Mercury (Hg) Speciation in the Soil–Plant System of Formerly Polluted Soils

Abstract

The distribution of mercury (Hg) species in soils, as affected by (i) the physicochemical and biological properties of the soils, (ii) the lithogenic vs. anthropogenic sources of soil Hg pollution, and (iii) soil–plant interactions, was investigated in a model pot experiment in which Sinapis alba (Brassicaceae) was planted. The pseudototal (aqua regia-soluble) Hg contents in soils originating from the vicinity of the former cinnabar mine varied between 16.6 and 44.7 mg/kg, whereas in soils from sites where the amalgamation technique had been used for the extraction of gold from ore-bearing materials, the pseudototal Hg values varied between 1.63 and 10.1 mg/kg. However, Hg accessibility was low, with mobilizable Hg(II) accounting for 3–11% of total soil Hg and mobilizable methylmercury (MeHg) remaining below 1%, indicating a limited bioavailable pool under the studied conditions. Mobilizable Hg(II) showed significant negative relationships with total soil carbon and cation exchange capacity (CEC), reflecting its strong association with charged functional groups of the soil sorption complex. The low Hg accessibility in the soil resulted in low Hg contents in plants, not exceeding the feed safety thresholds, with a significant proportion of Hg taken up by the plants being retained in the roots. The results of the determination of gaseous elemental mercury (GEM) indicated its relevance in soil mercury cycling, where further research on the role of plants in GEM emissions is necessary. In this context, the GEM concentrations increased in plants found in soils collected close to the former cinnabar mine. These aspects should be investigated further in future studies.

1. Introduction

Soil is a dynamic and complex system whose physical and chemical properties influence the retention of risk elements (REs) in the soil and their availability to plants. The main factors influencing the dynamics of RE availability include soil pH and the soil’s ion-exchange capacity, reflecting both cation and anion exchange, as well as surface complexation processes governing the retention of charged RE species. Additional controlling factors are the contents and quality of carbon (C), nitrogen (N), and sulfur (S), as well as their soluble fractions, which play an important role in regulating the biological availability of REs. Moreover, the C and N contents in soil microbial biomass serve as indicators of microbial abundance and overall soil biological quality [1,2]. Understanding the relationships between soil properties and the behavior of REs is the first and essential step in assessing the environmental impacts of soil contamination and designing effective methods for its decontamination [3].

Among the most intensively investigated REs, mercury (Hg) ranks among the most concerning due to its extreme toxicity to living organisms. The primary mineral source of Hg is cinnabar, and cinnabar mining can result in environmental pollution by Hg [4]. Among anthropogenic activities, gold mining based on amalgamation techniques constitutes one of the most significant global sources of Hg contamination. For instance, Domingues et al. [5] reported massive Hg pollution in such vulnerable areas as the Amazon River Basin. Additional significant sources of mercury emissions are coal combustion, metal smelting, cement production, and waste incineration. In the environment, soil serves as the largest terrestrial reservoir of mercury [6]. The average Hg content in soil is estimated to range from 0.03 to 0.3 mg/kg, with contaminated soils having contents up to four times higher [7]. The most important geochemical properties of mercury are the volatility of the metallic form, strong affinity for sulfur, the tendency to form organometallic compounds, low affinity for oxygen bound to nitrogen, and the ability to form gaseous species [8].

Mercury sorption in soil is controlled by a number of soil factors such as soil pH, soil organic matter (SOM), clay components, and Hg speciation in soil [6]. The mobility of Hg in the soil environment can be increased by sorption/complexation with fulvic acids and hydrophilic compounds, collectively referred to as colloidal organic matter. The binding of Hg to humic substances is considered the primary process in mercury sorption [7]. The importance of fulvic acids is primarily attributed to the early stages of mercury stabilization. Soils with higher pH values also show a higher percentage of mobile mercury fractions [9,10]. Moreover, the size of soil aggregates has a significant influence on the Hg content in the soil as well. When evaluating soil types, the highest mercury contents were found in clay soils with an average concentration of 0.170 mg/kg, followed by fine silt with 0.130 mg/kg and coarse silt with 0.080 mg/kg. The lowest average contents were recorded in fine sand, with a value of 0.032 mg/kg, and coarse sand, with 0.014 mg/kg [7].

Cinnabar (α-HgS) is formed under anaerobic conditions when sulfide combines with mercury. Cinnabar is chemically stable, often occurs in association with gold, and can represent the predominant form of Hg in soil; however, it can also dissolve, forming aqueous Hg sulfide species [11]. Thus, sulfur also controls the speciation of inorganic Hg and, consequently, its availability for methylation [12]. Oxidized mercury species can occur in the atmosphere in gaseous and aerosol forms and are subsequently deposited through precipitation or as dry deposition onto aquatic and terrestrial surfaces. A key aspect of the mercury biogeochemical cycle is the oxidation of elemental Hg(0). Gaseous elemental mercury (GEM) can persist in the atmosphere long enough to allow global transport. Hg(II), on the other hand, is more soluble and thus leaves the atmosphere more quickly [13].

Methylmercury (MeHg) is a persistent organic form of mercury known for its high capacity for absorption by organisms. It is the most common and also the most toxic organic form of this element. Mercury methylation occurs primarily in water, sediments, and, in some cases, on the roots of macrophytes. Methylmercury is formed mainly through the microbial conversion of divalent Hg by anaerobic bacteria, including sulfate- and iron-reducing bacteria and fermenters [14]. However, the soil physicochemical parameters can affect the Hg methylation rate. For instance, Zhang et al. [15] reported that elevated Mn contents in soil inhibit Hg methylation, and Wang et al. [16] found that soils treated with iron-based materials can alter MeHg formation. Microbial methylmercury formation depends on environmental conditions, chiefly the bioavailability of Hg(II), which is largely controlled by the ratio of reduced sulfur species [17].

Plants play an essential role in the Hg biogeochemical cycle; however, their involvement can also result in toxic effects on plant physiology. Mercury has no known beneficial role in animals or plants. The toxicity of this metal is determined by its binding to the thiol (–SH) groups of proteins, whereby proteins are disrupted and essential elements are displaced. Even low concentrations of mercury can lead to a reduction in the growth potential of plants. Among soil properties determining plant-available Hg in soils, the importance of total soil Hg content, pH, and SOM content has been highlighted [18]. No Hg-hyperaccumulating species have been identified within the plant kingdom [19]. However, elevated Hg contents in various plant species have been reported at extremely Hg-polluted locations. For instance, Fernández et al. [20] measured Hg contents of 84 mg/kg for Festuca rubra, 78 mg/kg for Leontodon taraxacoides, and 77 mg/kg for Equisetum telmateia. Similarly, pot experiments by Xun et al. [21] showed that the fern Cyrtomium macrophyllum can accumulate up to 36 mg/kg of mercury.

Moreover, rhizosphere processes and interactions between endophytes and Hg can affect the soil–plant relationships, Hg uptake and accumulation by plants, and Hg toxicity in plants [22]. As mentioned by Szponar et al. [23], Hg uptake by vegetation can effectively remove GEM from the atmosphere. They consider GEM to be a substantial source for Hg in terrestrial ecosystems. On the other hand, it is well known that dense, herbaceous, vegetative cover can effectively reduce GEM concentrations at the soil–air interface, as well as the presence of trees and litter [24]. The importance of foliar uptake of GEM by terrestrial plants, affecting the biogeochemical cycle of Hg, was highlighted by Liu et al. [25]. In this context, the principles governing Hg uptake, translocation, transformation, and re-emission are not yet fully elucidated. This study aimed to investigate (i) the transformation and speciation of Hg in the soil–air–plant interface in Hg-polluted soils, (ii) the assessment of the role of various soil physicochemical and biological parameters in the fate of Hg and its species in soils, and (iii) the comparison of Hg behavior in anthropogenically vs. lithogenically polluted soils.

2. Materials and Methods

2.1. Pot Experiment

Soils for the pot experiment were collected from two locations in the Czech Republic where enhanced Hg levels in these soils had previously been reported [26,27]. The first of these locations was an abandoned cinnabar mine in Horní Luby, which operated mainly in the 16th century. The source of Hg at this location was therefore lithogenic, with cinnabar being the predominant source of Hg. In contrast, the Libčice site was an abandoned gold mine where the amalgamation technique was used to extract gold from ore. The main source of Hg at this location was therefore anthropogenic. A detailed description of the locations, mining history, geological characteristics, the location of monitoring sampling points, physicochemical and biological characteristics of the soils, and the vegetation cover has already been presented in previous papers [26,27,28,29]. At each of these locations, three composite soil samples were collected from a depth of 0–25 cm at the previously monitored sites for the pot experiment. The choice of the sampling points was based on the Hg contents in the soils to achieve soils with different levels of total Hg. The soils were air-dried, ground, sieved through a 2 mm mesh sieve, homogenized, and labeled I, II, and III. The experiment therefore consisted of six variants, three (I, II, III) per location.

The pot experiment was carried out as follows: 500 g of soil was weighed into plastic pots and regularly watered up to 60% of the maximum water-holding capacity for 21 days to stimulate microbial activity. Each variant was prepared in four replicates. After the first GEM measurement (see details below), 15 white mustard (Sinapis alba L.) seeds were sown in each of the three pots within one variant. In each variant, one pot was always left without plants to evaluate the effect of the plants on mercury mobility in the soil. Thus, the experiment consisted of total 24 pots, 18 pots with S. alba plants and 6 pots without plants. The S. alba plants were chosen as a model organism commonly used in ecotoxicological experiments. After germination, the number of plants was thinned to the six strongest individuals to ensure optimal growth potential. The aboveground biomass of the plants was harvested at the flowering stage, washed, dried at laboratory temperature, and homogenized (IKA-Werke GmbH & Co. KG, Staufen, Germany) before analysis. After harvesting, the GEM concentrations were measured again, and soil samples and plant roots were taken from the whole soil profile. The roots were processed in the same way as the aboveground biomass. Soil samples were divided into two parts, where one fresh sample was immediately used for the determination of available nitrogen, total dissolved carbon, and microbial biomass C and N. The remaining soil was freeze-dried, homogenized, and sieved through a 2 mm sieve. These samples were used for the determination of risk elements and mercury species.

2.2. Analytical Methods

2.2.1. Main Soil Properties and Nutrient Contents

Cation exchange capacity (CEC) of the soils was determined based on the contents of basic cations determined in the extract with a 0.1 mol/L BaCl₂ solution according to the ISO standard [30]. The concentration of cations in the solution was determined using inductively coupled plasma optical emission spectrometry (ICP–OES, Agilent 720, Agilent Technologies Inc., Santa Clara, CA, USA). For the determination of soil pH, 5 g of the sample was weighed and mixed at a ratio of 1:10 with 50 mL of 0.01 mol/L CaCl₂ in a 250 mL plastic bottle. The prepared solution was further shaken on a horizontal shaker for 60 min, after which it was allowed to settle, and the pH was measured using a digital pH meter. The total C, N, and S contents in the soil were determined using the CHNOS Vario MACRO cube instrument (Elementar Analysensysteme GmbH, Langenselbold, Germany). To determine the easily mobilizable concentrations of selected elements, namely Fe, Mn, and S, the Mehlich III extraction procedure was performed by shaking the soil and the extractant (0.2 mol/L CH₃COOH, 0.25 mol/L NH₄NO₃, 0.013 mol/L HNO₃, 0.015 mol/L NH₄F, and 0.001 mol/L ethylenediaminetetraacetic acid, EDTA) at a ratio of 1:10 (w/v) for 10 min [31]. The total dissolved carbon and available nitrogen were determined in fresh soil samples spectrophotometrically using a SKALAR San Plus System continuous-flow segmented analyzer (Skalar, Breda, The Netherlands).

2.2.2. Microbial Biomass Carbon (MBC) and Microbial Biomass Nitrogen (MBN)

Microbial biomass carbon (MBC) and microbial biomass nitrogen (MBN) were determined by the fumigation method in version published by Gregorich et al. [32]. In this method, 10 g of fresh soil samples were placed into a vacuum desiccator, where the samples were exposed to chloroform vapors for 24 h. After that, the total organic carbon (TOC) was extracted using a 0.5 mol/L K₂SO₄ solution at a ratio of 1:4 (w/v), shaken for 60 min, and determined using the above-mentioned SKALAR San Plus System [33]. The MBC was calculated as C = EC/kEC, where EC = (organic C extracted from fumigated soils) − (organic C extracted from non-fumigated soils) and kEC = 0.45 [34]. The MBN was calculated as N = EN/kEN, where EN = (total N extracted from fumigated soils) − (total N extracted from non-fumigated soils) and kEN = 0.54 [35].

2.2.3. Risk Element Contents in Soils and Plants

For the determination of the pseudototal content of risk elements, soil samples were extracted as follows: homogenized soil samples (0.25 g) were weighed with an accuracy of 0.1 mg into quartz tubes (35 mL). To the samples, aqua regia consisting of 4.5 mL HCl and 1.5 mL HNO₃ (both Analpure^®, Analytika, Prague, Czech Republic) was added. The samples were subsequently closed with a Teflon cap and subjected to extraction in a microwave digestion system (Discover SP-D, CEM Corp., Matthews, NC, USA). The obtained acidic soil extracts were diluted with ultrapure water (≥18.2 MΩ/cm; Milli-Q system, Merck Millipore, Darmstadt, Germany) after cooling to a final volume of 45 mL.

Dry samples of plant biomass (0.1 g) were weighed with an accuracy of 0.1 mg into quartz tubes (35 mL). To the samples, a mixture of 4 mL HNO₃ (Analpure^®, Analytika, Czech Republic) and 2 mL H₂O₂ (Rotipuran^®, Carl Roth, Karlsruhe, Germany) was added. The samples were then closed with a Teflon cap and subjected to decomposition in a microwave decomposition system (Discover SP-D, CEM Corp., USA). The obtained acidic plant digests were diluted after cooling with ultrapure water (≥18.2 MΩ/cm; Milli-Q system, Millipore, SAS, Molsheim, France) to a final volume of 45 mL.

For the extraction of the mobile fraction of mercury in soil, a modified method validated by Han et al. [36] was used. Soil samples (1 g aliquots) were extracted in three successive steps using a solution of 10% (v/v) ethanol (p.a., Penta s.r.o., Prague, Czech Republic) in 2% (w/v) HCl (Analpure^®, Analytika s.r.o., Czech Republic). The first extraction (7 mL of extraction agent) was initiated by intensive manual shaking (10 s) and vortexing (10 s). Then the soil suspension samples were immersed in an ultrasonic bath (55 ± 3 °C) for 10 min. Subsequently, the samples were centrifuged (2700× g) for 5 min and filtered (KA-5, Papírna Perštejn, s.r.o., Perštejn, Czech Republic). The extraction was repeated twice, each time with a new portion of the extraction agent (4 mL). The individual filtrates were combined, 3-fold diluted with deionized water (≥18.2 MΩ/cm, Milli-Q system, Merck Milipore, Germany) and stored at −20 °C until analysis.

The elemental concentrations in the plant digests and soil extracts were measured by inductively coupled plasma mass spectrometry (ICP–MS; Agilent 8900, Agilent Technologies Inc., USA). The ICP-MS configuration included an autosampler SPS 4, a concentric nebulizer MicroMist, a double-walled Scott-type spray chamber, an octopole collision/reaction cell, and two quadrupole mass analyzers. Isotopes ⁹Be, ¹¹¹Cd, ²⁰¹Hg, and ²⁰⁸Pb were measured in no-gas mode, while other isotopes (⁵¹V, ⁵²Cr, ⁵⁹Co, ⁶⁰Ni, ⁶³Cu, ⁶⁶Zn, ⁷⁵As) were measured in helium collision mode, and the isotope ³²S, more precisely its oxidized species ⁴⁸SO⁺, was measured in reaction mode with oxygen. The instrument was calibrated using the external calibration method with suitably diluted mixed standard solution. For the correction of matrix effects and signal drift over time, the internal standard method was used, employing a multi-element solution containing 100 µg/L Sc, In, Ge, Lu, and Bi (ASTASOL-MIX, Analytika, Czech Republic), which was continuously introduced into the nebulizer together with the samples. For the determination of Hg, the digests and extracts were diluted before analysis with 0.18% (w/v) L-cysteine (≥98%, Carl Roth Ltd., Germany) in 2% HNO₃ [37], which facilitated rinsing of Hg from the analytical instrument and thus eliminated memory effects. The accuracy of ICP–MS measurements was verified by the analysis of certified reference material SRM 1641d (Mercury in water, NIST), CRM AN-BM02 (Green tea, Analytika s.r.o.), and RM Metranal^®1 (River sediment, Analytika s.r.o.).

2.2.4. Mercury Speciation in Soils

For the determination of Hg species, aliquots of the diluted soil extracts were additionally filtered (membrane filter, 0.45 µm, nylon), diluted with the mobile phase (see below) at a volume ratio of 1:1, and subjected to speciation analysis using a coupled technique of high-performance liquid chromatography (HPLC; Agilent 1260, Agilent Technologies Inc., USA) with ICP–MS as an element-specific detector [36]. The chromatograph was equipped with a C18 Eclipse Plus column (100 × 4.6 mm, 3.5 µm; Agilent Technologies Inc., USA). Elution was performed under isocratic conditions with a mobile-phase flow rate of 1.2 mL/min. The composition of the mobile phase was taken from Jagtap et al. [38] with slight modifications: 0.1% (v/v) 2-mercaptoethanol, 0.04 mol/L ammonium acetate (NH₄CH₃COO, ≥97%, Carl Roth, Germany) in 5% (v/v) methanol (ROTISOLV^®, Carl Roth, Germany), pH = 5.25 (adjusted with acetic acid). The sample injection volume was 20 µL, and the column was maintained at 30 °C. Calibration solutions of 0.05–10 µg/L Hg of individual Hg species, Hg(II) and MeHg, were prepared by dissolving and diluting HgCl₂ (≥99.5%, Honeywell-Fluka, Seelze, Germany) and CH₃HgCl (≥98%, Sigma-Aldrich, St. Louis, MO, USA).

2.2.5. Gaseous Elemental Mercury

GEM measurements were performed using a single-purpose portable atomic absorption spectrometer (AAS) Lumex RA-915+ (Lumex Ltd., St. Petersburg, Russia) in two stages, before sowing white mustard and one day after the harvest of the aboveground biomass. The measurement was performed at a wavelength of 253.7 nm. The air flow through the cuvette was 20 L/min, the detection limit was 2 ng/m³, the range was up to 20,000 ng/m³, and the method bias was ±20%. The instrument had an internal calibration set and was verified by the manufacturer. The concentration of mercury vapor in the greenhouse was measured before the start and after the end of the measurement (background). To obtain comparable results, all GEM measurements were performed using the same procedure: in each experimental vessel (pot), the soil was loosened to a depth of approximately 3 cm before the GEM measurement, and a cylindrical polyethylene chamber without a bottom, closed with a lid and with a volume of approximately 1.5 L, was immediately placed over the soil. After checking the zero signal, the gas supply hose to the analyzer was inserted into the sampling chamber, and after 4 min the maximum concentration value was measured, and after approximately 10 s, the stabilized GEM value. The air sample inside the sampler was drawn from a height of 1–2 cm above the soil surface. The samplers were decontaminated by washing with demineralized water and dried before use [39,40].

2.3. Data Analysis

The data were processed using Microsoft Excel version 2024 (Microsoft Corporation, Redmond, WA, USA) and Statistica version 14.0.0.15 (StatSoft, Inc., Tulsa, OK, USA) programs. Basic graphical analyses and visualization were carried out in Excel. Statistical analyses, including t-tests, one-way analysis of variance (ANOVA) followed by Tukey’s post hoc test, and Pearson correlation analyses, were conducted using Statistica. All statistical tests were performed at the significance level of p = 0.05.

3. Results and Discussion

3.1. Soil Properties and Elemental Contents

Among the soil physicochemical parameters, pH and CEC were determined (Figure 1). The pH levels at the Horní Luby location were strongly acidic (Figure 1A) and differed significantly (p < 0.05) from those in the Libčice soil, which had a near-neutral pH. Similarly to the case of soil pH, the CEC levels at the Libčice location were significantly (p < 0.05) higher compared with those at the Horní Luby location (Figure 1B). No significant differences in total or available N and C contents were observed between the two locations when the data were aggregated at the location level (mean ± SD). Total N contents were 0.36 ± 0.03% and 0.38 ± 0.03% at Horní Luby and Libčice, respectively, while available N contents were 60.2 ± 17.2 mg/kg and 31.0 ± 10.9 mg/kg (Figure 2A,B). Similarly, total C contents reached 4.92 ± 0.42% and 6.37 ± 0.67% at Horní Luby and Libčice, respectively, and available C contents were 87.2 ± 16.6 mg/kg and 75.4 ± 24.7 mg/kg (Figure 3A,B). However, site-specific differences were detected; for example, soils II and III from the Horní Luby location exhibited significantly higher available N contents (p < 0.05; Figure 2B) compared with soils from the Libčice location.

Figure 1. Soil pH (A) and cation exchange capacity (B) according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

Figure 2. Total (A) and available, i.e., 0.01 M CaCl₂ soluble (B) nitrogen content in soils according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

Figure 3. Total (A) and available, i.e., 0.01 mol/L CaCl₂ soluble (B) carbon content in soils according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

Sulfur is one of the important elements influencing the fate of Hg in soils. Total S contents were 0.07 ± 0.01% and 0.11 ± 0.02% at Horní Luby and Libčice, respectively, while available S contents were 37.9 ± 10.9 mg/kg and 16.8 ± 7.7 mg/kg (Figure 4A,B). In this respect, total S contents showed extremely high variability among samples from Libčice, with significant (p < 0.05) differences between individual samples, whereas S contents in soils from the Horní Luby location were comparatively homogeneous (Figure 4A). Although no significant (p < 0.05) differences were recorded between the aggregated available S contents at the two locations (Figure 4B), the available S contents at the Horní Luby location tended to show higher values.

Figure 4. Total (A) and available, i.e., Mehlich III [31] extractable (B) sulfur content in soils according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

To estimate the biological status of the soils, microbial biomass nitrogen (MBN) and microbial biomass carbon (MBC) were assessed (Figure 5), with no significant differences recorded (p < 0.05). Although MBN/MBC are only proxy approaches giving informative values only, the data can indicate no substantial effect of the soil Hg pollution on the soil microbial community as a whole. These data corresponded with the findings published by Száková et al. [28], in which no significant (p < 0.05) reduction in individual soil enzymatic activities was observed, indicating a negligible harmful effect of long-term soil Hg pollution on soil microbial activity.

Figure 5. Microbial biomass nitrogen (A) and carbon (B) content in soils according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

Table 1 shows the pseudototal contents of REs in soils. These values fully correspond to the previous findings [28], where Hg was identified as the main pollutant at both locations, exceeding the preventive soil mercury levels given in the Czech Republic by Public Notice No. 153/2016 [41], i.e., 0.3 mg/kg. Higher Hg values in the soils from the Horní Luby location, compared with Libčice, and the heterogeneous character of Hg within one location were also confirmed. Moreover, all the Hg contents exceeded the indicative values at which the soil element contents represented a potential risk for crop contamination (i.e., 1.5 mg/kg of Hg). In the soils from the Horní Luby location, the highest observed Hg contents could have directly threatened human and animal health (indicative values: 20 mg/kg of Hg). Corzo et al. [42] monitored the risk element contents close to an abandoned mercury mine that had been operated for almost four centuries. The contents of Hg in soils were lower compared with this study (they varied between 1.7 and 6.2 mg/kg), indicating that the mercury ores had been effectively extracted. On the contrary, this experiment indicated the presence of potential supplies of Hg ore deposits at the Horní Luby location. The correlation analysis did not confirm any significant (p < 0.05) relationships between pseudototal soil Hg content and the measured soil physicochemical and biological parameters.

Table 1. Pseudototal (Aqua regia soluble) contents of risk elements in soil according to the individual variants (n = 4); data presented as mean ± standard deviation.

		Be	V	Cr	Co	Ni	Cu	Zn	As	Cd	Hg	Pb
		mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg
Horní Luby	I	0.926 ± 0.068	24.6 ± 0.6	16.8 ± 0.2	8.19 ± 0.26	18.8 ± 0.0	25.7 ± 2.2	104 ± 7	10.8 ± 0.4	0.221 ± 0.027	18.6 ± 4.0	51.5 ± 1.8
	II	0.700 ± 0.018	18.8 ± 0.4	14.1 ± 0.6	13.4 ± 1.2	20.9 ± 1.1	27.2 ± 0.8	92.6 ± 4.5	10.8 ± 0.2	0.144 ± 0.012	16.6 ± 1.1	39.9 ± 1.8
	III	1.11 ± 0.04	20.4 ± 0.9	13.9 ± 0.2	10.4 ± 1.1	17.0 ± 0.7	21.1 ± 0.8	192 ± 6	13.2 ± 0.6	0.233 ± 0.007	44.7 ± 3.8	89.1 ± 1.6
Libčice	I	0.677 ± 0.051	52.1 ± 0.6	17.9 ± 1.3	11.5 ± 1.1	12.0 ± 0.7	48.4 ± 5.0	76.8 ± 2.1	68.6 ± 5.7	0.384 ± 0.019	10.1 ± 1.0	27.0 ± 1.8
	II	0.454 ± 0.025	45.9 ± 0.9	14.9 ± 0.5	9.86 ± 0.21	9.02 ± 0.31	27.1 ± 6.7	74.9 ± 2.4	23.2 ± 2.9	0.281 ± 0.025	1.63 ± 0.17	13.6 ± 0.8
	III	0.713 ± 0.002	47.8 ± 1.2	23.7 ± 0.9	10.2 ± 0.3	15.2 ± 0.7	43.6 ± 1.8	108 ± 2	43.8 ± 2.2	0.544 ± 0.029	3.71 ± 0.19	67.1 ± 9.7

However, mining activity could result in elevated contents of As, Cd, Pb, Zn, Ni, and Cu, accompanying the mercury ore deposit [42]. Soil contamination with multiple risk elements was previously recorded at these locations [28] and was also confirmed in this pot experiment. Table 1 shows elevated contents of Zn and Pb in the soils from the Horní Luby location (maximum levels exceeded the preventive values, i.e., 60 mg/kg for Pb and 120 mg/kg for Zn), and occasionally elevated Pb and Cd contents (preventive value: 0.5 mg/kg) at the Libčice location. In the case of As, the preventive value (i.e., 20 mg/kg) was exceeded in all three composite samples from the Libčice location, and the indicative value of 40 mg/kg of As, at which human and animal health could be directly threatened, was exceeded in two of these samples [41].

3.2. Growth Parameters and Hg Uptake by Plants

The growth and development of the S. alba plants are summarised in Figure 6. Regardless of the variability of the results, a difference between the two locations was apparent. Better plant growth and, therefore, a higher yield of dry matter biomass were recorded for the Libčice soil compared with the Horní Luby soil. However, the plants with limited growth did not manifest any other potential phytotoxicity symptoms. Although the total Hg contents in the Libčice soil were lower than in the Horní Luby soil, the reason for the differences between the growth parameters in these soils was most probably the soil parameters such as pH, CEC, and soil organic matter content, all of them more beneficial for plant growth in the Libčice soil. Umlaufová et al. [26] and Bauštein et al. [27] investigated the Hg uptake by plants at these two locations, where the plant communities at both locations were completely different, and only a few plant species occurred at both of them. These observations confirmed the specificity of the soil properties at these locations, advantageous for different plant communities. Although no plant species representing Brassicaceae occurred naturally at the locations, the preferable conditions for S. alba at the Libčice location were expected, at least with respect to higher pH levels, and available (Mehlich III extractable) Ca contents in the soil were 4874 ± 1390 mg/kg compared with only 925 ± 110 mg/kg at the Horní Luby location, because high soil Ca status is beneficial for this plant species. It could therefore be further concluded that the microbial community in historically contaminated sites was well adapted and equipped to deal with this kind of stress and that the vascular plants did not suffer from this kind of pollution, as shown in the study by Bauštein et al. [27], who showed that the spontaneously occurring vascular plants at these sites were not impacted by the presence of Hg.

Figure 6. Height of the stems (A) and biomass yield (B) of the S. alba plants according to the individual variants (n = 3); the bars marked by the same letter (lower-case letters for shoots, upper-case letters for roots) did not significantly differ at p < 0.05.

The risk element contents in plants are summarised in Table 2, showing (i) high variability of the results within one soil, and (ii) higher contents of most of the elements in the soils from the Horní Luby location, most probably also due to the lower biomass yield in the soils from this location, whereas a dilution effect affected the elemental contents in plants in the Libčice soils. The total elemental uptake by plants per pot (Table 3) confirmed the importance of the biomass yield in the estimation of the total elemental uptake by plants. The risk element contents in plants reflected the contents of these elements in soils, where the plant response to the elevated soil elemental contents differed. Whereas As and Pb were predominantly accumulated in roots (especially for As), the plants showed a low ability to take up these elements; however, Cd and Zn were more easily transported to the aboveground biomass. In the case of Zn, levels above 400 mg/kg of dry matter represented a toxic level of zinc [43]. These levels were exceeded in the plants growing in the Horní Luby soils (Table 2), and the role of high Zn contents in plants in the weak growth of these plants (Figure 6) cannot be excluded. For Cd, regardless of the low pseudototal levels of Cd in soils (Table 1), the contents of this element in most of the plants growing in the Horní Luby soils exceeded the limits given by Directive No. 2002/32/ES [44], where a maximum Cd value in raw feedstuffs was given at 1 mg/kg at a biomass moisture of 12%.

Table 2. The risk element contents in S. alba plants (dry matter) according to the individual variants (n = 4); data presented as mean ± standard deviation.

		Be	V	Cr	Co	Ni	Cu	Zn	As	Cd	Hg	Pb
		mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg	mg/kg
		shoots
Horní Luby	I	0.033 ± 0.009	0.099 ± 0.012	0.823 ± 0.450	1.03 ± 0.30	3.03 ± 0.14	15.7 ± 2.8	539 ± 86	0.288 ± 0.017	1.22 ± 0.06	0.015 ± 0.002	0.702 ± 0.228
	II	0.067 ± 0.016	0.257 ± 0.005	1.24 ± 0.30	6.41 ± 0.24	4.65 ± 0.13	28.6 ± 6.0	1219 ± 237	0.563 ± 0.069	1.94 ± 0.44	0.047 ± 0.007	2.49 ± 0.65
	III	0.149 ± 0.029	0.212 ± 0.015	0.719 ± 0.112	1.29 ± 0.67	3.88 ± 0.26	22.5 ± 3.3	2511 ± 317	0.359 ± 0.054	1.51 ± 0.28	0.034 ± 0.018	2.40 ± 0.68
Libčice	I	0.001 ± 0.000	0.104 ± 0.005	0.491 ± 0.035	0.815 ± 0.380	0.438 ± 0.116	3.45 ± 0.07	49.4 ± 3.6	0.836 ± 0.154	0.557 ± 0.031	0.013 ± 0.001	0.096 ± 0.003
	II	0.002 ± 0.001	0.109 ± 0.015	0.906 ± 0.446	0.672 ± 0.133	0.620 ± 0.205	3.76 ± 0.22	126 ± 10	0.915 ± 0.133	0.462 ± 0.013	0.015 ± 0.002	0.129 ± 0.024
	III	0.004 ± 0.001	0.298 ± 0.183	0.824 ± 0.424	0.961 ± 0.191	0.632 ± 0.192	4.96 ± 1.23	147 ± 10	0.875 ± 0.050	0.621 ± 0.098	0.014 ± 0.003	0.179 ± 0.075
		roots
Horní Luby	I	0.192 ± 0.032	4.58 ± 0.28	18.7 ± 2.0	2.41 ± 0.203	11.4 ± 1.1	22.7 ± 3.4	246 ± 26	2.80 ± 0.32	0.930 ± 0.044	1.51 ± 0.89	14.8 ± 4.2
	II	0.393 ± 0.066	9.65 ± 2.47	7.26 ± 1.65	18.1 ± 0.542	15.0 ± 3.6	42.2 ± 7.0	287 ± 116	7.53 ± 1.07	0.741 ± 0.043	5.56 ± 1.08	57.0 ± 7.4
	III	0.475 ± 0.066	8.64 ± 2.51	10.3 ± 1.0	4.63 ± 0.796	10.5 ± 1.1	34.7 ± 3.4	741 ± 45	8.96 ± 3.30	0.855 ± 0.081	7.42 ± 0.88	52.7 ± 6.9
Libčice	I	0.079 ± 0.035	9.22 ± 3.22	14.6 ± 2.4	3.09 ± 0.670	8.34 ± 2.33	21.9 ± 7.6	61.3 ± 16.3	13.9 ± 6.2	0.780 ± 0.188	1.56 ± 0.76	4.77 ± 0.60
	II	0.073 ± 0.025	8.00 ± 3.13	16.3 ± 2.5	3.22 ± 0.535	7.93 ± 1.33	20.7 ± 9.0	127 ± 34	7.07 ± 1.67	0.896 ± 0.328	0.702 ± 0.401	3.40 ± 0.63
	III	0.066 ± 0.012	5.43 ± 1.28	7.40 ± 1.32	2.67 ± 0.543	4.96 ± 0.83	26.8 ± 6.4	199 ± 81	7.50 ± 1.19	1.17 ± 0.13	0.583 ± 0.172	10.3 ± 2.3

Table 3. The total risk element uptake by S. alba plants per pot according to the individual variants (n = 4); data presented as mean ± standard deviation.

		Be	V	Cr	Co	Ni	Cu	Zn	As	Cd	Hg	Pb
		µg	µg	µg	µg	µg	µg	µg	µg	µg	µg	µg
		shoots
Horní Luby	I	0.091 ± 0.013	0.274 ± 0.022	2.31 ± 1.35	2.86 ± 0.77	8.48 ± 0.92	44.2 ± 11.1	1500 ± 246	0.803 ± 0.065	3.41 ± 0.52	0.041 ± 0.011	1.92 ± 0.37
	II	0.021 ± 0.008	0.078 ± 0.018	0.368 ± 0.073	1.95 ± 0.48	1.41 ± 0.28	8.58 ± 1.89	378 ± 145	0.171 ± 0.047	0.607 ± 0.261	0.014 ± 0.003	0.782 ± 0.374
	III	0.234 ± 0.018	0.340 ± 0.063	1.17 ± 0.37	1.97 ± 0.71	6.20 ± 0.82	35.5 ± 0.5	4046 ± 910	0.577 ± 0.123	2.38 ± 0.18	0.054 ± 0.028	3.74 ± 0.48
Libčice	I	0.006 ± 0.001	0.492 ± 0.016	2.31 ± 0.12	3.85 ± 1.85	2.07 ± 0.58	16.3 ± 0.1	233 ± 14	3.93 ± 0.66	2.62 ± 0.10	0.062 ± 0.002	0.450 ± 0.006
	II	0.007 ± 0.003	0.383 ± 0.106	3.14 ± 1.61	2.37 ± 0.72	2.19 ± 0.95	13.2 ± 2.5	441 ± 99	3.16 ± 0.48	1.61 ± 0.28	0.051 ± 0.006	0.456 ± 0.139
	III	0.031 ± 0.007	2.25 ± 1.38	6.25 ± 3.38	7.27 ± 1.66	4.78 ± 1.58	37.2 ± 8.2	1108 ± 104	6.59 ± 0.38	4.70 ± 0.88	0.102 ± 0.020	1.35 ± 0.56
		roots
Horní Luby	I	0.140 ± 0.044	3.32 ± 0.88	13.3 ± 2.54	1.72 ± 0.25	8.11 ± 1.12	16.2 ± 3.8	176 ± 36	1.98 ± 0.23	0.674 ± 0.173	1.11 ± 0.79	10.4 ± 3.1
	II	0.060 ± 0.035	1.52 ± 1.01	1.13 ± 0.709	2.70 ± 1.32	2.25 ± 1.20	6.28 ± 3.16	39.8 ± 15	1.17 ± 0.73	0.109 ± 0.050	0.869 ± 0.563	8.74 ± 5.16
	III	0.282 ± 0.116	5.09 ± 1.98	6.11 ± 2.16	2.79 ± 1.30	6.25 ± 2.43	20.6 ± 7.6	427 ± 96	5.24 ± 2.16	0.505 ± 0.178	4.42 ± 1.82	25.0 ± 1.7
Libčice	I	0.096 ± 0.046	11.2 ± 4.3	17.6 ± 3.19	3.74 ± 0.87	10.1 ± 2.8	26.6 ± 10.1	74.3 ± 21	17.0 ± 8.0	0.945 ± 0.248	1.91 ± 0.98	5.76 ± 0.85
	II	0.062 ± 0.029	6.75 ± 3.24	13.6 ± 2.01	2.68 ± 0.43	6.63 ± 1.32	18.1 ± 10.7	109 ± 46	6.12 ± 2.50	0.768 ± 0.385	0.548 ± 0.181	2.92 ± 1.04
	III	0.127 ± 0.027	10.4 ± 2.7	14.1 ± 2.17	5.09 ± 1.08	9.50 ± 1.82	51.2 ± 13.1	383 ± 165	14.4 ± 2.8	2.24 ± 0.29	1.12 ± 0.36	19.8 ± 5.0

In the case of Hg, the data confirmed a low ability of plants to take up mercury, where, if absorbed, the predominant proportion of mercury was retained in the roots. Hussain et al. [45] investigated the fate of Hg added to the soils in a soil–plant (Triticum aestivum) system. Their results showed increasing Hg uptake by T. aestivum plants with increasing Hg content in soil. These relationships were not clearly confirmed in this study, where only a weak correlation (not significant at p < 0.05) was recorded between soil and plant Hg contents. Weak soil–plant translocation of Hg was also confirmed by Meloni et al. [46] at an abandoned Hg mine area in Italy. In the case of tobacco (Nicotiana tabacum) plants growing in Hg-polluted soils, the Hg contents decreased predominantly in the order of leaves > root > stem [47], indicating that the predominant root accumulation of Hg was not typical for all plant species. The results also suggested the role of soil properties in the Hg uptake, where an increasing CEC level led to a decrease in Hg uptake by plants, as documented by a significant (p < 0.05) correlation between these two variables (r = −0.634 for aboveground biomass and r = −0.603 for roots). Moreover, Uwiringiyimana et al. [47] showed increasing Hg accumulation in tobacco plants with decreasing soil pH values. Hussain et al. [45] also observed that decreasing pH significantly increased Hg uptake by plants.

Zhang et al. [48] confirmed the negative impact of the mercury-polluted slag piles on the surrounding soil and, subsequently, on the crops grown in these soils. They found Hg contents exceeding the regional limits for rice, maize, and vegetables such as pepper, tomato, and eggplant, and hazard quotient levels indicating a potential human health risk from the pollution. Similarly, Yang et al. [49] reported a potential human health risk caused by Hg in soils and crops in the mining area. In this experiment, the Hg contents in the aboveground biomass of plants were low, not exceeding the threshold limit for Hg content in feedstuffs, i.e., 0.1 mg/kg at a biomass moisture of 12% (Table 2).

3.3. Mercury Speciation and Fractionation in Soils

Figure 7 shows a different pattern of mobile MeHg and Hg(II) species in the soils from different locations; regardless of the high variability of the results, the data indicated higher MeHg contents in the soil from the Libčice location, and higher contents of inorganic Hg(II) were determined in the soils from Horní Luby. A different methylation rate and behavior at these two locations were also observed by Stefanović et al. [29], where at the Libčice location MeHg dominated in the organic horizon, whilst at Horní Luby MeHg occurred preferentially in the mineral horizon due to the different source of Hg in both soils. In this experiment, the soils were taken from the whole soil column, and the MeHg contents in the whole soil were affected by the thickness of the individual horizons.

Figure 7. The contents of MeHg (A) and Hg(II) (B) extracted from the soils according to the individual variants (n = 4); the bars marked by the same letter did not significantly differ at p < 0.05.

De Souza et al. [50] investigated the distribution of Hg among the main fractions of contaminated tailings generated by artisanal small-scale gold mining activities, where a high bioaccessible proportion of Hg, representing 28% of the total Hg content, indicated a high environmental risk from the pollution. Dominating, semi-mobile Hg proportions were determined, for instance, in soils from the area surrounding an abandoned coalfield waste pile [51]. On the contrary, low accessibility of Hg in soils collected next to the abandoned cinnabar mine was confirmed by Silva-Gigante et al. [52]. The low accessibility of Hg was confirmed in this experiment as well. The proportion of mobilizable Hg(II) ranged from 3 to 11% of the total Hg, and the proportions of mobilizable MeHg did not exceed 1% of the total Hg contents in the soils. Thus, the potential environmental risk was considered low. Hussain et al. [45] spiked the soil with Hg as HgCl₂ before the start of the pot experiment; Hg was weakly mobile at the end of the experiment, where the water-soluble Hg proportion was less than 1%. Thus, even relatively soluble Hg compounds could be tightly bound to soil constituents over time, resulting in low bioaccessibility of this element in soil.

Płońska et al. [53] highlighted the role of soil moisture and the content and composition of SOM as the key factors affecting the formation and retention of MeHg in the soil. Hussain et al. [54] investigated the availability of spiked HgCl₂ in rhizosphere soil surrounding the roots of Capsicum annuum plants. They determined an enhanced Hg pool related to the soil organic matter in the rhizosphere soil compared with the bulk soil, whereas the accessible pool decreased. In this experiment, the mobile Hg(II) contents in the soil were significantly negatively correlated with the total carbon content (r = −0.505, p < 0.05), but no significant correlation was recorded between Hg(II) values and the soluble pool of carbon in the soils. Assuming that total soil C contributes to soil CEC, the observed correlations indicated that the Hg(II) pool was strongly bound to charged sites within the soil sorption complex, thereby limiting its mobility. This interpretation was further supported by the negative relationship between plant Hg uptake and soil CEC, reflecting reduced bioavailability of Hg(II) in soils with higher sorption capacity. In contrast, mobilizable MeHg was positively correlated with CEC, suggesting weaker retention by the sorption complex and a greater influence of exchange processes. These observations were consistent with previous findings, as Tipping [55] reported weaker binding of MeHg to model humic substances relative to Hg(II), and Schlüter [56] showed that MeHg solubility was more strongly controlled by ion-exchange mechanisms than that of Hg(II).

Du et al. [57] reported a significant positive correlation between organic matter content and MeHg production, whereas Yin et al. [58] found a negative correlation between MeHg content in soil and the ratio of organic matter to total mercury. Wu et al. [59] added easily degradable organic matter to the soil and found that adding this form of organic matter reduced the ability of the soil to adsorb mercury. However, in this experiment, in the case of MeHg, no interrelationships were recorded for either total or soluble C contents in the soils. However, the values of mobile Hg(II) were significantly negatively correlated with soil CEC (r = −0.577, p < 0.05), indicating the potential of the soil sorption complex to immobilize Hg in soils. Similarly, Huang et al. [60] concluded in their research working with thiol-modified biochar that cation exchange capacity helped to immobilize mercury. On the contrary, mobile MeHg contents in the soils were positively correlated with soil CEC (r = 0.619, p < 0.05). Thus, this experiment indicated the immobilization potential of soil organic matter for the bioaccessible pool of Hg, but the fate of MeHg remained ambiguous.

Baptista-Salazar et al. [61] investigated the methylation and demethylation processes of Hg in soils and found that the produced MeHg was sequestered in the soil and was not easily available for potential demethylation. Therefore, MeHg could be accumulated in the soil even if the soil conditions were not favorable for Hg methylation. However, Shuvaeva et al. [62] tested the bioavailability of different Hg species (HgCl₂, MeHgCl, and HgS) for S. alba plants in a soilless culture. They confirmed a higher plant accumulation ability of MeHg compared with the inorganic species. Thus, in principle, MeHg could be considered more plant-accessible compared with inorganic Hg species, but in soil, its accessibility could be hampered by the rapid sequestration of MeHg in the soil.

Hussain et al. [54] stated that the main factors controlling the plant availability of Hg were not the rhizosphere processes but simply the total Hg content in soil and the Hg pool associated with Mn/Fe oxides in soil. Carvalho et al. [63] also showed that sites with higher contents of these elements were more capable of binding soil mercury. The data concerning available proportions of Fe and Mn documented lower iron contents in soil at the Horní Luby location (varying between 330 and 400 mg/kg) than at Libčice (between 350 and 550 mg/kg), and an opposite pattern for manganese. For manganese, the available contents of this element varied between 100 and 230 mg/kg at the Horní Luby location, and between 90 and 110 mg/kg at Libčice. The presence of Mn/Fe oxides could affect the bioaccessibility of Hg in soil as well, including its potential methylation. Iron-based materials such as zero-valent iron showed the ability to reduce the soil–plant transport of both total Hg and MeHg [16]. They assumed that the formation of an iron plaque on the below-ground parts of plants could effectively inhibit the uptake of Hg by roots and subsequently reduce MeHg accumulation in plant tissues. In this context, Janeiro-Tato et al. [64] presented a novel, effective Hg immobilization amendment consisting of carbon foam impregnated with goethite, documenting the importance of Fe oxides in the remediation of Hg-polluted soils. Qin et al. [65] concluded that the presence of FeS₂ might promote mercury methylation in the soil and reduce mercury accumulation in plants. As already mentioned, a higher amount of MeHg was recorded in the soils from the Libčice location, with higher soil Fe content compared with the Horní Luby soil. It could therefore be concluded that the influence of iron mineral phases was also an active factor in the formation of MeHg in the soil in this experiment.

The potential role of sulfur in the mobility and transformation of Hg species could not be excluded as well. Sulfides could reduce the bioavailability of mercury for mercury-methylating bacteria. Despite sulfide addition in the vessel experiment, which led to Hg loss from the medium via reduction or transfer to the vessel walls, the initially reduced MeHg production in the culture medium stabilized [66]. In environments with sufficient sulfate availability, MeHg was produced primarily through microbial sulfate reduction. In this experiment, the correlation between mobile MeHg and total sulfur was significant and positive (r = 0.485, p < 0.05), suggesting that increasing total S promoted the methylation process.

For the comprehensive assessment of the Hg cycle in the soils, GEM had to be considered. The GEM concentrations at the soil–air interface, as affected by the presence of plants, are summarised in Figure 8, where significantly (p < 0.05) higher GEM concentrations were determined before sowing of the plants compared with the measurement performed immediately after harvest of the aboveground biomass. The data also confirmed that GEM represented an important part of the Hg cycle in the soils with elevated Hg content, regardless of the origin of soil Hg. Nadzir et al. [67] investigated GEM concentrations in ambient air across urban, suburban, rural, and industrial areas in Malaysia using an in situ measurement mode. The highest GEM concentrations were determined in industrial areas, with a maximum of 28.9 ng/m³, not reaching the maximum GEM levels determined in this experiment. However, they used a different measuring mode, where GEM was trapped on gold amalgamation tubes for 60 min and subsequently desorbed and determined by using cold vapor atomic absorption spectrometry.

Figure 8. GEM concentrations in the individual variants as affected by the presence of plants (n = 3 for pots with plants and n = 1 for pots without plants).

Szponar et al. [23] confirmed that GEM originating from artisanal and small-scale gold mining (where amalgamation is used for gold extraction) contributed to over 70% of total GEM concentrations in the entire area. The data from this experiment did not allow us to provide a detailed balance of soil Hg, but the importance of GEM in the whole Hg cycle was confirmed. Although not proven to be significant (because of the low number of data for statistics), the data indicated enhanced GEM concentrations in the presence of plants compared with the pots without plants in the Horní Luby soils. In the soils of Libčice, the data did not clearly indicate a similar pattern. The potential role of S. alba, the plant representing the Brassicaceae family, could have been discussed in this case. Brassicaceae are characterized by elevated demands for sulfur. Thus, it could be speculated that S release for plant uptake might also have mobilized Hg bound to various S-bearing phases in the soil. This assumption should be verified in further research. Moreover, more frequent measurements of GEM will be helpful for an assessment of the plant-mediated changes in GEM emissions in further research.

4. Conclusions

Location-specific differences in Hg behavior were observed between the investigated soils. Higher proportions of potentially mobilizable MeHg were detected in the Libčice soil, whereas inorganic Hg(II) predominated in the Horní Luby soil. Overall Hg accessibility was low, with mobilizable Hg(II) accounting for 3–11% of total soil Hg and mobilizable MeHg remaining below 1%, indicating a limited bioavailable pool under the studied conditions. Mobilizable Hg(II) showed significant negative relationships with total soil C and CEC, reflecting a strong association with the charged functional groups of the soil sorption complex. In contrast, mobilizable MeHg was independent of soil C pools but positively related to CEC, suggesting weaker binding to organic matter and a greater contribution of cation-exchange mechanisms.

The concentration of total Hg and the different Hg species did not affect the measured biological parameters of the soil in this experiment, as the different concentrations of total Hg and individual species (Hg(II) and MeHg) were not reflected in microbial biomass. Further, soil emissions were not related to microbial activity, as they did not correlate with the microbial biomass values. The height of the stems of the model plant species was not affected by the concentration of Hg or the measured individual species. Plant experiments confirmed low mercury uptake, with Hg largely retained in roots and contents in aboveground biomass remaining below feed safety thresholds. Measurements of GEM further indicated the presence of plant-mediated effects on GEM emissions, requiring further research.

Overall, these findings emphasize the presence of distinct controlling factors influencing Hg(II) and MeHg behavior in soils and highlight the need for mechanistic studies addressing the factors governing MeHg dynamics and fate in soil–plant systems. Further analysis of the microbial community, the presence of specific genes from the hgc and mer operons, and the activity of the microbial community in these soils is required to better understand the microbial community at these sites.