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Article

High Cadmium and Mercury Soil Contamination Outweighs the Effect of Soil Amendments When Growing Miscanthus x giganteus

by
Željka Zgorelec
1,*,
Lana Zubčić
1,*,
Silva Žužul
2,
Zorana Kljaković-Gašpić
2,
Marija Trkmić
3,
Marija Galić
1,
Iva Hrelja
1,
Ana Špehar Ćosić
4,
Aleksandra Perčin
1 and
Nikola Bilandžija
1,*
1
Faculty of Agriculture, University of Zagreb, Svetošimunska cesta 25, 10000 Zagreb, Croatia
2
Institute for Medical Research and Occupational Health (IMROH), Ksaverska cesta 2, 10000 Zagreb, Croatia
3
HEP Proizvodnja d.o.o. Central Laboratory for Chemical Technology, Zagorska 1, 10000 Zagreb, Croatia
4
Agroproteinka d.d., Strojarska cesta 11, Sesvetski Kraljevec, 10361 Sesvete, Croatia
*
Authors to whom correspondence should be addressed.
Appl. Sci. 2025, 15(16), 9075; https://doi.org/10.3390/app15169075
Submission received: 5 June 2025 / Revised: 25 July 2025 / Accepted: 4 August 2025 / Published: 18 August 2025

Abstract

This three-year study evaluated the effects of various soil amendments on growth parameters and heavy metal (HM) accumulation in above- and belowground biomass of Miscanthus x giganteus (MxG), assessing its phytoremediation potential. A randomised complete block design included four treatments: I—control, II—sludge, III—mycorrhiza, and IV—MxG ash. All experimental pots were filled with soil spiked with Cd (100 mg kg−1) and Hg (20 mg kg−1). Aboveground biomass yield ranged from 3.44 to 5.59 tDM ha−1, with Cd and Hg concentrations in biomass varying from 5.98 to 14.62 mg Cd kg−1 and 41.8 to 383.9 μg Hg kg−1, respectively. Belowground biomass mass ranged from 6.90 to 8.30 tDM ha−1, with Cd and Hg concentrations between 44.3 and 57.2 mg Cd kg−1 and 4.24 to 6.05 mg Hg kg−1, respectively. Enrichment coefficients (EC) in aboveground biomass ranged from 0.060 to 0.146 for Cd and 0.002 to 0.019 for Hg. Belowground biomass EC values ranged from 0.44 to 0.57 for Cd and 0.21 to 0.30 for Hg. The translocation factor (TF) varied from 0.104 to 0.145 for Cd and 0.008 to 0.024 for Hg. Our findings suggest that miscanthus is more effective for heavy metal phytostabilisation and biomass production in moderately contaminated soils than for phytoextraction.

1. Introduction

Heavy metals (HMs), characterised by diverse chemical traits and biological impacts, are often associated with toxicity and environmental contamination [1]. Soils are considered contaminated with HMs when their levels cause visible or measurable disturbances in soil functions [2], and these metals may be of natural or anthropogenic origin [3]. In agriculture, sources of HMs include organic and inorganic fertilisers (especially phosphate fertilisers), sewage sludge, irrigation water, and plant protection agents [1,4]. Once present in the soil, HMs resist degradation, resulting in bioaccumulation throughout the food chain [5]. In uncontaminated soils, average Cd concentrations typically range from 0.01 to 5 mg kg−1 [6], whereas average Hg concentrations are generally below 100 µg kg−1 [7]. Depending on soil pH, Cd typically occurs as CdCO3 (alkaline) or CdS (acidic) and can easily transform into plant-available forms [8]. Plants mainly absorb Cd as Cd2+ ions [9], while Hg is primarily taken up as Hg2+ [8]. Compared to many other HMs, Cd tends to accumulate in plants at higher levels [8], whereas Hg binds strongly to reduced sulphur groups and soil organic matter (OM) [10]. Various soil properties—such as pH, temperature, OM content, cation exchange capacity (CEC), redox potential, and the presence of other ions—affect HM bioavailability [11,12]. In this context, applying soil amendments (i.e., biochar, greenwaste compost) to multi-element contaminated soils can have contradictory effects on the mobility, bioavailability, and toxicity of specific elements, depending on the amendment used. For example, treatments with amendments may significantly increase dissolved organic carbon (DOC) and pH, while decreasing soil Cd concentrations [13]. HM accumulation in plants can cause growth disorders, including reduced germination and seed development, impaired water uptake and transpiration rates, inhibited photosynthesis and growth, poor nutrient uptake, and leaf damage [14,15,16]. However, certain plant species known as metallophytes, which may grow in highly polluted soils, have developed various adaptation mechanisms enabling hypertolerance to elevated HM concentrations [17]. Thus, contaminated soils can be repurposed for cultivation of tolerant non-food crops, including second-generation bioenergy crops (e.g., switchgrass, short rotation crops (Poplars, Robinia, Eucalyptus), and miscanthus), which thrive in contaminated and marginal soils [18]. These crops offer additional benefits such as erosion control and carbon sequestration [19], while accumulating pollutants through phytoremediation—a sustainable approach that relies on plants to remove, degrade, or retain harmful soil substances [20]. Miscanthus (Miscanthus x giganteus; MxG) has shown particular potential for phytoremediation due to its perennial growth, high biomass yield, resistance, and adaptability. The harvested biomass can be converted into thermal energy or electricity through direct combustion [21]. Phytoremediation mechanisms for HM removal from soil include phytostabilisation, phytoextraction, and rhizofiltration [18]. Previous research has examined the phytostabilisation potential of miscanthus for Cd and Hg, as well as its biomass production in soils with moderate Cd and Hg pollution [22,23]. Although soil contamination with Cd and Hg does not significantly affect the combustion properties of miscanthus biomass, Cd accumulation limits its safe use for solid biofuel. Therefore, miscanthus biomass may only be suitable for use in facilities with limited HM re-emissions. After burning the contaminated biomass, the ash composition must be analysed to ensure safe and environmentally friendly disposal. Additionally, soil amendments, such as sludge, mycorrhiza, and biomass ash, can be applied to the soil to reduce the negative impacts of HMs and support phytoremediation.
Waste sludge contains OM, nutrients, trace metals, micropollutants, and microorganisms, and its composition—determined by the wastewater source and treatment method—must be accurately characterised before application to soil [24,25]. Its high OM content, CEC [26], and pH can affect HM bioavailability [27]. Application of animal waste sludge, rich in N, P, and S, had no significant impact on lettuce yield or quality, suggesting its suitability for non-food crop cultivation [28]. Similarly, application of municipal sludge to miscanthus at varying doses (1.66, 3.32, and 6.64 tDM ha−1) resulted in minor changes to soil properties (decreased pH and available P, increased organic C and available K) but had limited effect on soil fertility and HM levels in soil and plant biomass [29]. In a five-year study, the application of food-industry sludge to tree species (Pinus silvestris L., Picea abies L., and Quercus robur L.; initially inoculated with mycorrhiza) growing near a zinc smelter with high Cd, Zn, and Pb concentrations improved soil properties and reduced HM uptake into aerial plant parts. In treated plants, Cd uptake remained stable over time, whereas untreated plants showed a doubling of Cd content annually, confirming the value of sludge amendments for phytoremediation purposes [26].
Mycorrhiza is a symbiotic relationship between higher plant roots and fungi, where fungi supply inorganic substances (water and nutrients) in exchange for organic substances (assimilates) [30]. Based on morphological characteristics, there are six types of mycorrhiza [31], including ectomycorrhiza, which was used in this study. Ectomycorrhiza forms a sheath around the host plant roots without penetrating the cells, creating a hyphal network called the Hartig net [32]. This symbiosis enhances host-plant resistance to various stresses [33] and extends root function through external mycelium, enabling greater nutrient uptake [34]. Ectomycorrhizal fungi have developed defence mechanisms against stress caused by HMs, enabling them to survive in contaminated soils [35]. These fungi can increase plant tolerance to HMs by binding metals to the fungal or plant cell walls, reducing root exposure, preventing root-to-shoot transfer, enhancing water and nutrient uptake, and diluting their content in the plant [32]. Conversely, ectomycorrhiza can also increase metal availability and uptake, thereby supporting phytostabilisation and phytoextraction strategies [36]. For example, a study that investigated the effects of three ectomycorrhizal fungi on Cd uptake in P. canadensis and S. viminalis grown in highly contaminated soil found that P. involutus significantly enhanced Cd extraction and translocation in P. canadensis [37]. Similarly, inoculation of Populus × canescens with P. involutus (strain MAJ) increased Cd accumulation in biomass, improved physiological performance, and reduced oxidative stress compared to the control [38].
Biomass ash, a solid by-product of biomass combustion, is frequently disposed of in landfills despite its high content of valuable micro- and macronutrients [39]. Its physical and chemical properties vary based on plant species and parts used, cultivation practices, soil characteristics, climatic conditions, power plant and technological processes, as well as transportation and storage methods [40]. In agriculture, ash can be recycled as a soil amendment, a fertiliser substitute for non-food crops, or as a material for soil calcification [41]. However, its chemical composition and HM content must be assessed before application [39]. Biomass ash serves as a source of essential elements such as Ca, Mg, K, and P [40] and can affect soil water retention due to its SiO2 content [42]. Most studies report that ash application has a positive effect on the abundance and diversity of soil bacteria and microorganisms, thereby supporting mineralisation and nitrification; however, some studies have observed reductions in organic matter mineralisation and microbial activity following ash application [41]. The use of biomass ash can improve plant growth and yield [40], with effects comparable to those of mineral fertilisers in certain grass species [42]. For example, a two-year study demonstrated that both biomass ash and biochar can enhance soil chemistry and increase miscanthus yield, with biochar showing greater effectiveness, highlighting their potential as alternatives to mineral fertilisers [43].
This study continues the evaluation of miscanthus as a sustainable tool for both the remediation and productive use of soils contaminated with Cd and Hg. Building on previous research, this work systematically examines the influence of various soil amendments—waste sludge, mycorrhiza, and biomass ash—on metal uptake, biomass yield, and overall phytoremediation performance under controlled conditions. Furthermore, by applying soil amendments, particularly by-products such as sludge and ash, for biomass cultivation on contaminated soils, this approach integrates phytoremediation with bioenergy production, thereby promoting a sustainable and circular system.
Cases of high mercury contamination have been reported; for example, Gosar et al. [44] observed severe environmental contamination in the western part of Slovenia, linked to the 500-year history of mining and ore processing in the Idrija mercury mine. The highest Hg concentration in the soil (5293 mg kg−1) was found in the Idrijca river valley. Extremely high levels of Hg were found in profile P4, containing 37,020 mg kg−1 Hg at a depth of 20–30 cm [45]. Kabata-Pendias and Mukherjee [46] reported up to 7.45 mg Hg kg−1 at Mount Etna, Italy, and up to 212 mg Hg kg−1 in the Monte Amiata mining area in Italy. With regard to Cd, the same authors observed soil concentrations of up to 345 mg Cd kg−1 and 3 410 mg Cd kg−1 when P fertiliser or sewage sludge was applied. These examples highlight the presence of highly contaminated sites across Europe. For this reason, our study focused on assessing MxG performance under extreme Cd and Hg contamination, addressing the lack of data on its cultivation in such conditions and evaluating its potential for phytoremediation.

2. Materials and Methods

2.1. Experimental Plot

The study was conducted under semi-controlled conditions in an open greenhouse at the Faculty of Agriculture, University of Zagreb, starting 8 May 2018. The experiment spanned three growing seasons: 2018/2019, 2019/2020, and 2020/2021. Meteorological records from the Maksimir station, Zagreb (Croatian Meteorological and Hydrological Service), reported annual precipitation values of 853.6 mm (2018), 1000.5 mm (2019), 950.4 mm (2020), and 772.2 mm (2021), while mean annual air temperatures were 13.0 °C (2018, 2019), 12.6 °C (2020), and 12.2 °C (2021). Weather conditions were natural, while soil moisture was regularly monitored and maintained at field capacity levels. The experiment consisted of four treatments with three replicates each, arranged in a completely randomised design. Seedlings were planted in truncated cone-shaped plastic experimental pots (EP; 28 cm top diameter, 19 cm base diameter, 29 cm height), each filled with 18 kg of prepared soil substrate.

2.2. Preparation of Contaminated Soil

As detailed in Šestak et al. [47], the soil used across all treatments was contaminated with 100 mg Cd kg−1 (as CdO(s), 99% purity) and 20 mg Hg kg−1 (as HgCl2(s), 99.5% purity). Four treatments were applied based on soil amendment: I—control (unamended soil), II—waste sludge from a biogas plant, III—mycorrhiza, and IV—MxG ash. Subsamples of uncontaminated soil were thoroughly homogenised with the designated concentrations of HMs before being combined with the total pot volume.

2.3. Soil Characteristics

Soil for the experiment was sourced from the faculty’s experimental field in Maksimir, and its physical and chemical properties were determined [48]. The mechanical composition of the soil was analysed in Na-pyrophosphate using the sieving and sedimentation method [49]. The soil contained 14.4% sand, 75.7% silt, and 9.9% clay, and its texture was classified as silty loam. It was slightly acidic (pH = 6.23; obtained in 1 M KCl in 1:2.5 (m/v); mod. [50]), and rich in organic matter (OM = 3.8%; determined by wet combustion method with sulfochromic oxidation; mod. [51]). Total carbon (C) (2.48%; [52]), nitrogen (N) (0.23%; [53]), and sulphur (S) (0.049%; [54]) were determined by the dry combustion method on the Vario Macro CHNS analyser (Elementar, Langenselbold, Germany). According to Woltmann’, the soil was classified as rich in total N, as referenced in Čoga and Slunjski [55]. The supply of physiologically active, plant-accessible forms of phosphorus (P2O5) and potassium (K2O) in the soil, determined using the AL method [56], was very low (PAL = 4.4 mg 100 g−1; KAL = 7.8 mg 100 g−1) [57]. The cation exchange capacity (CEC) was 18.9 cmol+ kg−1 [58]. Baseline concentrations of Cd and Hg in uncontaminated soil prior to the experiment (0.50 mg Cd kg−1 and 0.21 mg Hg kg−1 [59]) were substantially lower than the maximum allowable concentrations (MACs) set by Croatian legislation for agricultural soils with pH > 6 (MACCd = 2 mg kg−1; MACHg = 1.5 mg kg−1) [60]. An aliquot of this soil was also used as the control treatment (I—control, soil without amendments).

2.4. Soil Amendments

As described in Šestak et al. [47], the second treatment (II—sludge) involved animal waste sludge or digestate from the fermenter of the Agroproteinka-energija d.o.o. (Sesvete, Croatia) biogas plant, applied in liquid form. The composition of the sludge was analysed and confirmed to be bacteriologically safe and compliant with the Regulation on the protection of agricultural land against pollution regarding the permitted content of HMs in waste sludge intended for agricultural use [60]. A total of 340 g of waste sludge was added per EP, corresponding to the allowed annual amount of 1.66 tDM ha−1, as stipulated by the Regulation on sludge management from wastewater treatment plants when sludge is used in agriculture [61]. The sludge was alkaline (pHKCl = 8.5 in 1 M KCl in 1:2.5 (m/v) [50]) and contained 12.44% (bichromatic method), 0.72% nitrogen (N) [62], 80 mg 100 g−1 phosphorus (P2O5) (AL method), 65.83 mg 100 g−1 potassium (K2O), 400 mg kg−1 magnesium, and 384 mg kg−1 calcium (aqua regia extraction [63] and ICP-MS detection [64]). Total sulphur (TS) content was below 1.0% (dry combustion method [65]), while total Cd and Hg concentrations were below 0.100 mg kg−1 and 0.102 mg kg−1, respectively (mod. [66]). The third treatment involved the application of MYKOFLOR® (Konskowola, Poland), a commercial inoculum of live ectomycorrhizal mycelia suspended in AgroHydroGel® solution (Agroidea, Kraków, Poland). A total of 5 mL of mycorrhiza was applied per EP, which amounts to 15 mL for three replicates. AgroHydroGel® is a supersorbent polymer, capable of retaining water up to 300 times its weight. Available as 2–4 mm powder/granules, it forms a permanent gel after irrigation and has a near-neutral pH. Following the manufacturer’s recommendations for silt loam, 2 g AgroHydroGel® was applied per plant, amounting to 60 g for three replicates of 10 plants. The fourth treatment, which consisted of bottom ash generated from the combustion of MxG biomass, was incorporated into the soil at a 75:25 (soil:ash) ratio, with each pot receiving 13.5 kg of soil and 4.5 kg of ash.

2.5. Biomass Sampling and Growth Parameters

Aboveground biomass was collected from each experimental pot on 15 March 2019, 12 March 2020, and 18 March 2021 by manually cutting the plants at the base. Each sample included the entire aboveground biomass (including stems and dead leaves). For each sample, plant length (cm), yield (g per pot), and number of shoots (n per pot) were measured after the harvest. Samples were chopped, air-dried at room temperature, and later milled into a powder for further analysis. Moisture content was determined from subsamples after 24 h of drying at 105 °C in the oven to constant mass, after which corrections to dry matter (DM) were made. Belowground biomass (rhizomes) was sampled from each experimental pot only once, in the final year of the experiment. The mass of each sample was determined, and then the samples were cut, air-dried at room temperature, and milled into a powder for further analysis. Moisture content was determined as described for aboveground biomass.

2.6. Biomass Analysis

Sample preparation and determination of HMs in aboveground biomass were conducted at the Institute for Medical Research and Occupational Health (IMROH). Aliquots of dried and homogenised aboveground biomass (about 0.2 g) were digested in an UltraCLAVE IV digestion system (Milestone Srl, Sorisole, Italy) with the addition of 3 mL HNO3 (25%) and 7 mL HCl (6%). The final volume of all samples was 10 mL. The concentration of 111Cd in aboveground biomass was detected using ICP-MS (7500, Agilent Technologies, Tokyo, Japan) using rhodium as the internal standard and helium as collision gas. Concentrations of Hg in the aboveground biomass were detected using a Mercury Analyser (AMA-254, Leco Inc., Benton Harbor, MI, USA). The underground biomass (rhizomes) was analysed in the Central Laboratory for Chemical Technology, HEP-Proizvodnja d.o.o. For Cd determination, 0.4–0.5 g of each sample was digested (Multiwave PRO, Anton Paar, Graz, Austria) with 8 mL HNO3, and Cd concentrations were quantified by ICP-OES (ICPE-9000, Shimadzu, Kyoto, Japan) [67]. For Hg determination, samples were ground to a particle size of less than 1.0 mm, weighed into nickel containers (0.13 g per sample), and analysed using a Mercury Analyser (SMS 100, Perkin Elmer, Waltham, MA, USA) [67].

2.7. Mathematical Formulas

The enrichment coefficient (EC), or bioconcentration factor (BCF), was used to evaluate the plants’ ability to uptake and accumulate metals from soil into above- and belowground biomass:
EC   =   metal concentration in biomass [above/belowground] ,     mg DM   kg 1 metal concentration in soil ,   mg DM   kg 1
where DM stands for dry mass.
Based on the EC values, plants can be classified into four categories: high accumulators (EC 1–10), moderate accumulators (EC 0.1–1), low accumulators (EC 0.01–0.1), or non-accumulators (EC < 0.01) [68]. It can be used as a reliable way for quantifying relative differences in the bioavailability of heavy metals to plants [69]. In addition to the EC value, some studies (unlike ours) also calculate the enrichment factor (EF), which represents the ratio between the anthropogenic (e.g., agricultural or landfill) and geogenic contribution of an element, or more generally, the ratio between human- and naturally derived sources. For example, a study conducted in Morocco reported EF values for Cd ranging from 0.65 to 9.40, indicating strong anthropogenic enrichment of agricultural soils near the landfill [70].
The translocation factor (TF) was used to evaluate the plants’ ability to translocate metals from belowground to aboveground biomass:
TF   =   metal concentration in aboveground biomass ,     mg DM   kg 1 metal concentration in belowground biomass ,   mg DM   kg 1
When either EC or TF values are greater than 1, the plant is considered a hyperaccumulator and a suitable candidate for phytoremediation [71].

2.8. Statistical Analysis and Quality Control

All statistical analyses were performed using SAS software, version 9.1 (SAS Institute Inc., Cary, NC, USA). One-way analysis of variance (ANOVA) was employed to assess differences among treatments across all measured parameters. When ANOVA indicated statistically significant differences (p < 0.05), Fisher’s LSD post hoc test was performed. Quality control procedures were implemented throughout the study, including sample handling and laboratory analyses. Reference materials were used to verify the accuracy and precision of both soil (ISE 989) and plant material (IPE 225) measurements.

3. Results and Discussion

3.1. Growth Parameters

Figure 1, Figure 2 and Figure 3 and Table 1 illustrate the yield, plant length, and number of shoots observed during the study. Overall, aboveground biomass yields of MxG increased during the second and third years of cultivation across all treatments, with observed increases ranging from 15% to 61% compared to the first year. Regarding treatments, significant yield increases were observed only for treatments I (control) and III (mycorrhiza). Mycorrhizal application can have a positive impact on the reduction in stress caused by HMs and enhance nutrient availability and intake in plants [38]. During the first year of cultivation, the highest yield was achieved in treatment IV (MxG ash). In subsequent years, yield differences between treatments became less pronounced, with values ranging from 5.04 (I—control) to 5.59 tDM ha−1 (IV—MxG ash) in the second year and from 4.95 (IV—MxG ash) to 5.50 tDM ha−1 (III—mycorrhiza) in the third year. Comparable multi-year studies have reported a wide range of yields under HM contamination. For instance, Bilandžija et al. [23] and Zgorelec et al. [22] reported MxG yields between 6.3 and 15.5 tDM ha−1 in Cd- and Hg-contaminated soils, with yields declining by 37% to 55% in the second and third years compared to the first year. Such differences in yield between previous studies and the present work may likely be attributed to differences in soil contamination levels, amendment effects, and the use of different growing materials (seedlings vs. rhizomes). Szada-Borzyszkowska et al. [72] reported biomass yields of 16–17 tDM ha−1 in the third year and 20–24 tDM ha−1 in the fourth year (unpublished data) for MxG and novel seed-based hybrids with arbuscular mycorrhizal fungi (AMF) grown on soils contaminated with Pb, Cd, and Zn. In another three-year study [73], MxG and Sida hermaphrodita were cultivated in sand and loam soils contaminated with multiple HMs (Cd, Cu, Ni, Pb, and Zn) and fertilised with different NPK ratios across the years. The reported MxG yields were 1.85 tDM ha−1 (loam) and 4.78 tDM ha−1 (sand) in the first year, 6.76 tDM ha−1 (loam) and 19.35 tDM ha−1 (sand) in the second year, and 16.99 tDM ha−1 (loam) and 42.8 tDM ha−1 (sand) in the third year. These findings highlight the significant influence of soil type on biomass production. In our study, plant lengths ranged between 65 and 115 cm across treatments and years. Statistically significant increases in plant length were observed in treatments I (control), II (sludge), and III (mycorrhiza) during the second and third years in comparison to the first year. In the first year, plants in the treatment IV (MxG ash) were significantly taller than those receiving the mycorrhizal inoculant (III). Conversely, in the second year, MxG plants in treatment IV were significantly shorter than those in other treatments. By the third year, no significant differences in plant stem length were recorded among treatments. Supporting literature includes a pot experiment conducted in Obrenovac, Serbia, by Dražić et al. [74], where MxG was cultivated on Gleysol, ash from a thermal power plant disposal, and overburden from an open-pit coal mine containing various HMs (As, Cd, Co, Cr, Cu, Mn, Ni, Pb, Zn). Reported average plant stem heights in the first year were 54.44 cm (overburden), 73.67 cm (soil), and 87.12 cm (ash). Kocoń and Jurga [73] randomly selected and measured five MxG plant rootstocks per plot to assess the number of shoots per plant and the height of shoots before harvest in the third year, reporting shoot lengths of 177.2 cm (loam) and 272.5 cm (sand), with 43 (loam) and 69 (sand) shoots per plant. Comparatively, Dražić et al. [74] found lower tiller (side shoots) counts in plants grown on ash (3.78) and overburden (5.24) compared to those grown in soil (8.00), highlighting the significant impact of substrate on shoot development. In our study, shoot number ranged from 7 to 40 across all years: 7 to 14 in the first year, 19 to 40 in the second year, and 10 to 14 in the third year. A significant increase in the number of shoots was observed across all treatments in the second year.

3.2. Cadmium and Mercury Concentration in the Aboveground Biomass

As shown in Figure 4 and Figure 5 and Table 1, Cd and Hg concentrations in the aboveground biomass of MxG varied among treatments and across years. Statistical analysis confirmed that the vegetation year had a significant influence on the concentrations of both metals. Cd concentrations ranged from 5.98 (II—sludge, 2021) to 14.62 mg kg−1 (I—control, 2019), with a significant decrease observed in the third year across all treatments. Similarly, Hg concentrations varied from 41.8 μg kg−1 (II—sludge, 2021) to 383.9 μg kg−1 (I—control, 2019), with levels also generally declining in the third year for all treatments except IV (MxG ash). In the first year, plants grown in soils amended with sludge, mycorrhiza, or ash showed significantly lower Cd concentrations compared to the control treatment, though these differences were no longer significant in the second and third years. For Hg, only treatment IV (MxG ash) resulted in a significantly lower concentration in plants compared to other treatments in the first year. Ociepa-Kubicka et al. [75], who studied the impact of fertilisation with sewage sludge and various composts on HM content in MxG, reported Cd concentrations between 0.88 and 1.18 mg kg−1, with the highest levels found in plants cultivated on soils amended with 20 t ha−1 of sewage sludge. Cd concentrations in MxG aboveground biomass in our study were even higher and greatly exceeded both the typical value for combustion in solid biofuel production (TMxG—0.1 mg kg−1 [76]) and the regulatory limit for content in pellets (LMxG—0.5 mg kg−1 [77]). Likewise, Hg concentrations in all years and treatments surpassed the typical value for combustion of MxG in solid biofuel production (30 μg kg−1). Also, all treatments in the first and second years, except for treatment IV (MxG ash) in the first year, exceeded the limit value for the Hg in pellets (100 μg kg−1). In the third year, only treatment IV (MxG ash) slightly exceeded this limit. These results indicate that MxG biomass cultivated on highly contaminated soils is unsuitable for energy production without appropriate pretreatment. In cases of very high concentrations of HMs in the aboveground biomass, which was not the case in our study, the residual plant ash after combustion can be safely disposed of as hazardous waste in specialised landfills or processed for the biorecovery of precious and semiprecious metal, a practice known as phytomining [69].

3.3. Cadmium and Mercury Biomass Removal

Table 2 shows the removal of Cd and Hg by MxG aboveground biomass and the corresponding EC values for all treatments and years. Cd removal ranged from 31.1 to 70.5 g ha−1, while Hg removal ranged from 0.2 g to 1.7 g ha−1. These values are comparable to those reported by Barbu et al. [78], who measured heavy metal uptake between 35 and 55 g ha−1 in Miscanthus sinensis x giganteus cultivated on acidic soils polluted with Pb and Cd (Copșa Mică, Romania) and noted its ability to persist under such conditions. Lower annual Hg uptake (up to 0.79 g ha−1 per year) was reported in a three-year study by Bilandžija et al. [23] and Zgorelec et al. [22]. The EC values calculated in our study ranged from 0.060 to 0.146 for Cd, indicating low (EC 0.01–0.1) to moderate (EC 0.1–1.0) accumulation in aboveground biomass. Ociepa-Kubicka et al. [75] also reported moderate accumulation of Cd, Zn, and Ni in MxG aboveground biomass, with fertilisation exerting only a minor influence, while Pb accumulation remained low and largely unaffected. Similarly, Szada-Borzyszkowska et al. [72] found EC values for Cd consistently below 0.1 in the third and fourth growing seasons for MxG and novel seed-based miscanthus hybrids cultivated on an HM-contaminated field (Pb, Cd, Zn), suggesting limited soil-to-plant transfer of metals and supporting the use of miscanthus for phytostabilisation. These results are consistent with Barbu et al. [78], who concluded that miscanthus is not suitable for phytoextraction due to its low HM accumulation capacity. For Hg, EC values in this study ranged from 0.002 to 0.019, indicating that MxG is a very low accumulator or non-accumulator of Hg in aboveground biomass. Zhao et al. [71] reported a much broader EC range (0.365–225) for M. sinensis grown in EPs on soils spiked with varying Hg concentrations (1.48 mg kg−1 to 706 mg kg−1), with EC values decreasing as soil Hg concentration increased. Variations in plant species or cultivars, study duration, and the presence of added Cd and soil amendments may explain the differences in results between the studies.
The translocation of HMs from belowground to aboveground biomass is a key limiting factor for successful phytoremediation. In this study, the translocation factor (TF) was calculated in the final year after the removal and analysis of rhizomes. For Cd, TF values ranged between 0.104 (II—sludge) and 0.145 (IV—MxG ash) (Table 3). These values are slightly higher than those reported by Dražić et al. [74], who found Cd TFs of 0.065 and 0.090 for MxG grown on overburden and ash, respectively. Similarly, Szada-Borzyszkowska et al. [72] reported TF values below 1 for MxG and novel hybrids cultivated in HM-contaminated soils (Pb, Cd, Zn). These findings agree with Arduini et al. [79], who observed that the proportion of total Cd translocation to the aerial part is generally low in most plant species (generally less than 20%) and tends to decrease with increasing soil Cd concentrations. Regarding Hg, Zhao et al. [71] reported TF values between 0.998 and 2.14 for M. sinensis grown in soils spiked with varying concentrations of Hg. Specifically, for M. sinensis cultivated in soil with 24.4 mg kg−1 of Hg, which is comparable to the 20 mg kg−1 in our experiment, the TF was 2.13. In contrast, the Hg TF values for MxG in this study were substantially lower, ranging from 0.008 (II—sludge) to 0.024 (IV—MxG ash). These differences are likely attributable to varying tolerance mechanisms associated with species, applied amendments, and the concentrations of added metals.

3.4. Belowground Biomass (Rhizomes)

Table 4 presents the 2021 Cd and Hg concentrations in belowground biomass across all treatments. Compared to the corresponding aboveground values (Figure 4), belowground Cd levels were 7 to 10 times higher, depending on the treatment. These findings are consistent with Arduini et al. [79], who reported that Cd content in miscanthus typically follows the order: root > rhizomes > shoots > leaves. Similarly, Dražić et al. [74] reported significantly lower concentrations of As, Cd, Co, Cr, and Cu in stems of MxG compared to the rhizomes, with higher accumulation of Cd, Cr, and Pb in the belowground biomass relative to the substrate. Similarly, Szada-Borzyszkowska et al. [72] also found elevated HM concentrations in the roots and limited translocation to aerial parts, while Barbu et al. [78] reported higher Cd concentrations in rhizomes. Ociepa-Kubicka et al. [75] reported Cd concentrations in MxG roots ranging from 1.65 to 1.95 mg kg−1 in the third year of cultivation, depending on the type of fertilisation, with the highest levels reported for treatments amended with 20 and 40 t ha−1 of sewage sludge. A similar pattern was observed for Hg. Compared to aboveground biomass in 2021 (Figure 5), Hg concentrations in belowground biomass were substantially higher across all treatments: 63 times higher in treatment I (control), almost 132 times higher in treatment II (sludge), 59 times higher in treatment III (mycorrhiza), and nearly 42 times higher in treatment IV (MxG ash).
In the final year of the study, the mass of belowground biomass (Table 4) ranged from 41.31 g per pot (6.90 tDM ha−1) in treatment III (mycorrhiza) to 49.70 g per pot (8.30 tDM ha−1) in treatment I (control). No statistically significant differences were observed for any of the studied parameters shown in Table 4 (p ≥ 0.05). Arduini et al. [80] reported belowground biomass masses ranging from 27.9 g to 73.9 g per plant, with the lowest masses recorded under high Cd exposure (3.00 mg L−1). In our experiment, the mass of belowground biomass was 1.3 to 1.6 times greater than aboveground biomass yield in 2021 (Figure 1). ECs for Cd in belowground biomass ranged from 0.44 (IV—MxG ash) to 0.57 (II—sludge), and for Hg from 0.21 (IV—MxG ash) to 0.30 (I—control). All of these values fall within the moderate EC category (0.1–1), indicating moderate accumulation of both metals in MxG belowground biomass. This contrasts with the findings of Dražić et al. [74], who reported EC values greater than 1 and 3 for Cd in MxG belowground biomass grown on overburden and ash during the first year. However, it should be noted that the Cd concentrations in their substrates (0.71 mg kg−1 in overburden and 0.385 mg kg−1 in ash) were substantially lower than those used in our study (100 mg kg−1). Ociepa-Kubicka et al. [75] also observed high levels of Cd and Ni accumulation in MxG roots, with only minor differences between treatments.

4. Conclusions

A significant reduction in Cd and Hg concentrations in the aboveground biomass was observed in the third year across all treatments except IV (MxG ash). In most treatments and study years, Cd and Hg concentrations in the aboveground biomass exceeded both the typical values for combustion of MxG in solid biofuel production and the regulatory limits for pellet content. Although biomass yields slightly increased compared to the first year, they remained relatively low across treatments, suggesting that the extreme soil conditions outweighed the effects of the applied amendments. Annual removal rates were low (31.1–70.5 g ha−1 for Cd; 0.2–1.7 g ha−1 for Hg), and both EC and TF values were <1, indicating that MxG is unsuitable for phytoextraction or phytomining but shows potential for phytostabilisation and erosion minimisation (low aboveground biomass yield and up to 2.2 times higher belowground biomass). Cultivation of MxG under extreme soil Cd and Hg contamination may therefore achieve stabilisation and reduce environmental risks; however, yields will remain low, and biomass must undergo pretreatment before bioenergy use due to metal concentrations exceeding typical and regulatory limits. Further research, including field-based studies employing a range of HM concentrations and soil amendments, is recommended to better elucidate the effects of soil amendments on the phytostabilisation efficiency of MxG.

Author Contributions

Ž.Z.: conceptualisation (lead); data curation (lead); resources (lead); supervision (lead); methodology (lead); funding acquisition (lead); writing—review and editing (lead). L.Z.: formal analysis (supporting); investigation (supporting); visualisation (lead); writing—original draft preparation (lead); writing—review and editing (supporting). N.B.: conceptualisation (lead); data curation (lead); resources (supporting); methodology (lead); funding acquisition (lead); writing—review and editing (supporting). S.Ž.: validation (lead); formal analysis (supporting); investigation (supporting); writing—review and editing (supporting). A.Š.Ć.: data curation (supporting); M.T.: Validation (lead); formal analysis (supporting); investigation (supporting); writing—review and editing (supporting). Z.K.-G.: validation (lead); formal analysis (supporting); investigation (supporting); writing—review and editing (lead). M.G.: validation (supporting); formal analysis (supporting); investigation (supporting). I.H.: formal analysis (supporting); investigation (supporting). A.P.: validation (supporting); formal analysis (supporting); investigation (supporting); writing—review and editing (supporting). All authors have read and agreed to the published version of the manuscript.

Funding

This research received no external funding.

Data Availability Statement

The data supporting this study’s findings are available from the corresponding author upon reasonable request.

Conflicts of Interest

Author Marija Trkmić was employed by the company HEP Proizvodnja d.o.o. Author Ana Špehar Ćosić was employed by the company Agroproteinka d.d. The remaining authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Figure 1. Yield of MxG (tDM ha−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
Figure 1. Yield of MxG (tDM ha−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
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Figure 2. Length of MxG plants (cm) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
Figure 2. Length of MxG plants (cm) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
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Figure 3. Number of MxG shoots across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
Figure 3. Number of MxG shoots across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash.
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Figure 4. Cd concentrations in aboveground biomass (mg kg−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash. TMxG—Typical value of Cd for combustion of MxG in solid biofuel production; LMxG—Limit value for content of Cd in pellets.
Figure 4. Cd concentrations in aboveground biomass (mg kg−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash. TMxG—Typical value of Cd for combustion of MxG in solid biofuel production; LMxG—Limit value for content of Cd in pellets.
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Figure 5. Hg concentrations in aboveground biomass (μg kg−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash. TMxG—Typical value of Hg for combustion of MxG in solid biofuel production; LMxG—Limit value for content of Hg in pellets.
Figure 5. Hg concentrations in aboveground biomass (μg kg−1) across treatments and study years. Means sharing the same capital letter (across years) or lowercase letter (across treatments) are not significantly different (Fisher’s test, p < 0.05). Abbreviations: I—control; II—sludge; III—mycorrhiza; IV—MxG ash. TMxG—Typical value of Hg for combustion of MxG in solid biofuel production; LMxG—Limit value for content of Hg in pellets.
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Table 1. ANOVA results for measured parameters (aboveground biomass).
Table 1. ANOVA results for measured parameters (aboveground biomass).
ParameterFactorp-ValueLSD
YieldYear20190.0131.26
20200.7521.29
20210.6011.08
TreatmentI0.0080.71
II0.0881.27
III0.0201.46
IV0.5031.55
LengthYear20190.0168
20200.01125
20210.70027
TreatmentI0.00316
II0.00312
III0.00220
IV0.46237
No. of shootsYear20190.0023
20200.01512
20210.4616
TreatmentI0.01210
II0.0057
III0.0368
IV0.00017
CdYear20190.0372.83
20200.5514.12
20210.8132.77
TreatmentI0.0033.01
II0.0012.47
III0.0293.56
IV0.0064.61
HgYear20190.005125.9
20200.332229.0
20210.40487.5
TreatmentI0.003121.4
II0.009257.0
III0.005163.8
IV0.87580.4
p-values < 0.05 are considered statistically significant (bolded). LSD—Least Significant Difference.
Table 2. Removal of Cd and Hg by MxG aboveground biomass (in g ha−1) and enrichment coefficient (EC) according to treatments and studied years.
Table 2. Removal of Cd and Hg by MxG aboveground biomass (in g ha−1) and enrichment coefficient (EC) according to treatments and studied years.
Removal, g ha−1EC
Treatment201920202021Treatment201920202021
Cd
I55.955.635.9I0.1460.1100.071
II45.070.532.3II0.1060.1280.060
III39.456.734.7III0.1140.1020.063
IV52.365.631.1IV0.1070.1170.064
Hg
I1.51.10.5I0.0190.0110.005
II1.11.70.2II0.0130.0150.002
III1.01.40.5III0.0150.0130.005
IV0.50.60.5IV0.0050.0060.005
Table 3. Translocation factor (TF) for Cd and Hg for the final year of the study (2021).
Table 3. Translocation factor (TF) for Cd and Hg for the final year of the study (2021).
TFIIIIIIIV
Cd0.1250.1040.1270.145
Hg0.0160.0080.0170.024
Table 4. Mass of belowground biomass (rhizomes) (g per pot, tDM ha−1), Cd and Hg concentrations (mg kg−1), and enrichment coefficient (EC) in the final year of the study (2021).
Table 4. Mass of belowground biomass (rhizomes) (g per pot, tDM ha−1), Cd and Hg concentrations (mg kg−1), and enrichment coefficient (EC) in the final year of the study (2021).
Treatment
Belowground BiomassIIIIIIIV
Mass (g per pot)49.7044.3941.3144.01
Mass (tDM ha−1)8.307.406.907.30
Cd concentration (mg kg−1)57.157.249.644.3
Hg concentration (mg kg−1)6.055.545.634.24
EC (Cd)0.570.570.500.44
EC (Hg)0.300.280.280.21
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Zgorelec, Ž.; Zubčić, L.; Žužul, S.; Kljaković-Gašpić, Z.; Trkmić, M.; Galić, M.; Hrelja, I.; Špehar Ćosić, A.; Perčin, A.; Bilandžija, N. High Cadmium and Mercury Soil Contamination Outweighs the Effect of Soil Amendments When Growing Miscanthus x giganteus. Appl. Sci. 2025, 15, 9075. https://doi.org/10.3390/app15169075

AMA Style

Zgorelec Ž, Zubčić L, Žužul S, Kljaković-Gašpić Z, Trkmić M, Galić M, Hrelja I, Špehar Ćosić A, Perčin A, Bilandžija N. High Cadmium and Mercury Soil Contamination Outweighs the Effect of Soil Amendments When Growing Miscanthus x giganteus. Applied Sciences. 2025; 15(16):9075. https://doi.org/10.3390/app15169075

Chicago/Turabian Style

Zgorelec, Željka, Lana Zubčić, Silva Žužul, Zorana Kljaković-Gašpić, Marija Trkmić, Marija Galić, Iva Hrelja, Ana Špehar Ćosić, Aleksandra Perčin, and Nikola Bilandžija. 2025. "High Cadmium and Mercury Soil Contamination Outweighs the Effect of Soil Amendments When Growing Miscanthus x giganteus" Applied Sciences 15, no. 16: 9075. https://doi.org/10.3390/app15169075

APA Style

Zgorelec, Ž., Zubčić, L., Žužul, S., Kljaković-Gašpić, Z., Trkmić, M., Galić, M., Hrelja, I., Špehar Ćosić, A., Perčin, A., & Bilandžija, N. (2025). High Cadmium and Mercury Soil Contamination Outweighs the Effect of Soil Amendments When Growing Miscanthus x giganteus. Applied Sciences, 15(16), 9075. https://doi.org/10.3390/app15169075

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