Phytoremediation Efficiency of Hemp and Sorghum Grown in Contaminated Sediment: The Role of Organic Acids

Abstract

The sustainable management of dredged sediments contaminated with heavy metals represents a major environmental challenge. This study evaluated the phytoremediation potential of hemp (Cannabis sativa L.) and sorghum (Sorghum bicolor L.) cultivated in metal-enriched sediment from the Bega Canal (Cu = 204 mg kg⁻¹, Pb = 171 mg kg⁻¹, Cr = 281 mg kg⁻¹, Ni = 56 mg kg⁻¹, Cd = 6.8 mg kg⁻¹) and examined the effects of glutamic (GA) and tartaric (TA) acids (20 mmol kg⁻¹) on sediment properties and metal uptake. Pot experiments under natural conditions (n = 3, 6–8 weeks) showed that GA treatment resulted in cation exchange capacity (CEC) values ranging from 31.0 to 58.5 cmolc kg⁻¹, which were lower than in the initial sediment (60.7 cmolc kg⁻¹) but still higher than in the corresponding controls and TA treatments. GA also increased electrical conductivity from 435 to 1189 µS cm⁻¹, which may indicate enhanced ion mobility and be consistent with redox-related processes, whereas TA maintained near-neutral pH (8.0–8.2) and caused only minor changes in CEC and EC, preserving overall structural stability. Hemp produced up to 40% more biomass than sorghum and allocated a relatively larger share of Cu, Pb and Cd to shoots, whereas sorghum retained up to 80% of total Cr and Ni in roots. Bioaccumulation factors ranged from 4.3 for Cu in hemp (GA) to 20.8 for Cu in sorghum (GA), while translocation factors remained <1.0 in both species, indicating that root-based phytostabilization was the dominant mechanism. The results demonstrate that combining low-molecular-weight organic acids with energy crops can effectively enhance metal mobility and plant uptake, offering a viable route for sediment remediation and biomass valorization within circular economy strategies.

1. Introduction

The increasing accumulation of contaminated sediments in drainage and industrial canals represents a growing environmental challenge worldwide. These sediments often act as long-term reservoirs for various pollutants, including heavy metals and persistent organic compounds, which can be released back into the environment under changing physicochemical conditions [1,2]. The dredging of canals and waterways, although essential for maintaining navigability and preventing flooding, generates large quantities of dredged material that must be properly managed. Depending on the degree of contamination, such material can be classified as hazardous waste, creating serious disposal and cost-related challenges [3,4]. Consequently, developing sustainable and cost-effective strategies for the treatment and reuse of contaminated sediments has become a research priority.

Phytoremediation has emerged as a promising alternative to conventional physicochemical remediation methods, offering an environmentally friendly, low-cost, and aesthetically acceptable approach for the stabilization or removal of pollutants from contaminated sites [5,6]. The success of this technology depends primarily on the selection of plant species capable of tolerating and accumulating high concentrations of pollutants while producing significant biomass. Energy crops such as Sorghum bicolor L. (sorghum) [7,8] and Cannabis sativa L. (hemp) [8,9] have attracted considerable attention. These fast-growing species are characterized by a well-developed root system, high tolerance to heavy metals, and the ability to accumulate or stabilize pollutants in both root and shoot tissues [10,11]. The harvested biomass can be valorized for bioenergy production, providing an additional economic and environmental benefit.

The efficiency of phytoremediation is often limited by the low bioavailability of heavy metals in sediments and soils, as they are commonly bound to organic matter, clay minerals, or oxides in forms inaccessible to plants [12]. One of the most effective strategies to overcome this limitation involves the use of low-molecular-weight organic acids (LMWOAs), such as tartaric, glutamic, citric, oxalic, and malic acids [13,14,15]. Tartaric (TA) and glutamic (GA) acids were chosen in this experiment because they form moderately strong metal–ligand complexes. According to Martell and Smith [16], the stability constants (logK) for, for example, Cu–tartrate and Cu–glutamate typically range between 3.8 and 4.2 and 2.9–3.3, respectively, which are well below the corresponding values of stronger LMWOAs such as citric acid (logK ≈ 6.2–6.5) and oxalic acid (logK ≈ 7.8–8.2). This relatively moderate chelating power can increase metal solubility and phytoavailability, while also reducing the risk of mobilization and over-leaching [17,18]. In addition, both tartaric (TA) and glutamic (GA) acids belong to a group of low-molecular-weight organic acids, which are rapidly degraded in soil environments through microbial mineralization. This is consistent with Jones [19], who demonstrated that the mineralization of low-molecular-weight organic acids is rapid across a wide range of soil types, and with the findings of Ward [20], who reported very short half-lives for glutamic acid (2–5 h in surface soils and 8–25 h in subsurface soils), further confirming its fast microbial turnover and low environmental persistence. Recent studies have documented their effectiveness in improving the mobility of heavy metals and their plant uptake in contaminated soils and sediments [21,22,23,24]. These characteristics demonstrate that TA and GA can be considered suitable and environmentally compatible amendments for sediment phytoremediation. However, their use requires careful control of concentration and frequency, since even moderately strong chelators may induce excessive metal mobilization and potential leaching in groundwater if applied at elevated doses [25].

Although the use of LMWOAs as soil or sediment amendments is well documented, and the phytoremediation potential of sorghum (Sorghum bicolor L.) and hemp (Cannabis sativa L.) has been widely recognized, there is a lack of studies investigating their combined application. To the best of the authors’ knowledge, no previous research has evaluated interactions between glutamic acid and tartaric acid applied to these two plant species for the treatment of metal-contaminated dredged sediments, aiming to develop a first step to a sustainable and cost-effective strategy for the treatment and reuse of contaminated sediments.

This study provides a comprehensive evaluation of (1) the combined influence of low-molecular-weight organic acids (LMWOAs) and selected plant species on metal-contaminated sediments; (2) the balance between increased metal uptake and the challenges associated with dredged sediment; and (3) the overall phytoremediation efficiency, including a discussion of potential biomass valorization routes within a sustainable waste-management framework. In doing so, this research fills an important knowledge gap by connecting contaminated-sediment handling, phytoremediation performance, and opportunities for biomass utilization.

Accordingly, the objective of this work was to investigate how glutamic and tartaric acids influence metal mobility, plant accumulation, and key physicochemical properties of heavily polluted dredged sediments cultivated with hemp and sorghum. By combining sediment characterization, metal fractionation, and plant response analyses, this study offers new insights into the mechanisms and constraints of LMWOA-assisted phytoremediation. In addition, it explores the valorization potential of the resulting biomass, contributing to a broader, circular-economy-oriented perspective for sustainable sediment treatment and reuse.

2. Materials and Methods

2.1. Site Description and Sediment Sampling

The sediment used for this experiment originated from the Bega Canal, which forms part of the Danube–Tisa–Danube (DTD) Hydrosystem. This complex canal network connects the Danube and Tisa Rivers and serves multiple purposes, including flood control, irrigation, drainage, navigation, and wastewater reception. For decades, the canal has been affected by industrial and municipal discharges, as well as diffuse pollution from agricultural activities, resulting in the long-term accumulation of heavy metals and organic contaminants [26,27].

For the present study, sediment was collected from a confined disposal site situated approximately 300 m from the Srpski Itebej sluice, near the Serbian–Romanian border. The disposal site contains around 5900 m³ of dredged material from the Bega Canal, deposited over an area of about 3800 m² enclosed by earthen dikes. Approximately 250 kg of sediment was collected from several points within the site, transported to the laboratory, manually mixed, and homogenized prior to filling the experimental pots. The composite sample obtained in this way was used as the starting material (hereafter referred to as SS, i.e., start sediment). The results of its initial physicochemical characterization are presented in

Table 1. Physicochemical properties of contaminated start sediment (SS).

Indicator		Values
pH		7.46 ± 0.08
EC (µS/cm)		503.7 ± 54.8
Organic matter (%)		7.39 ± 0.035
Texture (%)	Send (50–2000 µm)	78.4 ± 0.49
	Silt (2–50 µm)	6.33 ± 0.37
	(Clay ˂ 2 µm)	15.2 ± 0.25
Totak K (%)		0.30 ± 0.07
Total N (%)		0.264 ± 0.02
Total P (%)		0.22 ± 0.05
CEC (cmolc/kg dry matter of soil)		60.75 ± 11.34
Metal concentration	Total (mg/kg)	BCR1 (%)	BCR2 (%)	BCR3 (%)	BCR4 (%)
Cu	204.2 ± 37.6	2.30	8.00	39.3	50.4
Pb	171.0 ± 29.7	0.185	8.72	6.10	85.0
Cd	6.79 ± 0.73	15.5	62.9	15.6	6.00
Cr	280.7 ± 63.7	0.148	0.710	25.4	73.8
Ni	56.4 ± 11.7	2.92	8.50	17.8	70.8

Standard deviations are based on three replications (n = 3).

.

The initial sediment was characterized as slightly alkaline (pH 7.46 ± 0.08) with a high sand fraction (78.4%) and moderate clay content (15.2%), corresponding to a sandy loam texture. It contained moderate levels of plant nutrients (N, P, K) and organic matter (7.39%), indicating moderate fertility and the presence of stabilized organic material typical of dredged sediments with limited microbial activity. The measured cation exchange capacity (CEC = 60.75 cmolc kg⁻¹ DW) falls within the range typical of organic soils (50–100 cmolc kg⁻¹), suggesting the presence of reactive mineral and organic components contributing to a high cation-retention potential. The electrical conductivity (503.7 µS cm⁻¹) indicated a non-saline sediment with a moderate level of soluble salts, typical of freshwater dredged materials.

Based on the concentrations of trace metals (Cu, Pb, Cd, Cr and Ni), the material is categorized as highly contaminated under the national sediment legislation [28]. Concentrations of Pb, Ni, Cd, Cu, and Cr exceeded the corresponding regulatory thresholds, while other metals were below the target values. Sequential extraction results (BCR procedure) indicated that most of the analyzed metals were predominantly associated with oxidizable and residual fractions, suggesting limited bioavailability under natural conditions. Metals bound to organic matter (BCR3) may be mobilized under oxidizing conditions, whereas those in the residual fraction (BCR4) remain tightly incorporated into the mineral matrix [29].

2.2. Experimental Design and Treatments

POT experiments were performed in the open air under natural weather conditions. The main objective was to investigate the effect of tartaric (TA) and glutamic (GA) acids on plant growth and the phytoextraction efficiency of selected crops: sorghum (Sorghum bicolor L.) and hemp (Cannabis sativa L.). Pots were filled with 5 kg of air-dried and homogenized sediment (SS). Five seeds were sown per pot, and after germination, the seedlings were thinned to maintain two plants per pot. Each treatment was conducted in triplicate, including an untreated control (no organic acid addition).

Organic acids were applied as aqueous solutions corresponding to 20 mmol per kg of sediment. Treatments with tartaric and glutamic acids were performed six and eight weeks after sowing, respectively. Each acid was dissolved in 100 mL of deionized water and evenly distributed on the surface of the sediment, followed by an additional 100 mL of deionized water to ensure uniform infiltration.

Table 2. The design of pot experiments.

Plant	Treatments	Abbreviation	Weeks for Amendment Application After Sowing	Weeks of Plant Harvesting After the Sowing
Hemp (Cannabis Sativa L.)	No treatment	H C 6 W	-	6 weeks and 4 days
	20 mmol/kg tartaric acid	H TA 6 W	6 weeks	6 weeks and 4 days
	20 mmol/kg glutamic acid	H GA 6 W	6 weeks	6 weeks and 4 days
	No treatment	H C 8 W	-	8 weeks and 4 days
	20 mmol/kg tartaric acid	H TA 8 W	8 weeks	8 weeks and 4 days
	20 mmol/kg glutamic acid	H GA 8 W	8 weeks	8 weeks and 4 days
Sorghum (Sorghum bicolor L.)	No treatment	S C 6 W	-	6 weeks and 4 days
	20 mmol/kg tartaric acid	S TA 6 W	6 weeks	6 weeks and 4 days
	20 mmol/kg glutamic acid	S GA 6 W	6 weeks	6 weeks and 4 days
	No treatment	S C 8 W	-	8 weeks and 4 days
	20 mmol/kg tartaric acid	S TA 8 W	8 weeks	8 weeks and 4 days
	20 mmol/kg glutamic acid	S GA 8 W	8 weeks	8 weeks and 4 days

provides an overview of the applied treatments, including the type and number of additions, as well as the timing of their application and plant sampling. Organic acids were added six and eight weeks after sowing to evaluate their effects at different stages of plant development. Hemp and sorghum were harvested four days after acid addition.

At the end of the experiment, one plant from each pot was used to determine biomass and heavy metal content. Plant samples were rinsed with distilled water, separated into shoots and roots, dried at 40 °C, weighed, and ground using a laboratory mill. Heavy metal concentrations were determined separately in the shoot and root fractions.

After harvesting the plant, composite sediment samples were prepared from each set of triplicates, air-dried, ground, and analyzed for heavy metal content and selected physicochemical parameters.

2.3. Methods for Analysis and Characterization

The total content of heavy metals in sediment and plant samples was determined after preparation of samples by microwave digestion (Milestone, Start E/0610030141, Sorisole, Italy) according to EPA 3051a [30] and analyzed for heavy metal content on AAS graphite furnace (G-AAS) (Thermo Scientific, Waltham, MA, USA) according to EPA 7010 [31].

The total amount of bioavailable fractions of heavy metals was determined by BCR sequential extraction, according to the procedure described in Arain et al. [32], and analyzed for heavy metal content on AAS graphite furnace (G-AAS) (Thermo Scientific, Waltham, MA, USA) according to EPA 7010 [31].

The total potassium content in sediment samples was determined after sample preparation by microwave digestion (Milestone, Start E/0610030141, Sorisole, Italy), and the concentration was subsequently measured by flame emission spectrophotometry using an atomic absorption spectrometer (Thermo Scientific, Waltham, MA, USA) according to APHA AWWA WPCF 3500-K D [33].

pH was determined according to ISO 10390:2005 [34] using a glass electrode (SenTix, Limburg, Germany) in a 1:5 (volume fraction) suspension of soil and water.

Electrical conductivity (EC) was determined according to ISO 11265:1994 [35] by mixing air-dried sediment with distilled water (1:5 ratio) and shaking for 30 min.

The organic matter content was determined according to the CEN standard EN 12879:2000 [36] by the loss-on-ignition method at 550 °C.

Total P determination was performed using an internal laboratory method, where phosphorus is extracted with ammonium lactate solution, then colored with ammonium molybdate and analyzed on a UV/VIS spectrometer (Shimadzu, Kyoto, Japan, UV-1800).

Total N was determined according to the Kjeldahl method (ISO 11261:1995 method [37]) via digestion of the sediment sample with ccH₂SO₄, distillation in H₃BO₃, and titration with HCl.

Texture was determined according to the ISO 11277:2009 [38] method by sieving soil samples through a series of sieves in the range of 2 mm to 0.063 mm, sedimentation method for fraction 0.063 m to <0.002 mm, and by withdrawing 25 mL of suspension at defined time and depth in the cylinder.

The cation exchange capacity (CEC) was determined according to the ammonium acetate extraction method (1 M NH₄OAc, pH 7.0) following the procedure modified from Reeuwijk [39]. Exchangeable Ca²⁺ and Mg²⁺ were analyzed using a graphite furnace atomic absorption spectrophotometer (G-AAS) according to EPA 7010 [31], while Na⁺ and K⁺ were determined by flame emission spectrophotometry using an atomic absorption spectrometer following APHA AWWA WPCF 3500-Na B [40] and 3500-K D [33] methods. The CEC was expressed in cmolc kg⁻¹ DW as the sum of exchangeable bases.

2.4. Data Analysis

All experiments were conducted in triplicate. Statistical data analysis was performed for results related to plant samples, using IBM SPSS Statistics software (version 30.0.0.0). The normality of data distribution was assessed using the Shapiro–Wilk test, and the homogeneity of variances was verified with Levene’s test. If the results of Levene’s test indicated unequal variances among treatments (p ≤ 0.05), Welch’s ANOVA was applied to evaluate differences between group means. When statistically significant differences were detected (p ≤ 0.05), the Games–Howell post hoc test was used to determine specific pairwise differences between treatments.

2.5. Evaluating the Phytoremediation Efficiency

The phytoremediation potential was estimated by calculating the bioaccumulation factor (BAF) and translocation factor (TF). The bioaccumulation factor (BAF) is defined as the ability of a plant to accumulate a certain metal in relation to its concentration in the substrate and is calculated as the ratio of the concentration of the metal in the root/above-ground parts to its concentration in the substrate, as shown in Equation (1) [41]. The translocation factor (TF) is used to estimate the relative translocation of metals from below-ground organs to above-ground organs and is calculated as the ratio of metal concentration in the above-ground part of plants to its concentration in the roots, as shown in Equation (2) [42].

$$\mathrm{BAF}={\displaystyle \frac{ C_{ above ground/root}}{{C }_{substrate}}}$$

(1)

$$\mathrm{TF}={\displaystyle \frac{ C_{ above ground}}{{C }_{root}}}$$

(2)

3. Results and Discussion

3.1. Physico-Chemical Properties of Sediments After Cultivation

After cultivation (

Table 3. Physicochemical properties of the sediment.

Treatments	Time	pH	EC	Organic Matter	Total N	Total P	CEC	Texture
			µS/cm	%	%	%	cmolc/kg Dry Matter of Soil	Send (50–2000 µm)	Silt (2–50 µm)	(Clay ˂ 2 µm)
H C 6 W	After harvest	8.08	424	8.96	0.268	0.095	21.01	76.2	6.8	17.0
H TA 6 W	After harvest	8.03	335	8.59	0.260	0.121	19.80	81.7	7.6	10.7
H GA 6 W	After harvest	7.7	1126	9.12	0.264	0.104	58.51	81.4	5.9	12.7
H C 8 W	After harvest	8.16	355	12.7	0.265	0.103	31.75	74.8	5.3	19.8
H TA 8 W	After harvest	8.12	319	8.62	0.255	0.099	34.16	78.2	8.3	13.5
H GA 8 W	After harvest	7.97	435	8.97	0.239	0.142	40.47	82.9	5.2	12.0
S C 6 W	After harvest	7.86	567	8.58	0.242	0.098	39.59	80.3	7.9	11.7
S TA 6 W	After harvest	8.12	44	10.1	0.224	0.103	20.39	79.3	6.4	14.3
S GA 6 W	After harvest	7.45	1189	9.29	0.253	0.099	44.44	82.9	7.1	10.0
S C 8 W	After harvest	8.07	334	4.66	0.255	0.112	27.40	83.0	5.0	12.0
S TA 8 W	After harvest	8.18	243	8.53	0.264	0.090	22.43	73.4	5.3	21.3
S GA 8 W	After harvest	7.83	602	9.45	0.264	0.103	31.02	73.4	7.8	18.8

), noticeable changes occurred in several parameters depending on the plant’s species, treatment, and duration. In general, pH increased slightly, reaching values between 7.7 and 8.2 across treatments, indicating a mild alkalization that may be related to cation uptake (Ca²⁺, Mg²⁺, K⁺) and the release of hydroxyl ions from root activity. This effect was particularly visible in control and TA-treated pots, while GA treatments maintained a slightly lower pH, consistent with the possibility that glutamic acid contributed to slight buffering effects through its weak acid properties [43,44]. Electrical conductivity (EC) values varied widely (from 44 µS/cm to 1189 µS/cm), indicating dynamic oxidation-reduction conditions influenced by plant metabolism and root exudates. The notably higher EC under GA treatments (e.g., H GA 6 W = 1126 µS/cm; S GA 6 W = 1189 µS/cm) suggests improved oxygenation and microbial stimulation in the rhizosphere, possibly due to enhanced root respiration and organic acid secretion [45]. Conversely, the extremely low EC in S TA 6 W (44 µS/cm) might indicate localized reducing conditions, potentially caused by metal–organic acid complexation and limited gas diffusion in fine-textured micro zones.

Organic matter content slightly increased in most treatments (from 7.39% to 8–12%), particularly after eight weeks in control pots (H CK 8 W = 12.7%), reflecting biomass decay and root residue accumulation. The increase was less pronounced in GA-treated pots, which may reflect enhanced decomposition processes influenced by GA. Similar trends were reported by Yang et al. [46], where amino acids accelerated organic matter mineralization and improved nitrogen cycling in treated soils.

Total N concentrations remained relatively stable after plant growth, ranging from 0.22 to 0.27%, with minor fluctuations among treatments. Slightly higher values were observed in some GA variants at 6 weeks, possibly reflecting enhanced N assimilation or microbial mineralization in the rhizosphere. Total P ranged from 0.09 to 0.14%, with generally stable or slightly variable values across treatments. The highest P concentration was recorded under the H GA 8 W treatment (0.142%), suggesting possible organic complexation or limited plant uptake under this condition. Conversely, the lowest p values in S TA 8 W (0.090%) may indicate plant uptake and local immobilization of phosphorus. This agrees with previous findings that low-molecular-weight organic acids enhance P bioavailability and subsequent plant assimilation [47,48].

Cation exchange capacity (CEC) values showed considerable variation among treatments, ranging from 19.8 to 58.51 cmolc kg⁻¹, compared with 50.75 cmolc kg⁻¹ in the initial sediment. The decrease in most treatments indicates a partial loss of exchangeable cations and mineral restructuring during plant growth, likely driven by root uptake and changes in the ionic composition of the pore water. This decline coincided with a slight reduction or fluctuation in organic matter content (4.66–12.7% compared with 7.39% initially), suggesting that the decomposition and mineralization of organic residues may have reduced the number of negatively charged exchange sites responsible for cation retention [49,50]. However, the highest CEC recorded in the H GA 6 W treatment (58.51 cmolc kg⁻¹) corresponded with relatively elevated organic matter (9.12%), indicating localized enrichment in exchangeable bases and organic residues, possibly associated with enhanced root activity and microbial turnover in the rhizosphere. The unusually wide range of CEC values observed across treatments (19.8–58.5 cmolc kg⁻¹) suggests that part of this variation may be attributed not only to biological or organic matter–related processes, but also to methodological and chemical effects introduced by the amendments [51,52]. In particular, the markedly elevated EC under GA treatments indicates an increase in ionic strength, which is known to influence the NH₄OAc extraction method by altering ion exchange equilibria and potentially leading to overestimation of extractable cations [53]. Furthermore, the strong complexation capacity of glutamic acid toward Ca²⁺, Mg²⁺, and other exchangeable cations may have modified their solubility and extractability, creating apparent increases or decreases in measured CEC depending on the predominant complex species formed [51,54].

In addition, transient redox shifts induced by root activity, especially in treatments with enhanced rhizosphere respiration, may have temporarily affected the charge density of organic matter and Fe/Mn oxide surfaces, thereby influencing their contribution to total exchange capacity [55]. All CEC measurements were performed in triplicate; however, given the chemical sensitivity of the ammonium acetate method to salinity, organic ligands, and rhizosphere-induced microenvironmental changes, some portion of the observed variability likely reflects methodological artifacts rather than true structural changes in sediment exchange sites. These factors have now been acknowledged to provide a more accurate interpretation of the observed CEC dynamics.

Sediment texture remained predominantly sandy (73–83%), with minor variations in silt and clay fractions due to particle reaggregation and drying–rewetting cycles during the experiment. Slightly higher clay percentages were observed in the control pots after eight weeks (e.g., H C 8 W = 19.8%), possibly because of micro-aggregate formation facilitated by root exudates and Ca²⁺ bridging.

3.2. Influence of LMWAO Application on Heavy Metals in Sediment

The total metal content (

Table 4. Total heavy metal content in sediment after the treatment.

Treatments	Cu	Pb	Cd	Cr	Ni
			mg/kg
H C 6 W	113.47 ± 26.21	113.02 ± 36.96	3.79 ± 1.24	159.30 ± 28.20	25.96 ± 4.91
H GA 6 W	142.17 ± 32.84	141.60 ± 46.30	4.53 ± 1.48	198.19 ± 35.08	28.48 ± 5.38
H TA 6 W	112.28 ± 25.94	122.19 ± 39.96	3.74 ± 1.22	161.43 ± 28.57	23.46 ± 4.43
H C 8 W	116.92 ± 27.01	123.49 ± 40.38	3.50 ± 1.14	161.00 ± 28.50	22.42 ± 4.24
H GA 8 W	116.65 ± 26.95	116.18 ± 37.99	3.71 ± 1.21	162.62 ± 28.78	23.37 ± 4.42
H TA 8 W	131.33 ± 30.34	141.57 ± 46.29	4.82 ± 1.58	274.63 ± 48.61	30.39 ± 5.74
S C 6 W	113.71 ± 26.27	109.45 ± 35.79	3.70 ± 1.21	148.40 ± 26.27	22.62 ± 4.28
S GA 6 W	110.76 ± 25.59	128.36 ± 41.97	3.83 ± 1.25	166.65 ± 29.50	24.22 ± 4.58
S TA 6 W	110.30 ± 25.48	117.51 ± 38.42	3.91 ± 1.28	131.57 ± 23.29	23.88 ± 4.51
S C 8 W	117.63 ± 27.17	123.22 ± 40.29	3.84 ± 1.26	208.50 ± 36.90	27.21 ± 5.14
S GA 8 W	100.90 ± 23.31	101.02 ± 33.03	3.55 ± 1.16	145.18 ± 25.70	22.94 ± 4.34
S TA 8 W	112.18 ± 25.91	114.22 ± 37.35	3.91 ± 1.28	163.16 ± 28.88	26.12 ± 4.94

Standard deviations are based on three replications (n = 3).

) in sediment or soil is often insufficient for a reliable assessment of heavy metal mobility and bioavailability, as it does not necessarily reflect the fraction that is accessible to organisms and/or plants [56]. As can be seen in Table 4, the total metal content in the sediment did not significantly change after the phytoremediation process. However, the chemical forms in which the metals occur were notably altered (Figure 1).

This redistribution among different fractions indicates that phytoremediation primarily affects the mobility and bioavailability of metals rather than their overall concentration [57,58].

This information is very important because metals that are not in a bioavailable form cannot be uptaken by plants, but plants can increase the bioavailability of heavy metals by releasing a variety of root exudates, which can change rhizosphere pH and increase heavy metal solubility [59]. An experimental approach frequently employed in studying the mobility, transport and bioavailability of metals in different types of environmental samples (soils, sediments) is the use of sequential extraction methods. Different sequential extraction methods have been developed over the years [60,61,62], but widely accepted and applied to metal fractionation in different environmental samples have been developed by the Community Bureau of Reference (BCR) [63,64]. Sequential extraction allows the determination of different forms of heavy metals in sediment using individual extraction agents, because the metals are distributed among the following fractions: exchangeable, carbonates-bound (BCR1), reducible, Fe-Mn oxides-bound (BCR2), oxidizable, organic matter-bound (BCR3), and residual (BCR4). The more mobilizable metals correspond to the first two fractions, which can be released simply by increasing the ionic strength and by slight pH changes [32,65]. The fractionation methods provide relevant information about the possible metal content that could be released in the environment.

Based on the obtained results of the BCR sequential extraction (Figure 1), the distribution of metals among different geochemical fractions varied considerably depending on the applied treatment. All the metals investigated, except Cd, were present in the less available oxidizable and residual fractions. The proportion of bioavailable metal forms was, however, strongly affected by the addition of LMWAO.

In the start sediment (SS), the bioavailable fraction of all analyzed metals accounted for approximately 0.148% for Cr, 2.92% for Ni, 2.30% for Cu, 0.185% for Pb, and 15.5% for Cd (Table 1). After the application of different treatments, notable shifts were observed. Treatments involving the addition of LMWAO, especially treatment with GA 8 W, for all metals except Ni, resulted in a significant increase in the exchangeable fraction of certain metals. For example, the highest share of exchangeable fraction was observed in treatment S GA 8 W (Cr increased to about 3.87%, Cu to 14.7%, Pb to 3.38%, and Cd to 44.3%). Also, the influence of the acid largely depends on the plant species used for phytoremediation [66]. It can be observed that the percentage of the exchangeable fraction for all metals was higher in the experiments with sorghum. Although LMWOAs treatments increased the proportion of metals in the exchangeable fraction, which may facilitate phytoextraction processes [17,67], these changes did not always result in greater metal accumulation in plant tissues. Such an outcome is not unexpected during short phytoremediation experiments and reflects the difference between potential bioavailability, which can be estimated using fractionation methods such as BCR, and effective bioavailability, i.e., actual uptake by the plant [68]. The transfer of metals from sediment to the root surface depends not only on their chemical speciation but also on kinetic factors, including diffusive transport through pore water, desorption kinetics from solid phases, and competition with the major cations Ca²⁺, Mg²⁺, Fe²⁺, and protons for transport sites in the root epidermis [1]. Furthermore, an exposure time of 6–8 weeks may be too short to establish equilibrium of mobilized metals with the rhizosphere, especially for metals that exhibit slow desorption rates, such as Pb and Cr [69]. All these factors, therefore, partially explain why, despite the increase in the exchangeable fraction, the actual metal uptake was rather modest and the translocation factors low (TF < 1.0), as shown in Table 3. The results therefore suggest that LMWOA increased the mobility and reactivity of metals in the sediment, but that most of the mobilized metals were retained at the sediment-root interface, rather than being absorbed and translocated into the shoots, which is consistent with phytostabilization as the predominant pathway. However, the overall effect varied among metals and treatments, reflecting the complex interactions between organic acids, sediment composition, and plant activity.

3.3. Influence of LMWAO Application on Heavy Metal Uptake

The results of the pot experiment demonstrated distinct differences in biomass production and heavy metal accumulation (Figure 2) between hemp (Cannabis sativa L.) and sorghum (Sorghum bicolor L.) under different treatments with TA and GA over 6- and 8-week periods.

As shown in Figure 2, one-way ANOVA confirmed significant treatment effects for both shoot biomass (F(11, 24) = 12.36, p < 0.001) and root biomass (F(11, 24) = 9.58, p < 0.001), while pairwise differences were evaluated using the Games–Howell post hoc test. Hemp produced significantly higher shoot and root biomass compared to sorghum across all treatments, indicating its stronger adaptability and tolerance to experimental conditions. The highest total biomass of hemp was observed after 8 weeks, in both TA and GA treatments. Similar observations have been reported by Farid et al. [23] and Khan et al. [70], where the addition of LMWOAs stimulated plant growth and concurrently enhanced metal uptake by increasing metal solubility and rhizosphere availability. In contrast, sorghum exhibited a slower biomass increase, with significantly lower values, especially in 6 weeks. An improvement was noticed after 8 weeks, which may reflect gradual physiological adaptation and increased root activity over time. The observed differences marked by letters (p < 0.05) confirm significant treatment effects, especially for hemp under TA and GA exposure.

The accumulation patterns of heavy metals (Cu, Pb, Cd, Cr, and Ni) shown in Figure 3 further illustrate the species-specific responses to treatments. Hemp generally accumulated higher quantities of Cu, Pb, and Cd in the above-ground biomass, indicating its potential for phytoextraction [71]. On the other hand, sorghum exhibited a greater accumulation of Cr and Ni in below-ground tissues, implying a stronger phytostabilization capacity [72]. The use of organic acids influenced the metal uptake differently depending on the metal type. For example, TA treatments appeared to enhance translocation of Cu and Pb to shoots, while GA treatments favored metal retention in roots, especially for Cr and Ni.

Although the graphs in Figure 3 show visible differences in metal accumulation between treatments, these patterns were not statistically significant (p > 0.05). This outcome is primarily due to the relatively high within-treatment variability and the limited number of replicates (n = 3), both of which reduce the statistical power of the analysis and may prevent the detection of significant differences despite the apparent visual trends. Nevertheless, the observed patterns still provide useful qualitative insight into species-specific responses and the distribution of metals between above- and below-ground plant tissues [71,72].

The differences between 6- and 8-week exposures highlight the importance of exposure duration for both biomass development and metal accumulation. Longer exposure (8 weeks) allowed plants to accumulate higher metal quantities, coinciding with the increased biomass observed for hemp. However, the variability in standard deviations and overlapping error bars indicates that metal accumulation efficiency was metal- and species-dependent rather than linearly correlated with biomass increase.

Hemp demonstrated a higher potential for phytoextraction due to its greater shoot biomass and ability to accumulate substantial amounts of metals in the above-ground part, while sorghum showed characteristics suitable for phytostabilization through root retention of certain metals (except Cd, than can be attributed to its high bioavailability and physicochemical properties similar to essential micronutrients such as Zn, which allows it to be easily absorbed and translocated by plants [73]. The differential influence of tartaric and glutamic acids suggests that organic amendments can modulate the uptake and distribution of metals within plant tissues, offering valuable insight into the optimization of chelate-assisted phytoremediation strategies.

3.4. Assessment of Phytoremediation Efficiency

The bioaccumulation (BAF) (

Table 5. Bioaccumulation factor (BAF) for hemp and sorghum metal uptake.

Treatments	Cr BAF			Ni BAF			Cu BAF			Cd BAF			Pb BAF
	Root	Shoot	Total	Root	Shoot	Total	Root	Shoot	Total	Root	Shoot	Total	Root	Shoot	Total
H C 6 W	0.080 ± 0.017	0.048 ± 0.011	0.053 ± 0.008	0.336 ± 0.185	0.112 ± 0.038	0.148 ± 0.056	0.175 ± 0.020	0.103 ± 0.026	0.113 ± 0.020	0.519 ± 0.315	0.050 ± 0.013	0.121 ± 0.063	0.096 ± 0.114	0.016 ± 0.003	0.014 ± 0.003
H TA 6 W	0.049 ± 0.018	0.041 ± 0.010	0.041 ± 0.010	0.037 ± 0.008	0.137 ± 0.014	0.145 ± 0.011	3.674 ± 1.179	0.084 ± 0.008	0.088 ± 0.007	0.383 ± 0.180	0.117 ± 0.017	0.166 ± 0.039	0.026 ± 0.008	0.013 ± 0.004	0.012 ± 0.005
H GA 6 W	0.140 ± 0.036	0.044 ± 0.010	0.059 ± 0.006	0.070 ± 0.031	0.108 ± 0.014	0.143 ± 0.025	4.321 ± 0.514	0.150 ± 0.029	0.150 ± 0.023	0.208 ± 0.032	0.050 ± 0.010	0.074 ± 0.011	0.034 ± 0.009	0.013 ± 0.002	0.009 ± 0.001
H C 8 W	0.031 ± 0.008	0.011 ± 0.002	0.014 ± 0.003	0.018 ± 0.006	0.056 ± 0.029	0.060 ± 0.028	2.284 ± 0.580	0.048 ± 0.007	0.050 ± 0.004	0.159 ± 0.112	0.032 ± 0.011	0.047 ± 0.014	0.009 ± 0.005	0.022 ± 0.029	0.221 ± 0.330
H TA 8 W	0.076 ± 0.049	0.016 ± 0.002	0.026 ± 0.007	0.048 ± 0.036	0.088 ± 0.024	0.113 ± 0.017	3.621 ± 1.446	0.070 ± 0.007	0.077 ± 0.006	0.327 ± 0.132	0.055 ± 0.006	0.098 ± 0.027	0.030 ± 0.014	0.009 ± 0.005	0.046 ± 0.015
H GA 8 W	0.032 ± 0.002	0.006 ± 0.002	0.010 ± 0.002	0.034 ± 0.009	0.050 ± 0.014	0.065 ± 0.018	3.067 ± 0.386	0.070 ± 0.013	0.076 ± 0.014	0.213 ± 0.065	0.029 ± 0.008	0.056 ± 0.015	0.015 ± 0.002	0.004 ± 0.001	0.032 ± 0.012
S C 6 W	0.710 ± 0.442	0.113 ± 0.029	0.224 ± 0.062	0.290 ± 0.175	0.260 ± 0.110	0.483 ± 0.147	17.537 ± 8.061	0.176 ± 0.094	0.250 ± 0.072	1.825 ± 0.367	2.804 ± 2.037	2.519 ± 1.457	0.237 ± 0.122	0.021 ± 0.008	0.002 ± 0.001
S TA 6 W	0.402 ± 0.002	0.075 ± 0.005	0.112 ± 0.009	0.228 ± 0.023	0.166 ± 0.011	0.262 ± 0.024	15.877 ± 2.908	0.117 ± 0.008	0.163 ± 0.003	1.568 ± 0.072	1.403 ± 0.284	1.417 ± 0.267	0.157 ± 0.010	0.016 ± 0.008	0.002 ± 0.000
S GA 6 W	0.799 ± 0.263	0.360 ± 0.382	0.437 ± 0.293	0.294 ± 0.074	0.302 ± 0.102	0.466 ± 0.013	18.663 ± 3.290	0.160 ± 0.002	0.236 ± 0.035	1.298 ± 0.232	1.389 ± 0.291	1.376 ± 0.281	0.278 ± 0.056	0.025 ± 0.003	0.002 ± 0.001
S C 8 W	1.009 ± 0.663	0.026 ± 0.012	0.146 ± 0.087	0.422 ± 0.265	0.089 ± 0.050	0.298 ± 0.177	14.967 ± 8.548	0.077 ± 0.023	0.127 ± 0.052	1.774 ± 0.237	1.078 ± 0.146	1.167 ± 0.104	0.323 ± 0.095	0.008 ± 0.002	0.011 ± 0.005
S TA 8 W	1.652 ± 0.871	0.033 ± 0.003	0.213 ± 0.098	0.511 ± 0.232	0.080 ± 0.006	0.320 ± 0.118	16.754 ± 10.940	0.084 ± 0.023	0.140 ± 0.027	1.561 ± 0.426	0.916 ± 0.427	0.988 ± 0.406	0.411 ± 0.328	0.016 ± 0.014	0.015 ± 0.007
S GA 8 W	1.008 ± 0.387	0.027 ± 0.007	0.153 ± 0.079	0.291 ± 0.114	0.065 ± 0.009	0.216 ± 0.097	20.786 ± 6.847	0.119 ± 0.006	0.195 ± 0.044	1.695 ± 0.444	0.879 ± 0.234	0.981 ± 0.278	0.419 ± 0.157	0.007 ± 0.002	0.018 ± 0.006

Standard deviations are based on three replications (n = 3).

) and translocation factors (TF) (

Table 6. Translocation factor (TF) for hemp and sorghum metal uptake.

Treatments	Cu	Pb	Cd	Cr	Ni
H C 6 W	0.602 ± 0.210	0.400 ± 0.314	0.120 ± 0.066	0.616 ± 0.218	0.363 ± 0.131
H TA 6 W	0.787 ± 0.336	0.517 ± 0.211	0.338 ± 0.109	0.926 ± 0.421	0.771 ± 0.183
H GA 6 W	1.065 ± 0.284	0.414 ± 0.178	0.241 ± 0.031	0.339 ± 0.167	0.366 ± 0.172
H C 8 W	0.764 ± 0.315	3.107 ± 4.339	0.283 ± 0.225	0.338 ± 0.048	0.602 ± 0.203
H TA 8 W	0.714 ± 0.411	0.298 ± 0.123	0.187 ± 0.079	0.287 ± 0.185	0.553 ± 0.437
H GA 8 W	0.620 ± 0.075	0.237 ± 0.061	0.136 ± 0.010	0.199 ± 0.061	0.349 ± 0.090
S C 6 W	0.373 ± 0.289	0.107 ± 0.073	1.616 ± 1.251	0.226 ± 0.189	0.253 ± 0.242
S TA 6 W	0.217 ± 0.031	0.101 ± 0.044	0.891 ± 0.144	0.187 ± 0.013	0.160 ± 0.018
S GA 6 W	0.247 ± 0.041	0.094 ± 0.033	1.067 ± 0.041	0.583 ± 0.737	0.249 ± 0.158
S C 8 W	0.172 ± 0.042	0.027 ± 0.010	0.619 ± 0.152	0.029 ± 0.009	0.050 ± 0.005
S TA 8 W	0.200 ± 0.133	0.075 ± 0.089	0.595 ± 0.214	0.024 ± 0.011	0.040 ± 0.014
S GA 8 W	0.179 ± 0.070	0.019 ± 0.009	0.524 ± 0.087	0.030 ± 0.013	0.057 ± 0.021

Standard deviations are based on three replications (n = 3).

) varied notably among treatments and between the two plant species studied. Overall, the application of TA and GA influenced both metal uptake and redistribution within plant tissues, reflecting their potential role as low-molecular-weight organic chelators that can enhance metal mobility and bioavailability in sediments.

Effect on metal accumulation and BAF. In hemp, the highest root BAF values (Table 5) were generally observed for Cu, particularly under GA treatment after six weeks (BAF_root 4.32), indicating pronounced metal accumulation in the rhizosphere. Similar trends, though of lower intensity, were observed for Cd (BAF_root up to 0.52 in control) and Ni (BAF_root 0.07 under GA). Although hemp produced a high shoot biomass and accumulated notable amounts of metals in the above-ground parts (Figure 3), indicating potential for phytoextraction, the BAF values showed that metal accumulation was generally higher in roots than in shoots. This suggests that, under the applied conditions, hemp primarily acted through a phytostabilization mechanism, retaining most metals in the rhizosphere while still exhibiting a moderate capacity for translocation to aerial tissues. Such dual behavior has also been reported by Placido and Lee [74] and Cleophas et al. [75], who observed limited translocation of heavy metals from roots to aerial parts in hemp grown on contaminated soils.

In sorghum, similar patterns were observed (Table 5). The BAF values for Cu and Ni were slightly higher than those in hemp, particularly under the GA treatments, indicating a somewhat greater capacity of sorghum to accumulate metals in the presence of LMWOAs. This trend is consistent with the concentration data (Figure 3), where sorghum showed pronounced metal retention in roots—especially for Cu and Ni—confirming its predominant role in phytostabilization. The limited metal transfer to shoots and the relatively low shoot BAF values further support the notion that sorghum effectively immobilizes metals within the root zone. These findings align with those of Balafrej et al. [76], Aryal [15], and Tamma et al. [77], who reported enhanced Cu and Zn uptake in sorghum under organic acid amendments, suggesting that root exudates or externally supplied organic acids can increase metal solubility and promote root absorption.

Translocation factor (TF) and metal mobility. TF values (the ratio of metal concentration in shoots to that in roots) were generally below 1.0 for all metals and treatments (except for H GA 6 W for Cu, H C 8 W for Pb, and S C 6 W and S GA 6 W for Cd), confirming limited metal translocation to aboveground tissues. The consistently low TF values (<1 in most cases) further support the phytostabilization potential of both species rather than phytoextraction. However, glutamic acid, as a naturally occurring amino acid, can form stable complexes with divalent cations and facilitate their transport through the xylem [43,44]. In contrast, tartaric acid exhibited a milder effect on translocation. This is consistent with the findings of Khan et al. [78], who reported that tartaric acid, although capable of increasing metal solubility in soil, tends to promote root retention rather than shoot transport due to its relatively weak chelating stability constants compared to amino acids.

Overall, the combined assessment of bioaccumulation and translocation factors indicates that both hemp and sorghum primarily rely on phytostabilization mechanisms, with only limited metal transfer to above-ground tissues. The presence of low-molecular-weight organic acids, particularly glutamic acid, slightly increased metal mobility and uptake but did not alter the overall behavior toward effective phytoextraction. These findings confirm that both crops can be effectively used in the phytomanagement of contaminated sediments, emphasizing metal immobilization rather than extraction, and highlighting their potential for stabilizing dredged materials prior to their reuse in environmental applications.

In summary, the BAF and TF results show clear and consistent differences between plant species and organic acid treatments. Sorghum showed a stronger root-based stabilization capacity, especially for Cr and Ni, which remained predominantly retained below ground. Hemp, in contrast, tended to allocate a slightly larger fraction of Cu, Pb and Cd to the shoots, although translocation remained limited (TF < 1.0), confirming that both species predominantly stabilized metals in the rhizosphere. Among the two acids, glutamic acid produced more pronounced effects, especially by improving the mobility of Cu and Cd and increasing their uptake compared to controls. It is important to note that even in treatments where GA increased the fraction of exchangeable metals, most of the mobilized metals were still confined to the root zone, indicating a dominance of phytostabilization rather than phytoextraction.

3.5. Environmental Risk Assessment: Leaching Potential Under GA and TA-Induced Metal Mobilization

Although the application of LMWOAs is intended to enhance phytoextraction efficiency, their use may also introduce risks related to excessive metal mobilization in the sediment matrix [79,80]. In particular, the results of the BCR fractionation analysis (Figure 1) indicate that GA treatment substantially increased the proportion of metals in the exchangeable fraction, which represents the most labile and environmentally mobile pool [23,57]. This effect was especially pronounced for Cu, Pb, Cr, and Cd in the S GA 8 W treatment, where exchangeable fractions reached 14.7% for Cu, 44.3% for Cd, and 3.87% for Cr, compared with <3% in the initial sediment. Such increases suggest that GA facilitates desorption of metals from organic matter and oxide-bound phases into more soluble, weakly bound forms [81]. Given that exchangeable metals are highly susceptible to displacement through percolating water, these findings highlight a potential risk of downward leaching and subsequent contamination of groundwater [82]. This risk is particularly relevant for light-textured sediments such as those used in this study (73–83% sand), where lower water retention and reduced sorption capacity may further promote the vertical transport of mobilized metals. Although the POT setup did not allow for direct measurement of leachates, the strong increase in metal mobility under GA indicates that enhanced phytoextraction may come at the cost of increased leaching potential, especially in field conditions with natural rainfall or irrigation.

The contrasting behavior of TA-treated variants supports this interpretation: tartaric acid maintained relatively stable exchangeable fractions, suggesting a milder chelating effect and lower risk of unintended metal mobilization. These observations emphasize the need for a cautious, site-specific evaluation of LMWOA-assisted phytoremediation strategies, particularly when amino-acid-based amendments are used. Future field trials should include monitoring of percolating water or lysimeter-based assessments to quantify leaching fluxes under different treatments. Additionally, controlled dosing strategies or split applications may help minimize excessive mobilization while still improving root-zone availability for plant uptake. Overall, while GA-enhanced treatments showed clear benefits in terms of metal solubility and potential bioavailability, the associated increase in exchangeable metal fractions underscores a critical environmental trade-off. Integrating mobility risk assessments into phytoremediation protocols is therefore essential to ensure that improvements in uptake efficiency do not lead to secondary contamination pathways.

4. Discussion of Future Directions in the Phytoremediation Process Management and Valorization of Post-Phytoremediation Biomass

Beyond soil restoration, phytoremediation generates significant amounts of plant biomass, whose proper management and valorization are key to ensuring the overall sustainability of the technology. The integration of biomass into circular economy concepts, through energy production or material recovery, has become increasingly important to offset the economic costs of remediation and to mitigate waste generation [83].

The management and valorization of post-phytoremediation biomass represent a critical step in ensuring the sustainability and long-term feasibility of phytoremediation technologies [57]. Once plants have accumulated or stabilized contaminants, the resulting biomass often contains elevated levels of potentially toxic (inorganic and/or organic) elements, raising concerns regarding its safe disposal or further utilization [70]. Consequently, researchers have explored diverse strategies for handling such biomass, including thermal conversion (e.g., combustion, pyrolysis), biological processes (e.g., composting, anaerobic digestion), and innovative approaches such as phyto-mining or nanoparticle synthesis. The table below (

Table 7. Post-phytoremediation biomass management, highlighting their main findings, challenges, and potential applications.

	Results	Challenges	Applications	Future Research
Mukherjee et al., [84]	Pyrolysis is identified as the most effective method for converting post-phytoremediation biomass into biofuel while minimizing toxic metal transfer. S. marianum showed a higher biogas yield (190 mL g−1) than H. annuus (130 mL g−1).	Safe disposal of metal-contaminated biomass, toxicity of hyperaccumulators, and lack of clear regulations limit practical use.	Biomass can be used for bioenergy, nanomaterial production, or as ash in agriculture and landfills.	Improve efficiency via gene editing, study contaminant transformations, and develop sustainable biomass management.
Wang et al., [85].	The review highlights the crucial role of plant growth-promoting bacteria (PGPB) in enhancing plant growth and resistance under metal stress, emphasizing their contribution to phytoremediation efficiency and soil health improvement. It also notes the impact of climate change on plant–metal interactions and the importance of understanding rhizosphere processes.	Poor genetic stability of PGPB, competition with native microbes, and sensitivity to pH and temperature limit field application. Limited research on PGPB metabolomics and behavior under environmental stress conditions.	Use of PGPB and genetically modified organisms (GEMs) for improving phytoremediation of metal-contaminated soils; bioaccumulating plants and microbial consortia can enhance pollutant removal in controlled environments.	Focus on understanding plant–microbe–metal interactions under climate stress, optimizing PGPB–plant combinations, improving metal recovery and safe disposal of contaminated biomass, and developing molecular-level insight into rhizosphere ecology.
Gomez et al., [86]	Thermal analysis of biomass from plants grown on contaminated soils showed that Salvia rosmarinus and R. sphaerocarpa had the best combustion performance, while B. juncea from highly contaminated soils left more residue and showed lower activation energy. The study confirms that thermal characteristics determine suitability of phytoremediation biomass for energy recovery.	Variations in thermal behavior depending on contamination level complicate uniform energy valorization of phytoremediation biomass.	Integration of phyto-remediation with bioenergy production; recommendation of R. sphaerocarpa and S. rosmarinus for highly contaminated soils, and B. juncea for mildly contaminated soils as solid biofuels.	Further investigation of combustion and energy recovery potential of phytoremediation biomass under different contamination levels and soil conditions.
Kowalska and Biczak, [83]	The paper emphasizes the importance of plant biomass and microorganisms in phyto-remediation for regenerative agriculture, highlighting their role in improving soil quality, biodiversity, carbon sequestration, and bioenergy production. It underlines the relevance of EU environmental policies such as the Green Deal and Renewable Energy Directive III.	Implementing large-scale phytoremediation while balancing food security and economic feasibility; variability in biomass composition affects uniform biofuel production.	Post-phytoremediation biomass can be transformed into biofuels and bioproducts, supporting circular economy goals and sustainable agricultural systems.	Further development of bio-based technologies for biomass valorization and assessment of policy-driven impacts on soil restoration and energy sustainability.
Jiang et al., [87]	The study evaluated the economic and technical feasibility of phytoremediation and element recovery, finding returns in the order Ni > Pt > As, with profits of £1265–2975 ha−1 (Ni) and £887–2124 ha−1 (Pt). High biomass-yielding species improved profitability, and success was linked to high bioaccumulation and translocation factors.	Long treatment periods, site-specific contamination, and biomass disposal challenges limit large-scale application. Economic risks and uncertain profit margins reduce financial appeal. Ignoring dependencies in risk models may lead to biased interpretations.	Integration of land remediation with biomass-to-energy conversion and element recovery (“phytomining”) enhances economic and environmental value. Monte Carlo simulations support feasibility assessment.	Focus on maximizing by-product utilization and integrating energy recovery with metal reclamation. Transition from phytoremediation to phytomining to improve profitability; develop efficient conversion technologies and safe biomass use.
Edgar et al., [88]	Phytoremediation effectively removes heavy metals using non-edible, metal-tolerant plants. Biomass produced can be used for bioenergy or converted into metal-enriched materials for catalysts or reuse. Engineering strategies can improve retention of metals in fewer fractions, minimizing remobilization.	Technical inefficiencies in thermochemical conversion, microbial remobilization of metals, and limited field-scale success hinder application. Selecting optimal plant species and balancing economic and ecological objectives remains complex.	Valorization of phyto-remediation biomass through bioenergy, biogas, and catalyst production. Use of biowaste-derived polymers (e.g., canola oil–sulfide) for mercury removal in sludge remediation. Biotechnological approaches like bioaugmentation and biostimulation reduce pollutant toxicity and costs.	Develop improved engineering designs for metal retention, expand genomics-based control of metal uptake, and explore metal-rich fractions as sources for catalysts. Strengthen safety screening before environmental deployment of phytotechnologies.
Bernal et al., [89]	The study evaluated biomass from plants grown on contaminated soils for bioenergy potential. Silybum marianum and Piptatherum miliaceum showed the best biogas yields via anaerobic digestion, while Nicotiana glauca exhibited low production due to high Pb, lignin, and C/N ratio. High mineral and ash contents may cause fouling during combustion.	Trace element (TE) transfer to aerial plant parts limits bioenergy use due to potential pollution. High TE levels in biomass exceed composting limits. Fouling and slagging in combustion systems and slow degradation of organic compounds hinder efficient processing.	Application of phyto-remediation biomass for anaerobic digestion, composting, and pyrolysis-derived biochar production. Energy crops such as S. marianum and P. miliaceum are proposed for renewable energy use and soil phytostabilization.	Long-term validation and techno-economic assessment of phytostabilization-based bioenergy systems. Further research on TE effects on microbial populations in composting and digestion, and potential biochar production from pyrolysis.

) summarizes selected studies addressing various aspects of post-phytoremediation biomass management, highlighting their main findings, challenges, and potential applications. These studies collectively emphasize that while post-remediation biomass offers opportunities for renewable energy production, material recovery, and circular economy integration, significant limitations remain, particularly related to contaminant transfer risks, technological efficiency, and economic feasibility. Future research directions, therefore, focus on optimizing treatment processes, improving contaminant recovery, and developing regulatory and technological frameworks to ensure environmentally sound and economically viable biomass utilization.

Integrating circular economy principles into phytoremediation can significantly enhance the sustainability and overall efficiency of remediation systems. Biomass valorization transforms phytoremediation from a waste-generating process into a resource-recycling system, where the harvested plant material becomes a valuable input for secondary applications rather than disposal [83]. Renewable energy generation through thermal (e.g., combustion, pyrolysis) and biochemical (e.g., anaerobic digestion, fermentation) conversions can partially offset remediation costs and contribute to bioenergy production [90]. Metal recovery through phytomining provides opportunities to extract and recycle valuable metals accumulated in plant tissues, turning contaminated biomass into a resource for the circular economy [91]. Finally, soil health enhancement via composting or biochar production can increase soil organic matter, improve nutrient availability, and support sustainable land restoration [92].

5. Conclusions

This study provides clear experimental evidence that low-molecular-weight organic acids can substantially influence the physicochemical behavior of contaminated sediments and improve the overall performance of phytoremediation systems. The addition of GA induced pronounced redox activation, increasing the sediment’s electrical conductivity up to 1189 µS cm⁻¹ and cation exchange capacity to 58.5 cmolc kg⁻¹, indicating enhanced ion exchange and metal mobility. In contrast, TA acted more conservatively, maintaining stable pH conditions with gradual nutrient transformation. Both Cannabis sativa and Sorghum bicolor contributed to organic matter turnover and nutrient cycling, confirming their potential in the phytomanagement of nutrient- and metal-enriched sediments; hemp favored metal accumulation in shoots (notably Cu, Pb, Cd) and sorghum showed strong retention of Cr and Ni in roots (up to 80% of total uptake). However, for the observed differences in metal uptake and distribution between the two species, hemp showed higher stabilization capacity and sorghum showed higher translocation potential, highlighting the importance of crop selection based on site-specific remediation goals. Despite encouraging outcomes, the transition from laboratory-scale experiments to practical biomass valorization remains challenging. Risks of metal volatilization during thermal treatment, inhibition of biological conversion processes, and limitations in the agricultural reuse of contaminated by-products present significant technical and regulatory constraints. Addressing these barriers through improved process optimization, risk assessment, and regulatory harmonization will be essential to integrate phytoremediation with sustainable biomass utilization and circular economy strategies.