Science of the Total Environment 497–498 (2014) 345–352

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Impact of fresh organic matter incorporation on PAH fate in a contaminated industrial soil Audrey Pernot a,b,c,d, Stéphanie Ouvrard a,b,⁎, Pierre Leglize a,b, Françoise Watteau a,b,e, Delphine Derrien f, Catherine Lorgeoux e,g,h, Laurence Mansuy-Huault c,d, Pierre Faure c,d a

Université de Lorraine, LSE, UMR 1120, Vandœuvre-lès-Nancy, F-54518, France INRA, LSE, UMR 1120, Vandœuvre-lès-Nancy, F-54518, France Université de Lorraine, LIEC, UMR 7360, Vandœuvre-lès-Nancy, F-54506, France d CNRS, LIEC, UMR 7360, Vandœuvre-lès-Nancy, F-54506, France e CNRS, UMS 3562, Vandœuvre-lès-Nancy, F-54501, France f INRA, BEF, UR 1138, Centre Nancy-Lorraine, Champenoux, F-54280, France g Université de Lorraine, Géoressources, UMR 7359, Vandœuvre-lès-Nancy, F-54506, France h CNRS, Géoressources, UMR 7359, Vandœuvre-lès-Nancy, F-54506, France b c

H I G H L I G H T S • • • •

Fresh OM input in an industrial soil leads to aggregation. TC and δ13C increase in fine silts. Fine silts store both the natural and anthropogenic OM. PAH concentration and availability are not impacted by an addition of OM.

a r t i c l e

i n f o

Article history: Received 4 June 2014 Received in revised form 22 July 2014 Accepted 2 August 2014 Available online 17 August 2014 Editor: Adrian Covaci Keywords: PAH availability Anthropic OM Maize addition Soil structure 13 C natural abundance

a b s t r a c t The impacts of fresh organic matter (OM) incorporation in an industrial PAH-contaminated soil on its structure and contaminant concentrations (available and total) were monitored. A control soil and a soil amended with the equivalent of 10 years maize residue input were incubated in laboratory-controlled conditions over 15 months. The structure of the amended soil showed an aggregation process trend which is attributable to (i) the enhanced microbial activity resulting from fresh OM input itself and (ii) the fresh OM and its degradation products. Initially the added organic matter was evenly distributed among all granulodensimetric fractions, and then rapidly degraded in the sand fraction, while stabilizing and accumulating in the silts. PAH degradation remained slight, despite the enhanced microbial biomass activity, which was similar to kinetics of the turnover rate of OM in an uncontaminated soil. The silts stabilized the anthropogenic OM and associated PAH. The addition of fresh OM tended to contribute to this stabilization process. Thus, in a context of plant growth on this soil two opposing processes might occur: rhizodegradation of the available contaminant and enhanced stabilization of the less available fraction due to carbon input. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Soils contaminated by organic contaminants, especially polycyclic aromatic hydrocarbons (PAH), are a serious environmental and health concern. Such soils are found on areas associated with former coal exploitation activities such as coking plants. Several remediation scenarios can be applied depending on the contamination level and nature, the ⁎ Corresponding author at: Laboratoire Sols et Environnement, Université de Lorraine/ INRA, 2 avenue de la Forêt de Haye, TSA 40602, 54518 Vandœuvre-lès-Nancy, France. Tel.: +33 383 595 762; fax: +33 383 595 791. E-mail address: [email protected] (S. Ouvrard).

http://dx.doi.org/10.1016/j.scitotenv.2014.08.004 0048-9697/© 2014 Elsevier B.V. All rights reserved.

size of the site and on financial and societal pressures. Most frequently used remediation techniques are thermal desorption, in situ chemical oxidation (ISCO) or bioremediation, including phytoremediation (Gan et al., 2009). Despite this large array of efficient remediation techniques, none has proved to be without drawback or to be applicable in all situations. Chemical- or thermal-based techniques have harmful consequences on the soil properties and seriously impair its reuse (Colombano et al., 2010; Gan et al., 2009; Khan et al., 2004). Biological treatments tend to be privileged for moderate to low contamination levels of a large surface area. These are more respectful of the environment and may be applied on site and even in situ. Phytoremediation, in providing a vegetation cover, additionally offers an esthetic benefit

346

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

and a better social acceptability of the treatment (Ouvrard et al., 2014). In some cases the biomass produced may even generate a profit, thereby contributing to the treatment costs. However this technique is only applicable when the soil conditions allow vegetation settlement, i.e. for low to moderate pollution levels, and for plant-accessible contamination, i.e. located in the upper soil horizon. Moreover, the low availability of contaminants tends to restrict the efficiency of this treatment (Reid et al., 2000; Usman et al., 2012). A better understanding of the processes controlling contaminants availability is the key to better phytoremediation efficiency, whether for degradation or stabilization purposes (Ouvrard et al., 2014). Phytoremediation, more precisely called rhizodegradation in the case of PAH, uses the increased microbial activity of the rhizosphere to accelerate degradation (Joner and Leyval, 2003). Moreover, root exudates are also thought to favor PAH desorption due to their complexing properties and capacity to enable co-metabolism (Subramaniam et al., 2004; Volkering et al., 1995). These processes occur throughout the plant's life cycle. However, when several successive cultivation cycles are considered, an additional phenomenon of plant tissue and litter input into the soil, which results in increased soil carbon content, should also be considered. Sorption on the organic matter (OM) is the major process governing PAHs behavior in a soil and thereby determines their availability (Ockenden et al., 2003). The nature and the composition of OM is a critical parameter: fulvic and humic acids do not have the same affinity for PAH and sorption generally increases with aromaticity (Doick et al., 2005; Pan et al., 2007). Black carbon, a condensed form of organic matter, has also been identified as a high-affinity adsorbent for PAH in soils (Agarwal and Bucheli, 2011; Cornelissen et al., 2005; Jeong et al., 2008). In the context of phytoremediation, the specific effect of the organic matter input due to litter deposition has not yet been clearly quantified in a real and unfavorable low PAH-availability industrial soil. The objective of this study was therefore to assess the effect of a natural C input on PAH degradation and availability. Moreover, the incorporation of this fresh carbon, as opposed to the “old” fossil carbon of the industrial soil, was also monitored. The methodology combined both agronomic and organic geochemistry tools and aimed at tracking carbon, both pollutant and of natural origin, in the various soil fractions, as is commonly performed when studying the carbon cycle. An incubation experiment over 15 months was performed in controlled conditions with an industrial soil, containing almost exclusively anthropogenic carbon, mixed with a large quantity of maize residue. A water granulodensimetric fractionation protocol and organic geochemistry tools were used to respectively monitor particle size distribution and PAH concentrations and availability. The first was used to obtain water stable aggregates. The second consisted in a solvent extraction of the extractable OM (EOM) to quantify the total PAH as well as an extraction with Tenax® resin to quantify the available PAH. Maize incorporation in the different particle size classes was determined by 13C isotope analysis. 2. Materials and methods 2.1. Materials The material studied was collected from a former industrial site (Neuves-Maisons), in the Northeast of France. This site had been

impacted by the steel industry and coking activities, resulting in a multicontamination by trace elements and organic pollutants (Ouvrard et al., 2011; Usman et al., 2012). The fresh material was sieved at 2 mm and stored at 50% of its water holding capacity. It was characterized for its agronomic properties by the Laboratoire d'Analyse des Sols (Arras, France) according to standard normalized methods (Table 1). The 14C activity measurement (Poznan Radiocarbon Laboratory, Poland) gave a modern carbon percentage (pMC) of 5.87%. The maize was collected from a field at the ENSAIA (Ecole Nationale Supérieure d'Agronomie et des Industries Alimentaires) experimental farm “La Bouzule” in Champenoux (France) in September 2011 just before the harvest. Only the aerial parts (stems and leaves) were used. They were washed with distilled water, dried at 105 °C for 48 hours, crushed and sieved at 500 μm. Maize is a C4 plant and its organic matter has a distinctive δ13C value that can be distinguished from the bulk soil organic matter. 2.2. Incubation of maize in the soil The amended M10 treatment mix was sieved at 2 mm ten times over, to insure optimal homogenization between maize and soil. The control M0 treatment was prepared similarly without maize addition. For both treatments, incubations were conducted in glass bottles filled with 150 g of material at 50% water holding capacity (WHC) and placed in the dark at 20 °C. Once a week, the moisture level was monitored and adjusted when necessary and the incubators were gently stirred with a glass rod and aerated to insure optimal microorganism activity. The amount of maize added corresponded to a quantity ten times higher than the expected annual input in order to amplify the subsequent reactions. This quantity was calculated on the basis of estimated C levels returning to the soil after 1 year of cultivation, as established by Balesdent et al. (1998) and the C content of maize (44.5%). The density of our soil (Table 1) was compatible with the cultivated soil used for the Balesdent et al. (1998) study. Thus, for 100 g of moist soil (50% WHC), we added 1.8 g of maize dry matter corresponding to 0.77 g of carbon. The incubation was stopped on 6 pre-defined dates, after 0, 3, 6, 9, 12 and 15 months, by freezing the incubators (−20 °C). These dates were called respectively t0, t3, t6, t9, t12 and t15. For each treatment and sampling date, four replicates were prepared. At each sampling time, the material was separated into five aliquots. Two of them were used for global microbial biomass estimation using the fumigation–extraction method. Two were used for water granulodensimetric separation. The last was used for bulk soil characterization. Bulk soil and soil fractions were fully characterized with (i) δ13C measurements, (ii) solvent extractable organic matter (EOM) isolation and quantification and (iii) total and available PAH determination. 2.3. Microbial biomass The fumigation–extraction method was performed on each replicate, according to the method described by Vance et al. (1987). Briefly, two aliquots of 10 g moist soil sample were used, one for the fumigated and the other for the non-fumigated treatment. Each aliquot was cleaned with 20 mL of 0.05 M K2SO4. The non-fumigated aliquot was stored at 4 °C. The other was placed in a desiccator with a beaker of

Table 1 Main properties of the industrial material. pH 7.67

Total CaCO3 (g kg−1) 19.4

Corg (g kg−1) 66

Olsen P2O5 (mg kg−1) 116

CEC (cmol kg−1) 12.4

Water holding capacity (%) 23.4

Coarse sands (g kg−1) 570

Fine sands (g kg−1) 80

Coarse silts (g kg−1) 80

Total nitrogen (g kg−1) 2.9

C/N (−) 22.8 Bulk density (−) 1.22

Fine silts (g kg−1) 150

Clays (g kg−1) 120

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

chloroform to lyse the soil microorganism cells. The desiccator was placed at 25 °C in the dark for 24 hours. Afterwards, both aliquots were extracted with 40 mL of 0.5 M K2SO4. The dissolved organic carbon was measured in the K2SO4 extract with a TOC-VSCN total organic carbon analyzer (Shimadzu). In parallel, the soil moisture was determined by drying 2 g at 105 °C for 48 hours. The soil microbial biomass carbon (MB) expressed in mg C kg−1 was calculated according to the following equation: MB ¼ ðOCfum – OCnon‐fum Þ=kEC

ð1Þ

where OCfum is the organic C extracted from the fumigated treatment (mg kg−1) and OCnon-fum the C extracted from the non-fumigated treatment (mg kg−1) and kEC is a constant (0.45) corresponding to the extractable part of the microbial biomass C after fumigation (Joergensen, 1996). 2.4. Water granulodensimetric fractionation A water granulodensimetric fractionation was applied to obtain five organo-mineral fractions (neither carbonate nor organic matter destruction was carried out as for the evaluation of particle-size distribution). These fractions consisted of mineral particles, organic matter and organo-mineral aggregates. The five fraction sizes corresponded to the granulometry of coarse sand (200–2000 μm), fine sand (50–200 μm), coarse silt (20–50 μm), fine silt (2–20 μm) and clays (0–2 μm). In our material, the clay quantity was too small (b0.1%) to be recovered and was not further characterized in this study. The protocol used was adapted from Watteau et al. (2006). Briefly, 30 g of the material and 200 mL of distilled water were placed in a glass bottle. The suspension was shaken at 20 °C on a rotating shaker at 15 rpm for 1 hour. In the first step, coarse and fine sands were separated by wet sieving and freeze-dried for 48 hours. In the second step, the remaining fraction suspension was shaken for 16 hours then placed in a sedimentation column where the suspension volume was adjusted to 1 Liter. The fraction distribution was quantified by sequential sedimentation and 25 mL aliquots sampled at pre-defined time intervals following Stokes law. Samples were dried at 70 °C and then weighed (NF X31107). In order to recover enough material for further characterization, the fraction collection was performed by siphoning off the complete suspension volume after pre-defined sedimentation periods as defined by Stokes law. Excess water was removed by centrifugation and fractions were freeze-dried. The 8 analytical replicates were pooled and stored at − 20 °C in glass bottles. For further analysis, except for PAH availability measurements, the coarser organo-mineral fractions (i.e., coarse and fine sands) were crushed and sieved to 500 μm. 2.5. Chemical characterization 2.5.1. Isotopic analysis The δ13C composition and the total carbon were obtained for the M0 and M10 bulk materials and for the granulodensimetric fractions using an elemental analyzer (Carlo Erba, NA 1500- NC, Milano, Italy), coupled to an isotope-ratio mass spectrometer (Finnigan, Delta-S, Bremen, Germany) with a precision of 0.2‰, at the Technical Platform of Functional Ecology at INRA Forest Ecology and Ecophysiology Unit, INRA Nancy, France. Measurements were carried out in triplicate for each sample. Approximately 5 mg of accurately weighed finely grounded material (b500 μm), supplemented with approximately 2.5 mg of accurately weighed V2O5, to promote combustion characteristics (Bol et al., 2005). The isotopic signatures were interpreted with a mixing model with two end-members, allowing the proportion of each source to be distinguished, according to the following equation: δmix ¼ f A δA þ ð1 – f A ÞδB

ð2Þ

347

where δmix is the δ value in the mixed soil, δA and δB are the isotopic values of sources A and B and fA is the fraction of the total contributed by source A (Dawson et al., 2002). In our case, δmix corresponded to the value of the M10 treatment and the two carbon sources were the fresh carbon derived from the maize and the old carbon already present in the M0 treatment. Thus, after equation rearrangement, we calculated fmaize, the contribution of the “maize” source with the following equation: f maize ¼ ðδM10 – δM0 Þ=ðδmaize – δM0 Þ

ð3Þ

with δM10 and δM0 the isotopic values of the two soils of our experiment. 2.5.2. Solvent extractable organic matter Solvent extractions were performed with an automated solvent extractor Dionex® ASE 350. Activated copper powder (1 g) and Na2SO4 (1 g) were added to 1 g of soil in the extraction cells prior to the extractions, in order to remove the molecular sulfur and the residual water respectively. Extractions were performed twice at 100 °C and 100 bars with HPLC grade dichloromethane (DCM), using a static time of 5 minutes. Organic extracts were diluted with DCM to obtain 20 mL. A 3 mL aliquot was sampled and left to dry in a fume hood for 36 hours, in order to quantify the extractable organic matter (EOM) mass (Biache et al., 2008). The remaining extract volume was stored in a glass vial at 4 °C before further analysis. 2.5.3. Molecular analysis The EOM was analyzed at molecular scale using a gas chromatography–mass spectrometer (GC–MS), with a GC-2010 plus (Shimadzu) instrument equipped with a DB 5-MS column (60*0.25 mm), coupled to a QP2012 Ultra (Shimadzu) mass spectrometer, operated in full scan mode for each fraction. The GC oven temperature was programmed from 70 °C (held 2 min) to 130 °C at 15 °C min− 1, then from 130 °C to 315 °C (held 2 min) at 4 °C min− 1. For 16 PAH (USEPA list) quantification, an internal PAH standard mix (naphthalened8, acenaphtene-d10, phenanthrene-d10, chrysene-d12 and perylened12) was added to the EOM. 2.6. PAH availability PAH availability was estimated by the Tenax® resin extraction method, using a protocol defined by (Barnier et al., 2014). Two grams of freeze-dried, non-crushed material were shaken with 2 g of Tenax and 300 mL of a 0.01 M CaCl2 and 200 mg L−1 NaN3 solution for 30 h. The Tenax grains were removed, cleaned with distilled water and airdried. PAH recovery was then performed by sonication with 10 mL of 1:1 mix of acetone hexane. This was repeated twice. Both extracts were pooled, evaporated under nitrogen and finally brought to 5 mL with the same solvent mix. The extracts were analyzed by GC–MS in full scan and SIM (single ion monitoring) modes, using the same analytical procedure as for total 16 PAH. 2.7. Statistical analysis Statistical analyses were performed using R software (R Development Core Team, 2011). Microbial biomass data were processed using a twoway ANOVA followed by multiple comparisons among means, using Tukey's (HSD) test. Differences among treatments as a function of time for water stable aggregate distributions were tested by the nonparametric Kruskal–Wallis and post-Hoc tests. All the statistical analyses were performed on treatments grouped together. Significant differences (p b 0.05) between data are indicated by different letters.

348

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

3. Results 3.1. Soil microbial biomass C estimation No algal or fungal development was observed for either treatment, whether inside the incubators or on their cover. The evolution curve of the microbial biomass carbon (MB) of M0 represented that of the native microbial biomass of the soil (Fig. 1). At t0, the M10 treatment presented a microbial biomass twice as high as that of the M0 treatment, indicating that a significant number of microorganisms already present in the maize was added with the amendment (Potthoff et al., 2005). After 3 months of incubation (t3) and for both treatments, the microbial biomass C strongly increased. For the M0 treatment, this rose by 71 mg kg− 1 (from 119 to 190 mg kg− 1) and for the M10 treatment, by 121 mg kg−1 (from 216 to 337 mg kg−1). These increases were induced by the optimal incubation conditions for the microorganisms, i.e. incubation temperature (20 °C), moisture at 50% of water holding capacity and soil homogenization. In addition, for the M10 treatment, the microbial activity was stimulated by the large input of fresh OM, which is consistent with the priming effect commonly observed in incubation experiments (Jenkinson, 1966; Kuzyakov and Gavrichkova, 2010). For both treatments, the microbial biomass C decreased after t3 and remained stable at a level close to the initial value, throughout the experiment after 6 months (t6). 3.2. Granulodensimetric fractionation The granulodensimetric distribution of our systems did not change significantly, either with time or maize addition (Table 2). Only slight tendencies could be identified. The initial granulodensimetric fractionation (t0) showed a surprising, though statistically significant, difference between the two treatments for the proportions of the sands (coarse and fine). The value for the coarse sands was higher for M0 than for the M10 treatment and conversely, the proportion of the fine sands of the M0 treatment was lower than that of the M10 treatment. These differences at the beginning of the experiment, due to the heterogeneity of the samples, did not allow a statistically significant evolution to be established between treatments, solely due to the maize input. If considered individually, the time series of the M0 and M10 treatments displayed opposing trends, namely aggregation for M10 and disaggregation for M0. A statistically significant increase in the coarse sand proportion was observed on M10 at t3 that was preserved

for the rest of the experiment. It was compensated for by a decrease in the fine sand proportion, the global sand fraction remaining globally constant over the incubation period. For M0, the fine sand proportion increased between t3 and t6 and remained stable there onward. Data for coarse sand displayed an opposite trend, although not statistically significant in this case. Here again, the overall sand fraction (coarse and fine) did not change significantly with time. The coarse silt fraction had a tendency to decrease for both treatments, though it was not statistically significant. For fine silts, no consistent evolution could be seen. 3.3. Carbon distribution and evolution For the M0 treatment, the total carbon content did not change significantly with time neither in the bulk nor within each class fraction (Table 3). The bulk carbon content remained steady at around a mean value of 68 ± 3 mg g−1. The addition of all class fraction contributions, referred to as “M0 fractions” and “M10 fractions,” gave a slightly higher mean value of 76 ± 5 mg g−1 (Fig. 2a). The discrepancy between these two values did not seem excessive when considering the number of data measured and their respective errors used for the mass balance calculation. Carbon was unevenly distributed between fractions with a clear enrichment of the fine sand and fine silt fractions with average values of around 125 and 93 mg g− 1 respectively (Table 3). Coarse silts and sands displayed a similarly slight depletion compared to the bulk, with average values of 60 mg g−1. For the M10 treatment, as expected from the maize addition, the initial carbon content was clearly higher than for M0. It rapidly decreased and reached the M0 value after 3 months. A decreasing trend, probably statistically not significant, was further observed until the end of the experiment. The initial maize-C input preferentially enriched the fine sands and, to a lesser extent, coarse sands fractions, whose carbon contents increased respectively to 146 and 66 mg g−1. After three months of incubation, both class fractions presented a total carbon content similar to the control (M0). Coarse silts did not display any significant evolution for total carbon. Total carbon of fine silts, initially unaffected by maize addition, increased at 3 months and remained steady until the end of the incubation period. Exploitation of the δ13C measurements enabled the maize-origin carbon content also called fresh carbon to be monitored, allowing repartition between fractions with time (Table 3). This evolution confirmed the global trends observed on total carbon. Although initially distributed in all fractions, maize was mostly initially found in the sand fractions (80%) (Fig. 3b). Over the 15 months of incubation, half of this fresh carbon disappeared from our system, most of it during the first three months (Fig. 2b). The remaining carbon was mostly found in the fine silt fraction and accumulated there. This fraction held more than 50% of this remaining fresh carbon at 15 months. In contrast, it almost disappeared from the coarse sand fraction. 3.4. Pollution distribution and evolution

Fig. 1. Evolution of microbial biomass C during the 15 months of incubation in the M0 and M10 modes (data with same letter are not statistically different, p b 0.05).

3.4.1. Extractable organic matter (EOM) At the beginning of the experiment, the EOM content was similar for both treatments with bulk values at around 10 mg g−1 (Table 4). The EOM content of maize was 25.5 mg g− 1, corresponding to a nonsignificant additional input of 0.45 mg of EOM for 1 g of soil for M10 treatment. EOM represented around 13% of the total organic carbon. This was 30 times higher than that commonly found in unpolluted soil (Kélomé et al., 2012), but in the same value range as other coking plant soils (Biache et al., 2008; Usman et al., 2012). The organic matter initially present in the material was mostly of industrial origin (i.e., resulting from fossil material). Indeed modern carbon was estimated at 5.9%, thereby justifying the term “fresh carbon” when referring to the maize addition (Pernot et al., 2013). For all fractions, EOM contents

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

349

Table 2 Evolution of the water stable granulodensimetric distribution (wt.%) for the M0 and M10 treatments over time (mean ± standard deviation, n = 4). Material fraction

Treatment

t0

200–2000 μm (coarse sands)

M0 M10 M0 M10 M0 M10 M0 M10

61.9 59.2 16.1 18.7 8.4 9.1 13.5 13.0

50–200 μm (fine sands) 20–50 μm (coarse silts) 2-20 μm (fine silts)

t3 ± ± ± ± ± ± ± ±

0.4 0.8 0.2 0.7 0.0 0.5 0.3 0.7

t6

59.9 61.6 17.3 17.1 8.0 7.2 14.9 14.1

± ± ± ± ± ± ± ±

0.9 1.1 0.6 0.8 0.2 0.6 0.2 0.2

were in the same range except for the fine silts, which presented a significant enrichment with a 3 times higher EOM concentration. For the bulk, the EOM content of both treatments evolved in the same way with a progressive and slight decrease of 14.3% and 13.0% for M0 and M10 respectively (Fig. 4a). For the granulodensimetric fractions, a similar decreasing trend was observed for both treatments. It did not appear very significant, given the small value of EOM content compared to analytical error, except in the case of the fine silts (Fig. 4b).

t9

60.8 61.3 16.9 17.0 7.8 8.0 14.5 13.7

± ± ± ± ± ± ± ±

0.5 0.6 0.2 0.2 0.2 0.7 0.1 0.3

t12

60.0 61.6 17.2 17.0 6.9 7.6 15.9 13.8

± ± ± ± ± ± ± ±

0.6 0.5 0.3 0.6 0.7 0.2 0.5 0.1

59.4 62.3 17.2 16.7 7.7 7.7 15.6 13.3

t15 ± ± ± ± ± ± ± ±

0.9 1.3 0.7 0.9 0.4 0.2 0.3 0.5

60.6 62.7 17.4 16.7 7.1 6.8 14.9 13.7

± ± ± ± ± ± ± ±

0.9 1.0 0.4 1.0 0.5 0.7 0.3 0.8

stable aggregate distribution, an indicator of soil structure, and (ii) C distribution and storage among these fractions. Regarding soil structure, fresh organic matter addition to our contaminated soil led to an increase in coarse particles concomitant with a decrease in fine particles. The particles of maize residues sieved at 500 μm could not be held responsible for this dual phenomenon, which would rather show evidence of an aggregation process. This is not new and has been already observed on non-polluted soils (Bravo-Garza et al., 2009; Denef et al., 2002). The increase in microbial activity due to fresh substrate addition can lead to biofilm production and exopolymer secretion. These compounds contribute to organo-mineral associations (Qurashi and Sabri, 2012). Additionally, organic residues produced with the added natural organic matter can also play a role of cement between particles in the soil (Bronick and Lal, 2005; Six et al., 2004). The evolution of the added fresh organic matter displayed a traditional pattern, in agreement with other studies carried out on natural soils (Angers et al., 1997; Coppens et al., 2006; Puget et al., 1995), with a progressive decrease in the sand fractions and a stabilization/ accumulation in silt fractions. The initial TC increase in coarse fractions was due to the fresh added-C preferential incorporation, which then decreased with the degradation processes taking place. Fine fractions acted as preservation sites, accumulating degradation residue produced from larger particles and protecting them from further degradation processes. This resulted in differentiated residence time for organic matter with particle size in agreement with data given by Puget et al. (2000). Organic carbon had a lifetime of a few years in the coarse aggregates (N500 μm) and in the order of 100 years in fine aggregates (b500 μm). The preservation process in the fine fraction illustrated here for fresh organic matter has already been observed for the anthropogenic organic matter of this material (Pernot et al., 2013). Indeed, a preservation process occurred in the fine silt fraction compared to the other fractions with the accumulation of PAH. The combination of these two results, the preservation of pollution and the fresh C storage in fine silts, clearly demonstrate that the fine silt fraction acts as a protective sink for organic matter, whether its origin be natural or anthropogenic.

3.4.2. PAH repartition and availability Initially, the PAH concentration was close to 1100 μg g−1 for both bulk treatments, which represented 12% of the EOM content (Table 4). These concentrations were similar to those obtained by previous studies of the same material (Biache et al., 2008; Laurent et al., 2012; Usman et al., 2012). No significant evolution was observed with incubation time on any treatment or granulodensimetric fraction. The bulk soil displayed an overall degradation of 9.3% and 6.6% for M0 and M10 respectively. These variations fell within standard error and were not otherwise systematically confirmed by data on individual fractions. Detailed analysis of the evolution of PAH molecule taken separately did not show any additional specific trend. Given the limited reactivity of the system, any potentially differentiated behavior was masked by the variability of the heterogeneity of the material and fell within standard error. The PAH availability of both bulk treatments was very low (1.5% of the total PAH) and remained stable during the experiment (Table 4). In the different granulodensimetric fractions, PAH availability in both treatments tended to decrease with time. However the measured variation remained within standard variation range and might not have been statistically significant if analysis had been performed on separate experimental replicates. 4. Discussion 4.1. Fresh organic matter incorporation

4.2. Pollution status Fresh organic matter incorporation, simulating litter deposit, has a direct immediate impact on C content, but might also induce soil structure modifications. These two factors were analyzed through (i) water

The pollution status was estimated on the basis of the total carbon (TC) content, the solvent extractable organic matter (EOM) and the

Table 3 Evolution of the total carbon (mg g−1) in the bulk and granulodensimetric fractions (data are presented as mean value of three analytical replicates ± standard deviation). Material fraction

Treatment

t0

200–2000 μm (coarse sands)

M0 M10 M0 M10 M0 M10 M0 M10 M0 M10

52 66 122 146 59 60 92 90 66 92

50–200 μm (fine sands) 20–50 μm (coarse silts) 2–20 μm (fine silts) Bulk

t3 ± ± ± ± ± ± ± ± ± ±

4 2 2 2 4 4 1 1 3 16

64 59 117 115 64 65 93 101 67 76

t6 ± ± ± ± ± ± ± ± ± ±

4 4 11 4 11 3 1 1 2 3

60 64 123 130 55 60 93 103 74 72

t9 ± ± ± ± ± ± ± ± ± ±

1 6 6 3 5 7 1 2 2 2

62 59 129 130 56 74 93 101 68 69

t12 ± ± ± ± ± ± ± ± ± ±

3 4 4 5 3 5 2 1 2 3

67 53 117 123 63 78 94 103 67 66

t15 ± ± ± ± ± ± ± ± ± ±

12 3 2 5 1 2 2 1 3 1

53 55 141 117 62 54 95 89 69 53

± ± ± ± ± ± ± ± ± ±

2 3 8 4 9 5 1 1 5 6

350

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

a

b

Fig. 2. Evolution with time of a) total carbon and b) maize-derived carbon. Each graph presents data for bulk soil analysis (“bulk,” mean value of three analytical replicates ± standard deviation) and calculated value resulting from average data obtained from individual fraction analyses (“fractions”).

total and available PAH concentrations. Although not strictly speaking pollution, TC and EOM were in this case good indicators. The first because of the small contribution of recent carbon to total carbon (only 5.9%) and the second because of the nature of the solvent used and the strong positive correlation between this data and organic pollution. The total carbon remained stable in the control, M0, whereas it slightly decreased in the maize-amended treatment, M10. Indeed, maize was clearly degraded and the amount of degraded maize alone could explain the TC evolution, on this scale. The low TC value of the M10 bulk at t15 might not be significant given the discrepancy with the corresponding fraction mass-balance value. Indeed, the global trend occurred mainly in the coarse fractions, whereas fine fractions displayed no degradation pattern with only transitory TC accumulation. For EOM, a degradation trend was observed on both bulk soil treatments. The decrease over the 15 months was around 13–14%, which was above the analytical error on EOM (4%) and greater than the error between direct and fraction mass-balance estimates of the EOM bulk content (7%). Since EOM represented only 13% of TC, this variation would not be seen on the TC value. Therefore, for the M10 treatment, a concomitant degradation of fresh and anthropogenic (“old”) carbon occurred. This was further confirmed by the evolution of fresh carbon contribution to TC in the bulk that passed from 8.9% to 7.4%. The decrease was indeed much smaller than would be expected from the maize-carbon degradation alone. PAH degradation, 9% at most, was in the same order as observations conducted in situ on the same site. We showed that 5% of PAH was annually degraded with or without plant growth at a mean temperature of 10 °C (extreme values: −21.6 °C to 37.6 °C) (Ouvrard et al., 2011).

Such values are compatible with the decomposition rate of natural organic matter in temperate ecosystems or even higher (Schmidt et al., 2011) (annual decomposition rate: 1 to 5%). One may have expected a slightly higher mineralization of PAH in the favorable incubation conditions, in comparison with the in situ observations. The microbial communities of our polluted soil are known for their PAH degrading ability (Cébron et al., 2009; Thion et al., 2012). However it is likely they also fed on fresh maize input, which slowed down the PAH degradation (Bamforth and Singleton, 2005; Peng et al., 2008). The preferential degradation of organic residue by the microorganisms is consistent with the results of Hultgren et al. (2009). These authors amended a creosote-contaminated soil with wheat residue (5 g for 100 g of soil) and they observed an increase in the microbial biomass activity and an absence of PAH degradation. Moreover, the availability of PAH, already small in this system, had a tendency to decrease. Nevertheless, this global tendency of the bulk hid differentiated behavior within class fractions. If coarse fractions behaved as the bulk, the fine fractions and especially the fine silts displayed an accumulation behavior towards added maize-carbon. The silts generally preserved the organic matter of this soil, both from natural and anthropogenic origin. This suggests that the silts stabilized the anthropogenic OM. The freshly-added OM tended to accelerate this process. In such a stable system where PAH availability is very limited, the key to increased degradation is an increase in availability. The addition of plant biomass succeeded in enhancing microbial activity. However, the degradation mostly occurred on the easily accessible and freshly added organic matter. It did not significantly affect pollution. Moreover, this organic matter acted as an additional binding agent for contaminants and between

Fig. 3. Evolution with time of the contribution of the different fractions to a) total carbon and b) maize-derived carbon.

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

351

Table 4 Evolution of the EOM content (mg g−1), the total PAH concentration (μg g−1) and the available PAH (%) for the M0 and M10 treatments over time (data are presented as mean value of three analytical replicates ± standard deviation when available).a Material fraction

Parameter

Treatment

t0

200–2000 μm (coarse sands)

EOM

M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10 M0 M10

7.4 ± 6.6 ± 972 849 1.2 ± 1.4 ± 8.9 ± 8.8 ± 985 1002 1.6 ± 1.5 ± 7.8 ± 6.9 ± 858 720 1.5 ± 1.8 ± 29.3 ± 34.1 ± 2645 2268 1.3 ± 1.2 ± 10.4 ± 10.3 ± 1163 ± 1144 ± 1.5 ± 1.2 ±

Total PAH Available PAH 50–200 μm (fine sands)

EOM Total PAH Available PAH

20–50 μm (coarse silts)

EOM Total PAH Available PAH

2–20 μm (fine silts)

EOM Total PAH Available PAH

Bulk

EOM Total PAH Available PAH

t3 0.4 0.2

0.2 0.1 0.7 0.2

0.1 0.2 0.3 0.5

0.2 0.1 1.5 0.7

0.0 0.1 0.4 0.6 63 19 0.1 0.0

6.9 ± 6.2 ± 948 903 1.3 ± 1.2 ± 8.1 ± 8.2 ± 987 952 1.8 ± 1.5 ± 6.5 ± 6.8 ± 724 771 1.6 ± 1.8 ± 28.1 ± 26.6 ± 2515 2324 1.2 ± 1.2 ± 9.6 ± 9.3 ± 1033 1142 1.4 ± 1.4 ±

t9 0.1 0.7

0.1 0.1 0.1 0.1

0.2 0.1 0.5 0.4

0.2 0.1 0.1 0.4

0.1 0.1 0.4 0.2

0.2 0.0

6.7 ± 7.0 ± 829 824 1.3 ± 1.6 ± 8.7 ± 9.2 ± 856 943 1.6 ± 1.6 ± 6.7 ± 7.8 ± 674 794 1.7 ± 1.4 ± 27.5 ± 30.5 ± 2707 2593 1.1 ± 0.8 ± 9.3 ± 10.0 ± 1056 1038 1.4 ± 1.4 ±

t12 0.5 0.3

0.2 0.1 0.2 0.1

0.1 0.1 0.6 0.4

0.1 0.0 0.1 0.5

0.1 0.0 0.3 0.5

0.2 0.1

6.3 ± 6.4 ± 837 773 1.2 ± 1.5 ± 8.1 ± 8.2 ± 1027 950 1.3 ± 1.7 ± 7.3 ± 6.5 ± 755 658 1.3 ± 1.4 ± 26.2 ± 27.4 ± 2849 2703 0.9 ± 0.8 ± 8.9 ± 9.1 ± 1046 ± 1070 ± 1.6 ± 1.3 ±

t15 0.2 0.4

0.1 0.2 0.2 0.2

0.1 0.2 0.1 0.2

0.1 0.0 0.2 0.2

0.0 0.0 0.3 0.3 83 80 0.4 0.1

6.4 ± 7.6 ± 824 923 1.0 ± 1.1 ± 8.7 ± 8.6 ± 1248 1130 1.3 ± 1.1 ± 5.7 ± 8.1 ± 714 791 1.4 ± 1.2 ± 26.3 ± 26.9 ± 2654 2521 0.9 ± 1.0 ± 9.0 ± 8.9 ± 1055 ± 1068 ± 1.3 ± 1.3 ±

0.7 0.3

0.4 0.1 1.0 0.9

0.0 0.1 1.1 0.9

0.1 0.0 1.1 0.2

0.0 0.0 0.6 0.1 4 42 0.1 0.1

a No systematic analytical replicates were done for total PAH due to lack of material. Tendencies were however discussed based on the expected variability usually obtained for this parameter on this material measured in previous works.

particles. PAH transfer to the aqueous phase is mostly dependent on sorption on organic matter. Here, the added OM and its residues accumulated in the fine fractions enhancing the already strong PAH retention. Thus, in a context of plant growth (e.g. phytoremediation or natural attenuation) on this type of soil, the litter turnover on successive cultivation cycles would mostly lead to additional carbon sequestration. This total carbon increase would provide additional sinks for the pollution. 5. Conclusion The incorporation of fresh maize in a soil highly contaminated by PAH did not result in any significant additional degradation of the

a

contamination. On the contrary, different processes tended to further limit PAH availability. The freshly added organic matter was preferentially degraded in the sand fraction and accumulated in the fine silts. This fraction protected PAH from further degradation. Moreover, the promoted microbial activity was mostly dedicated to the easily degradable carbon and not to pollution. These processes were similar to those observed on natural soil and should be seriously considered when phytoremediation is proposed as a remediation treatment over several years of plant cultivation. The rhizosphere activity, generally presented as beneficial for PAH degradation, might be partly counterbalanced over the years by availability limitation due to C input coming from litter decomposition. Over the long term, the initial PAH might follow two opposite paths: degradation of the available

b

Fig. 4. EOM evolution for a) bulk soil and b) granulodensimetric fractions. Graph a) presents data for bulk soil analysis (“bulk,” mean value of three analytical replicates ± standard deviation) and calculated value resulting from average data obtained on individual fraction analyses (“fractions”).

352

A. Pernot et al. / Science of the Total Environment 497–498 (2014) 345–352

fraction and increased retention of the remaining pool, which would prevent any total removal of the contamination, but guarantee its stabilization and limit the risk of dissemination. These phenomena should be further studied in situ and with real plant cultivation. Indeed, rooting systems and living plants have additional effects that might counterbalance the processes observed in this study. This is all the more so as root exudates are known to act as extracting agents, thereby locally increasing PAH availability. Acknowledgements This work was conducted within the framework of the GISFI program and was financially supported by ICEEL and CNRS EC2CO “DECORES” program. The authors wish to thank Adeline Bouchard, Romain Goudon, Jean-Claude Bégin and Stéphane Colin from LSE and Lucas Charrois from GISFI for their technical help. They are grateful to Emile Benizri from LSE for his help on extraction–fumigation analysis and to the technical staff of the functional ecology technical platform (OC 081) at the INRA Forest Ecology and Ecophysiology Unit for the δ13C analysis. References Agarwal T, Bucheli TD. Is black carbon a better predictor of polycyclic aromatic hydrocarbon distribution in soils than total organic carbon? Environ Pollut 2011;159:64–70. Angers DA, Recous S, Aita C. Fate of carbon and nitrogen in water-stable aggregates during decomposition of (CN)-C-13-N-15-labelled wheat straw in situ. Eur J Soil Sci 1997; 48:295–300. Balesdent J, Besnard E, Arrouays D, Chenu C. The dynamics of carbon in particle-size fractions of soil in a forest-cultivation sequence. Plant Soil 1998;201:49–57. Bamforth SM, Singleton I. Bioremediation of polycyclic aromatic hydrocarbons: current knowledge and future directions. J Chem Technol Biotechnol 2005;80:723–36. Barnier C, Ouvrard S, Robin C, Morel JL. Desorption kinetics of PAHs from aged industrial soils for availability assessment. Sci Total Environ 2014;470–471:639–45. Biache C, Mansuy-Huault L, Faure P, Munier-Lamy C, Leyval C. Effects of thermal desorption on the composition of two coking plant soils: impact on solvent extractable organic compounds and metal bioavailability. Environ Pollut 2008;156:671–7. Bol R, Eriksen J, Smith P, Garnett MH, Coleman K, Christensen BT. The natural abundance of 13C, 15N, 34S and 14C in archived (1923–2000) plant and soil samples from the Askov long-term experiments on animal manure and mineral fertilizer. Rapid Commun Mass Spectrom 2005;19:3216–26. Bravo-Garza MR, Bryan RB, Voroney P. Influence of wetting and drying cycles and maize residue addition on the formation of water stable aggregates in Vertisols. Geoderma 2009;151:150–6. Bronick CJ, Lal R. Soil structure and management: a review. Geoderma 2005;124:3–22. Cébron A, Beguiristain T, Faure P, Norini MP, Masfaraud JF, Leyval C. Influence of vegetation on the in situ bacterial community and polycyclic aromatic hydrocarbon (PAH) degraders in aged PAH-contaminated or thermal-desorption-treated soil. Appl Environ Microbiol 2009;75:6322–30. Colombano S, Saada A, Guerin V, Bataillard P, Bellenfant G, Beranger S, et al. Quelles techniques pour quels traitements — analyse coûts-bénéfices; 2010. Coppens F, Merckx R, Recous S. Impact of crop residue location on carbon and nitrogen distribution in soil and in water-stable aggregates. Eur J Soil Sci 2006;57:570–82. Cornelissen G, Gustafsson O, Bucheli TD, Jonker MTO, Koelmans AA, van Noort PCM. Extensive sorption of organic compounds to black carbon, coal, and kerogen in sediments and soils: mechanisms and consequences for distribution, bioaccumulation, and biodegradation. Environ Sci Technol 2005;39:6881–95. Dawson TE, Mambelli S, Plamboeck AH, Templer PH, Tu KP. Stable isotopes in plant ecology. Annu Rev Ecol Syst 2002;33:507–59. Denef K, Six J, Merckx R, Paustian K. Short-term effects of biological and physical forces on aggregate formation in soils with different clay mineralogy. Plant Soil 2002;246: 185–200. Doick KJ, Burauel P, Jones KC, Semple KT. Distribution of aged 14C-PCB and 14C-PAH residues in particle-size and humic fractions of an agricultural soil. Environ Sci Technol 2005;39:6575–83. Gan S, Lau EV, Ng HK. Remediation of soils contaminated with polycyclic aromatic hydrocarbons (PAHs). J Hazard Mater 2009;172:532–49.

Hultgren J, Pizzul L, Castillo MdP, Granhall U. Degradation of PAH in a creosotecontaminated soil. A comparison between the effects of willows (Salix viminalis), wheat straw and a nonionic surfactant. Int J Phytoremediation 2009;12:54–66. Jenkinson DS. Studied on the decomposition of plant material in soil II. Partial sterilization of soil and the soil biomass. J Soil Sci 1966:280–302. Jeong BS, Wander MM, Kleineidam S, Grathwohl P, Ligouis B, Werth CJ. The role of condensed carbonaceous materials on the sorption of hydrophobic organic contaminants in subsurface sediments. Environ Sci Technol 2008;42:1458–64. Joergensen. The fumigation–extraction method to estimate soil microbial biomass: calibration of the kEC value. Soil Biol Biochem 1996;28:25–31. Joner EJ, Leyval C. Rhizosphere gradients of polycyclic aromatic hydrocarbon (PAH) dissipation in two industrial soils and the impact of arbuscular mycorrhiza. Environ Sci Technol 2003;37:2371–5. Kélomé NC, Faure P, Mathieu O, Lévêque J, Oyédé LM. Evidence of organic contamination in urban soils of Cotonou town (Benin). Br J Environ Clim Chang 2012:99–112. Khan FI, Husain T, Hejazi R. An overview and analysis of site remediation technologies. J Environ Manag 2004;71:95–122. Kuzyakov Y, Gavrichkova O. REVIEW: time lag between photosynthesis and carbon dioxide efflux from soil: a review of mechanisms and controls. Glob Chang Biol 2010;16:3386–406. Laurent F, Cebron A, Schwartz C, Leyval C. Oxidation of a PAH polluted soil using modified Fenton reaction in unsaturated condition affects biological and physico-chemical properties. Chemosphere 2012;86:659–64. Ockenden WA, Breivik K, Meijer SN, Steinnes E, Sweetman AJ, Jones KC. The global re-cycling of persistent organic pollutants is strongly retarded by soils. Environ Pollut 2003;121:75–80. Ouvrard S, Barnier C, Bauda P, Beguiristain T, Biache C, Bonnard M, et al. In situ assessment of phytotechnologies for multicontaminated soil management. Int J Phytoremediation 2011;13:245–63. Ouvrard S, Leglize P, Morel JL. PAH phytoremediation: rhizodegradation or rhizoattenuation? Int J Phytoremediation 2014;16:46–61. Pan B, Xing B, Tao S, Liu W, Lin X, Xiao Y, et al. Effect of physical forms of soil organic matter on phenanthrene sorption. Chemosphere 2007;68:1262–9. Peng R-H, Xiong A-S, Xue Y, Fu X-Y, Gao F, Zhao W, et al. Microbial biodegradation of polyaromatic hydrocarbons. FEMS Microbiol Rev 2008;32:927–55. Pernot A, Ouvrard S, Leglize P, Faure P. Protective role of fine silts for PAH in a former industrial soil. Environ Pollut 2013;179:81–7. Potthoff M, Dyckmans J, Flessa H, Muhs A, Beese F, Joergensen RG. Dynamics of maize (Zea mays L.) leaf straw mineralization as affected by the presence of soil and the availability of nitrogen. Soil Biol Biochem 2005;37:1259–66. Puget P, Chenu C, Balesdent J. Total and young organic matter distributions in aggregates of silty cultivated soils. Eur J Soil Sci 1995;46:449–59. Puget P, Chenu C, Balesdent J. Dynamics of soil organic matter associated with particle‐ size fractions of water‐stable aggregates. Eur J Soil Sci 2000;51:595–605. Qurashi AW, Sabri AN. Bacterial exopolysaccharide and biofilm formation stimulate chickpea growth and soil aggregation under salt stress. Braz J Microbiol 2012;43: 1183–91. R Development Core Team. R: A Language and Environment for Statistical Computing. Vienna, Austria: Computing. R Foundation for Statistical Computing; 2011. Reid BJ, Jones KC, Semple KT. Bioavailability of persistent organic pollutants in soils and sediments — a perspective on mechanisms, consequences and assessment. Environ Pollut 2000;108:103–12. Schmidt MWI, Torn MS, Abiven S, Dittmar T, Guggenberger G, Janssens IA, et al. Persistence of soil organic matter as an ecosystem property. Nature 2011;478:49–56. Six J, Bossuyt H, Degryze S, Denef K. A history of research on the link between (micro)aggregates, soil biota, and soil organic matter dynamics. Soil Tillage Res 2004;79:7–31. Subramaniam K, Stepp C, Pignatello JJ, Smets B, Grasso D. Enhancement of polynuclear aromatic hydrocarbon desorption by complexing agents in weathered soil. Environ Eng Sci 2004;21:515–23. Thion C, Cébron A, Beguiristain T, Leyval C. Long-term in situ dynamics of the fungal communities in a multi-contaminated soil are mainly driven by plants. FEMS Microbiol Ecol 2012;82:169–81. Usman M, Faure P, Ruby C, Hanna K. Remediation of PAH-contaminated soils by magnetite catalyzed Fenton-like oxidation. Appl Catal B Environ 2012;117–118:10–7. Vance ED, Brookes PC, Jenkinson DS. An extraction method for measuring soil microbial biomass C. Soil Biol Biochem 1987;19:703–7. Volkering F, Breure AM, Vanandel JG, Rulkens WH. Influence of nonionic surfactants on bioavailability and biodegradation of polycyclic aromatic hydrocarbons. Appl Environ Microbiol 1995;61:1699–705. Watteau F, Villemin G, Burtin G, Jocteur-Monrozier L. Root impact on the stability and types of micro-aggregates in silty soil under maize. Eur J Soil Sci 2006;57:247–57.