Open AccessArticle

Induced Phytomanagement of Multi-Metal Polluted Soil with Conocarpus erectus Supported by Biochar, Lignin, and Citric Acid

Hafiz Muhammad Tauqeer

¹,

Karolina Lewińska

Muhammad Umar

³,

Faisal Mahmood

⁴

Tanvir Shahzad

⁴

Faiqa Sagheer

⁵,

Hina Sajid

⁵,

Iqra Chaudhary

⁵ and

Muhammad Iqbal

^4,*

Department of Environmental Sciences, University of Gujrat, Gujrat 50700, Pakistan

Department of Soil Science and Remote Sensing of Soils, Faculty of Geographical and Geological Sciences, Adam Mickiewicz University in Poznan, ul. Bogumiła Krygowskiego 10, 61-680 Poznań, Poland

Department of Neurosurgery, Allied Hospital Faisalabad, Faisalabad 38000, Pakistan

⁴

Department of Environmental Sciences, Government College University, Faisalabad 38000, Pakistan

⁵

Department of Biochemistry, Institute of Biochemistry, Biotechnology and Bioinformatics, The Islamia University of Bahawalpur, Bahawalpur 63100, Pakistan

Author to whom correspondence should be addressed.

Minerals 2024, 14(11), 1149; https://doi.org/10.3390/min14111149

Submission received: 27 August 2024 / Revised: 1 November 2024 / Accepted: 9 November 2024 / Published: 13 November 2024

(This article belongs to the Special Issue Potentially Toxic Elements in Soils Affected by Metal Mining and Processing, 2nd Edition)

Download

Browse Figures

Figure 1
The influences of biochar (BC), lignin (LN), and citric acid (CA) on shoot dry weight (SDW) (a) and root dry weight (RDW) (b) of C. erectus grown on shooting-range soil (SS). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by one-way ANOVA, followed by an LSD test. "> Figure 2
Influences on plant height (A), chlorophyll a (Chl-a) (B), chlorophyll b (Chl-b) (C), relative water content (RWC) (D), water uptake capacity (WUC) (E), protein (F), amino acids (G), ascorbic acid (AsA) (H), and total phenolics (I) of C. erectus grown on shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test. "> Figure 3
The concentrations of Pb, Cr, Cd, Ni, and Cu (a,c,e,g,i) in C. erectus tissues and labile portions of Pb, Cr, Cd, Ni, and Cu (b,d,f,h,j) in post-harvest shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test. "> Figure 4
The contents of Pb, Cr, Cd, Ni, and Cu (a,c,e,g,i) in C. erectus tissues and total removal of Pb, Cr, Cd, Ni, and Cu (b,d,f,h,j) by C. erectus from the shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test. "> Figure 5
The influences of biochar (BC), lignin (LN), and citric acid (CA) applications on Pb (a), Cr (b), Cd (c), Ni (d), and Cu (e) concentrations in the leachates collected at 1st (30 days), 2nd (60 days), 3rd (90 days), and 4th (120 days) sampling periods. The standard errors are calculated from the triplicates of each treatment. Dashed lines depict the critical limits of Pb, Cr, Cd, Ni, and Cu for safe water reuse and agricultural purposes [<a href="#B39-minerals-14-01149" class="html-bibr">39</a>,<a href="#B40-minerals-14-01149" class="html-bibr">40</a>]. "> Figure 6
The influences of biochar (BC), lignin (LN), and citric acid (CA) on protease (a), chitinase (b), urease (c), acid phosphatase (d), and catalase (e) activities in shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated using a one-way ANOVA, followed by an LSD test. ">

Versions Notes

Abstract

Induced heavy metals (HMs) phytoextraction from heavily contaminated soils is challenging, as high HM bioavailability causes phytotoxicity and leaching. This study introduces a novel approach for HM immobilization with biochar (BC) and lignin (LN), and later their controlled mobilization with citric acid (CA) in soil. Conocarpus erectus was grown for 120 days in shooting-range soil (SS) polluted with Pb, Cr, Cd, Ni, and Cu. HM concentrations in parts of the plants, their percentage removal, and leaching from SS were measured. Moreover, plant biochemical parameters such as the contents of chlorophyll a (Chl-a), chlorophyll b (Chl-b), protein, ascorbic acid (AsA), amino acids, and total phenolics, along with biophysical parameters such as relative water content (RWC) and water uptake capacity (WUC), were also inspected. Adding BC, LN, and BC+LN to SS improved biomass, as well as the biophysical and biochemical parameters of plants, while efficiently reducing HM concentrations in plant parts, DTPA extract, and leachates compared to the control (CK). However, the greatest amplifications in plant height (82%), dry weight of root (RDW) (109%), and dry weight of shoot (SDW) (87%), plant health, and soil enzymes were noted with the BC+LN+CA treatment, compared with the CK. Moreover, this treatment resulted in Pb, Cr, Cd, Ni, and Cu removal by 68, 30, 69, 59, and 76% from the SS compared to the CK. Surprisingly, each HM concentration in the leachates with BC+LN+CA was below the critical limits for safer water reuse and agricultural purposes. Initial HM immobilization in HM-polluted soils, followed by their secured mobilization during enhanced phytoextraction, can enhance HM removal and reduce their leaching without compromising plant and soil health.

Keywords:

phytoextraction; heavy metals; mobilization; contents; leachates

1. Introduction

Soil pollution through anthropogenic sources such as mining and shooting ranges has become a swift environmental concern due to huge HMs concentrations released from them [1,2]. Primarily, the HMs found in these soils are Pb, Cd, Cu, Ni, Cr, Sb, and Zn [3,4,5]. For instance, approximate quantities of Pb (up to 521 mg kg⁻¹), Cd (up to 25 mg kg⁻¹), Zn (up to 1365 mg kg⁻¹), and Cu (up to 452 mg kg⁻¹), as well as other HMs, were found in Pb–Zn mining soils (MSs) [3]. Likewise, due to intense shooting activities at numerous shooting ranges worldwide, they exhibited higher Pb concentrations, up to 54,600 mg kg⁻¹ soil, owing to the dissociation of bullets and their fragments [1,6,7].

Both MSs and shooting ranges pollute the surrounding ecosystem with HMs through aerial deposition of windblown dust to nearby arable lands, erosion, and leaching, resulting in HM accumulation in the food cycle [1,3,5]. Additionally, HM mobility from shooting ranges has resulted in the contamination of nearby streams and sediments, leading to their accumulation in Salmo trutta [8]. Higher HM concentrations in shooting-range soil (SS) negatively affect microbial ecology, earthworms, and plants [5]. Recently, Pb poisoning in many terrestrial and aquatic birds, raptors, and scavengers due to ingesting Pb pellets from soil and mud has resulted in their demise [9]. Unfortunately, ruminants grazing on the grass growing on SS exhibited higher Pb accumulation in their liver, which compromised human health after consuming their offal [10].

Conventional remediation approaches to decontaminate SS are not practiced due to limitations such as their high cost, disturbance of soil functions, and air pollution from the dust particles harboring HMs [5]. Contrary, cheap and eco-friendly remediation techniques such as HM immobilization with different organic and inorganic soil amendments [4,7,11,12] and HMs removal/stabilization through HMs hyperaccumulator/stabilizing plants [1,6] have been of great choice for the management of SS. For the plants to be considered hyperaccumulators of HMs, they must have shoot concentrations of 1000 mg kg⁻¹ for Pb and Ni, 100 mg kg⁻¹ for Cd, and 300 mg kg⁻¹ for Cr and Cu. Moreover, such plants also exhibit values of bioconcentration factor [BCF (shoot to soil HMs concentration ratio)] and translocation factor [TF (shoot to root HMs concentration ratio)] >1 [13]. Apart from it, plants utilized for HM phytoextraction also have several features, including (1) extreme tolerance against HM phytotoxicity and other harsh environmental conditions; (2) high biomass production, ensuring higher contents of HM removal from the soil over time; and (3) mitigation of the risk of secondary pollution [11,14,15]. Interestingly, the removal of HMs from HMs-polluted soils through enhanced phytoextraction is an even more effective, and cheaper option [14,15]. In enhanced phytoextraction, the bioavailability of HMs in the soil is accelerated through several organic acids and inorganic additives that facilitate the solubility of HMs in soil and their uptake in plant-harvestable parts [15,16,17]. However, growth retardation in plants due to HMs phytotoxicity can limit their establishment on the remediation site, resulting in ground and surface water contamination through HM leaching during their enhanced phytoextraction [1,14].

To address these limitations, we propose a new management option for the removal of HMs from SS by HM immobilization with biochar (BC) and lignin (LN) in the first step, and subsequently induced phytoextraction by C. erectus through the supplementation of SS with citric acid (CA). Biochar and LN immobilize HMs in soil, reduce their prevalence in leachates, support plant growth, and restore soil health [2,18,19]. On the other hand, CA positively enhances the growth of plants, mobilizes HMs in polluted soils, and improves the soil environment [15]. C. erectus is a fast-growing, hardy, deeply rooted, and densely branched shrub that can easily grow in soils facing different stresses, including HMs pollution, poor drainage, and salinity, and thrives in extreme weather conditions [20,21]. Previously, C. erectus accumulated Pb up to 2800 mg kg⁻¹ dry weight of shoot (SDW), with BCF and TF values >1, showing its potential as a Pb hyperaccumulator. However, it phytostabilized Cd, Ni, and Cr by accumulating them up to 73, 502, and 159 mg kg⁻¹ dry weight of root (RDW), respectively, with BCF and TF values <1 for these HMs [21].

Thus, the current research aimed to test the efficacy of the following procedures: (i) initially amending SS with BC and LN for HMs immobilization to support the growth and establishment of C. erectus; (ii) resolubilizing HMs in the amended SS after plant establishment by adding CA; (iii) assessing the potential of C. erectus to remove HMs from SS; (iv) observing the effects of HM immobilization–resolubilization approach on plant growth, health, and oxidative stress; and (v) monitoring HMs leaching from SS during this process.

2. Materials and Methods

2.1. SS and Its Characteristics

The top surface (0–20 cm) of SS was acquired from “Arabian Sea Shooting Range, Karachi, Pakistan”, with an average summer temperature of 31 °C and 19 °C in winter, and an average annual precipitation of 295 mm. The SS was picked up using a shovel and thoroughly mixed to achieve a composite sample. Afterward, the SS was filled in plastic pouches and transported to the laboratory. The SS was air-dried and screened to remove the spent bullets, stones, and other debris. Later, the SS was subjected to numerous physicochemical analyses by adopting the recommended protocols defined in our earlier studies [21,22]. The descriptions of all physicochemical characteristics of the SS are as follows: sand = 35%, silt = 27%, clay = 38%, texture = clay loam, pH = 6.1, organic matter (OM) = 0.38%, calcium carbonate (CaCO₃) = 7.3%, electrical conductivity (EC) = 5.92 dS m⁻¹, cation exchange capacity (CEC) = 13.2 cmol_c kg⁻¹, N = 112 mg kg⁻¹, P = 9.7 mg kg⁻¹, K = 112 mg kg⁻¹, diethylenetriaminepentaacetic acid (DTPA)-Pb = 28.1 mg kg⁻¹, Cr = 1.48 mg kg⁻¹, Cd = 2.11 mg kg⁻¹, Ni = 2.21 mg kg⁻¹, Cu = 1.84 mg kg⁻¹; total Pb = 4350.1 mg kg⁻¹, Cr = 15.0 mg kg⁻¹, Cd = 23.5 mg kg⁻¹, Ni = 47.6 mg kg⁻¹, and Cu = 270.9 mg kg⁻¹.

2.2. Additives for SS

2.2.1. HM Stabilizing Additives

Feedstock consisting of pruned waste of bougainvillea vine (bougainvillea alba) was used to prepare BC, according to our previous publication [22]. BC has a specific surface area (SA) = 287 m² g⁻¹, EC = 0.74 dS m⁻¹, pH = 8.2, CEC = 31.7 cmol_c kg⁻¹, C = 72.9%, N = 0.71%, H = 2.9%, O = 16.1%, H:C = 0.04, O:C ratio = 0.22, P = 1.81 mg kg⁻¹, K = 1.9 mg kg⁻¹, Ca = 1.71 mg kg⁻¹, Mg = 1.21 mg kg⁻¹, S = 0.06 mg kg⁻¹, Fe = 0.19 mg kg⁻¹, Mn = 0.08 mg kg⁻¹, and Zn = 0.43 mg kg⁻¹. Similarly, the LN was ordered from Jinan Yuansheng Chemical Technology Co., Ltd., China. The physiochemical properties of LN are illustrated as follows: pH = 9.8, C = 60.4%, H = 6.1%, O = 33.2%, N = 0.3%, CEC = 27.9 cmol_c kg⁻¹, H:C = 0.1, O:C = 0.55 [23].

2.2.2. HMs Mobilizing Additive

For the solubilization of HMs in SS, the analytical grade (99% pure) CA was purchased from Sigma-Aldrich, Merck KGaA, Darmstadt, Germany. The chemicals used for other analyses were also of analytical grade.

2.3. Soil Treatments and Pot Experiment

2.3.1. Setting of Plant Experiment

The SS was mixed with BC and LN by using a mechanical mixer to acquire eight treatments (Table 1). The treated SS masses were moistened (65%) with deionized water, tightly packed in zip-lock plastic pouches, and placed for incubation (dark room, 25 °C, and 42 days). Afterward, the moisture content in these incubated treatments was maintained by spraying deionized water and then shaking with a glass rod at seven-day intervals. The amended SS masses were placed in plastic pots (depth = 18 cm, diameter = 15 cm) in triplicates, each pot containing 2.5 kg of soil. The pots were carefully transferred to a wire house (humidity: 40–50%, light: 9 h, temperature: 35 °C (day) and 26 °C (night)) and positioned in a completely randomized design (CRD). Before placing the pots, a plastic tray (length = 20 cm, width = 18 cm, and height = 5 cm) was laid beneath each pot for the collection of leachates. The fresh and healthy cuttings of C. erectus (length ≈ 15 cm, diameter ≈ 0.7 cm) were acquired from a plant nursery. Five cuttings were inserted in each pot. Once the sprouting of cuttings had occurred, three healthy plants were kept in every pot. The plants were fed once with Nutrigold^® NPK 20-20-20+TE (MILAGRO International, Parabiago, Milan, Italy). Routine irrigation was performed with distilled water to maintain 65% soil water holding capacity (WHC). The plants were grown for 120 days.

2.3.2. Citric Acid Application and Collection of Leachates

CA (5 mm kg⁻¹ soil) was applied with irrigation water to the pots, according to the treatment plan (Table 1), at 30, 60, 90, and 120 days indicated as 1st, 2nd, 3rd, and 4th sampling, respectively. After the CA application, the leachates from each tray were carefully collected with sterile syringes, transferred to polyethylene sampling bottles, acidified by adding HNO₃ (100 µL, 65%, purchased from Merck), and preserved in a refrigerator (4 °C) until their analysis.

2.3.3. Termination of the Plant Experiment and Collection of Plant and Soil Samples

Plant height was recorded with the help of a tapeline. Afterward, the plants were clipped right above the soil surface by using sharp pruning shears. The harvested aerial plant biomass was stored in polyethylene bags. The coarser and finer roots were also collected from each pot after overturning the soil on a polyethylene sheet. The coarser roots were collected manually, whereas the finer roots were removed through sifting (2 mm). Afterward, the roots and shoots were individually preserved in polyethylene bags. Likewise, a subsequent amount of soil from every pot was also collected with the help of a spatula and stored in plastic bags. Immediately after the plant harvest, the plant (shoots and roots) and soil samples were transported to the laboratory. The soil samples were sifted (2 mm), air-dried and stowed for analysis. The above- and below-ground plant biomass was first rinsed with running tap water until the removal of soil and later with distilled water. Sub-samples of fresh leaves were also separated for estimating the activities of antioxidants, the contents of reactive oxygen species (ROS), and various other biochemical compounds in plants. Afterward, the plant biomass was kept at 70 °C for 3 days in an oven (Memmert, Beschickung-loading, model 100–800, Schwabach, Germany) to achieve constant weight. The weight of roots and shoots gathered from every pot was measured on weighing balance. Afterward, both plant parts were ground, passed through a mesh (0.5 mm), and saved for their analysis.

2.4. SS and Plant Analysis

2.4.1. HMs in SS, Leachates, Plant Parts, and Their Removal by the Plants

A subsequent amount (1 g) of the dried C. erectus biomass (roots and shoots) was taken, transferred to a 100 mL digestion flask containing a di-acid blend (HNO₃: HClO₄ = 2:1 v/v), and kept overnight [24]. The flasks were put on a temperature-adjustable hot plate and gradually heated (up to 350 °C) till the appearance of dense white fumes. Later, the digests were allowed to cool and then transferred to a 50 mL volumetric flask. The final volume (50 mL per digest) was made by the addition of distilled water. Likewise, to measure the labile fractions of HMs, soil samples were extracted with 5 mM DTPA extractant [25]. Finally, the concentrations of Pb, Cr, Cd, Ni, and Cu in plant parts, DTPA extract, and leachates were estimated on ICP-MS (PerkinElmer’s NexION^® 2000). Concentrations of HMs in DTPA extract, leachates, and plant dry biomass were represented as mg kg⁻¹ soil, mg L⁻¹, and mg kg⁻¹ dry weight (DW), respectively. Each HM content in roots and shoots was calculated by Formula (1) and was presented as mg pot⁻¹.

Concentration of HM in plant part (mg kg⁻¹) × plant part DW (kg) harvested from the pot

(1)

On the other hand, the removal of each HM from the SS was calculated by Formula (2) and was presented as %.

HM content (root) + HM content (shoot)/total HM content in soil × 100

(2)

In Formula (2), the total HM content in soil depicts the content of HM in the whole pot soil (mg pot⁻¹).

2.4.2. Soil Enzymes in SS

The urease activity in SS was estimated by following the protocol recommended by Kandeler and Gerber [26]. Likewise, the catalase activity was assessed by recording the decline in the rate of hydrogen peroxide (H₂O₂) consumed by the soil [27]. The activity of acid phosphate was estimated by measuring the absorbance at 400–420 nm in the reaction mixture using a spectrophotometer (SPECORD 200 PLUS, Analytik Jena GmbH, Jena, Germany) after adjusting the ρ-nitrophenyl phosphate contents [28]. The protease activity was evaluated by following the gelatin hydrolyzation method, as described by Tabatabai [29] and Guan [30]. Likewise, the chitinase activity in post-harvest SS was measured after calculating the release of para–nitrophenyl (ρ–NP) from para–nitrophenyl N-acetyl glucosaminide as a substrate. The spectrophotometer (400 nm) was used to gauge the absorbance [31].

2.4.3. Plant Biophysical Parameters and Biochemical Compounds

The method of Hiscox and Israelstam was used to quantify the contents of chlorophyll a (Chl-a) and chlorophyll b (Chl-b) in C. erectus leaves [32]. To attain this, 1 g of fresh leaf tissue was taken and extracted in a methanol–chloroform–water (20 mL) mixture. The absorbances in the extract for Chl-a (664.5 nm) and Chl-b (647.4 nm) were recorded on a spectrophotometer. Similarly, for the estimation of the water uptake capacity (WUC) and relative water content (RWC), recommended protocols were used [21]. The protein content was assessed using the Bradford protocol [33] by taking bovine serum albumin as the standard. Further, the ascorbic acid (AsA) content in the extract was examined by following Huang et al.’s [34] procedure, with some alterations [35]. Subsequently, the absorbance value was taken (265 nm) on a spectrophotometer after the addition of 1.0 U ascorbate oxidase. Similarly, the extract of dry leaf in ethanol was used to evaluate the contents of amino acids, with ninhydrin as the standard [36]. Likewise, for estimating the total phenolics, the method recommended by Folin Ciocalteu [37] was deployed, and the absorbance in the standardized curve of gallic acid was documented on a spectrophotometer (760 nm).

2.4.4. Plant Antioxidants and Reactive Oxygen Species

The activities of catalase (CAT), ascorbate peroxidase (APX), and superoxide dismutase (SOD); the contents of H₂O₂ and malondialdehyde (MDA); and the generation of superoxide radical (O₂^•−) in C. erectus leaves were measured following various protocols as detailed in Tauqeer et al. [21].

2.5. Statistical Analysis

The outcomes of the current investigation were interpreted by using a one-way analysis of variance (ANOVA) through Statistix 8.1.1 (Analytical Software, Tallahassee, FL, USA). The standard error (SE) from the three repeats of each treatment was computed in Microsoft Excel, 2013. The significant difference between the mean values of the treatments was calculated by a least significant difference (LSD) test (p < 0.05) [38] and is indicated by different lowercase letters.

3. Results

3.1. Plant Growth, Biophysical and Biochemical Traits

The RDW and SDW values were up to 2.50 g pot⁻¹ and 11.1 g pot⁻¹, respectively (Figure 1), whereas the plant height was up to 73.7 cm (Figure 2). Compared to the CK, significant improvements in the plant height, SDW, and RDW of C. erectus were observed in all treatments, while no change was observed in the CA. However, the most significant increases in plant height, by 71 and 82%, and in SDW, by 77 and 87%, were noted in the BC and the BC+LN+CA treatments, respectively, compared to the CK. On the other hand, the maximum improvement in the RDW by 109% was seen with the BC+LN+CA treatment, compared to the CK. When compared with BC, significant reductions in RDW by 52 and 14% were observed in the CA and BC+LN treatments, respectively. On the other hand, an improvement by 10% in the RDW was observed with the BC+LN+CA treatment, compared to the BC treatment. Statistically, comparable SDWs in the BC and BC+LN+CA treatments were noted. However, significant reductions in the SDW, by 48 and 18%, were noticed in the CA and BC+LN treatments, compared to the BC treatment.

The leaf contents of Chl-a, Chl-b, RWC, and WUC were up to 0.23 mg g⁻¹ FW, 0.20 mg g⁻¹ FW, 21.8%, and 0.66 mg g⁻¹ DW, respectively. However, protein, amino acids, AsA, and total phenolic content were up to 16.9 mg g⁻¹ FW, 19.6 mg g⁻¹ FW, 0.38 mg g⁻¹ FW, and 3.07 µmol g⁻¹ FW, respectively (Figure 2). The maximum significant increases in the contents of Chl-a, by 64 and 80%, WUC, by 83 and 100%, protein, by 33 and 38%, amino acids, by 33 and 36%, and AsA, by 89 and 102%, were achieved with the BC and BC+LN+CA treatments, compared to the CK. Moreover, the most pronounced reductions in phenolic content, by 44% with BC and 48% with BC+LN+CA, compared to CK, were noted. The highest significant increment, by 104%, in the Chl-b content, compared to the CK, was found with the BC+LN+CA treatment. Moreover, the highest increases in RWC, by 27, 21, and 31%, were observed in the BC, BC+LN, and BC+LN+CA treatments, respectively, compared to the CK. No significant differences in plant height and in the contents of Chl-a, RWC, WUC, protein, amino acids, AsA, and total phenolics were seen in the BC and BC+LN+CA treatments. However, the BC+LN+CA treatment significantly improved the Chl-b content by 31%, compared to the BC treatment.

3.2. Phytoavailable HMs in Soil, Their Plant Concentrations, and Removal from SS

The Pb, Cr, Cd, Ni, and Cu concentrations were up to 4896.7, 6.77, 10.6, 23.9, and 263.1 mg kg⁻¹ DW, respectively, in shoots, whereas the roots had up to 2885.3, 13.1, 26.1, 54.2, and 419.0 mg kg⁻¹ DW of Pb, Cr, Cd, Ni, and Cu concentrations, respectively (Figure 3). It is noteworthy that the Pb concentration in plant shoots was higher than in roots, whereas a contrasting pattern was noticed for Cr, Cd, Ni, and Cu. Similarly, the DTPA-extracted concentrations were up to 29.4, 1.48, 2.33, 2.40, and 1.97 mg kg⁻¹ soil for Pb, Cr, Cd, Ni, and Cu (Figure 3). The sole application of CA resulted in the highest increments in shoot concentrations of Pb, Cr, Cd, Ni, and Cu, by 7, 16, 10, 12, and 19%, while these concentrations were 9, 14, 11, 9, and 14% in roots, respectively, compared to the CK. The highest compressions in the concentration of Pb, by 36 and 33%, Cr, by 66 and 63%, Cd, by 59 and 57%, and Ni, by 59 and 56% in shoots were seen with the BC and BC+LN, treatments, respectively, compared to the CK. On the other hand, with the BC and BC+LN treatments, the concentration of Pb in roots was reduced by a maximum of 37 and 34%, Cr by 69 and 64%, Cd by 55 and 54%, and Ni by 52 and 46%, compared to the CK. Moreover, relative to the CK treatment, Cu concentrations in shoots and roots were reduced by up to a maximum of 74 and 39%, with BC.

The application of the CA treatment significantly enhanced Pb, Cr, Cd, Ni, and Cu concentrations in DTPA extract by a maximum of 10, 8, 14, 13, and 11%, compared to the CK. The concentration of each HM extracted from the post-harvest soil was always significantly lower in the case of the BC treatment compared to BC+LN+CA and statistically equivalent to the BC+LN treatment, except for Cd (Figure 3). The greatest decreases in the extractable Pb, by 52 and 50%; Cr by 95 and 93%; Ni by 35 and 30%; and Cu by 38 and 32% were found with the BC and BC+LN treatments, respectively, in comparison to the CK. Adding BC resulted in the highest decrease in the extractable Cd concentration, by 48%, compared to the CK.

The contents of Pb, Cr, Cd, Ni, and Cu in the shoots were calculated to reach up to 44.7, 0.04, 0.09, 0.19, and 2.13 mg pot⁻¹, respectively, while in roots, the estimated contents were up to 5.91, 0.02, 0.05, 0.11, and 0.79 mg pot⁻¹ for Pb, Cr, Cd, Ni, and Cu (Figure 4). The effects of each treatment on the contents of every HM in shoots and roots were variable. Notably, the highest significant increases in Pb, Cd, Ni, and Cu contents were noted in the BC+LN+CA treatment, with increases of 66, 67, 49, and 72% in shoots, and 89, 75, 82, and 84% in roots, respectively, compared to the CK. In the BC+LN+CA treatment, the Cr content in shoots was statistically equivalent to that in the CK but was significantly 83% higher in roots, compared to the CK.

The calculated values of percentages of HM removal from SS extended to 0.47% for Pb, 0.17% for Cr, 0.24% for Cd, 0.25% for Ni, and 0.43% for Cu, respectively (Figure 4). Variations in the percentage removal of each HM were seen across different treatments compared to the CK. Remarkably, the BC+LN+CA treatment resulted in a significant maximum rise in Pb, Cr, Cd, Ni, and Cu removal, by 68, 30, 69, 59, and 76%, in comparison to the CK.

3.3. HMs in Leachates

No changes in the concentrations of HMs in the leachates were observed at all sampling times (first, second, third, and fourth) in the CK treatment (Figure 5). Citric acid application gradually increased HM concentrations in the leachates over time. In the BC and BC+CA treatments, reductions in HM concentrations from the first to the third sampling point were observed, with no further changes noted thereafter. However, a gradual decrease in HM concentrations with the LN and LN+CA treatments was observed over time. Interestingly, HM concentrations in the leachates of the BC+LN and BC+LN+CA treatments were also followed by a temporal decrease. Pb, Cr, and Cu concentrations in leachates of BC, LN, and BC+LN treatments were lower than the treatments where CA was added with these amendments. Similarly, higher concentrations of Cd, in the case of the LN and BC+LN treatments, and Ni in the BC and LN treatments were noted with the addition of CA, compared to their counter treatments without CA. Except for CA, the concentration of each HM at several sampling times was below the critical limits for safe water reuse and agricultural purposes [39,40] with the BC, LN, and BC+LN treatments, both alone and in combination with CA.

3.4. Antioxidants and ROS

The values of CAT, SOD, and APX were up to 44.8 μmol min⁻¹ mg⁻¹ protein, 61.8 U min⁻¹ mg⁻¹ protein, and 0.53 μmol min⁻¹ mg⁻¹ protein, respectively. The MDA and H₂O₂ contents and the O₂^•− generation rate were up to 85.4 nmol g⁻¹ FW, 91.0 nmol g⁻¹ FW, and 79.0 nmol min⁻¹ g⁻¹ FW, respectively, in the plants (Table 2). The BC and BC+LN+CA treatments exhibited the greatest improvements in the activities of CAT by 48 and 56%, SOD by 41 and 48%, and APX by 84 and 96%, while reducing H₂O₂ content by 44 and 48% and the O₂^•− generation rate by 45 and 49%, respectively, compared to the CK. Moreover, the greatest reductions in MDA content, by 47, 44, and 50% compared to CK were noted with the BC, BC+CA and BC+LN+CA treatments, respectively.

3.5. Soil Enzymatic Activities

The protease, chitinase, urease, acid phosphatase, and catalase activities were up to 44.0 mg kg⁻¹ 24 h⁻¹, 11.60 mg ρ-NP kg⁻¹ soil h⁻¹, 2.53 µg N-N(H₄ kg⁻¹ h⁻¹), 30.4 mg p-NP kg⁻¹ h⁻¹, and 0.34 Vol. of 0.1 M KMnO₄ g⁻¹, respectively (Figure 6). The BC+CA and BC+LN+CA treatments showed the greatest improvements in the activities of protease by 58 and 71%, chitinase by 164 and 174%, and acid phosphatase by 59 and 68%, respectively, compared to the CK. Interestingly, compared to CK, enhancements up to 145 and 232% in urease and catalase activities with the BC+LN+CA treatment, respectively, were seen.

4. Discussion

4.1. Plant Biomass, Growth, and Affiliated Traits

A decline in growth and biomass of Miscanthus [11] and in the total chlorophyll content (Chl-a and Chl-b) of M. sativa growing in SS [1] was observed, while the biomass and biochemical traits of C. erectus [21] in the HM-contaminated soil were reported. In this study, the BC+LN+CA treatment resulted in the most significant (p < 0.05) improvements in the RDW and Chl-b contents compared to the CK. In addition, significantly (p < 0.05) higher plant height, SDW, Chl-a, WUC, protein, amino acids, and AsA levels were observed, while reduced phenolic content was noticed with the BC and BC+LN+CA treatments compared to the CK (Figure 1 and Figure 2). Supplementing Pb-polluted soil with BC promoted biomass, as well as the biophysical and biochemical attributes of S. oleracea [22]. Moreover, biomass and growth enhancement of L. sativa grown in Cd-polluted soil [41] and Z. mays in soil with Pb pollution were reported with the LN treatment [18].

The growth, biophysical, and biochemical traits of C. erectus were enhanced because the LN and BC treatments (1) improved soil structure and its physicochemical properties [3,18] and (2) increased soil porosity, which supported root development, especially fine roots, for better exploration of BC-induced nutrient hotspots [18,42]. Lignin and BC also favor plant growth by increasing soil moisture content, enhancing aerobic respiration, and releasing nutrients [22,43]. Citric acid supports plant growth and development through (1) solubilizing nutrients from soil [44,45] and BC [46] and (2) enhancing the plant’s water status and chlorophyll contents via improving osmotic balance and membrane stability [14,47]. In addition, HMs immobilization with BC and LN [2,18], CA-mediated formation of HM chelates in soil [16] and phytochelatins in plants [47] reduced HMs phytotoxicity, which improved plant growth and related parameters.

Compared to BC, the CA treatment significantly reduced SDW and RDW (Figure 1). In soils with low HMs pollution, CA application efficiently chelates HMs, leading to their reduced phytotoxicity [44,45]. However, applying CA in the soils with higher concentrations of HMs leads to their higher solubility, which causes reduced plant growth and biomass production due to severe HM phytotoxicity [44,48]. The SDW and RDW in the BC+LN treatment were significantly lower than in the BC treatment but higher than in the LN treatment (Figure 1). As apparent from our results, LN has a smaller effect on promoting plant biomass compared to BC. To formulate the BC+LN treatment, half doses of both BC and LN were used, which led to a reduction in essential nutrients and other growth-promoting mechanisms provided by BC, reducing plant biomass by half, compared to the BC treatment. In the BC+LN+CA treatment, the SDW was statistically comparable to that in the BC treatment (Figure 1). Despite higher HMs bioavailability with the BC+LN+CA treatment compared to BC alone in SS (Figure 3), this comparable plant biomass is attributed to reduced HMs toxicity to plants in the BC+LN+CA treatment. The BC+LN (present in the BC+LN+CA treatment) efficiently reduced HMs phytotoxicity through several immobilization mechanisms (Section 4.2). After this major reduction in HMs phytotoxicity, the additive effects of CA present in this treatment further reduced HMs phytotoxicity through their efficient chelation [44,45].

4.2. Phytoavailability of HMs, Their Removal, and Leaching from SS

The incorporation of BC and BC+LN resulted in significantly (p < 0.05) lower HM concentrations in the DTPA extract, plant roots, shoots (Figure 3), and leachates in most cases, compared to the CK (Figure 5). Previously, BC reduced the mobility of Pb, Cu, and Cd mobility in MS and their uptake in L. perenne [3] and S. lycopersicum [2]. Moreover, reduced Pb mobility in Pb-polluted soil, as well as in Hordeum vulgare [23] and Zea mays [18] were also reported with LN. Biochar reduced HMs mobility in SS, C. erectus, and leachates because of (i) HM adsorption onto its larger surface area [19], (ii) the formation of insoluble HM precipitates with OH⁻, CO₃²⁻, PO₄³⁻, and Fe–Mn oxides [3,49,50], and (iii) complexation of HMs with −OH and carboxyl −COOH functional groups [2,50]. In addition, LN binds HMs with −COOH, phenol, and OH⁻ groups and forms HMs–OM complexes, which have reduced HMs mobility in SS, plants, and leachates [18,51]. An increase in soil pH due to BC and LN also reduces HM mobility in soil and plants [18,49].

Significantly (p < 0.05) higher concentrations of HMs in the DTPA extract, plant roots, and shoots (Figure 3), as well as in the leachates (Figure 5), were noted with CA addition in SS, compared to CK. Previously, CA amplified the solubility of Pb and Cd in soil and increased their concentrations in P. hortroum [17]. Moreover, enhanced soil bioavailability and uptake of Cd, Cr, Ni, and Cu in G. max were also noted with the CA treatment [45]. Adding CA to SS (i) increased H⁺ ions in soil, which desorbed HMs from the sorption complex [16,17] and (ii) enhanced HM desorption from insoluble HM fractions [16]. In addition, CA forms soluble and thermodynamically stable HMs–citrate complexes in soil, which increase HM concentrations in soil, plant parts, and leachates [4].

In the BC+LN+CA treatment, the HM concentrations in the DTPA extract and plant parts (roots and shoots) were significantly (p < 0.05) greater than in treatments with stabilizing additives (BC, LN, and BC+LN) but lower than in the CA treatment, than CK (Figure 3). Whereas, CA addition in SS amended with BC, LN, and BC+LN showed significantly (p < 0.05) more pronounced concentrations of HMs in leachates compared to their counter treatments without CA (Figure 5). Such moderate HM mobility in SS, plants, and leachates is due to special behaviors of BC and LN upon soil acidification with CA. Surprisingly, the HM concentrations in the leachates of this treatment were below the critical limits set by FAO [39] and USEPA [40] for safe water reuse and agricultural purposes [51]. Biochar neutralizes soil acidity by (1) raising soil pH, (2) promoting decarboxylation, and (3) increasing exchangeable cations, as well as CO₃²⁻ and HCO₃⁻ contents in soil [19], thus restricting HM mobility in soil, plants and groundwater [3,6]. Contrarily, acidic conditions delignify LN and break aryl ether bonds (α−O−4, β−O−4) [52], thus releasing HMs in soil and enhancing their uptake in C. erectus. Moreover, CA solubilized Fe and K from soil, which replaced HMs from LN functional groups, leading to higher HM mobility in the soil–plant system [22,53].

C. erectus exhibited higher Pb concentrations in the above-ground parts than in the roots, whereas a contrasting tendency was observed for Cr, Cd, Ni, and Cu (Figure 3). In this context, (i) higher Pb uptake via a denser root system, (ii) the activity of Pb binding and transporter genes (PbATPases and CPx-type), and (iii) the presence of organic acids in xylem may have improved Pb movement towards the aerial parts [21,45]. Contrarily, the retention of Cr, Ni, and Cd in the roots is linked to their precipitation in the root membrane [6]. HM contents in plant roots and shoots were significantly (p < 0.05) higher in the BC+LN+CA treatment compared to the CK (Figure 4). The higher HM contents in the roots and shoots were due to (i) the role of BC+LN in improving plant biomass (Figure 1) through several mechanisms (see Section 4.1) and subsequently, (ii) the moderate solubilization of pre-immobilized HMs with CA (Figure 3). Moreover, higher root biomass and the binding of HMs with LN and BC reduced their leaching [11]. Significantly (p < 0.05), the most pronounced percentages of HMs removed from SS were noted in the BC+LN+CA treatment, compared to the CK (Figure 4). Previously, soil acidification significantly enhanced the percentage of Ni removed from the soil with Salix alba [15]. These findings highlight the potential of C. erectus to efficiently decontaminate SS polluted with different HMs.

4.3. Antioxidant Defense in C. erectus

Heavy metal uptake in plants produces ROS, which are neutralized by plant antioxidants [14,20]. Both the BC and BC+LN+CA treatments resulted in significantly (p < 0.05) higher activities of CAT, SOD, and APX, as well as in reduced H₂O₂ content and O₂^•− generation rate in leaves, compared to the CK. Apart from this, the MDA content in leaves was also significantly (p < 0.05) lower, compared to the CK, in the BC, BC+CA, and BC+LN+CA treatments (Table 2). Previously, higher activities of CAT, APX, and SOD, as well as reduced O₂^•−, H₂O₂, and MDA contents were seen in Solanum lycopersicum growing in Pb-, Cu-, and Cd-polluted mining/smelting soil treated with BC [2]. Adding LN in Pb-contaminated soils improved CAT, SOD, and APX activities, while decreasing the generation of O₂^•−, H₂O₂, and MDA contents in barely [23]. Additionally, CA application in Cd-polluted soil enhanced the activities of CAT, SOD, and APX, while reducing the MDA content in A. rosea [14]. Supplementing Pb-contaminated soil with CA reduced the O₂^•− generation rate and the contents of MDA and H₂O₂, while augmenting CAT and SOD activities in Rhus chinensis [54].

The improved antioxidant activities and lower ROS contents in C. erectus are because of BC, as it improves the metabolisms of vital nutrients in plants, supports intracellular osmotic adjustments, and reduces the accumulation of free radicals [2]. The cell membrane is the first fence against HMs-linked lipid peroxidation [14]. CA maintains cell membrane stability through proline formation, thus supporting HM detoxification and improving plant antioxidants [45]. CA promotes the AsA-glutathione (GSH) pool in cell organelles, which helps scavenge H₂O₂. In this process, AsA directly reacts with O₂^•− and decreases its cellular content [47]. Further, LN and BC mediated HM immobilization [19,51], and the formation of HM complexes with CA in SS also alleviated HM toxicity to plants, enhancing antioxidant activities [16,17].

Significantly lower activities of antioxidants and higher production of ROS in plants were noted with the CA treatment, compared to BC (Table 2). Citric acid enhances HMs solubility in HMs-polluted soils [44]. Such soil conditions lead to higher uptake of HMs by the plant, causing the production of excessive free radical ions, damage to cell membranes, increased electrolyte leakage, and oxidative damage [14,47]. In the case of BC+LN, higher contents of ROS and lower antioxidant activities were noted in plants, compared to BC (Table 2). Since the contents of essential nutrients present in the BC were reduced by half in the BC+LN treatment, BC-associated mechanisms, such as nutrient metabolism, osmotic adjustments, and a decline in the accumulation of free radicals [2] were also reduced, leading to higher contents of ROS and lower antioxidant activities in this treatment, compared to BC. Similarly, the BC+LN+CA treatment showed comparable results with the BC treatment for these parameters (Table 2). Though the bioavailability of HMs in the BC+LN+CA treatment was higher than in BC alone (Figure 3), the comparable antioxidant activities and ROS contents in the BC+LN+CA treatment are linked to reduced HM phytotoxicity [17,19,50]. In the BC+LN+CA treatment, the initial immobilization of HMs with BC and LN (Section 4.2) reduced the HM toxicity to plants. Subsequently, the CA addition remobilized the HMs in soil, chelated them, and reduced their toxicity to the plants [4,44,45].

4.4. Enzymatic Activities in SS

Soil enzymes are secreted by various microorganisms, which improve soil health and nutrient cycling [3]. In SS, high HM concentrations negatively affect enzymatic activities through the toxicity of HM to microorganisms [5,11]. Interestingly, the activities of protease, chitinase, and acid phosphatase were significantly (p < 0.05) higher in the BC+CA and BC+LN+CA treatments compared to the CK. However, the urease and catalase activities were the most prominent in the BC+LN+CA treatment, with a statistically significant difference (p < 0.05), when compared with the CK (Figure 6). Adding BC to different MSs polluted with Cd, Cu, and Pb raised the activities of protease, catalase, and urease within them [2,55]. Mixing LN into Pb-polluted soil upgraded the urease and catalase activities within it [21]. Interestingly, supplementing Pb- and Cd-polluted soils with CA boosted soil protease and catalase activities [56].

In SS, adding BC and LN raised enzyme activities, as both materials are porous, which enhanced the growth, metabolic activities, and respiration of microbial communities by (i) providing niches to them, (ii) the sustainable release of mineral nutrients, and (iii) maintaining a favorable soil moisture content [2,3,18]. Similarly, CA improved the microbial population and their activities in SS by providing essential nutrients via (i) enhancing nutrient solubility from the soil matrix [16,45] and (ii) liberating C, N, and O from BC [46] and C from LN through acidification [57]. In addition, CA decomposition in soil also releases C, H, and O, which are consumed by soil microorganisms to increase their biomass [44]. The immobilization of HMs with BC and LN and HM chelation with CA improved enzyme secretion and their activities through (i) reducing HM toxicity to soil microorganisms [16,49,50] and (ii) minimizing HM reactions with sulfhydryl groups and enzyme substrates [3].

The sole application of CA significantly reduced the activities of soil enzymes, compared to the BC treatment (Figure 6). Since the SS used in this experiment has higher concentrations of HMs, adding CA solubilized these HMs [44], which reduced the activities of soil enzymes through (1) HMs toxicity to microorganism [5,11] and (2) HMs reactions with sulfhydryl groups and enzyme substrates [3,44]. The soil enzymatic activities were significantly lower in the BC+LN treatment compared to BC (Figure 6). In this context, the BC content in the BC+LN treatment was half that of the BC treatment alone. Since the contents of essential nutrients present in the BC were reduced to one half in the BC+LN treatment, this reduction in the contents of the nutrients available to the soil microorganisms resulted in the secretion of smaller quantities of soil enzymes [3,58]. The activities of soil enzymes were significantly higher in the BC+LN+CA treatment compared to BC (Figure 6). Despite higher HMs bioavailability in the BC+LN+CA treatment compared to the BC treatment (Figure 3), such significant soil enzymatic activities in the BC+LN+CA treatment are attributed to a reduction in the HMs toxicity to soil microorganisms [2,58]. The initial HMs immobilization by BC and LN reduced HMs toxicity to soil microorganisms [2,3,23]. Subsequently, this alleviation in toxicity was also boosted by the formation of HMs chelates with CA. Such alleviation of HMs toxicity to microorganisms improved their abilities to secrete more soil enzymes [44,47,48].

5. Conclusions

The initial immobilization of HMs with BC+LN, followed by their periodic mobilization with CA during enhanced phytoextraction from SS seems practical. This finding may be important, as the initial immobilization with BC+LN can (i) positively influence the growth and establishment of phytoextraction plants on HMs-contaminated SS (ii) and limit HM leaching into groundwater. After the plants have become established on SS, periodic solubilization of HMs in SS with CA can enhance phytoextraction efficiency. In this pot study, the initial immobilization of Pb, Cr, Cd, Ni, and Cu with BC+LN and their periodic remobilization with CA positively enhanced the phytoextraction capability of C. erectus and the percentages of HMs removed from SS. Moreover, the HM concentrations in the leachates were also below the critical limits for safe water reuse and agricultural purposes. The significant results of enhanced HMs phytoextraction with the BC+LN+CA treatment were due to (i) enhanced plant biomass and (ii) adequate HM uptake by the plants. Our results confirm that the phytoextraction of HMs with C. erectus in the presence of cheap soil amendments (BC+LN+CA) is a cost-effective and green technique that can remediate different HMs-polluted soils, including SS at field scale. Further research should also address the feasibility of this technique to decontaminate soils containing both HMs and organic pollutants, either by inoculating these soils with microorganisms that can degrade organic pollutants or through some other cheap soil amendment with similar functions. Interestingly, the expenses of soil additives (BC, LN, and CA) could be offset later by utilizing the biomass of C. erectus for several purposes, such as the production of renewable energy, packing, and building material.

Author Contributions

Conceptualization, M.I. and H.M.T.; methodology, H.M.T., F.S. and H.S.; software, M.U., H.M.T. and I.C.; validation: K.L. formal analysis, H.M.T. and M.I.; investigation, K.L., F.M., T.S. and M.I.; resources: M.I. and H.M.T.; data curation: H.M.T. and M.I.; writing—original draft, H.M.T., F.M., I.C. and M.I.; writing—review and editing, M.U., H.S., T.S., F.S., M.I. and K.L.; visualization, F.S.; supervision, M.I. and K.L.; project administration, M.I.; funding acquisition, K.L. All authors have read and agreed to the published version of the manuscript.

Funding

Hafiz Muhammad Tauqeer arranged the SS and plant cuttings, bought the pots and trays, oversaw the experiment, and conducted various analyses. The cost of chemicals and other reagents was shared by Hafiz Muhammad Tauqeer and Muhammad Iqbal. Apart from it, there was no other funding source.

Data Availability Statement

All data supporting the findings of this experiment are presented in this manuscript.

Acknowledgments

We are grateful to Zaheer Abbas Virk, the Department of Environmental Sciences, Government College University, Faisalabad, 38000, Pakistan, for his input in improving English language.

Conflicts of Interest

The authors declare no conflicts of interest.

References

Mankė, J.; Praspaliauskas, M.; Pedišius, N.; Sujetovienė, G. Evaluation of phytoremediation efficiency of shooting range soil using the bioaccumulation potential and sensitivity of different plant species. Ecol. Eng. 2024, 198, 107134. [Google Scholar] [CrossRef]
Ghani, M.I.; Ahanger, M.A.; Sial, T.A.; Haider, S.; Siddique, J.A.; Fan, R.; Liu, Y.; Esmat, F.A.; Kumar, M.; Yang, X.; et al. Almond shell-derived biochar decreased toxic metals bioavailability and uptake by tomato and enhanced the antioxidant system and microbial community. Sci. Total Environ. 2024, 929, 172632. [Google Scholar] [CrossRef]
Cheng, Y.; Bu, X.; Ji, Z.; Wang, C.; Xiao, X.; Li, F.; Wu, Z.-H.; Wu, G.; Jia, P.; Li, J.-T. Application of biochar and compost improved soil properties and enhanced plant growth in a Pb–Zn mine tailings soil. Environ. Sci. Pollut. Res. 2023, 30, 32337–32347. [Google Scholar] [CrossRef]
Silwamba, M.; Ito, M.; Tabelin, C.B.; Park, I.; Jeon, S.; Takada, M.; Kubo, Y.; Hokari, N.; Tsunekawa, M.; Hiroyoshi, N. Simultaneous extraction and recovery of lead using citrate and micro-scale zero-valent iron for decontamination of polluted shooting range soils. Environ. Adv. 2021, 5, 100115. [Google Scholar] [CrossRef]
Zhu, Y.; Che, R.; Tu, B.; Miao, J.; Lu, X.; Li, J.; Zhu, Y.; Wang, F. Contamination and remediation of contaminated firing ranges—An overview. Front. Environ. Sci. 2024, 12, 1352603. [Google Scholar] [CrossRef]
Česynaitė, J.; Praspaliauskas, M.; Sujetovienė, G. Bioaccumulation of trace metal (loid)s and toxic response of Lactuca sativa grown in shooting range soil. Toxicol. Environ. Chem. 2023, 105, 28–41. [Google Scholar] [CrossRef]
Lewińska, K.; Karczewska, A. A release of toxic elements from military shooting range soils as affected by pH and treatment with compost. Geoderma 2019, 346, 1–10. [Google Scholar] [CrossRef]
Mariussen, E.; Heier, L.S.; Teien, H.C.; Pettersen, M.N.; Holth, T.F.; Salbu, B.; Rosseland, B.O. Accumulation of lead (Pb) in brown trout (Salmo trutta) from a lake downstream a former shooting range. Ecotoxicol. Environ. Saf. 2017, 135, 327–336. [Google Scholar] [CrossRef]
Maruyama, M.; Ushine, N.; Watanabe, Y.; Ishii, C.; Saito, K.; Sakai, H.; Kuritani, T.; Doya, R.; Ogasawara, K.; Ikaneka, Y.; et al. Current situation of lead (Pb) exposure in raptors and waterfowl in Japan and difference in sensitivity to in vitro lead exposure among avian species. Environ. Pollut. 2024, 349, 123907. [Google Scholar] [CrossRef]
Johnsen, I.V.; Aaneby, J. Accumulation of copper and lead in ruminants grazing on a contaminated shooting range in Nordland County, Norway. Environ. Sci. Pollut. Res. 2024, 31, 11026–11036. [Google Scholar] [CrossRef]
Alasmary, Z.; Hettiarachchi, G.M.; Roozeboom, K.L.; Davis, L.C.; Erickson, L.E.; Pidlisnyuk, V.; Stefanovska, T.; Trögl, J. Phytostabilization of a contaminated military site using Miscanthus and soil amendments. J. Environ. Qual. 2021, 50, 1220–1232. [Google Scholar] [CrossRef]
Kicińska, A.; Wikar, J. The effect of fertilizing soils degraded by the metallurgical industry on the content of elements in Lactuca sativa L. Sci. Rep. 2021, 11, 4072. [Google Scholar] [CrossRef]
Egendorf, S.P.; Groffman, P.; Moore, G.; Cheng, Z. The limits of lead (Pb) phytoextraction and possibilities of phytostabilization in contaminated soil: A critical review. Int. J. Phytoremediation 2020, 22, 916–930. [Google Scholar] [CrossRef]
Khan, A.A.; Wang, T.; Nisa, Z.U.; Alnusairi, G.S.; Shi, F. Insights into cadmium-induced morphophysiological disorders in Althea rosea Cavan and its phytoremediation through the exogeneous citric acid. Agronomy 2022, 12, 2776. [Google Scholar] [CrossRef]
Virk, Z.A.; Al Farraj, D.A.; Iqbal, M.; Lewińska, K.; Hussain, S. Inoculation with the pH Lowering Plant Growth Promoting Bacterium Bacillus sp. ZV6 Enhances Ni Phytoextraction by Salix alba from a Ni-Polluted Soil Receiving Effluents from Ni Electroplating Industry. Sustainability 2022, 14, 6975. [Google Scholar] [CrossRef]
Manzoor, M.; Shafiq, M.; Gul, I.; Kamboh, U.R.; Guan, D.X.; Alazba, A.A.; Tomforde, S.; Arshad, M. Enhanced lead phytoextraction and soil health restoration through exogenous supply of organic ligands: Geochemical modeling. J. Environ. Manag. 2023, 348, 119435. [Google Scholar] [CrossRef]
Gul, I.; Manzoor, M.; Hashmi, I.; Bhatti, M.F.; Kallerhoff, J.; Arshad, M. Plant uptake and leaching potential upon application of amendments in soils spiked with heavy metals (Cd and Pb). J. Environ. Manag. 2019, 249, 109408. [Google Scholar] [CrossRef]
Li, T.; Lü, S.; Zhang, S.; Gao, C.; Liu, M. Lignin-based multifunctional fertilizer for immobilization of Pb(II) in contaminated soil. J. Taiwan Inst. Chem. Eng. 2018, 91, 643–652. [Google Scholar] [CrossRef]
Bolan, N.; Sarmah, A.K.; Bordoloi, S.; Bolan, S.; Padhye, L.P.; Van Zwieten, L.; Sooriyakumar, P.; Khan, B.A.; Ahmad, M.; Solaiman, Z.M.; et al. Soil acidification and the liming potential of biochar. Environ. Pollut. 2023, 317, 120632. [Google Scholar] [CrossRef]
Rehman, S.; Azam, F.; Rehman, S.; Rehman, T.U.; Mehmood, A.; Gohar, A.; Samad, A. A review on botanical, phytochemical and pharmacological reports of Conocarpus erectus. Pak. J. Agric. Res. 2019, 32, 212–217. [Google Scholar] [CrossRef]
Tauqeer, H.M.; Hussain, S.; Abbas, F.; Iqbal, M. The potential of an energy crop “Conocarpus erectus” for lead phytoextraction and phytostabilization of chromium, nickel, and cadmium: An excellent option for the management of multi-metal contaminated soils. Ecotoxicol. Environ. Saf. 2019, 173, 273–284. [Google Scholar] [CrossRef]
Naeem, I.; Masood, N.; Turan, V.; Iqbal, M. Prospective usage of magnesium potassium phosphate cement combined with Bougainvillea alba derived biochar to reduce Pb bioavailability in soil and its uptake by Spinacia oleracea L. Ecotoxicol. Environ. Saf. 2021, 208, 111723. [Google Scholar] [CrossRef]
Khan, M.A.; Ramzani, P.M.A.; Zubair, M.; Rasool, B.; Khan, M.K.; Ahmed, A.; Khan, S.A.; Turan, V.; Iqbal, M. Associative effects of lignin-derived biochar and arbuscular mycorrhizal fungi applied to soil polluted from Pb-acid batteries effluents on barley grain safety. Sci. Total Environ. 2020, 710, 136294. [Google Scholar] [CrossRef]
Jones, J.B., Jr.; Case, V.W. Sampling, Handling and Analyzing Plant Tissue Samples. In Soil Testing and Plant Analysis, 3rd ed.; SSSA Book Series; Westerman, R.L., Ed.; Soil Science Society of America: Madison, WI, USA, 1990; pp. 389–427. [Google Scholar]
Lindsay, W.L.; Norvell, W. Development of a DTPA soil test for zinc, iron, manganese, and copper. Soil Sci. Soc. Am. J. 1978, 42, 421–428. [Google Scholar] [CrossRef]
Kandeler, E.; Gerber, H. Short-term assay of soil urease activity using colorimetric determination of ammonium. Biol. Fertil. Soils 1988, 6, 68–72. [Google Scholar] [CrossRef]
Xu, G.H.; Zheng, H.Y. Handbook of Analysis of Soil Microorganism; Agriculture Press: Beijing, China, 1986; pp. 249–291. [Google Scholar]
Tabatabai, M.A.; Bremner, J.M. Use of p-nitrophenyl phosphate for assay of soil phosphatase activity. Soil Biol. Biochem. 1969, 1, 301–307. [Google Scholar] [CrossRef]
Tabatabai, M.A. Soil enzymes. In Methods of Soil Analysis, Part 2; Albert, L., Ed.; Chemical and Microbiological Properties Soil Science Society of America: Madison, WI, USA, 1982; pp. 903–943. [Google Scholar]
Guan, S.Y. Research Methods on Soil Enzymes; China Agricultural: Beijing, China, 1986; pp. 182–266. [Google Scholar]
Hayano, K. A method for determination of β-glucosidase activity in soil. Soil Sci. Plant Nutr. 1973, 19, 103–108. [Google Scholar] [CrossRef]
Hiscox, J.D.; Israelstam, G.F. A method for the extraction of chlorophyll from leaf tissue without maceration. Can. J. Bot. 1979, 57, 1332–1334. [Google Scholar] [CrossRef]
Bradford, M.M. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein-dye binding. Anal. Biochem. 1976, 72, 248–254. [Google Scholar] [CrossRef]
Huang, C.; He, W.; Guo, J.; Chang, X.; Su, P.; Zhang, L. Increased sensitivity to salt stress in an ascorbate-deficient Arabidopsis mutant. J. Exp. Bot. 2005, 56, 3041–3049. [Google Scholar] [CrossRef]
Hasanuzzaman, M.; Hossain, M.A.; Fujita, M. Nitric oxide modulates antioxidant defense and the methylglyoxal detoxification system and reduces salinity-induced damage of wheat seedlings. Plant Biotechnol. Rep. 2011, 5, 353–365. [Google Scholar] [CrossRef]
Hamilton, P.B.; Van Slyke, D.D.; Lemish, S. The gasometric determination of free amino acids in blood filtrates by the ninhydrin-carbon dioxide method. J. Biol. Chem. 1943, 150, 231–250. [Google Scholar] [CrossRef]
Singleton, V.L.; Orthofer, R.; Lamuela-Raventós, R.M. Analysis of total phenols and other oxidation substrates and antioxidants by means of folin-ciocalteu reagent. In Methods in Enzymology; Packer, L., Ed.; Academic Press: Cambridge, MA, USA, 1999; Volume 299, pp. 152–178. [Google Scholar]
Steel, R.G.; Torrie, J.H.; Dickey, D.A. Principles and Procedures of Statistics: A Biometrical Approach, 3rd ed.; McGraw Hill Book Co. Inc.: New York, NY, USA, 1997; p. 672. [Google Scholar]
Ayers, R.S.; Westcot, D.W. Water Quality for Agriculture; Food and Agriculture Organization: Rome, Italy, 1989; p. 174. [Google Scholar]
USEPA. Guidelines for Water Reuse 600/R-12/618; Environmental Protection Agency: Washington, DC, USA, 2012; p. 643. [Google Scholar]
He, L.; Yu, Y.; Lin, J.; Hong, Z.; Dai, Z.; Liu, X.; Tang, C.; Xu, J. Alkaline lignin does not immobilize cadmium in soils but decreases cadmium accumulation in the edible part of lettuce (Lactuca sativa L.). Environ. Pollut. 2022, 310, 119879. [Google Scholar] [CrossRef]
Song, X.; Razavi, B.S.; Ludwig, B.; Zamanian, K.; Zang, H.; Kuzyakov, Y.; Gunina, A. Combined biochar and nitrogen application stimulates enzyme activity and root plasticity. Sci. Total Environ. 2020, 735, 139393. [Google Scholar] [CrossRef]
Lu, J.; Cheng, M.; Zhao, C.; Li, B.; Peng, H.; Zhang, Y.; Hassan, M. Application of lignin in preparation of slow-release fertilizer: Current status and future perspectives. Ind. Crops Prod. 2022, 176, 114267. [Google Scholar] [CrossRef]
Hosseini, S.S.; Lakzian, A.; Halajnia, A.; Razavi, B.S. Optimization of EDTA and citric acid for risk assessment in the remediation of lead contaminated soil. Rhizosphere 2021, 17, 100277. [Google Scholar] [CrossRef]
Hussain, A.; Kamran, M.A.; Javed, M.T.; Hayat, K.; Farooq, M.A.; Ali, N.; Manghwar, F.J.; Chaudhary, H.J. Individual and combinatorial application of Kocuria rhizophila and citric acid on phytoextraction of multi-metal contaminated soils by Glycine max L. Environ. Exp. Bot. 2019, 159, 23–33. [Google Scholar] [CrossRef]
Nazari, S.; Rahimi, G.; Nezhad, A.K.J. Effectiveness of native and citric acid-enriched biochar of Chickpea straw in Cd and Pb sorption in an acidic soil. J. Environ. Chem. Eng. 2019, 7, 103064. [Google Scholar] [CrossRef]
Al Mahmud, J.; Hasanuzzaman, M.; Nahar, K.; Bhuyan, M.B.; Fujita, M. Insights into citric acid-induced cadmium tolerance and phytoremediation in Brassica juncea L.: Coordinated functions of metal chelation, antioxidant defense and glyoxalase systems. Ecotoxicol. Environ. Saf. 2018, 147, 990–1001. [Google Scholar] [CrossRef]
Li, Y.; Wang, Y.; Khan, M.A.; Luo, W.; Xiang, Z.; Xu, W.; Zhong, B.; Ma, J.; Yr, Z.; Zhu, Y.; et al. Effect of plant extracts and citric acid on phytoremediation of metal-contaminated soil. Ecotoxicol. Environ. Saf. 2021, 211, 111902. [Google Scholar] [CrossRef]
Farhad, M.; Noor, M.; Yasin, M.Z.; Nizamani, M.H.; Turan, V.; Iqbal, M. Interactive suitability of rice stubble biochar and arbuscular mycorrhizal fungi for improving wastewater-polluted soil health and reducing heavy metals in peas. Sustainability 2024, 16, 634. [Google Scholar] [CrossRef]
Wu, P.; Cui, P.; Fang, G.; Gao, J.; Zhou, D.; Wang, Y. Sorption mechanism of zinc on reed, lignin, and reed-and lignin-derived biochars: Kinetics, equilibrium, and spectroscopic studies. J. Soils Sediments 2018, 18, 2535–2543. [Google Scholar] [CrossRef]
Sanga, V.F.; Fabian, C.; Kimbokota, F. Heavy metal pollution in leachates and its impacts on the quality of groundwater resources around Iringa municipal solid waste dumpsite. Environ. Sci. Pollut. Res. 2023, 30, 8110–8122. [Google Scholar] [CrossRef] [PubMed]
Chen, H.; Liu, N.; Qu, X.; Joshee, N.; Liu, S. The effect of hot water pretreatment on the heavy metal adsorption capacity of acid insoluble lignin from Paulownia elongata. J. Chem. Technol. Biotechnol. 2018, 93, 1105–1112. [Google Scholar] [CrossRef]
Fu, Y.; Qin, M.; Wang, Z.; Li, Z. Application of Lignin and Its Derivatives as Slow/Controlled Release Materials in Agricultural Fields. Pap. Biomater. 2018, 3, 53–62. [Google Scholar] [CrossRef]
Shi, X.; Wang, S.; He, W.; Wang, Y. Lead accumulation and biochemical responses in Rhus chinensis Mill to the addition of organic acids in lead contaminated soils. RSC Adv. 2023, 13, 4211–4221. [Google Scholar] [CrossRef]
Tang, J.; Zhang, L.; Zhang, J.; Ren, L.; Zhou, Y.; Zheng, Y.; Luo, L.; Yang, Y.; Huang, H.; Chen, A. Physicochemical features, metal availability and enzyme activity in heavy metal-polluted soil remediated by biochar and compost. Sci. Total Environ. 2020, 701, 134751. [Google Scholar] [CrossRef]
Mao, L.; Tang, D.; Feng, H.; Gao, Y.; Zhou, P.; Xu, L.; Wang, L. Determining soil enzyme activities for the assessment of fungi and citric acid-assisted phytoextraction under cadmium and lead contamination. Environ. Sci. Pollut. Res. 2015, 22, 19860–19869. [Google Scholar] [CrossRef]
Wang, F.; Gao, Y.; Li, X.; Luan, M.; Wang, X.; Zhao, Y.; Zhou, X.; Du, G.; Wang, P.; Ye, C.; et al. Changes in microbial composition explain the contrasting responses of glucose and lignin decomposition to soil acidification in an alpine grassland. Sci. Total Environ. 2024, 930, 172671. [Google Scholar] [CrossRef]
Pandey, B.; Suthar, S.; Chand, N. Effect of biochar amendment on metal mobility, phytotoxicity, soil enzymes, and metal-uptakes by wheat (Triticum aestivum) in contaminated soils. Chemosphere 2022, 307, 135889. [Google Scholar] [CrossRef]

Figure 1. The influences of biochar (BC), lignin (LN), and citric acid (CA) on shoot dry weight (SDW) (a) and root dry weight (RDW) (b) of C. erectus grown on shooting-range soil (SS). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by one-way ANOVA, followed by an LSD test.

Figure 2. Influences on plant height (A), chlorophyll a (Chl-a) (B), chlorophyll b (Chl-b) (C), relative water content (RWC) (D), water uptake capacity (WUC) (E), protein (F), amino acids (G), ascorbic acid (AsA) (H), and total phenolics (I) of C. erectus grown on shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test.

Figure 3. The concentrations of Pb, Cr, Cd, Ni, and Cu (a,c,e,g,i) in C. erectus tissues and labile portions of Pb, Cr, Cd, Ni, and Cu (b,d,f,h,j) in post-harvest shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test.

Figure 4. The contents of Pb, Cr, Cd, Ni, and Cu (a,c,e,g,i) in C. erectus tissues and total removal of Pb, Cr, Cd, Ni, and Cu (b,d,f,h,j) by C. erectus from the shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated by a one-way ANOVA, followed by an LSD test.

Figure 5. The influences of biochar (BC), lignin (LN), and citric acid (CA) applications on Pb (a), Cr (b), Cd (c), Ni (d), and Cu (e) concentrations in the leachates collected at 1st (30 days), 2nd (60 days), 3rd (90 days), and 4th (120 days) sampling periods. The standard errors are calculated from the triplicates of each treatment. Dashed lines depict the critical limits of Pb, Cr, Cd, Ni, and Cu for safe water reuse and agricultural purposes [39,40].

Figure 6. The influences of biochar (BC), lignin (LN), and citric acid (CA) on protease (a), chitinase (b), urease (c), acid phosphatase (d), and catalase (e) activities in shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA). The standard errors are calculated from the triplicates of each treatment. The treatments exhibiting distinct alphabetic characters are statistically different from each other at p < 0.05, calculated using a one-way ANOVA, followed by an LSD test.

Table 1. Treatment plan for this experiment.

Treatments	Abbreviations	Amount Added in Soil
		Biochar/Lignin	Citric Acid
		(% of Soil)	(mmol kg⁻¹ Soil)
Shooting-range soil	CK	0/0	0
Shooting-range soil + biochar	BC	5/0	0
Shooting-range soil + lignin	LN	0/5	0
Shooting-range soil + biochar + lignin	BC+LN	2.5/2.5	0
Shooting-range soil + citric acid	CA	0/0	5
Shooting-range soil + biochar + citric acid	BC+CA	5/0	5
Shooting-range soil + lignin + citric acid	LN+CA	0/5	5
Shooting-range soil + biochar + lignin + citric acid	BC+LN+CA	2.5/2.5	5

Table 2. Influences on catalase (CAT), ascorbate peroxidase (APX), and superoxide dismutase (SOD) activities, while contents of hydrogen peroxide (H₂O₂), malondialdehyde (MDA), and superoxide radical (O₂^•−) generation rate in C. erectus growing on shooting-range soil (SS) amended with biochar (BC), lignin (LN), and citric acid (CA) (mean ± SE, n = 3). For every column, the numbers having different lower-case letters statistically differ from each other and are calculated by using a one-way ANOVA, followed by an LSD test.

Treatments	Antioxidant Enzymes			Reactive Oxygen Species
	CAT	APX	SOD	H₂O₂	MDA	O₂^•−
	(μmol min⁻¹ mg⁻¹ Protein)		(U min⁻¹ mg⁻¹ Protein)	(nmol g⁻¹ FW)		(nmol min⁻¹ g⁻¹ FW)
CK	28.8 ± 0.9 ef	0.27 ± 0.01 de	41.6 ± 1.4 de	86.4 ± 2.9 a	83.3 ± 2.8 a	75.8 ± 2.5 a
BC	42.7 ± 1.4 ab	0.50 ± 0.02 a	58.7 ± 1.9 a	48.0 ± 1.6 d	43.7 ± 1.5 d	41.6 ± 1.4 d
LN	35.2 ± 1.2 d	0.37 ± 0.01 c	47.0 ± 1.6 c	69.4 ± 2.3 b	56.6 ± 1.9 b	57.6 ± 1.9 b
BC+LN	39.5 ± 1.3 bc	0.42 ± 0.02 b	52.3 ± 1.8 b	60.8 ± 2.0 c	50.1 ± 1.7 c	52.3 ± 1.8 bc
CA	26.7 ± 0.9 f	0.25 ± 0.01 e	38.4 ± 1.3 e	90.7 ± 3.0 a	85.4 ± 2.9 a	79.0 ± 2.7 a
BC+CA	37.4 ± 1.3 cd	0.40 ± 0.02 bc	43.7 ± 1.5 cd	58.7 ± 1.9 c	47.0 ± 1.6 cd	49.1 ± 1.6 c
LN+CA	30.9 ± 1.0 e	0.31 ± 0.01 d	40.2 ± 1.3 de	65.1 ± 2.2 bc	52.3 ± 1.8 bc	53.4 ± 1.8 bc
BC+LN+CA	44.8 ± 1.5 a	0.53 ± 0.02 a	61.8 ± 2.1 a	44.8 ± 1.5 d	41.6 ± 1.4 d	38.4 ± 1.3 d

Disclaimer/Publisher’s Note: The statements, opinions and data contained in all publications are solely those of the individual author(s) and contributor(s) and not of MDPI and/or the editor(s). MDPI and/or the editor(s) disclaim responsibility for any injury to people or property resulting from any ideas, methods, instructions or products referred to in the content.

© 2024 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https://creativecommons.org/licenses/by/4.0/).

Share and Cite

MDPI and ACS Style

Tauqeer, H.M.; Lewińska, K.; Umar, M.; Mahmood, F.; Shahzad, T.; Sagheer, F.; Sajid, H.; Chaudhary, I.; Iqbal, M. Induced Phytomanagement of Multi-Metal Polluted Soil with Conocarpus erectus Supported by Biochar, Lignin, and Citric Acid. Minerals 2024, 14, 1149. https://doi.org/10.3390/min14111149

AMA Style

Tauqeer HM, Lewińska K, Umar M, Mahmood F, Shahzad T, Sagheer F, Sajid H, Chaudhary I, Iqbal M. Induced Phytomanagement of Multi-Metal Polluted Soil with Conocarpus erectus Supported by Biochar, Lignin, and Citric Acid. Minerals. 2024; 14(11):1149. https://doi.org/10.3390/min14111149

Chicago/Turabian Style

Tauqeer, Hafiz Muhammad, Karolina Lewińska, Muhammad Umar, Faisal Mahmood, Tanvir Shahzad, Faiqa Sagheer, Hina Sajid, Iqra Chaudhary, and Muhammad Iqbal. 2024. "Induced Phytomanagement of Multi-Metal Polluted Soil with Conocarpus erectus Supported by Biochar, Lignin, and Citric Acid" Minerals 14, no. 11: 1149. https://doi.org/10.3390/min14111149

APA Style

Tauqeer, H. M., Lewińska, K., Umar, M., Mahmood, F., Shahzad, T., Sagheer, F., Sajid, H., Chaudhary, I., & Iqbal, M. (2024). Induced Phytomanagement of Multi-Metal Polluted Soil with Conocarpus erectus Supported by Biochar, Lignin, and Citric Acid. Minerals, 14(11), 1149. https://doi.org/10.3390/min14111149

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Menu