCB congeners concentrations in sediments at the different timepoints of analysis. The same letters on the bars indicate values not significantly different at the 5 % level by LSD test (P>0.05). Abbreviations: BE, beginning of the experimentation; AcS6, acetate-spiked sediments incubated for 6 months in not-vegetated conditions; AcS6V3, acetate-spiked sediments incubated for 6 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS9, acetatespiked sediments incubated for 9 months in not-vegetated conditions; AcS9V3, acetate-spiked sediments incubated for 9 months in notvegetated conditions and vegetated for 3 months with Sparganium sp.; AcS12, acetate-spiked sediments incubated for 12 months in notvegetated conditions; nAcSnV, not-biostimulated sediments: not spiked with acetate and not vegetated with Sparganium sp.

(AcS9) (0.03
Fig. 3 Percentage of abatement of the single PCB congeners at the successive time-points of analysis. The same letters on the bars indicate values not significantly different at the 5 % level by LSD test (P>0.05). Abbreviations: AcS6, acetate-spiked sediments incubated for 6 months in not-vegetated conditions; AcS6V3, acetatespiked sediments incubated for 6 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS9, acetate-spiked sediments incubated for 9 months in not-vegetated conditions; AcS9V3, acetate-spiked sediments incubated for 9 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS12, acetate-spiked sediments incubated for 12 months in not-vegetated conditions; nAcSnV, not-biostimulated sediments: not spiked with acetate and not vegetated with Sparganium sp.

PCBs removed by plant absorption was negligible with reference to the portion removed by transformation (Table 3). In not-vegetated sediments, the redox potential was constantly negative (≈Eh≤−230±0.56 mV). In the case of vegetation of sediments with Sparganium sp., the related values were between Eh ≤ −227± 0.43 mV and Eh≤−203±0.67 mV.

Bacterial community structure Dehalococcoides sp., Chloroflexi o-17/DF-1 strains, Acidobacteria, β-Proteobacteria, Actinobacteria, αProteobacteria, Bacteroidetes, and Firmicutes have been quantified in biostimulated sediments with reference to the not-biostimulated one. Results obtained are shown in Fig. 4. The taxonomic bacterial groups were all detected at the BE. Biostimulation of sediments induced changes in the bacterial community structure. In acetate-spiked sediments, after 6 (AcS6) and 9 (AcS9) months of not-vegetated incubation, a progressive enrichment of the Dehalococcoides sp. up to a total of 8.6 (AcS6) and 16.1 (AcS9) times higher than at the BE has been observed (0.02
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Table 3 Contribution of plant absorption to low-chlorinated PCB depletion PCB congeners

AcS6V3 in plant

AcS6V3 removed in total

AcS9V3 in plant

AcS9V3 removed in total

(%) of depletion PCB-52 PCB-81 PCB-77

3.4±0.04 2.7±0.09 4.6±0.06

91.99±0.78 93.89±0.82 93.02±0.76

4.7±0.05 5.2±0.04 7.2±0.08

98.87±0.98 99.67±0.99 96.99±0.79

AcS6V3 acetate-spiked sediments incubated for 6 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp., AcS9V3 acetate-spiked sediments incubated for 9 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp

after 9 months in the same condition (AcS9; 6.7 instead of 16.1 times higher than at BE; P00.02). In the rhizospheric and bulk portion of sediments, vegetation with Sparganium sp. determined a decrement in fractional copy numbers of Dehalococcoides sp. with reference to the values reached after 6 (AcS6) and 9 (AcS9) months of not-vegetated incubation (0.01
Acidobacteria ß-Proteobacteria Actinobacteria a-Proteobacteria Bacteroidetes Firmicutes Dehalococcoides sp. o-17/DF-1 group

Fractional copy number

100 80 60

b b

bb b b b

40

bb

a aa a a

20

nAcSnV

AcS12

AcS9V3 rhizo

AcS9V3 bulk

AcS9

AcS6V3 rhizo

AcS6V3 bulk

AcS6

BE

0

Fig. 4 Fractional copy numbers of bacterial groups at the different time-points of analysis. The same letters on the bars indicate values not significantly different at the 5 % level by LSD test (P>0.05). Abbreviations: BE, beginning of the experimentation; AcS6, acetatespiked sediments incubated for 6 months in not-vegetated conditions; AcS6V3 bulk, bulk portion of sediments in acetate-spiked sediments incubated for 6 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS6V3 rhizo, rhizospheric portion of sediments in acetate-spiked sediments incubated for 6 months in notvegetated conditions and vegetated for 3 months with Sparganium sp.; AcS9, acetate-spiked sediments incubated for 9 months in notvegetated conditions; AcS9V3 bulk, bulk portion of acetate-spiked sediments incubated for 9 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS9V3 rhizo, rhizospheric portion of acetate-spiked sediments incubated for 9 months in not-vegetated conditions and vegetated for 3 months with Sparganium sp.; AcS12, acetate-spiked sediments incubated for 12 months in notvegetated conditions; nAcSnV, not-biostimulated sediments: not spiked with acetate and not vegetated with Sparganium sp.

fractional copy number of Dehalococcoides sp. was the highest recorded during the experimentation (25.8 times higher than at BE) (0.01
Discussion The sediment analyzed in this experimentation was contaminated by high-chlorinated PCBs. The contamination was mainly restricted to PCB-180 and PCB-156 and few other highchlorinated congeners that were detected in trace amounts only at the beginning of the experimentation. Historical data

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collected on the industrial activity performed on the site indicated that it was dedicated to the production of different end products like paints and pastes that definitely contained PCBs. However, the last period of the activity was mainly dedicated to the production of additives for the above-mentioned final products. PCB production was restricted specifically to the PCB180 and PCB-156. In this context, it is reasonable to assume that the presence of the latter in sediments was mainly the result of the contamination of the site occurred during the last industrial productive period. They do not necessarily represented the residual portion of PCB contamination still present after the series of physical-chemical treatments adopted, during the past 10 years, for the reclamation of the area. At the same time, the level of contamination ascribable to the two high-chlorinated PCBs was reasonably originally very high and consequently not completely depleted by the intervention. With the aim to adopt an in situ bio-based protocol for the depletion of the residual PCB contamination in sediments, we tentatively restored the favorable physiological conditions for the occurring of the bacterial reductive dechlorination of the high-chlorinated PCBs still present in the matrix. The maintenance of the sediment as saturated in water ensured the negative redox potentials necessary for the dechlorinating process. Nevertheless, at the beginning of the experimentation, a pool of electron donors necessary to sustain the reductive dehalogenation was not measurable. Consequently, acetate was spiked in sediment. Furthermore, to counteract the expected production of low-chlorinated congeners, a phyto-based approach capable to promote their depletion by rhizodegradation has been adopted. The selected plant species was Sparganium sp. because growing on the site and described as a low-transpiring plant species (Calhoun and King 1997). The exploitation of lowtranspiring plants to eventually elicit PCB rhizodagradation has been described as mandatory to avoid the inhibition of PCB dechlorination by the shifting of the redox potential of the vegetated matrices to positive values (Smith et al. 2007). Results here obtained showed that Sparganium sp. induced a shift of redox potentials of sediments to values still compatible with anaerobic environments, where reductive dehalogenation can occur (DeLaune and Reddy 2005). Thus, Sparganium sp. resulted to be exploitable for the highchlorinated PCB depletion, even though its direct intervention by absorption was here excluded. In fact, any absorption of PCB-156 and PCB-180 has been recorded. PCB-52, PCB-77, and PCB-81 were absorbed by the root apparatus; however, the percentage of absorbed PCBs accounted for marginal values with reference to the percentage depleted by their transformation. Thus, it is reasonable to consider that the microbial transformation of the contaminants is the main mechanism responsible for the here detected PCB depletion. Accordingly, the spiking of acetate in anaerobic condition elicited the depletion

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of PCB-156 and PCB-180 at the expenses of acetate and its consumption was accompanied by the rapid interruption of the high-chlorinated PCB depletion. Bacterial reductive dechlorination of high-chlorinated PCBs in sediments has been actually extensively described and, in most of the cases, Dehalococcoides spp. were indicated as the dechlorinators capable of dechlorination on metaand para- positions on aromatic rings, with the ortho-substitute congeners as the dominant accumulating dechlorination products (Bedard et al. 2007; Quensen et al. 1990). However, microbial enrichment cultures, capable of ortho-dechlorination processes, have been also described and the Chloroflexi o17/DF-1 have been identified as the ortho-dechlorinating strains (Cutter et al. 2001; Wu et al. 2002). In this context, it is reasonable to assume that the dechlorination of the PCB-156 and PCB-180 in para- and/or metapositions on the aromatic rings determined the accumulation in sediments of the ortho-substituted congener PCB-52. The process can be ascribed to the here detected Dehalococcoides sp. that actually increased in fractional copy number concomitantly to the progressive PCB-52 accumulation in sediment. On the other hand, it is reasonable to assume that the non ortho-PCB-77 and PCB-81 accumulation in sediments derived from the dechlorination of PCB-180 and PCB-156 on ortho-positions. The process can be reasonably ascribed to the here detected Chloroflexi o-17/DF-1 strains. However, in this case, a positive correlation between their fractional copy number and the accumulation of PCB-77 and PCB-81 was not recorded. In fact, despite the progressive accumulation of the PCB-77 and PCB-81 congeners, starting from the spiking of acetate, an increase in fractional copy number of the o-17/DF-1 strains was recorded only in the bulk portion of sediments vegetated after 9 months of notvegetated incubation. Thus, the possibility of the involvement of bacterial groups capable of ortho-dechlorination of PCB-180 and PCB-156, diverse from o-17 and DF-1 strains cannot be excluded, even though we were not able to detect the candidates. Even though acetate was here mandatory for the depletion of PCBs, the volatile fatty acid had no effect on the dechlorinating Chloroflexi o-17/DF-1 strains, as well as it had no effect on Acidobacteria, β-Proteobacteria, Actinobacteria, αProteobacteria, Bacteroidetes, and Firmicutes that actually, as well as the o-17/DF-1 strains, increased in fractional copy number only in response to vegetation. The positive effect of vegetation on diverse microbial communities is actually known and a stronger effect in the rhizospheric portion of the vegetated matrix is also expected. In fact, plants produce exudates (Van Aken et. 2010) that among other effects increase the pool of bioavailable carbonaceous sources for the microorganisms colonizing the rhizosphere. The concentration of exudates follows a gradient that reaches lower values with the increase of the distance from the

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root. Consequently, the bacterial load in the rhizospheric portion of the vegetated matrices is always higher than the bacterial load in the bulk one. On the other hand, in the rhizospheric portion of vegetated sediments, the inhibition of strictly anaerobic microorganism is also expected because of the capacity of the plant roots to transport oxygen. Accordingly, in consequence of vegetation of sediments, we observed an impressive increase in the fractional copy number of Acidobacteria, β-Proteobacteria, Actinobacteria, αProteobacteria, Bacteroidetes, and Firmicutes, with reference to the beginning of the experimentation and to not-vegetated sediments. The increase was higher in the rhizospheric portion of sediments if compared to the bulk one and it was higher when vegetation occurred in sediments after 9 instead of 6 months of not-vegetated incubation. On the other hand, as expected, vegetation of sediment induced a decrease in fractional copy number of the strictly anaerobic Dehalococcoides sp. in the rhizospheric portion of the matrix. However, the observed drastic decrease of the fractional copy number of Dehalococcoides sp. in the bulk portion of sediment vegetated after 6 months of not-vegetated incubation, was not expected and actually did not occur in the bulk portion of sediments vegetated after 9 months of notvegetated incubation. At this time-point of analysis either Dehalococcoides sp., Acidobacteria, β-Proteobacteria, Actinobacteria, α-Proteobacteria, Bacteroidetes, or Firmicutes increased in fractional copy number and the Dehalococcoides sp. accounted for the highest value observed in the frame of this experimentation. On the contrary, the decrease of the Dehalococcoides sp. fractional copy number, recorded in sediments vegetated after 6 months of not-vegetated incubation, determined the inhibition of PCB180 and PCB-156 depletion. In this context, it is worth mentioning that PCB dehalogenation is a cometabolic process that needs carbon source diverse for the chlorinated aromatics to proceed (Abraham et al. 2002). The dramatic increase in the fractional copy number of Acidobacteria, β-Proteobacteria, Actinobacteria, αProteobacteria, Bacteroidetes, and Firmicutes in the bulk portion of sediment in response to vegetation could have created the conditions for a competition for carbonaceous sources between the dechlorinator Dehalococcoides sp. and the above-mentioned numerical dominant taxa, which might have suppressed the Dehalococcoides sp. growth and metabolic activity and consequently the further dechlorination of the high-chlorinated PCBs. In this context, it is worth mentioning that the concentration of PCBs in sediments after 6 months of not-vegetated incubation was higher than the one recorded after 9 months in the same incubation condition. Considering that PCBs exert a toxic effect on bacterial cells (Cámara et al. 2004), higher level of contamination can be associated to a lower capacity of the autochthonous microbial community to positively react to the external stimuli, comprising root

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exudates and metabolic intermediates of microbial origin. In fact, the level of contamination of a matrix is one of the physical–chemical parameters that distress the microbial community composition and physiology, affecting the activation of different metabolic profiles in response to different stimuli. This scenario might have been responsible for the different responses of Dehalococcoides sp. to Sparganium sp., when vegetation occurred at different degree of contamination of sediments, as well as it might have been responsible for the different responses of the o-17/DF-1 strains to the same biostimulation approach. In fact, when vegetation occurred after 9 months of anaerobic incubation, the concomitant increase of o-17/DF-1 strains and Dehalococcoides sp., in the bulk portion of sediments, has been observed and actually it matched with the highest percentage of PCB depletion, accounting for the 91.5 % of the total. In this context, it is reasonable to suggest that the increase in the copy number of both meta-, para-, and ortho- dechlorinators enlarged the specificity of the dehalogenators for chlorine substitution, facilitating the depletion of the here-detected high-chlorinated PCBs. However, when vegetation occurred at higher level of contamination such as in sediments after 6 months of not-vegetated incubation, the o-17/DF-1 strains were not capable to positively respond to the biostimulating effect of Sparganium sp. and any increment in the corresponding fractional copy number has been observed. It is interesting to note that when vegetation occurred after 9 months of not-vegetated incubation, a higher increment in fractional copy number of Acidobacteria, βProteobacteria, Actinobacteria, and α-Proteobacteria with reference to Bacteroidetes and Firmicutes has been also observed. The effect was present either in the bulk or rhizospheric portion of vegetated sediments. The diverse response of Bacteroidetes and Firmicutes with reference to Acidobacteria, β-Proteobacteria, Actinobacteria, and αProteobacteria to vegetation excludes the occurring of a generalized eliciting effect of the plant on the taxa indigenous of the sediment. Eventually the difference suggests a specific response of the taxa to the biostimulation strategy. Moreover, it is worth mentioning that the increment in fractional copy number of Acidobacteria, β-Proteobacteria, Actinobacteria, α-Proteobacteria, Bacteroidetes, and Firmicutes was concomitant to the depletion of the lowchlorinated PCB-52, PCB-77, and PCB-81 that occurred only in presence of Sparganium sp. Actually, Acidobacteria, β-Proteobacteria, Actinobacteria, αProteobacteria, Bacteroidetes, and Firmicutes have been already described as enriched by the spiking of PCBs in soil by Correa et al. (2010). The authors reported also a higher increment in fractional copy number of Acidobacteria, β-Proteobacteria, Actinobacteria, and α-Proteobacteria with reference to Bacteroidetes and Firmicutes, and suggested an involvement of the different taxa in the low-chlorinated PCB depletion. Aguirre

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de Cárcer et al. (2007) observed an enrichment of βProteobacteria in historically PCB-contaminated soil when compared to non-contaminated soil. Nogales et al. (1999, 2001) identified a large majority of α-, β-, and γ- Proteobacteria, Acidobacteria, and Actinobacteria in PCB-contaminated soil. Tillmann et al. (2005) detected the predominance of bacteria belonging to α- and β-Proteobacteria onto the Bacteroidetes and Firmicutes in PCB-contaminated soil. Unlike what was reported by the other authors, we actually did not detect predominant bacterial groups at the beginning of the experimentation; however, either Acidobacteria, βProteobacteria, Actinobacteria, α-Proteobacteria, Bacteroidetes, or Firmicutes were present in the contaminated sediments. On the other hand, PCB depletion was concomitant to the increase in their fractional copy number and, accordingly to what reported by the other authors, a predominance of Acidobacteria, β-Proteobacteria, Actinobacteria, and α-Proteobacteria on Bacteroidetes and Firmicutes in presence of PCB has been observed. The enrichment of the same taxa and the enrichment of the latter with the same reciprocal ratio in PCB-contaminated matrices of different origins are noteworthy, as it may represent selection of specific bacterial taxa and/or population due to the presence of PCBs. In this context, it should be mentioned that the increment of Actinobacteria, β-Proteobacteria, or α-Proteobacteria represents the increment of bacterial taxa described as harboring biphenyl dioxygenases and involved in the oxidation of low-chlorinated PCBs (Furukawa and Fujihara 2008). In fact, genera like Comamonas, Rhodococcus, Burkholderia, and Corynebacterium spp. belonging to the Actinobacteria group, Alcaligenes, Achromobacter, and Ralstonia belonging to the β-Proteobacteria group, Sphimgomonas belonging to α-Proteobacteria, constitute the major bacterial groups where biphenyl dioxygenases genes have been cloned (Borja et al. 2005; Furukawa and Fujihara 2008). A direct involvement of the above-mentioned taxa in PCB depletion by oxidation of the aromatic rings cannot be excluded. At the same time, they could also be indirectly involved in the process of low-chlorinated PCB transformation, singularly or as components of a PCB-selected bacterial population, as bacterial species that produce metabolic favorable milieu for the activation of the essentially cometabolic process of PCB transformation and depletion.

Conclusions The combinatorial biostimulation of dehalogenation of highchlorinated PCBs and rhizodegradation of low-chlorinated PCBs resulted to be successful in the depletion of PCBs in the contaminated sediments. For the first time, Sparganium sp. has been described as capable to elicit PCB depletion. The

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peculiar capacity of Sparganium sp. to elicit either dechlorination or oxidation of the different PCB congeners in saturated sediments resulted to be essential for the success of the approach. However, the bacterial capacity to transform the contaminants has been acknowledged as the principal mechanism of PCB depletion. For the first time, a detailed description of the changes in the microbial community structure in response to biostimulation approaches dedicated to either highchlorinated PCB reductive dechlorination or low-chlorinated PCB transformation has been described in terms of the relative abundance of the bacterial taxonomic groups eventually responsible for PCB depletion. Dehalococcoides sp. and the Chloroflexi o-17 and DF-1 were presumably responsible for the depletion of the high-chlorinated PCBs by the dechlorination of either para-, meta-, or ortho- positions of the biphenyl rings. The increment in Acidobacteria, β-Proteobacteria, Actinobacteria, α-Proteobacteria, Bacteroidetes, and Firmicutes fractional copy number was concomitant to the depletion of the low-chlorinated PCBs. A direct and/or indirect role in the low-chlorinated PCB depletion has been envisaged. The effectiveness of the approach herein described requires to be verified at field scale; however, results obtained suggested that the integration of biostimulation of the autochthonous microbial community by the spiking of electron donors and the vegetation with Sparganium sp. can be exploited as a sustainable in situ approach for the management of watersaturated dredged sediments, eventually dedicated to the site of their disposal and containment, before their eventual reallocation as decontaminated matrices.

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