Introduction

Heavy metal and metalloid soil pollution through anthropogenic activities such as mining and smelting operations, burning of fossil fuels, applications of insecticides or fertilizers, and waste disposal are an increasing problem (Khan 2005; Yoon et al. 2006). High metal loads affect soil functions including soil structure and its productivity and may lead to contamination of ground- and surface waters (Ali et al. 2013; Vamerali et al. 2009). Decontamination of metal-polluted soils (Baker et al. 1994; Raskin et al. 1997) may be achieved by conventional remediation approaches like ex situ application of physicochemical methods; however, these techniques are both destructive and costly (Arthur et al. 2005; Saraswat and Rai 2009). Alternatively, bioremediation provides a sustainable and cost-efficient solution with phytoremediation aiming to apply metal accumulation in harvestable plant biomass (phytoextraction) with subsequent burning and ash deposition or to decrease metal mobility and toxicity (phytostabilization) (Brunetti et al. 2011; Dushenkov et al. 1997).

Phytoremediation strategies offer several benefits: they can be performed in situ and at low cost, prevent destroying soil structure and function, provide a vegetative ground cover abating erosion, and even are permissible for future land use and biorecovery of valuable metals (McGrath et al. 2001; Yang et al. 2005). Limitations that have been encountered are connected to soil properties, level of contamination, and bioavailability of pollutants (Pilon-Smits 2005). To overcome such limitations, organic or inorganic amendments have been applied which, however, may result in nonpredictable results (Raskin et al. 1997). An alternative strategy is to modulate phytoremediation by addressing the microbial activities in the rhizosphere (Bolan et al. 2014; Sullivan et al. 2013).

Soil bacteria and mycorrhizal fungi can alter physicochemical properties in the rhizosphere and affect plant growth, thus changing metal uptake, e.g., by secretion of phytohormones (Zhuang et al. 2007b), production of chelators and siderophores (Dimkpa et al. 2009a; Raskin et al. 1997), acidification, and biomineralization (Abou-Shanab et al. 2008; Lasat 2002). It has been shown that specifically Gram-positive bacteria such as streptomycetes are ubiquitous in metalliferous soils where they thrive due to specific metal resistance traits aiding, in turn, plant growth (Abbas and Edwards 1989; Dimkpa et al. 2008, 2009a; Haferburg and Kothe 2007; Schmidt et al. 2005, 2009). These interactions between metal-tolerant soil microorganisms and plant roots play a significant role in remediation of heavy metals. Their beneficial effects on plant growth through nitrogen fixation, solubilization of phosphate, or acting as biocontrol agents (Ahemad and Kibret 2014) are well-studied features of plant-associated microorganisms with which they improve efficiency of phytoremediation. Already improved growth, increased metal bioavailability, and protection of plants against phytotoxic metal effects are among the desired characteristics of microbial bioinoculants for improved phytoremediation (Lasat 2002; Weyens et al. 2009).

The potential of heavy metal-resistant bacteria for enhancing the growth of host plants in contaminated soil has been reported (Nogueira et al. 2007; Sessitsch et al. 2013) For instance, Streptomyces mirabilis has been found to improve biomass productivity of Sorghum bicolor in metal-contaminated soil (Schütze et al. 2014). Fast growing crop plants, like S. bicolor, offer several advantages for phytoremediation processes because of its high biomass production, stress tolerance, and metal accumulation potential (Ciura et al. 2005; Epelde et al. 2009; Marchiol et al. 2007; Murillo et al. 1999; Zhuang et al. 2009). In the work presented here, we evaluated the application of two metal-resistant Streptomyces strains, isolated from a former uranium mining site and the arbuscular mycorrhizal fungus Rhizophagus irregularis for microbially assisted phytoremediation approaches. The study investigated the impact of microbial amendment on plant performance and metal extraction by S. bicolor and examined metal mobility in contaminated soil in pot experiments and with field trials.

Material and methods

Site description and soil analysis

Pot and field experiments were carried out using contaminated soil material from the test site Gessenwiese installed by the University of Jena in 2004 on the basement of the former uranium leaching heap Gessenhalde near Ronneburg in Eastern Thuringia, Germany. Between 1952 and 1990, low-grade uranium ores were leached by irrigation with acid mine drainage (AMD) waters or diluted sulfuric acid (Büchel et al. 2005; Neagoe et al. 2005). After uranium mining operations were stopped in 1990, remediation started for restoration of this contaminated site. However, the drainage waters within this experimental site (Gessenwiese) still show high concentrations of heavy metals resulting in a spatially heterogeneous, but comparatively low, multimetal contamination (Schindler et al. 2012).

Soil samples were air-dried and sieved to a grain size up to 2 mm for determination of soil pH and total digestions and for sequential extractions. Soil pH was measured after shaking a 1:4 suspension for 1 h, left to settle for 24 h, and measured using pH330 (WTW). The same solution was used to determine electrical conductivity (EC; TetraCon 325 and LF320, WTW). Total heavy metal contents were determined using a pressure digestion system (DAS 30, PicoTrace). The bioavailable fraction of soil elements was determined following sequential extraction (Zeien and Brümmer 1989). The mobile fraction (F1) was extracted with 1 M NH4NO3 (p.a., Merck; compare Grawunder et al. 2009). Element contents were analyzed using inductively coupled plasma-optical emission spectrometry (ICP-OES; 725 ES, Varian) and inductively coupled plasma-mass spectrometry (ICP-MS, X-Series II, Thermo Fisher Scientific) in triplicates. The metal concentrations for total contents and bioavailable fractions are added as values before planting (t 0 at day 0) in the respective experiments, where t 0 was compared to soil concentrations after planting and inoculation. The sandy silt (53.93 % silt, 46.07 % sand) showed a cation exchange capacity of 9.07 mol/kg with a water content of 5.55 to 17.57 % and very low values for carbon, nitrogen, and sulfur as main nutrients (below detection limit for N and S, 1.01 to 1.17 % C).

Preparation of microbial inocula

The two multiresistant strains Streptomyces acidiscabies E13 and Streptomyces tendae F4 isolated from the former uranium mining site near Ronneburg, Germany (Amoroso et al. 2000), were used as bacterial inoculum. These strains are known to tolerate high concentrations of toxic metals and further for their plant growth promotion traits (Dimkpa et al. 2008; Schmidt et al. 2005). To prepare the bacterial inoculum for pot and field experiments, strains were cultivated in fermenters (7-L BIOSTAT B-DCUII, Sartorius Stedim Systems, or 300L Braun Biotech International). S. acidiscabies E13 was grown in medium 3 (glucose monohydrate, 5 g/l; soluble starch, 25 g/l; casein-peptone, 10 g/l; yeast extract, 5 g/l; (NH4)2SO4, 1.5 g/l; KH2PO4, 1.5 g/l; trace element solution, 1 ml [ZnCl2 40 mg/l, FeCl3⋅6 H2O 200 mg/l, CuCl3⋅6 H2O 10 mg/l, MnCl2⋅4 H2O 10 mg/l, Na2B4O7 10 mg/l, (NH4)2Mo7O24⋅6 H2O 10 mg/l, pH 7.0]) and S. tendae F4 was grown in medium 2 (replacing the C sources of medium 3 with glucose monohydrate, 30 g/l; casein-peptone, 10 g/l; cornsteep [Roquette]). Precultures for inoculating the fermenters were grown in the same media with additional 5 g/l CaCO3. Fermentation conditions were 25 °C, 500 rpm, pO2 > 20 %, aeration 2 slpm, and pH > 6 controlled with 10 % NaOH (only for S. tendae F4). After 42 h of growth, mycelium was harvested by centrifugation (6000 rpm, 15 min, Avanti J-20 XP, Beckman) or separation (300–400 l/h, CSA8, Westfalia) and resuspended in tap water. Dead biomass was obtained by autoclaving. The arbuscular mycorrhizal inoculum was obtained as expanded clay containing spores of R. irregularis (Biofa AG, Münsingen, Germany) with 100 spores per gram.

Pot experiments

Pot experiments were carried out from May to October, 2012, on S. bicolor plants grown in a greenhouse (Thüringer Landesanstalt für Landwirtschaft, Jena). The setup consisted of 40 polyethylene pots (12 × 12 × 16 cm) filled with 2.5 kg contaminated soil from the test site. Each pot was sowed with 23 seeds of S. bicolor. After germination, seedlings were thinned to 12 plants per pot. The experimental design included four treatments: a negative control (unamended, C), amended with Streptomyces strains (S), amended with mycorrhizal fungus (M), and amended with a mixture of the two streptomycetes and the mycorrhizal fungus (MS). All treatments were carried out in five replicates.

Microbial inoculation was performed by mixing 20 ml of bacterial suspension and/or 4 g of R. irregularis granulate at the time of seeding. The pots were arranged in a randomized pattern and randomly rearranged every 4 days. Plants grew with natural day/night rhythm at ambient temperature between 15 and 30 °C. All plants were irrigated daily with distilled water. Aboveground biomass was harvested at 3 and 6 months after planting.

Field experiment

The field experiment was carried out from May to September, 2013, on the test site Gessenwiese (50° 51′ 27″ N and 12° 08′ 82″ E) in the former uranium mining district Ronneburg, Germany (Büchel et al. 2005). S. bicolor was cultivated in two different plots of 12 × 12 m each, one of which had been amended with 5 cm of calcareous topsoil in 2004 (topsoil plot), while the second plot was left unamended (control plot). Sorghum plants were subjected to three experimental treatments in three replicates at each plot: unamended control (C), inoculated with mycorrhizal R. irregularis (M), and inoculated with a mixture of mycorrhiza and streptomycetes (MS). For microbial inoculation, a volume of 20 l of bacterial suspension and/or granulate of R. irregularis as recommended were applied per subplot (Neagoe et al. 2014; Schindler et al. 2012). Harvest occurred after 17 weeks.

Plant analyses

After harvesting, plant shoots were thoroughly washed with deionized water and oven-dried at 40 °C until constant weight to determine shoot dry weight. Plants were then ground to a fine powder using an ultracentrifugal mill (ZM100, Retsch). Up to 200 mg of plant material was weighted and digested with 5 ml HNO3 (65 %, supra, Merck) in a microwave pressure system (Mars 5 XPRESS, CEM, Germany). The digested samples were transferred into 25 ml flasks filled up with ultrapure water (PureLab Plus, USF Elga) and analyzed for heavy metals by ICP-OES (725 ES, Varian) and ICP-MS (X-Series II, Thermo Fisher Scientific) in triplicates. The precision and accuracy of the ICP-MS and ICP-OES measurements were proven by analyzing standard reference material SPS-SW2 (Spectrapure Standards AS) and NIST 1643e (NIST) and by measuring multielement standard solution (500 mg/l Ca, K, Mg, Bernd Kraft) each in dilution 1:5 (v/v) and comparison to the certified values. Typical precision for triplicate measurements was ≤2 % for ICP-MS and ≤5 % for ICP-OES.

Statistical analyses

All statistical analyses were performed with R 3.0.3. The data were analyzed for variance (ANOVA) with a confidence level of 95 %. Significant differences between treatment means were confirmed by Tukey’s test or, for nonparametric data, by Kruskal-Wallis test (P < 0.05). Means and standard deviations were calculated using Microsoft Excel 2007 (Microsoft Corporation) for Windows 7.

Results

Plant performance on contaminated substrate under glasshouse conditions

After 3 and 6 months of plant growth, the influence of microbial inoculation on biomass production of S. bicolor was evaluated by measuring shoot weight (Fig. 1). The biomass productivity of inoculated plants that were treated with both mycorrhiza and streptomycetes showed a slight, albeit statistically significant increase after 3 months.

Fig. 1
figure 1

Effects of treatments on shoot dry weight of Sorghum bicolor per pot after 3 and 6 months of growth in potting substrate. Values represent means ± SD (n = 5). Two-way ANOVA was performed to determine the effects of treatments and time. Significance levels of time, treatments (Treat), and the interaction treat × time are shown: *P < 0.05; **P < 0.01; ***P < 0.001. C control treatment without inoculation, S amended with streptomycetes, M amended with mycorrhiza, MS amended with mycorrhiza and streptomycetes

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The uptake of metals into shoots of Sorghum plants (Table 1) showed significant differences between treatments. Highest concentrations of Al, Co, and Ni were observed in the shoots of noninoculated Sorghum plants, while highest amounts of Cd, Mn, Sr, and Zn were found in plants with microbial amendment after 3 months.

Table 1 Metal concentrations in shoots of Sorghum bicolor grown in greenhouse pots
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After 6 months, Sorghum plants without microbial inoculation accumulated significantly higher levels of Al, Co, Mn, Ni, and Zn, while a significant contribution of bacterial and mycorrhizal inoculation could be observed for Cd and Sr uptake. The low bioavailability of U resulted in very low concentrations (0.01 without standard deviation) and was not further considered.

Treatment effects on metal availability and contents in the potting substrate

Both total metal contents and bioavailability were examined in order to evaluate the potential of microbially assisted phytoremediation under controlled conditions. The substrate showed multimetal contamination with high Al bioavailability at pH 4.4 to 4.6 and an electrical conductivity of 439 ± 12 μS cm−1. Soil bacteria and mycorrhizal fungi can change soil pH and, hence, alter bioavailability. Additional metal tolerance mechanisms including chelator or siderophore production may lead to changes in metal transfer from soil into plant biomass. Thus, the changes in bioavailable metal contents within the soil after the plant growth season were checked with respect to the different microbial treatments. For most metals, a decrease in mobile fraction was seen after the pot experiment, most prominent for Al, while a slight mobilization of Co and Mn had occurred for both inoculated and noninoculated pots after 6 months. In contrast, Cd showed very low mobilization for all treatments (Table 2). The inoculation did show a statistically significant effect on reduction of mobile Al and Ni contents.

Table 2 Alteration of soil metal bioavailability in pots after plant growth
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Biomass production and metal uptake in the field trial

S. bicolor was grown in both field substrates which featured pH 5.2 to 5.4 and conductivity of 276 ± 60 μS cm−1 for the topsoil field site and pH 4.4 to 5.2 and 249 ± 99 μS cm−1 for the control soil plot. The microbial amendments of either mycorrhiza or a mixture of mycorrhiza and streptomycetes were applied to evaluate microbial impact on plant growth. This was following the results of the pot experiments in which either mycorrhiza alone or, in most parameters, combined streptomycete and mycorrhiza application had induced changes in metal bioavailability.

As the field site had been amended in 2004 with 5 to 10 cm topsoil to allow for better plant performance, this effect was evaluated in addition to microbial inoculation. Generally, an effect on biomass production by adding low amounts of topsoil in 2004 was not observed. In addition, for plants grown on the topsoil plot, microbial inoculation did not enhance biomass production (Fig. 2). In contrast, in the nonamended control soil, aboveground biomass was significantly increased by combined inoculation with mycorrhiza and streptomycetes. Thus, an effect of topsoil addition was seen, albeit only with the help of microbial amendments. While microbial addition could help plant growth on the unamended control soil, the topsoil addition had been sufficient—potentially even by adding the soil microbial community—to support plant growth in a way that made additional microbial inoculation superfluous.

Fig. 2
figure 2

Effects of treatments on shoot dry weight of Sorghum bicolor grown in two different field substrates. Values represent means ± SD (n = 9). One-way ANOVA was performed for each field substrate. Means with different letters are significantly different (P < 0.05) according to Tukey’s test. C control treatment without inoculation, S amended with streptomycetes, M amended with mycorrhiza, MS amended with mycorrhiza and streptomycetes

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In line with the lack of a measurable effect of inoculation on the topsoil-treated field site, microbial inoculation had no significant effect (P < 0.05) on general metal accumulation in S. bicolor shoots (Table 3). This was found for both substrates. In a more detailed analysis, the topsoil field-grown plants showed higher levels of U and Zn in inoculated subplots. However, the combined application of mycorrhiza and streptomycetes decreased the uptake of Ni into shoots. On the unamended control soil, Co and Mn were accumulated in high amounts into shoot biomass of inoculated Sorghum, while the concentration of Ni was lowest in the shoots of plants treated with mycorrhiza.

Table 3 Metal concentrations in shoots of Sorghum bicolor grown in the field trial
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To test the effects of planting and inoculation on the respective substrate, metal mobility was scored by sequential extraction before and after planting (Table 4). In the topsoil plot, neither plant growth nor inoculation induced visible changes in metal mobility recorded at the end of the growing season for most metals. However, there was a significant increase in Sr in the mobile fraction, while U availability was slightly reduced. Only few significant changes in bioavailable metal concentrations were detected for the control soil, with increases in Al and U and decreases in Co and Mn after plant growth.

Table 4 Alteration of soil metal bioavailability in the field after plant growth
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Discussion

Phytoextraction of heavy metals by using crop species with high biomass production is a promising approach to remediate low to moderately contaminated soils (Ernst 2005). In our study, we evaluated the role of microbial inoculation on metal extraction and uptake by S. bicolor grown on multimetal-contaminated soil. Besides metal accumulation capacity, plant biomass production was measured to define a phytoextraction potential for this particular plant species known to tolerate heavy metals including Zn, Cu, Cd, Ni, and Pb (Hernández-Allica et al. 2008; Zhao et al. 2003; Zhuang et al. 2007a). We did not consider root concentrations which are often included in phytoextraction calculations. The crop harvestable biomass exclusively consists of the aerial part, and thus, roots may not be considered for phytoextraction or phytomining as long as the root is not harvested (which would be the case for, e.g., potato or beets).

The diffusive metal translocation in soil and root apoplast reflects the bioavailability of a given metal, making it essential to analyze the bioavailable fraction. The impact of soil microbes on bioavailability is an essential part influencing plant uptake. Soil bacteria and mycorrhizal fungi facilitate an increase in soil metal mobility (Zhuang et al. 2007b) and can significantly promote heavy metal uptake by plants (Rojas-Tapias et al. 2012; Usman and Mohamed 2009).

Metal-resistant streptomycetes were applied to enhance plant performance. These Gram-positive, aerobic soil bacteria have been found to promote plant growth on metalliferrous soils (Dimkpa et al. 2008, 2009b). The effects of streptomycetes on plant development can be triggered by various mechanisms including phosphate solubilization, production of phytohormones, and siderophore excretion (Dimkpa et al. 2009a; Langella et al. 2014). Additionally, metal-resistant arbuscular mycorrhizal fungi have been extensively investigated for application in soil remediation (Griffioen 1994; Khan 2005; Khan et al. 2000; Turnau et al. 2001). They can support growth of host plants in metal-contaminated environments by enhanced uptake of nutrients and water and by modification of metal toxicity via complexation or precipitation (Ernst 2005; Gaur and Adholeya 2004; Wang et al. 2007). In mycorrhizal plants, toxic elements were found to be either more highly concentrated or reduced through fungal metal-binding processes within the rhizosphere (Toler et al. 2005; Usman and Mohamed 2009).

A dual inoculation with arbuscular mycorrhizal fungi and rhizospheric bacteria, specifically streptomycetes, showed enhanced plant biomass productivity and increased levels of mycorrhization (Abdel-Fattah and Mohamedin 2000). In contrast, antagonistic interactions between AM fungi and actinomycetes have also been reported (Adriano-Anaya et al. 2006; Ames et al. 1984; Schreiner and Koide 1993), underlining the necessity to first score the microbial interactions in pot experiments.

The potential of metal removal of a given plant species is mainly influenced by its metal accumulation capacity and biomass productivity (Lasat 2002; Zhuang et al. 2007a). In our experiment, microbial inoculation could partially enhance shoot metal concentration of Sorghum plants under greenhouse conditions, while due to high variability, the effect was less clear under field conditions. These differences in metal uptake between small-scale and open environment, large-scale field trials may be caused by different physiological states of the plants and heterogeneous soil conditions.

The bioconcentration factor (BCF; metal concentration in harvested shoots/soil content in mobile fraction F1) is one of the most important variables for a successful phytoextraction process (McGrath and Zhao 2003). Sorghum plants showed relatively high BCF values for Cd and Zn (Table 5). Shoot metal concentrations, including BCF values, decreased during growth except for Zn (compare Epelde et al. 2009).

Table 5 Bioconcentration factors for metal accumulation into Sorghum shoots (shoot metal concentration/soil content in mobile fraction F1)
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Since only a small fraction of heavy metals is bioavailable for plant uptake, it is necessary to follow metal mobility in soil (Violante et al. 2010). Besides physicochemical properties like soil pH, redox potential, or metal speciation, which strongly influence bioavailability of heavy metals, soil microorganisms can significantly promote metal solubility and mobilization in the soil through acidification or by producing chelators (Marques et al. 2013; Sheng et al. 2012). The microbial inoculation had a significant impact on reduction of bioavailable soil fractions of Al and Ni after 6 months of plant growth under controlled conditions (see also Schütze et al. 2014). Thus, our study support the use of Sorghum for phytoextraction specifically for Cd and Co, while microbial inoculation can lead to higher plant survival by minimizing the toxic effects of other metals like Ni in the multimetal-contaminated substrate that is usually found at anthropogenically contaminated, metalliferous sites. In our experiments, we were able to extract, choosing the right conditions, approximately 0.5 g Co or Cd per hectare and 4.5 and 1.2 g Ni and Sr per hectare, respectively, and at the same time, 15 mg of highly detrimental U and 120 g Mn per hectare. This multielement remediation can provide a suitable method to stabilize contaminated land and provide future land use, potentially with alternating extraction and renewable energy plant production cycles.