Environmental Pollution 188 (2014) 124e131

Contents lists available at ScienceDirect

Environmental Pollution journal homepage: www.elsevier.com/locate/envpol

Impact of soil amendments and the plant rhizosphere on PAH behaviour in soil Geoffrey Marchal a, Kilian E.C. Smith b, Philipp Mayer c, Lis Wollesen de Jonge d, Ulrich G. Karlson e, * a

Center for Energy Resources Engineering, Technical University of Denmark, Søltofts Plads Building 229, 2800 Lyngby, Denmark Korean Institute of Science and Technology Europe, Campus E7.1, Universität des Saarlandes, 66123 Saarbrücken, Germany Department of Environmental Engineering, Technical University of Denmark, Miljøvej Building 113, 2800 Kgs. Lyngby, Denmark d Aarhus University, Department of Agroecology, Blichers Allé, Postbox 50, DK 8830 Tjele, Denmark e Department of Environmental Science, Aarhus University, Frederiksborgvej 399, 4000 Roskilde, Denmark b c

a r t i c l e i n f o

a b s t r a c t

Article history: Received 16 October 2013 Received in revised form 6 February 2014 Accepted 8 February 2014

Carbonaceous amendments reduce PAH dissolved concentrations (Cfree), limiting their uptake and toxicity. A soil contaminated with PAHs was mixed with activated carbon (AC), charcoal or compost and planted with radish (Raphanus sativus L.), and Cfree, chemical activities and diffusive uptake of the PAHs measured over 2 months. For AC, Cfree and diffusive uptake were decreased by up to 94% compared to the unamended soil within one week. In addition, the sum chemical activity of the PAHs remained below the threshold for baseline toxicity. In contrast, charcoal and compost only led to modest reductions in Cfree and diffusive uptake, with sum chemical activities that could potentially result in baseline toxicity being observed. Furthermore, both Cfree and diffusive uptake were lower in the planted compared to unplanted soils. Therefore, only AC successfully reduced PAH acute toxicity in the soil, but plant-promoted microbial degradation may also play an important role in PAH attenuation. Ó 2014 Elsevier Ltd. All rights reserved.

Keywords: Radish (Raphanus sativus L.) Freely dissolved concentration Chemical activity Diffusive uptake Passive sampling

1. Introduction Soils become contaminated with organic contaminants such as PAHs via industrial activities, pollution events or atmospheric deposition (Haritash and Kaushik, 2009). In the soil, PAHs partition to the soil matrix (Semple et al., 2003), such that soils are one of the main repositories of PAHs and other similar contaminants (Alexander, 2000). This raises concern about the negative effect of PAHs on ecosystem health (White and Claxton, 2004), and in agricultural soils the risk of contaminants entering the food chain (Agarwal et al., 2009). Therefore, there is interest in understanding PAH fate in soil, and assessing that portion that is available for biotic uptake (GomezEyles et al., 2011). Soil treatment methods for reducing this available fraction are being proposed (Megharaj et al., 2011), including the in-situ application of carbonaceous materials. Here, strong sorption of the soil contaminants to these amendments reduces their uptake and potential for ecotoxicological effects (Kookana,

* Corresponding author. E-mail addresses: [email protected], [email protected] (U.G. Karlson). http://dx.doi.org/10.1016/j.envpol.2014.02.008 0269-7491/Ó 2014 Elsevier Ltd. All rights reserved.

2010). Several studies have investigated the influence of different types of carbonaceous soil amendments on PAH bioavailability in soil systems without plants (Beesley et al., 2010; Chen and Yuan, 2011; Jakob et al., 2012). These show that adding amendments such as activated carbon (AC) or biochar (charcoal formed from biomass pyrolysis) dramatically reduce PAH accumulation in organisms, and therefore reduce the risks (Jakob et al., 2012). Less information is available about their effectiveness in “living” soils with an active rhizosphere. This is critical since planting soils brings additional benefits including improving the structure and stabilizing against erosion which is a problem for contaminated sites (Zheng et al., 2012). Furthermore, planting contaminated soils has itself been suggested as a strategy for reducing soil contamination (de-Bashan et al., 2012), under the premise that exudates lead to an increased rhizosphere microbial community (Cheema et al., 2009). For instance, the degradation of PAHs was found to increase in soils with an active rhizosphere (Fu et al., 2012). Therefore, the simultaneous application of carbonaceous amendments and planting of the soil could form part of a holistic approach for remediating contaminated sites. The most positive scenario is that the active microbial community due to the plant rhizosphere additionally reduces bioavailable PAH concentrations

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

further to that resulting from sorption to the amendment. Alternatively, increased secretion of organic matter by the plant rhizosphere might lead to competitive sorption effects or blockages of the micro-pores of the carbonaceous amendment (Pignatello et al., 2006), thus decreasing their sorption efficiency. This study aimed to determine in a planted soil the effectiveness of AC, charcoal and compost in reducing the availability of PAHs compared to an unamended soil. The integral role of contaminant bioavailability in environmental risk is accepted, and recently some authorities have started to adopt a risk-based approach for assessing soil, sediment and groundwater pollution that includes considerations of compound bioavailability (Fujinaga et al., 2012). Therefore, in this study temporal changes in dissolved concentrations, chemical activities and diffusive uptake into a silicone sink of phenanthridine, phenanthrene, anthracene, fluoranthene, and pyrene were measured in spiked soils planted with radish (Raphanus sativus L.) and amended with 1% w/w of soil amendment. 2. Materials and methods 2.1. Chemicals and solvents All chemicals were of analytical grade or better, and all media and aqueous solutions prepared using Milli-Q water (Super Q treated, Millipore, MA, USA). Ethylacetate (p.a. grade) and methanol (HPLC grade) were obtained from Merck (Darmstadt, Germany), and acetone (HPLC grade) from Rathburn (Walkerburn, UK). Glass scintillation vials (Mikrolab Aarhus A/S, Aarhus, Denmark) and 120 mL brown pill glasses (Apodan Nordic A/S, Copenhagen, Denmark) were used as supplied. Whatman n
41 filter paper was used for germinating the radish seeds (Whatman International Ltd, England). Phenanthridine (99%, Aldrich, St Louis, USA), phenanthrene (96%, Aldrich), anthracene (97%, Steinheim, Germany), fluoranthene (99%, Aldrich) and pyrene (99%, Aldrich) spiking solutions were prepared in acetone. Their physicochemical properties are given in Table S1. Polydimethylsiloxane (PDMS)-coated SPME fibres were used to measure the freely dissolved concentrations of PAHs in the soil. SPME fibres with a glass core diameter of 200 mm and a PDMS coating thickness of 20 mm were supplied by Polymicro Technologies Inc. (Phoenix, USA). The volume of the PDMS coating was calculated to be 13.55 mL per metre of fibre. These were cut into 80 mm pieces, cleaned by soaking three times for 24 h in methanol, and rinsed three times for 24 h in Milli-Q water. Food-grade silicone O-rings with an outer diameter of 12 mm, inner diameter of 9.6 mm, mass of 231 mg (C.V. 1%, n ¼ 10) and volume of 0.171 mL (Altec, Cornwall, United Kingdom) were used to estimate the diffusive uptake of PAHs from the soils. The O-rings were cleaned as described in Smith et al. (2010). AC (untreated powder, 100e400 mesh. SigmaeAldrich, St Louis, USA) had a particle size of less than 75 mm, and was prepared from wood charcoal and then chemically activated. A large surface area of between 1000 and 1500 m2 per gram has been reported for AC (Foo and Hameed, 2012; Rivera-Utrilla and Sánchez-Polo, 2011). Charcoal wood powder (Merck, Darmstadt, Germany) was made at 600
C and had a particle size of below 150 mm. For a similar wood charcoal, Cyganiuk et al. (2012) reported an average surface area of 204 m2 g1 (range 5e358 m2 g1). Compost was made from garden waste by a municipal composting company (Solum Gruppen, Roskilde, Denmark). 2.2. Spiking of the soil with PAHs A sandy loam soil from Flakkebjerg, Denmark (55.3125
N, 11.3952
E) was sampled from the plough layer at a depth of between 2 and 22 cm and stored at 4
C in plastic buckets (properties given in Table S2). Native black carbon represent less than 3% of the total organic carbon (Llorente et al., 2010), a concentration 1 order of magnitude lower than amount of black carbon to the soil in this study. Prior to use, it was passed through a 2 mm gauge sieve and homogenized by hand. Two soil treatments were prepared according to Northcott and Jones (2000) and Doick et al. (2003): soil spiked with a mixture of 5 PAHs and an unspiked soil to provide information on the blank levels. For each treatment, a 10% portion of the soil was air-dried at room temperature for 2 days prior to spiking. For the spiked soil treatment, the air dried portion was spiked with an acetone solution (60 mL kg1 of dry soil) containing phenanthridine, phenanthrene, anthracene, fluoranthene and pyrene each at 6.67 g L1, giving final nominal concentrations for each PAH of 30 mg kg1 soil. For the unspiked treatment soil portion pure acetone was used. After evaporation of the acetone for 24 h under a fumehood, the PAH-spiked and unspiked air-dried soil portions were thoroughly mixed with the remainder (90%) of the soil from each treatment in a large glass mixing bowl using a stainless steel spoon, and then passed three times through a 2 mm steel gauge sieve for further homogenization. Finally, the PAH-spiked and unspiked soils were rehydrated with deoinized water to the original field-moisture content of 35% by weight. A portion of the PAH-spiked soil was treated with HgCl2 at 1 g kg1 soil to prevent biodegradation. The different soils were separately stored in aluminium lined plastic bucket

125

(10 L), closed with an airtight lid, and aged at 4
C for a further 1 month. This temperature was chosen to limit biodegradation and/or volatilization losses. Subsequently, fractions of the PAH-spiked (without HgCl2) and unspiked soils were mixed with 1% (dry weight basis) of either AC, charcoal, or compost. Previous studies have indicated that 1% w/w of soil amendment was sufficient to dramatically reduce PAH desorption and bioavailability (Chen and Yuan, 2011; Marchal et al., 2013a; Marchal et al., 2013b; Rhodes et al., 2010). The soil plus amendment mixtures were thoroughly mixed and passed 3 times though a 2 mm steel gauge sieve to ensure a uniform distribution of the soil amendment. In parallel, additional portions of the PAH-spiked soil without any amendment were prepared. All soils were stored and aged in the dark at 4
C for an additional 1 month.

2.3. Preparation of the planted soil treatments Four planted treatments were prepared using the different PAH-spiked and (un) amended soils: soil only and soil amended with either AC, charcoal or compost (each n ¼ 4). To determine blank levels, four planted treatments were also prepared using the unspiked and (un)amended soils prepared above: unspiked soil only and unspiked soil amended with either AC, charcoal or compost (each n ¼ 2). A schematic overview of the treatments is given in Fig. S1. Radish plants (Raphanus sativus L., commercial seeds) were selected as they are small, grow fast, and reach an adult stage within 2 months. Prior to planting, the radishes were germinated for 2 days on filter paper moistened with deionized water and placed in a growth chamber with a day/night temperature of 19/17
C and under a 16 h photoperiod (110 mmol m2 s1 PFD, 400W PlantastarÒ fluorescent tubes (E40/ES)). After protrusion of the radicle, 10 young seedlings of uniform size (2 cm; 2 leaf stage) were transferred to aluminium foil-lined pots (240 mL) containing the different soil treatments (200 g). During the filling, clean silicone O-rings were placed in the soil at 3 cm and 6 cm depths (n ¼ 6) to act as a sink for measuring PAH diffusive uptake. After planting, 4 cleaned SPME fibres were carefully inserted into the soil to a depth of 4e6 cm, the exact depth was determined at sampling (Fig. S2). The replicate planted pots were grouped per treatment and placed in glass aquaria (Fig. S1). To minimize losses by diffusion of PAHs into the silicone sealant of the aquaria, joins were covered using aluminium tape. Air exchange was minimized by partially closing the aquaria with a glass sheet. Every week the plants were watered from the bottom with 500 mL of deionized water. To differentiate the role of an active rhizosphere, four unplanted treatments were also prepared using the PAH-spiked and unamended soils but without radish plants. These treatments were prepared in 120 mL brown pill glasses filled with 100 g of soil (each n ¼ 3). These had a similar surface area to volume ratio as the pots used for planting (1.10 versus 1.09 m2 kg1 soil, respectively). During preparation, three clean silicone O-rings were placed at a depth of 3 cm and three cleaned SPME fibres were inserted into the soil to a depth of 3 cm. The glass jars were placed in an aquarium under the same conditions as above. The four unplanted treatments were: (i) PAH-spiked soil without HgCl2 to determine biodegradation and other losses in a “living” soil without an active rhizosphere, (ii) as in the previous treatment but in closed jars to investigate volatile losses, (iii) PAH-spiked soil with added HgCl2 to determine abiotic losses in a soil without an active rhizosphere and (iv) as in the previous treatment but in closed jars to investigate volatile losses. The equivalent open and closed jar treatments showed similar losses indicating that volatilization was not a major losses route (data not shown), and the data from the closed glass jars treatments have been left out of the below discussion.

2.4. SPME fibre and O-ring sampling and HPLC analysis SPME fibres were withdrawn at 7, 26, 42 and 60 days (7, 42, and 60 days for the treatment without plants), wiped using lint free tissue to remove adhering soil and cut into pieces and stored in 2 mL brown HPLC glass vial at 20
C until analysis. The O-rings were removed at experiment completion (t ¼ 60d), rinsed with Milli-Q water, wiped using lint free tissue and stored in 20 mL glass vial at 20
C until analysis. SPME fibres and O-rings were sequentially extracted twice overnight at room temperature at 200 rpm with 0.5 mL and 1 mL of methanol, respectively and the extracts combined. PAHs were separated on a CP-EcoSpher 4 PAH column (Varian Inc., Palo Alto, CA) and detected by fluorescence detection on an Agilent 1100 HPLC equipped with a G1321A FLD operated at Ex, 260 nm, and Em, 350, 420, 440, and 500 nm. Full details are given in Marchal et al. (2013a).

2.5. Plant sampling For each treatment, 2 plants from each of the PAH-spiked and planted soil treatments were randomly harvested after 26, 42 and 60 days. Only the leaves and hypocotyl were sampled. The plants were cut at the basis of the hypocotyl, with root staying in the soil. Hypocotyls, stems and leaves were pooled per treatment, and weighed for fresh weight determination, after quickly rinsing with deionized water and gently blotting dry with lint-free tissue to remove soil traces from the plant. Note the hypocotyl of the radish was fully developed only at experiment completion (t ¼ 60d).

126

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

2.6. Calculation of the dissolved soil concentrations The success of PDMS fibre in acting as a passive sampler depends on (i) sorption to the PDMS sink being sufficiently fast to equilibrate with the compounds from the freely dissolved phase and (ii) the capacity of the PDMS sink being sufficiently small to induce only an insignificant desorption from the soil (also called “nondepletive” extraction). A minimum exposure time of 7 days was selected to ensure the PDMS was close to equilibrium at each sampling point for the 3- and 4ringed PAHs (log KOW between 3.48 and 5.10) used in this study (see discussion below). Dissolved pore water concentrations were determined from the PAH concentrations in the PDMS as shown in Eq. (1): Cfree ¼ CPDMS =KPDMS;water

(1)

Where Cfree (mg L1) is the dissolved concentration in the soil pore water, CPDMS (mg L1) is the concentration in the PDMS and KPDMS, water (L L1) is the equilibrium partitioning ratio between the PDMS and dissolved phases. Values of KPDMS, water for phenanthrene, anthracene, fluoranthene and pyrene were from Gouliarmou et al. (2012). A value for phenanthridine was not available. The log KPDMS, water values from Gouliarmou et al. (2012) were therefore linearly regressed against log KOW (log KPDMS, water ¼ 0.841 log KOW; log KOW range 3.0e5.5, n ¼ 5 r2 ¼ 0.99), and this used to derive a phenanthridine log KPDMS, water of 2.93, using the log KOW value of 3.48 from Kobeti cová et al. (2011). 2.7. Cumulative diffusive uptake into a silicone sink Techniques for measuring the diffusive flux towards a sink are well developed for cationic metals (Degryse et al., 2009; Zhang et al., 1998). In this study, the cumulative diffusive uptake of the PAHs into a large silicone sink in the form of Orings was measured. This proceeds in three steps: 1) uptake of PAHs from the surrounding pore water by the silicone creates a concentration gradient, 2) PAHs are transported along this concentration gradient to the silicone sink via diffusion in the soil solution, and 3) further PAHs are released from the soil matrix into the pore water. Therefore, PAHs sorbed to the soil also contribute to the total diffusive uptake into the sink. Further details of this approach are given in Marchal et al. (2013b). 2.8. Quantification of chemical activity in the soil pore water The Cfree of the PAHs were divided by their respective subcooled liquid solubilities in water (SL,water, mg L1) to derive their chemical activity (a, unitless). Values of SL,water at the average experimental temperature of 18
C were calculated from their solubility in water (Swater, mg L1) divided by the respective maximum chemical activity (amax, unitless) (Eq. (2)). amax ¼ Swater =SL;water

(2)

PAH values for Swater were taken from Rojo-Nieto et al. (2012), and for phenanthridine from Kobeti cová et al. (2011). The amax of a PAH at aqueous solubility, i.e., at saturation concentrations, is equal to the chemical activity of the PAH in its pure solid crystal state and calculated as described in Mayer and Holmstrup (2008). The calculated values of SL,water were used to convert the Cfree values determined by SPME into their respective values of a: a ¼ Cfree =SL;water

(3)

3. Results and discussion 3.1. Cfree in the unspiked soils and in the PAH-spiked but unamended and unplanted soil controls Cfree levels in all the unspiked soil treatments with or without amendments were several orders of magnitude lower than those determined in the different PAH-spiked soil treatments (data not shown). In the PAH-spiked but unamended and unplanted soil controls, similar trends were observed with or without HgCl2 (Figs. S3 and S4). Initial Cfree levels were highest for phenanthridine and phenanthrene, followed by fluoranthene, pyrene and finally anthracene. Subsequently, Cfree levels similarly decreased in both, with the exception of phenanthrene which showed a more pronounced decrease in the absence of HgCl2. After 60 days, Cfree in the soil treated with HgCl2 to prevent biodegradation had decreased by 54, 84, 46, 64 and 58% for phenanthridine, phenanthrene, anthracene, fluoranthene and pyrene, respectively. As discussed above volatilization did not play a role, and it is likely that these are partly explained by continued ageing during the experiment.

3.2. Changes in Cfree in the PAH-spiked but unamended soils with or without radish plants One aim was to investigate the effect of the plant rhizosphere in combination with different soil amendments on changes in PAH Cfree in soil, requiring that the PDMS was close enough to equilibrium to properly reflect the soil Cfree levels. Recently, Witt et al. (2013) showed for a static sediment that a partitioning equilibrium the 4 most hydrophobic PAHs used in this study was reached in less than 3 days for fibres with a PDMS surface area to volume ratio of 833 cm2 cm3 and from 3 to 16 days for fibres with a ratio of 450 cm2 cm3. The fibres in this study had a surface area to volume ratio of 550 cm2 cm3 and thus would have required an intermediate amount of time to reach equilibrium. Therefore, the initial sampling after 7 days might have resulted in a slight underequilibration for some of the PAHs, and thus the calculated Cfree values represent a lower estimate. However, this effect will have at least been partly offset by the observed decrease in Cfree. In any case, equilibrium sampling devices with shorter response times but with sufficient analytical sensitivity would definitely be useful for this type of study. Fig. 1 shows the changes in Cfree for the different PAH-spiked and planted soil treatments, as well as for the corresponding unplanted soil. The different amendment regimes resulted in quite different effects on the Cfree temporal trends, and these also depended on the PAH compound. In the unamended soils, the initial Cfree values for all compounds were consistently higher in the unplanted compared to the planted soils. In the absence of plants, initial Cfree values were higher by 38, 98, 39, 19 and 17% for phenanthridine, phenanthrene, anthracene, fluoranthene and pyrene, respectively. Subsequently, the Cfree levels decreased in both, but after 60 days still remained higher in the unplanted soil by 51, 89, 90 and 92% for phenanthridine, anthracene, fluoranthene and pyrene, respectively (Fig. 2A). Phenanthrene was the exception, with similar Cfree final values in unplanted and planted soils (Fig. 2A). Interestingly, for phenanthrene highly reduced Cfree levels were observed from the outset in the planted soil (Fig. 1). Therefore, the growth of the roots in the soil resulted in an immediate reduction in the phenanthrene Cfree. This suggests that revegetation could be considered as a supplementary approach to any chemical, mechanical or biodegradation based decontamination practices. The above observations indicate that the growth of radish roots in the soil enhanced the reduction of the Cfree levels for all PAHs compared to an unplanted soil. This could be due to higher microbial activity around the rhizosphere leading to increased biodegradation, increased sorption to plant organic material or enhanced water movement through the soil due to transpiration increasing losses. It appears that even a simple strategy of planting a soil can lead to faster reductions in the PAH dissolved concentrations. It would be interesting to investigate whether the Cfree levels in the unplanted soil would eventually be reduced to the same extent as observed for the planted soil over longer time periods. Similar effects of plant growth on PAH dissipation in soil have been observed in other studies. For example, Cheema et al. (2009) studied phenanthrene (11.5e344.2 mg kg1 dw soil) and pyrene (15.6e335.8 mg kg1 dw soil) biodegradation over 65 days in unplanted soil and soil planted with tall fescue (Festuca arundinacea). At experiment completion, tall fescue additionally decreased the concentration of phenanthrene and pyrene by 1.9e3.2% and 8.9e20.7%, respectively. Fu et al. (2012) investigated PAH dissipation in contaminated agricultural soil (PAH concentrations from 15 to 222 mg kg1 dw soil) planted with perennial ryegrass (Lolium perenne L.). After 7 months, concentrations of anthracene, fluoranthene, and pyrene decreased by 13, 18, and 18%, respectively,

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

127

Fig. 1. Measured freely dissolved pore water concentrations of PAHs (Cfree, mg L1) over time in the unplanted and unamended soil (soil only minus plants) as well as planted soils that were either unamended (soil only plus plants) or amended with either 1% (w/w) of activated carbon, charcoal or compost. The initial total soil concentration of each PAH was 30 mg kg1 w/w. Error bars indicate the standard error of the mean (SEM) (n ¼ 4).

compared to the unplanted soil. Interestingly, phenanthrene concentrations did not decrease whether plants were present or not, which is in contrast to the efficient removal observed in this study. 3.3. Changes in Cfree in the PAH-spiked and amended soils with radish plants For the planted soils with different amendment regimes, initial PAH Cfree values depended on the amendment (Fig. 1). Despite differences in the absolute values of the initial Cfree between the PAHs, for a given compound similar Cfree levels were found between the unamended, and compost or charcoal amended soils (within a factor of 2). In contrast, initial Cfree values were lower by more than

2 orders of magnitude when AC was present. Therefore, of the three amendments only AC rapidly reduced Cfree to very low levels ranging between 1.3 and 0.2 mg L1. These concentrations are in the same range as those measured in the unspiked soils, and furthermore were maintained over the 60 days. This indicates that despite any growth of the radish plant rhizosphere and release of organic material, AC is still able to maintain its strong sorption propensity. Again, phenanthrene displayed a different behaviour, with low Cfree values of less than 51 mg L1 maintained in the unamended, as well as the compost and charcoal amended, planted soils. Nevertheless, an additional effect on the Cfree levels was still observed for AC, with concentrations being further reduced by a factor of around 40 to less than 0.5 mg L1.

128

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

The above findings are in good agreement with other studies, where adding AC to different soils consistently has resulted in a reduction in Cfree of different PAH (Hale et al., 2012; Rhodes et al., 2008; Yang et al., 2009). Therefore, it is positive to note that in this study AC remains highly effective in reducing Cfree even in a planted soil. Charcoal on the other hand had a relatively minor effect on the time profiles of Cfree. Compared to the unamended and planted soil, charcoal led to rather modest reductions in the initial values of Cfree of only between 37 and 61% depending on the compound (Fig. 1). Furthermore, charcoal resulted in similar final Cfree values to the planted but unamended soil (Fig. 2B). This is unexpected as it has been postulated that even small additions of charcoal might increase sorption and thus reduce Cfree (Kookana, 2010; Marchal et al., 2013b; Rhodes et al., 2008). As an example, biochar improved the sorption affinity of soil, with the total sorption of naphthalene, phenanthrene, and pyrene being enhanced by up to 90% when the biochar content of the soil was not greater than 0.5% (w/w) (Chen and Yuan, 2011). Several reasons could explain the reduction of sorption efficiency of the charcoal including differences in biochar characteristics, the low PAH concentrations in this study, different soil Kd values or an active root system leads to increased organic matter competitive sorption on the charcoal surface. In any case, the reasons for the low effectiveness of charcoal in the planted soil need to be investigated. Compost had no consistent effect on either the initial or final values of Cfree when compared to the planted but unamended soil (Figs. 1 and 2B), suggesting that compost did not lead to an improved biodegradation of PAHs over the 60 days as has been postulated in other studies (Beesley et al., 2010; Ke et al., 2009). Nevertheless, it should not be forgotten that adding compost can have other beneficial effects such as adding nutrients and improving the soil texture, and thus plant growth. These results suggest that natural attenuation is greatly improved by the presence of plants, and in fact that plants were as efficient as charcoal or compost in decreasing PAH Cfree levels. 3.4. Chemical activity (a) of the soil pore water PAHs Fig. 2. Measured freely dissolved pore water concentrations of PAHs (Cfree, mg L1) at experiment completion (t ¼ 60d). In (A) the values of Cfree for the unplanted and unamended soil (soil only minus plants) as well as planted and unamended soil (soil only plus plants) are shown. In (B) the values of Cfree for the planted soil without amendment (soil only plus plant) or soils amended with either 1% (w/w) of activated carbon, charcoal or compost are shown. Arrows indicate Cfree in the activated carbon amended soil, with these also shown in the inset graph. The initial total soil concentration of each PAH was 30 mg kg1 w/w. Error bars indicate the standard error of the mean (SEM) (n ¼ 4).

Final Cfree values for the planted and amended soils are shown in Fig. 2B. AC resulted in final Cfree values that were lower by a factor of more than 10 compared to the other soil treatments. However, given the relatively large variation it is not possible to differentiate between the effectiveness of the unamended, compost and charcoal treatments in reducing Cfree. The final Cfree values for all compounds were consistently different between the 4 replicates. This is not due to a poor distribution of the amendment as the initial Cfree values showed a good reproducibility (see Fig. 1). Furthermore, in the unplanted soils reproducibility was also good (Fig. 2A), suggesting a uniform microbial activity in the soil as well as similar abiotic losses. Therefore, it appears that in some way the growth of the radish plants resulted in pockets of increased microbial activity in the soil despite the intensive efforts at homogenization. This implies that in field situations there can be considerable spatial differences in Cfree, which in turn would mean that exposure to the soil contaminants can vary with location.

PAHs accumulate in biological membranes, disturbing their structure, permeability and function (Yu, 2002). This mode of acute toxic action is called narcosis or baseline toxicity (Escher and Schwarzenbach, 2002), and at the membrane level is concentration-additive (Escher et al., 2002). The partitioning of PAHs into biological membranes is a spontaneous process driven by differences in chemical activity (a) between the membrane and external environment (Reichenberg and Mayer, 2006). Since the narcotic toxicity of PAHs is also a-additive, the sum chemical activity (Sa) of all PAHs in a mixture is an indicator of the potential narcotic toxicity of the mixture (Di Toro and McGrath, 2000). Although equilibrium will be reached relatively quickly for small soil organisms such as the springtail Folsomia candida (e.g., Schmidt et al., 2013), this will take longer for larger organisms as earthworms and these will also be more affected by metabolic losses (e.g., van der Wal et al., 2004). The Sa in the soil pore water therefore provides a “worst case” scenario for the narcotic potential of the PAHs. Values of a for the different PAHs in the soil pore water were calculated as described above, with changes in Sa over time shown in Fig. 3. The dotted area represents the range in a of 0.01e0.1 that has consistently been observed for the onset of narcotic baseline toxicity (Ferguson, 1939; Mayer and Reichenberg, 2006). For the unamended and unplanted soil, the initial value of Sa was 0.29 and exceeded the upper limit of the above range by a factor of 3. Even after 60 days, Sa was 0.06, which although reduced still could result

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

129

the soil will likely have resulted in initial Cfree levels higher than those when PAHs are introduced associated with black carbon or non-aqueous phase liquids. On the other hand, in the field considerably more organic contaminants will contribute to such narcotic toxicity. Therefore, similar studies in various field situations would be useful.

3.5. Cumulative diffusive uptake of PAHs into a silicone sink

P Fig. 3. Comparison of the sum chemical activity ( a) over time of the five PAHs in the unplanted and unamended soil (soil only minus plants), planted and unamended soil (soil only plus plants) and planted soils amended with 1% (w/w) of activated carbon, charcoal or compost. The shaded area indicates the range in chemical activities observed in the literature for the onset of narcotic baseline toxicity. Error bars indicate the standard error of the mean (SEM) (n ¼ 4).

in narcotic toxicity. For the unamended, as well as the compost and charcoal amended soils with plants, the initial values of Sa were 0.15, 0.12 and 0.08, respectively. Therefore, these soils would similarly have the potential to result in narcotic toxicity. However, after 60 days the values of Sa in these soils were reduced to 0.006, 0.017 and 0.009, respectively. These values are at the lower limit or below the above range in a. Note however, that the above 0.01 to 0.1 range in a has been established for acute narcotic toxicity, and a lower range in Sa is expected for sub-lethal effects (Mayer and Holmstrup, 2008; Reichenberg and Mayer, 2006). In contrast, adding AC immediately reduced Sa by over 99% to 0.0005. This is below the lower limit for narcotic toxicity by 2 orders of magnitude. Therefore, only the addition of AC resulted in Sa values that from the outset posed no risk for narcotic toxicity. Although the trends in Cfree will be similar, it is of course difficult to directly extrapolate the magnitude of these changes in Sa to different field scenarios. On the one hand, spiking of the PAHs into

Measuring the Sa of compounds such as PAHs in the soil pore water provides an indication of maximum narcotic potential. This may be relevant for small soil dwelling organisms which can relatively rapidly reach equilibrium with the surrounding environment (Schmidt et al., 2013). However, for larger organisms, such as earthworms or plants, equilibrium may not apply (Bergknut et al., 2007). Here, toxicity will be determined by the extent of PAH accumulation. This makes it interesting to complement the above SPME approach with an approach measuring the cumulative diffusive uptake of the PAHs. The diffusive uptake measurements given by the silicone O-rings therefore indicate how accumulation differs between the PAH compounds and soil treatments, and therefore gives an estimate of the potential accumulation of the contaminants throughout the 60 day period. The O-ring silicone concentrations at experiment completion are shown in Fig. 4. The relative trends in Cfree between the different soil treatments were directly reflected in their cumulative accumulation into the silicone. Compared to the other PAHs, phenanthrene was only accumulated in the silicone O-rings to low levels, in line with the consistently low Cfree levels in the planted and amended soil treatments. For the remaining PAHs, highest silicone concentrations were consistently observed in the unplanted and unamended soil. Compared to the planted but unamended soil, these were higher by 45, 94, 43, 46 and 41% for phenanthridine, phenanthrene, anthracene, fluoranthene and pyrene, respectively. This higher accumulation follows from the higher Cfree values observed in the

Fig. 4. Measured concentrations of PAHs in the silicone O-rings (mg L1) at experiment completion (t ¼ 60d) in the unplanted and unamended soil (soil only minus plants), planted and unamended soil (soil only plus plants) and planted soils amended with 1% (w/w) of activated carbon, charcoal or compost. Notice the different scales for the Y axes. The initial total soil concentration of each PAH was 30 mg kg1 w/w. Error bars indicate the standard error of the mean (SEM) (n ¼ 4).

130

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131

unplanted and unamended soil for all PAHs, providing a larger driving force for diffusive uptake into the silicone. For both unamended soil treatments (i.e., unplanted and planted), the highest concentrations were observed for fluoranthene and pyrene, which is in line with their higher hydrophobicity (Table S1). Addition of charcoal and compost only led to rather modest reductions in the cumulative PAH accumulation in planted soils. On average for charcoal and compost, these were lower by 50, 53, 31, 71 and 72% for phenanthridine, phenanthrene, anthracene, fluoranthene and pyrene, respectively. Interestingly, for these two amendments decreases in the cumulative accumulation were more pronounced for fluoranthene and pyrene, perhaps due to a proportionally higher sorption on the soil amendments. Compared to the other soil treatments, AC addition dramatically decreased the accumulation in the silicone O-rings. For example, compared to the unamended but planted soil, in the presence of AC the accumulation was reduced by 98, 96, 99 and 99% for phenanthridine, anthracene, fluoranthene and pyrene, respectively. Similarly, the cumulative accumulation of phenanthrene into the silicone sink was 2 orders of magnitude lower compare to the other treatments. Note that at experiment completion, the masses of PAHs in the silicone O-rings were 3 orders of magnitude below the nominal masses of PAHs added to the soil and thus did not deplete the bulk soil to any great extent. These results support the previous suggestion that AC and plants can reduce the transfer and accumulation of such contaminants into organisms. Simple transformation of the cumulative PAH accumulation into a meaningful diffusive uptake flux (e.g., mg m2 day1) would require that Cfree be constant, and that uptake remain in the initial linear part of the uptake curve. However, from Fig. 1 it is clear that Cfree was certainly not constant. This means that it is difficult, if not impossible, to extrapolate the cumulative accumulation determined in this study to other scenarios. Nevertheless, within this particular study the relative amounts in the silicone between the different soil treatments still provide a useful indication of the effectiveness of the amendments in reducing their potential accumulation. Moreover, in a historically contaminated soil, the Cfree of contaminants will be more constant, and thus using a silicone sink to measure a diffusive uptake flux might be more meaningful. 3.6. Effect of the soil amendments on plant growth In the PAH-spiked soils, plant growth was higher in presence of AC, followed by charcoal, and compost when compared to the unamended soil (Fig. S5). However, growth in the equivalent unspiked but (un)amended treatments was higher (Fig. S6). For charcoal and compost this might be partly explained by their limited effectiveness in reducing PAH toxicity. However, for AC the reasons are unclear given its documented ability to reduce the available PAH concentrations. In fact, no PAHs were detected in leaf or hypocotyl tissues regardless of the presence or absence of amendment. This is expected as hydrophobic organic compounds (log KOW > 3.5) are strongly bound to the soil or root surfaces, making accumulation within the above-ground parts unlikely (Morel et al., 2014; Simonich and Hites, 1995). 4. Concluding remarks In a PAH contaminated and unamended soil, the presence of a rhizosphere reduced the PAH Cfree, Sa and diffusive uptake compared to the equivalent unplanted soil. Charcoal and compost had a rather modest effect and furthermore did not significantly reduce Sa to levels below those resulting in narcotic toxicity. In contrast, adding AC to a PAH contaminated soil with an active plant rhizosphere dramatically reduced the Cfree values of all PAHs,

lowered Sa to below the threshold for baseline toxicity and also the diffusive uptake into a silicone sink. This further supports the application of AC as an amendment to minimize the uptake and toxicity of contaminants such as PAHs even when soils are planted. A minor downside to this was the slight decrease in plant growth observed upon adding powdered AC to the soil. However, this might be avoided by using granular AC, provided that the sorption kinetics are not negatively impacted. Taken together, the results from this study suggest that revegetation could form part of a remediation strategy with sorbents such as AC, to reduce the risks associated with PAH contamination whilst at the same time bringing other benefits such as reducing erosion. This study also highlights the use of equilibrium sampling and associated techniques for investigating changes in soil contamination metrics that are more relevant for assessing risk. Acknowledgements We thank Margit Fernqvist and Ellen Christiansen. This research project was financially supported by the Danish Council for Strategic Research, “Innovative REMediation and assessment TEChnologies for contaminated soil and groundwater” (REMTEC) (http:// www.remtec.dk/). Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2014.02.008. References Agarwal, T., Khillare, P.S., Shridhar, V., Ray, S., 2009. Pattern, sources and toxic potential of PAHs in the agricultural soils of Delhi, India. J. Hazard. Mater. 163, 1033e1039. Alexander, M., 2000. Aging, bioavailability, and overestimation of risk from environmental pollutants. Environ. Sci. Technol. 34, 4259e4265. Beesley, L., Moreno-Jiménez, E., Gomez-Eyles, J.L., 2010. Effects of biochar and greenwaste compost amendments on mobility, bioavailability and toxicity of inorganic and organic contaminants in a multi-element polluted soil. Environ. Pollut. 158, 2282e2287. Bergknut, M., Sehlin, E., Lundstedt, S., Andersson, P.L., Haglund, P., Tysklind, M., 2007. Comparison of techniques for estimating PAH bioavailability: uptake in Eisenia fetida, passive samplers and leaching using various solvents and additives. Environ. Pollut. 145, 154e160. Cheema, S.A., Khan, M.I., Tang, X., Zhang, C., Shen, C., Malik, Z., Ali, S., Yang, J., Shen, K., Chen, X., Chen, Y., 2009. Enhancement of phenanthrene and pyrene degradation in rhizosphere of tall fescue (Festuca arundinacea). J. Hazard. Mater. 166, 1226e1231. Chen, B., Yuan, M., 2011. Enhanced sorption of polycyclic aromatic hydrocarbons by soil amended with biochar. J. Soils Sediments 11, 62e71. Cyganiuk, A., Gorska, O., Olejniczak, A., Lukaszewicz, J.P., 2012. Pyrolytic production of microporous charcoals from different wood resources. J. Anal. Appl. Pyrolysis 98, 15e21. de-Bashan, L.E., Hernández, J.-P., Bashan, Y., 2012. The potential contribution of plant growth-promoting bacteria to reduce environmental degradation e a comprehensive evaluation. Appl. Soil. Ecol. 61, 171e189. Degryse, F., Smolders, E., Zhang, H., Davison, W., 2009. Predicting availability of mineral elements to plants with the DGT technique: a review of experimental data and interpretation by modelling. Environ. Chem. 6, 198e218. Di Toro, D.M., McGrath, J.A., 2000. Technical basis for narcotic chemicals and polycyclic aromatic hydrocarbon criteria. II. Mixtures and sediments. Environ. Toxicol. Chem. 19, 1971e1982. Doick, K.J., Lee, P.H., Semple, K.T., 2003. Assessment of spiking procedures for the introduction of a phenanthrene-LNAPL mixture into field-wet soil. Environ. Pollut. 126, 399e406. Escher, B.I., Eggen, R.I.L., Schreiber, U., Schreiber, Z., Vye, E., Wisner, B., Schwarzenbach, R.P., 2002. Baseline toxicity (narcosis) of organic chemicals determined by in vitro membrane potential measurements in energytransducing membranes. Environ. Sci. Technol. 36, 1971e1979. Escher, B.I., Schwarzenbach, R.P., 2002. Mechanistic studies on baseline toxicity and uncoupling of organic compounds as a basis for modeling effective membrane concentrations in aquatic organisms. Aquat. Sci. e Res. Across Bound. 64, 20e 35. Ferguson, J., 1939. The use of chemical potentials as indices of toxicity. Proc. R. Soc. Lond. Ser. B 127, 387e404.

G. Marchal et al. / Environmental Pollution 188 (2014) 124e131 Foo, K.Y., Hameed, B.H., 2012. Mesoporous activated carbon from wood sawdust by K2CO3 activation using microwave heating. Bioresour. Technol. 111, 425e432. Fu, D., Teng, Y., Shen, Y., Sun, M., Tu, C., Luo, Y., Li, Z., Christie, P., 2012. Dissipation of polycyclic aromatic hydrocarbons and microbial activity in a field soil planted with perennial ryegrass. Front. Environ. Sci. Eng. 6, 330e335. Fujinaga, A., Uchiyama, I., Morisawa, S., Yoneda, M., Sasamoto, Y., 2012. Methodology for setting risk-based concentrations of contaminants in soil and groundwater and application to a model contaminated site. Risk Anal. 32, 122e 137. Gomez-Eyles, J.L., Jonker, M.T.O., Hodson, M.E., Collins, C.D., 2011. Passive samplers provide a better prediction of PAH bioaccumulation in earthworms and plant roots than exhaustive, mild solvent, and cyclodextrin extractions. Environ. Sci. Technol. 46, 962e969. Gouliarmou, V., Smith, K.E.C., de Jonge, L.W., Mayer, P., 2012. Measuring binding and speciation of hydrophobic organic chemicals at controlled freely dissolved concentrations and without phase separation. Anal. Chem. 84, 1601e1608. Hale, S.E., Elmquist, M., Brändli, R., Hartnik, T., Jakob, L., Henriksen, T., Werner, D., Cornelissen, G., 2012. Activated carbon amendment to sequester PAHs in contaminated soil: a lysimeter field trial. Chemosphere 87, 177e184. Haritash, A.K., Kaushik, C.P., 2009. Biodegradation aspects of polycyclic aromatic hydrocarbons (PAHs): a review. J. Hazard. Mater. 169, 1e15. Jakob, L., Hartnik, T., Henriksen, T., Elmquist, M., Brändli, R.C., Hale, S.E., Cornelissen, G., 2012. PAH-sequestration capacity of granular and powder activated carbon amendments in soil, and their effects on earthworms and plants. Chemosphere 88, 699e705. Ke, L., Bao, W., Chen, L., Wong, Y.S., Tam, N.F.Y., 2009. Effects of humic acid on solubility and biodegradation of polycyclic aromatic hydrocarbons in liquid media and mangrove sediment slurries. Chemosphere 76, 1102e1108.  Kobeti cová, K., Simek, Z., Brezovský, J., Hofman, J., 2011. Toxic effects of nine polycyclic aromatic compounds on Enchytraeus crypticus in artificial soil in relation to their properties. Ecotoxicol. Environ. Saf. 74, 1727e1733. Kookana, R.S., 2010. The role of biochar in modifying the environmental fate, bioavailability, and efficacy of pesticides in soils: a review. Aust. J. Soil. Res. 48, 627e637. Llorente, M., Glaser, B., Belen Turrion, M., 2010. Storage of organic carbon and Black carbon in density fractions of calcareous soils under different land uses. Geoderma 159, 31e38. Marchal, G., Smith, K.E., Rein, A., Winding, A., Trapp, S., Karlson, U.G., 2013a. Comparing the desorption and biodegradation of low concentrations of phenanthrene sorbed to activated carbon, biochar and compost. Chemosphere 90, 1767e1778. Marchal, G., Smith, K.E.C., Rein, A., Winding, A., Wollensen de Jonge, L., Trapp, S., Karlson, U.G., 2013b. Impact of activated carbon, biochar and compost on the desorption and mineralization of phenanthrene in soil. Environ. Pollut. 181, 200e210. Mayer, P., Holmstrup, M., 2008. Passive dosing of soil invertebrates with polycyclic aromatic hydrocarbons: limited chemical activity explains toxicity cutoff. Environ. Sci. Technol. 42, 7516e7521. Mayer, P., Reichenberg, F., 2006. Can highly hydrophobic organic substances cause aquatic baseline toxicity and can they contribute to mixture toxicity? Environ. Toxicol. Chem. 25, 2639e2644. Megharaj, M., Ramakrishnan, B., Venkateswarlu, K., Sethunathan, N., Naidu, R., 2011. Bioremediation approaches for organic pollutants: a critical perspective. Environ. Int. 37, 1362e1375. Morel, J.L., Leglize, P., Ouvrard, S., 2014. PAH phytoremediation: rhizodegradation or rhizoattenuation? Int. J. Phytoremediat. 16, 46e61.

131

Northcott, G.L., Jones, K.C., 2000. Developing a standard spiking procedure for the introduction of hydrophobic organic compounds into field-wet soil. Environ. Toxicol. Chem. 19, 2409e2417. Pignatello, J.J., Kwon, S., Lu, Y., 2006. Effect of natural organic substances on the surface and adsorptive properties of environmental black carbon (Char): attenuation of surface activity by humic and fulvic acids. Environ. Sci. Technol. 40, 7757e7763. Reichenberg, F., Mayer, P., 2006. Two complementary sides of bioavailability: accessibility and chemical activity of organic contaminants in sediments and soils. Environ. Toxicol. Chem. 25, 1239e1245. Rhodes, A.H., Carlin, A., Semple, K.T., 2008. Impact of black carbon in the extraction and mineralization of phenanthrene in soil. Environ. Sci. Technol. 42, 740e745. Rhodes, A.H., McAllister, L.E., Chen, R., Semple, K.T., 2010. Impact of activated charcoal on the mineralisation of 14C-phenanthrene in soils. Chemosphere 79, 463e469. Rivera-Utrilla, J., Sánchez-Polo, M., 2011. Adsorbent-adsorbate interactions in the adsorption of organic and inorganic species on ozonized activated carbons: a short review. Adsorption 17, 611e620. Rojo-Nieto, E., Smith, K.E.C., Perales, J.A., Mayer, P., 2012. Recreating the seawater mixture composition of HOCs in toxicity tests with Artemia franciscana by passive dosing. Aquat. Toxicol. 120e121, 27e34. Schmidt, S.N., Smith, K.E.C., Holmstrup, M., Mayer, P., 2013. Uptake and toxicity of polycyclic aromatic hydrocarbons in terrestrial springtailsdstudying bioconcentration kinetics and linking toxicity to chemical activity. Environ. Toxicol. Chem. 32, 361e369. Semple, K.T., Morriss, A.W.J., Paton, G.I., 2003. Bioavailability of hydrophobic organic contaminants in soils: fundamental concepts and techniques for analysis. Eur. J. Soil. Sci. 54, 809e818. Simonich, S.L., Hites, R.A., 1995. Organic pollutant accumulation in vegetation. Environ. Sci. Technol. 29, 2905e2914. Smith, K.E.C., J.Oostingh, G., Mayer, P., 2010. Passive dosing for producing defined and constant exposure of hydrophobic organic compounds during in vitro toxicity tests. Chem. Res. Toxicol. 23, 55e65. van der Wal, L., Jager, T., Fleuren, R.H.L.J., Barendregt, A., Sinnige, T.L., van Gestel, C.A.M., Hermens, J.L.M., 2004. Solid-phase microextraction to predict bioavailability and accumulation of organic micropollutants in terrestrial organisms after exposure to a field-contaminated soil. Environ. Sci. Technol. 38, 4842e4848. White, P.A., Claxton, L.D., 2004. Mutagens in contaminated soil: a review. Mutat. Res. Rev. Mutat. Res. 567, 227e345. Witt, G., Lang, S.-C., Ullmann, D., Schaffrath, G., Schulz-Bull, D., Mayer, P., 2013. Passive equilibrium sampler for in situ measurements of freely dissolved concentrations of hydrophobic organic chemicals in sediments. Environ. Sci. Technol. 47, 7830e7839. Yang, Y., Hunter, W., Tao, S., Crowley, D., Gan, J., 2009. Effect of activated carbon on microbial bioavailability of phenanthrene in soils. Environ. Toxicol. Chem. 28, 2283e2288. Yu, H.T., 2002. Environmental carcinogenic polycyclic aromatic hydrocarbons: photochemistry and phototoxicity. J. Environ. Sci. Health Part C-Environ. Carcinog. Ecotoxicol. Rev. 20, 149e183. Zhang, H., Davison, W., Knight, B., McGrath, S., 1998. In situ measurements of solution concentrations and fluxes of trace metals in soils using DGT. Environ. Sci. Technol. 32, 704e710. Zheng, Y., Luo, X., Zhang, W., Wu, B., Han, F., Lin, Z., Wang, X., 2012. Enrichment behavior and transport mechanism of soil-bound PAHs during rainfall-runoff events. Environ. Pollut. 171, 85e92.