The Potential of Hydroxyapatite for the Remediation of Lead-Contaminated Territories: A Case Study of Soils in Primorsky Krai

Abstract

Finding ways to enhance the resilience of soil ecosystems in the context of heavy metal contamination remains an important and urgent challenge. This work is devoted to assessing the impact of the soil composition in Primorsky Krai on the efficiency of using hydroxyapatite to decrease lead intake into plants. The physicochemical characteristics of Luvic Anthrosol and Gleyic Cambisol and their absorption properties with respect to lead have been studied. Adsorption, distribution of forms, and biotesting were carried out under lead saturation of soils conditions. It has been found that soil composition determines sorption properties and the proportion of mobile lead. The high organic carbon content in Gleyic Cambisol explains its high adsorption capacity and low content of water-soluble lead fraction. The addition of hydroxyapatite reduces the water solubility of lead in Luvic Anthrosol by three orders of magnitude and in the ion mobile form by one order. The capacity of hydroxyapatite decreases by more than thirty times when added to Luvic Anthrosol. With a ratio of hydroxyapatite/soil 0.2, oat germination increases by 18.7%, average seedling length increases by 7 cm, and lead uptake into tissues decreases by 83%.

1. Introduction

The stability of soil ecosystems is a crucial prerequisite for maintaining soil fertility and ecological integrity. The entry of pollutants into the soil as a result of anthropogenic activity leads to degradation and disrupts the stable functioning of these ecosystems. Such disturbances disrupt the critical ecological functions provided by soils, resulting in the death of organisms in trophic networks and a decrease in overall biodiversity.

Soil pollution by heavy metals is a serious environmental problem of modern times. The content of metals in soils can be increased due to natural factors, such as the geological composition of parent rocks and volcanic activity [1,2]. Anthropogenic activities also contribute to this problem [3,4]. Urban soils become polluted through the deposition of dust and aerosols and by industrial activity, as well as by the use of fertilizer and paint [5,6,7]. Lead is released from car paint and exhaust fumes. In agricultural areas, pollution occurs through atmospheric transport and the combustion of fuel and due to the use of fertilizers and plant protection chemicals [8,9,10,11].

According to research, areas of the world that are most contaminated with heavy metals include Linfen, Shenyang in China; Haina in the Dominican Republic; Ranipet in India; Mailuu-Suu in Kyrgyzstan; and cities in Zambia that are near lead and cadmium mines [12,13]. In Russia, the issue of heavy metal soil contamination is a serious problem, particularly in Primorsky Krai (Primorye Territory). The eastern coast of the region has experienced a prolonged environmental crisis due to the activities of industries such as the Mining and Chemical Company “BOR” LLC, the lead plant in the village of Rudnaya Pristan, and the Khrustalnensky mining and processing plant [14,15]. These facilities created persistent geochemical abnormalities, especially in Dalnegorsk, where there were systematic excesses that exceeded the hygiene standards (since 2007–2014) for heavy metals in the soil, affecting the Terney District [16]. Residential areas, agricultural land, and garden plots here are located in zones of high and medium pollution, with a significant portion of the population living in the sanitary protection zone of class 1 hazardous enterprises [14].

The main sources of toxic elements are tailing ponds containing lead, cadmium, and copper, as well as other heavy metals [15]. In the village of Rudnaya Pristan, soil contamination is associated with natural factors and a century-old history of lead mining and processing [16]. Lead concentrations around tailings dams exceed permissible levels by 18 times [17], and in the area influenced by the third tailing dump of the Khrustalnensky mining and processing plant, soil, sediment, and vegetation pollution exceed background levels by 1.5–32 times over a distance up to 5 km [15,18]. From 2012 to 2018, lead concentrations in slurry waters from the Kavalerovsky District and the Rudnaya River exceeded regional background levels by 619 and 1280 times, respectively, which is associated with the oxidation of sulfide minerals [19].

Coal mining operations, such as the Lipovtsy coal field mine No. 4, further amplify the problem: soils adjacent to waste dumps exhibit lead levels as high as 22.09 mg/kg, with elevated concentrations persisting even at considerable distances [20]. Soil degradation driven by pollution manifests through reduced fertility within a 1–2 km radius of tailings ponds, alongside disrupted soil structure and suppressed vegetation growth [17]. Lead bioaccumulation in plants reaches 2–30 times permissible thresholds, causing teratological deformities, impaired reproductive capacity, and pollen sterility (reaching 25% in heavily contaminated zones) [15,18]. The threat to agriculture is acute; for example, soybean yields directly depend on the concentrations of lead and cadmium in the soil [21].

Environmental contamination extends beyond industrial areas. In Vladivostok, urban soil is polluted with chromium, copper, zinc, and lead, and the situation is exacerbated by topsoil degradation [22]. In Ussurisk and its surrounding area, plant tissue exhibits abnormally high concentrations of lead [23]. The concentration of heavy metals in the sediments of the Ussuri River is increasing due to anthropogenic impact, which creates risks for wetlands [24]. Elevated levels of lead have been detected near Ussuriysk, in the Razdoly River, although the source of contamination remains unknown [25]. A government report from 2022 Primorsky territory linked transport infrastructure with lead pollution: in Nakhodka and Partizansk, the average regional values were exceeded, while the proportion of soil samples exceeding regulatory limits increased from 13.1% in 2020 to 25% in 2022 [26].

Thus, the long-term effects of heavy metals on ecosystems and human health necessitate the implementation of innovative techniques to restore soil coverage and reduce pollution. In order to restore ecological functions and reduce pollution levels to acceptable levels, it is essential to carry out soil remediation. A safe, technically feasible solution is to introduce additives that reduce levels of water-soluble, exchangeable forms of lead (immobilization), which reduces its availability.

Liming of soils reduces the bioavailability of Pb and Cd [12]. The use of lime fertilizers or composite additives reduces the content of Pb and Cd in rice sprouts and grains [27]. Additives based on oyster shells and eggshells reduce the concentration of Cd by 63–77% and Pb by 47–75% in corn sprouts. The addition of CaO and MgO reduces the bioavailability of Cd, Cu, and Pb. However, an assessment of long-term stability is necessary for limestone fertilizer [28]. At high dosages (5–10 wt.%), this can negatively affect plant growth due to a decrease in the availability of essential trace elements such as Fe, Mn, Zn, Cu, and B [29] and an increase in the proportion of mobile forms of heavy metals [30]. The addition of fly ash to contaminated soils reduces the bioavailability of Pb in rape seed [31] and rice [32] by forming insoluble hydroxide compounds and increasing soil pH [32]. However, fly ash can itself be a potential source of contamination [33,34] and can pose a risk to soil, water, and plants [35]. Despite the effectiveness of liming and its availability, its effects are only temporary, as soil pH decreases over time. For this reason, it is essential to continuously monitor, as when soil becomes acidic Pb becomes more bioavailable in its gross form.

The presence of ion-exchange groups in compost allows it to effectively bind heavy metals such as Pb, Cu, Cd, and Ni [36]. Animal-derived compost reduces the bioavailability of Pb by forming low-solubility phosphate minerals, thereby improving plant growth and microbial activity [37]. The addition of 10% compost reduces Pb availability by about 92% but increases arsenic mobility.

Commercial compost has been demonstrated to reduce arsenic absorption in clay soils, diminishing its overall positive impact as an additive [38]. Biochar is a popular material that is used as a complex soil additive, improving soil quality and adsorbing heavy metals, including Pb [39,40,41]. It has been shown that the bioavailability of Pb in soil treated with biochar is reduced due to the formation of two specific minerals, chlorpyromorphite (Pb₅(PO₄)₃Cl) and hydroxypyromorphite (Pb₅(PO₄)₃(OH)) [42].

Despite the numerous advantages, biochar tends to “age”, which is manifested in changes in its properties due to external factors [43]. These changes affect the elemental composition, aromatic group, specific surface area, morphology, pH, and ion exchange capacity [44]. The aging of biochar can increase the mobile and water-soluble fraction of Pb in the soil due to acidification, and also the amount of Pb involved in the formation of contaminated colloidal particles which enter the groundwater [44]. The effectiveness of using biochar for immobilizing pollutants depends on several factors: the type of plant, the type of biochar used, the climate, and the soil conditions [3].

Aluminosilicate materials, such as bentonite, have been shown to reduce the mobility of Pb by more than 40% compared to a control experiment [45]. Bentonite has been successfully used to immobilize Cd and Pb in contaminated floodplain soils [31]. Zeolite increases soil pH and reduces concentrations of Cd and Pb in rice grains by 27.9% and 63.5%, respectively [46]. Expanded clay can also be used as a soil additive [47]. The addition of gibbsite or iron crumbs reduces the mobility of Pb in soil [48,49]. The disadvantage of these materials is their low selectivity for Pb. Manganese oxides are more effective at immobilizing Pb than iron oxides [50,51]. However, manganese oxides are sensitive to redox potential, which can change the properties of the additive due to the reduction of Mn(IV) to Mn(II) [52,53].

Apatites, hydroxyapatites, diammonium phosphate, and phosphate-based rocks have been successfully used as additives to immobilize heavy metals, such as Pb. The use of phosphate additives causes the precipitation of pyromorphite, which is a mineral that has poor solubility over a wide pH range and under various redox conditions [54,55]. The use of organic fertilizers with a high phosphorus content, such as poultry waste, can lead to the immobilization of Pb into PbCO₃, followed by its transformation into chlorpyromorphite [56]. However, the excessive application of certain types of phosphate fertilizers, such as triple superphosphate and phosphorite, may increase the uptake of Cd by plants due to a reduction in soil pH [57].

Additives based on hydroxyapatite (HAP), a substance that is insoluble and nontoxic, can be used to immobilize Pb in soils [58]. Hydroxyapatite is used for purifying water from pesticides [59], heavy metals [60], and radionuclides [61]. Hydroxyapatite does not exhibit the disadvantages associated with other soil additives and has a high selectivity for Pb. As a result, alkaline and alkaline earth cation have a minimal impact on the efficiency of Pb extraction from water solutions [62,63]. Hydroxyapatite can be used independently or in combination with biochar to enhance selectivity [64,65,66]. However, the characteristics of these complex soil additives are determined by the properties of the biochar, which depend on the type of raw materials, pyrolysis conditions, and other factors [67]. The addition of nanoscale hydroxyapatite into soil reduces the percentage of water soluble Pb by 72% and Cd by 90% [58]. Hydroxyapatite can be used to immobilize Pb and Cd in the long-term, when combined with biochar and bentonite [68]. The combination of HAP, bentonite, and biochar, as a soil additive, has been shown to effectively reduce the uptake of Cd and Pb by edible plant parts [69].

Nanocomposites based on natural polymers (chitosan, alginate, and cellulose) have been developed to reduce lead bioavailability. High efficiency is achieved due to the large specific surface area and presence of additional amino and hydroxyl functional groups. The biodegradable and low-toxicity nature of these polymers reduces their negative impact on the environment [70]. Hybrid systems with fungal hyphae (Aspergillus niger) with n-HFP/n-HFS nanoparticles reduce the water-soluble fraction of lead by 63% and the bioavailable fraction by 60%. Hyphae prevent the aggregation of nanoparticles and transform mobile metal forms into more stable ones. However, this method requires a lengthy equilibrium period of 42 days and is difficult to implement in heterogeneous soils [71]. ANHP biocomposite (hydroxyapatite + biochar) immobilizes lead due to the formation of stable minerals and fungal hyphae fixation. Fungal hyphae also prevent the aggregation of HAP nanoparticles and increase the stability of the composites. The immobilization efficiency of lead is 42.6%, which is higher than that of pure HAP. ANHP biocomposite can be used for high lead concentrations (2548–6120 mg/kg), but its effectiveness depends on the fungal type [72]. Modified biochar (montmorillonite + chitosan) reduces lead mobility by 73% with the addition of 10%. It improves soil porosity but requires an assessment of its long-term impact on ecosystems [73].

Despite great success, the use of nanomaterials is associated with a number of limitations: scaling difficulties, high cost of synthesis, the negative impact of metal nanoparticles on the soil microbiota, and competition with the agricultural sphere (for example, the use of corn for PLA production). Although nanomaterials based on natural polymers for soil lead purification are a promising area, their implementation requires technology optimization and environmental risk assessment [70].

The aim of this study is to investigate the potential of hydroxyapatite in reducing the concentration of bioavailable Pb in the soil of Primorsky Krai. Due to its several advantages, such as high selectivity, stability, and biocompatibility, the use of hydroxyapatite-based additives appears to be the most promising approach for restoring soils in Primorsky Krai contaminated with Pb.

2. Materials and Methods

2.1. Materials and Reagents

The study is devoted to analyzing the effectiveness of hydroxyapatite (HAP) in reducing the bioavailability of lead in contaminated soils of Primorsky Krai, as well as assessing the role of soil composition in the success of remediation measures. In the experimental part, model lead solutions were prepared using lead(II) nitrate(Pb(NO₃)₂), which was also used to saturate soil samples during the investigation of metal form distribution depending on the soil-to-hydroxyapatite ratio. Control of ionic strength in the model solutions during adsorption experiments was ensured by adding sodium nitrate (NaNO₃). To characterize the physicochemical parameters of the soils, including organic carbon and humus content, analyses were conducted using sulfuric acid (H₂SO₄), sodium pyrophosphate (Na₄P₂O₇), sodium hydroxide (NaOH), and potassium permanganate (KMnO₄).

Hydroxyapatite was synthesized via a chemical precipitation method, with ammonium hydrogen phosphate ((NH₄)₂HPO₄) and calcium chloride tetrahydrate (CaCl₂ × 4H₂O) used as precursors. pH adjustment during synthesis was performed using ammonium hydroxide (NH₄OH). All chemical reagents meeting the “chemically pure” qualification were obtained from NevaReaktiv LLC and used without further purification.

2.2. Sampling and Preparation of Soils

Two types of soil were studied in Primorski Krai: Gleyi Cambisols (common in urban recreation areas) and Luvic Anthrosol (predominantly in arable land) [74]. Luvic Anthrosols were selected in May of 2022 at the Primorskaya Vegetable Experimental Station of VNIIO, LLC (Surazhevka, Russia). Gleyi Cambisols were collected in July of 2019 from Cape Vyatlina (Russky Island). Samples were collected using the envelope method (5 points) from the top horizon (0–5 cm) of 2 kg of each type. The size of the envelope was 5 × 5 m. The samples were then dried, cleaned, and crushed, and fractions of 0.25–1.0 mm and 2.0 mm isolated.

2.3. Evaluation of the Main Soil Characteristics

The moisture content of soil samples was determined by repeating the experiment three times with drying to a constant weight [75]. Moisture content (W, wt.%) was calculated using Formula (1):

$$W={\displaystyle \frac{A-B}{B-C}}\times 100,$$

(1)

where A is the weight of the initial soil sample with a vial (g), B is the weight of the dried soil sample with a vial (g), and C is the weight of the vial (g).

The density of the solid phase of soil (p_s) was determined using the pycnometric method [76]. An amount of 10 g of soil was placed in a pycnometer, filled with distilled water, and kept for 10–12 h. The suspension was then boiled for 30–60 min, cooled, and brought to the desired mark with water. The weight of the suspension was measured, and the ps value (g/cm³) was calculated using Formula (2).

$$p_s={\displaystyle \frac{v_s}{v_s+P{W}_1-P{W}_2}}\times d{H}_{2}O,$$

(2)

where Vs is the weight of the air-dried soil sample (g), PW₁ is the weight of the vial with water (g), PW₂ is the weight of the vial with water and soil (g), and dH₂O is the density of water at a given temperature (g/cm³).

The granulometric analysis of soil was carried out according to standard methodology [77]. Sampling was conducted taking into account density of solid phase and temperature of water.

The determination of the pH of water (pH(H₂O)) and salt extracts at 1 M concentration of KCl (pH(KCl)) was carried out in accordance with GOST 26423-85 [78]. Soil extracts were prepared by mixing 25 mL of water with 10 g of the sample, followed by stirring. After 5 min, the pH value of the resulting aqueous extract was measured. To determine the pH of a salt extract, 1 M of KCl solution was added, and the pH measurement was taken after 24 h.

Organic carbon content (C_(org)%) as a percentage of air-dry soil was determined using Formula (3):

$$C_{\left(org\right)}={\displaystyle \frac{\left(a-b\right)\times n\times 0.03\times 100}{p}},$$

(3)

where a is the amount of Mohr’s salt used to titrate 10 mL of the chromium mixture in a blank experiment (mL) and b is the amount of molarity of Mohr’s salt used in the titration of excess chromium in the soil experiment (mL). n is the molar concentration of the Mohr’s salt solution, determined by titrating it with a 0.05 N solution of KMnO₄ and 0.003 is the weight of 1 milligram equivalent of carbon. p is the weight of the soil sample (g) and 100 represents the factor for converting the analysis results to 100 g of soil.

The group composition of humus was determined using the method of Kononova and Belchikova [79], which involves the extraction of humic and fulvic acids. To 5 g of soil, 100 mL of a mixture of Na₄P₂O₇ (0.1 M) and NaOH (0.1 M) was added, incubated for 12 h in a sealed flask, and then filtered. To determine the total carbon of humic and fulvic acids (C₁), an aliquot of the extract was neutralized with 1 M H₂SO₄, evaporated to dryness, and the carbon content was measured using wet burning methods [80]. C₁ was calculated using Formula (3), where p₁ represents the sample weight corresponding to the aliquot volume of the original extract (g).

For analyzing the carbon of humic acids (C_(hum)), an aliquot of the pyrophosphate extract was neutralized with H₂SO₄, heated to 80 °C, cooled, and incubated for 12 h. The humic acid gel was separated using a “blue ribbon” filter, dissolved in 0.05 M NaOH, and the carbon content was determined by method [80]. C_(hum) was calculated using Formula (3), where p₂ denotes the sample weight corresponding to the aliquot volume of the pyrophosphate extract (g).

The carbon content of fulvic acid (C_(fulvo)) in alkaline extracts is calculated from the difference between the total carbon content of humic substance and the carbon of humic acid (Formula (4)):

$$C_{\left(fulvo\right)}=C_1-C_{\left(hum\right)},$$

(4)

2.4. Synthesis of Hydroxyapatite

Hydroxyapatite was prepared according to the method described in [81]. A 0.25 M solution of (NH₄)₂HPO₄ was mixed with 0.5 M CaCl₂ for 60 min with continuous stirring at room temperature. The molar ratio of Ca/P in the final mixture was 1.67. The pH of the mixture was maintained at pH 12 using concentrated NH₄OH. The completed mixture was heated to a boil with continuous stirring and then kept for one day. The precipitate was separated from the solution using filtration, and washed with hot distilled water under vacuum. The material was then dried at 150 °C for six hours and calcined at 400 °C in an air atmosphere.

2.5. Adsorption Experiments

The process of Pb adsorption on HAP and soil samples was carried out under static conditions. Air-dried samples were placed in plastic tubes and 50 mL of the solution was added. After that, the mixture was shaken for a specified period. At the end of the test, the solution was separated from the sample by centrifugation at 4000 rpm. The value of the static exchange capacity (SEC) was calculated using Formula (5).

$$SEC={\displaystyle \frac{\left(C_o-C_t\right)*V_s}{m_{HAP}}},$$

(5)

where C_o is the concentration of Pb in the initial solution (mg/L), C_t is the residual equilibrium concentration of Pb in the solution (mg/L), V_s is the volume of the solution (L), and m_HAP is the HAP weight (g).

The adsorption isotherm was obtained by putting HAP or soil samples into contact with a series of solutions with various concentrations of Pb. To control the ionic strength, NaNO₃ with a concentration of 0.01 M was added to the model solutions.

The standard Freundlich Equation (6), Langmuir Equation (7), and Sips Equation (8) were used to describe sorption isotherms.

$$\Gamma =K_f*C_e^m,$$

(6)

$$\Gamma ={\displaystyle \frac{G_{max}*K_l*C_e}{1+K_l*C_e}},$$

(7)

$$\Gamma ={\displaystyle \frac{G_{max}*K_{lf}*C_e^m}{1+K_{lf}*C_e^m}},$$

(8)

where G_max is the maximum sorption capacity (mmol/g); C_e is the equilibrium concentration of Pb (mmol/L); and K_f is the Freundlich constant. It characterizes the relative adsorption capacity. It represents the value of the adsorption at an equilibrium concentration of unity and K_l and K_lf are the constants of adsorption equilibrium, which characterize binding energy between adsorbent and adsorbate, and m is the index of heterogeneity of exchange centers, which characterizes changes in heat of adsorption, depending on degree of filling.

2.6. Determination of Pb Forms in Soil Samples

To evaluate the ability of HAP to reduce the proportion of mobile and water-soluble forms of Pb, we used an approach based on complete saturation of soil samples. We took into account calculated maximum adsorption values (Equations (7) and (8)). When maximum soil saturation was reached, Pb concentrations exceeded established standards significantly. This simplified and improved the accuracy of analysis by eliminating the need for further sample concentration steps.

The evaluation of the efficacy of HAP for the distribution of Pb concentration in Luvic Anthrosols has been evaluated in the following studies:

“Experiment A”: an assessment of HAP effectiveness when Pb enters soil that has not been contaminated.

“Experiment B”: evaluation of HAP effectiveness in soil already contaminated with Pb.

In “Experiment A”, the ability of HAP to reduce the available forms of lead in uncontaminated soils was studied. Soil samples (5 g, <0.25 mm fraction) were treated with HAP and then incubated with a lead nitrate solution (0.016 M, 50 mL, 7 days). The lead concentration required to fully saturate the soil samples was calculated using previously determined maximum adsorption capacity values (G_max). After centrifugation, lead forms were analyzed. For Gleyic Cambisols, only this scenario was investigated.

In “Experiment B”, the effect of HAP on already contaminated soils was evaluated. Samples (100 g) were saturated with Pb by incubation with a lead nitrate solution (0.058 M, 250 mL, 7 days), followed by drying. Then, 5 g of contaminated soil was treated with HAP and incubated with 0.01 M NaNO₃ (7 days). Lead forms were analyzed similarly.

The proportion of the water-soluble fraction was determined according to RD 52.18.286-91 (water extraction, 60 min) [82], while mobile forms were assessed using RD 52.18.289–2022 with ammonium-acetate buffer (AAB) [83].

2.7. Biological Testing

Biotesting was conducted in accordance with GOST R ISO 18763-2019 [84]. Soil samples were sieved to a 2–3 mm fraction, placed in Petri dishes (50 g each), and seeded with 100 oat (Avena sativa) seeds. The experiment was carried out under natural lighting in 4 series, each with three replicates (

Table 1. Experimental legend.

Legend	Addition of Pb	HAP/Soil Mass Ratio
Control	No	0
H(0)	Yes	0
H(10)	Yes	0.1
H(20)	Yes	0.2

). An experiment without the addition of lead or HAP was used as the control group.

In experiments H(10) and H(20), HAP was mixed with soil mechanically and distributed into Petri dishes in layers of 5–7 mm. All samples (except controls) were treated with lead at a concentration of 1.25 g: 50% on the first day and the rest on days 4–5. The dishes were closed for 48 h to allow germination of seeds, then moistened periodically with distilled water. After 2 weeks, the germination rate was evaluated, as well as biomass, sprout lengths, and lead concentrations in plant tissue.

Lead content in plant tissue was determined using a method described in [85]. Dried plant samples were washed at 525 ± 25 °C, and ash residue was dissolved in 15 mL of HNO₃ (1:1), heated in a water bath for 30 min, filtered (“white ribbon” filter), and adjusted to 50 mL final volume. Lead concentration was then measured by atomic absorption spectrometry.

The effects of lead and HAP were quantified by calculating the difference in mean values between control groups (control, H(0)) and experimental groups (H(10) and H(20)). Statistical significance was determined using Student t-test (p < 0.05) and Welch test was used for unequal variances. Due to the non-normal distribution of data (Shapiro–Wilk test, p < 0.001), sprout length was compared using Mann–Whitney U-test (p < 0.05). To minimize variability (likely caused by uneven HAP and lead distribution), replicates under identical conditions were compared. Outliers (p < 0.05) were excluded, and remaining data were pooled.

The phytoeffect size (PhyE) was calculated using Formula (9):

$$PhyE={\displaystyle \frac{L_{contr}-L_{eff}}{L_{contr}}}\times 100,$$

(9)

L_contr is the average sprout length (mm) in the control group. When assessing the effect of lead, a control without its addition was used. To assess the effect of lead in the presence of HAP, negative control was used (without HAP, experiment H(0)); L_eff is the average sprout length (mm) in the experimental group with the addition of HAP.

2.8. Equipment and Software

The study used: a pH meter “FiveEasy F20” (Mettler Toledo, Greifensee, Switzerland) with a plastic electrode LE438 to measure the acidity of aqueous and salt extracts; atomic absorption spectrometers SOLAAR M6 (Thermo Scientific, Waltham, MA, USA) to determine the concentration of lead (Pb) in solutions, and AA6800 (Shimadzu, Kyoto, Japan) for analysis of Pb in plant tissues; D8 ADVANCE diffractometer (Bruker, Berlin, Germany) for phase analysis in the range of 2θ 3–85° (step 0.02°, exposure time 0.6 s/point). The phase composition was identified using the Qualx 2.0 program [86] based on the Crystallography Open Database (COD, version SQLITE3, May 2022). Statistical data processing, including the Shapiro–Wilk (normality assessment) and Mann–Whitney (sample comparison) criteria, was performed in the GNU PSPP program (GNU PSPP ver. 2.0.1. for Linux, GNU General public licence).

3. Results

Table 2. Characteristics of soil samples.

Indicator		Soil
		Luvic Anthrosols	Gleyic Cambisols
pH(H2O)		6.27 ± 0.16	5.23 ± 0.03
pH(KCl)		5.66 ± 0.07	4.35 ± 0.16
Solid phase density (g/cm3)		2.32	2.39
Content of granulometric fraction (%)	2–0.05 mm	23.5	46
	0.05–0.002 mm	63.5	45
	<0.002 mm	13.0	9
Classification of soils based on granulometric composition		Clay	Sandy clay
Humidity (W)		2.2 ± 0.1	1.6 ± 0.2

presents the results of a comprehensive analysis of the characteristics of soil samples that were significant for the study.

The pH values for Luvic Anthrosol and Gleyc Cambisol are markedly different. For Gleysic Cambisol, pH(H₂O), and especially pH(KCl), is characterized by an acidic environment, indicating the presence of many protonated functional groups, corresponding to its classification criteria [74].

According to the results, the surface horizon of Gleyic Cambisols contains 46.0% sand and 9.0% clay, which can be classified as “Sandy clay”, according to the Ferre triangle [87]. Luvic Anthrosols contains 23.5% sand and 13.0% fine clay fraction and can be classified as “Clay”.

Table 3. The content of total and organic carbon (%), as well as the group composition of humus in soil samples.

Sample	Humus	C(org)	C1	C(hum)	C(fulvo)
Luvic Anthrosols	4.27 ± 0.13	2.48 ± 0.08	0.83 ± 0.06	0.49 ± 0.05	0.34 ± 0.01
Gleyic Cambisols	14.28 ± 0.08	8.28 ± 0.04	1.72 ± 0.04	0.51 ± 0.07	1.20 ± 0.14

shows the results of a comparative analysis of the group composition of humus, which demonstrates significant differences between the studied soil types.

Figure 1 shows the X-ray diffraction patterns of the original soil samples. The samples are characterized by the presence of quartz as the main mineral component, as well as muscovite aluminosilicate, belonging to the mica group, and albite, belonging to the feldspar group.

Figure 2 shows the isotherms of Pb adsorption by the studied soils. According to Giles, the sorption isotherm can be classified as L-type, indicating a high affinity for Pb under the studied conditions. The adsorption isotherms are well described by the Sips equations (R² > 0.99).

Figure 3 illustrates the X-ray diffraction patterns of the synthesized hydroxyapatite before and after the adsorption of lead. The X-ray pattern of the initial sample clearly exhibits peaks that correspond to the hydroxyapatite phase, indicating successful synthesis. This approach was subsequently employed to produce additional batches of hydroxyapatite.

Work was carried out to evaluate the adsorption characteristics of HAP in relation to lead. Figure 4 shows adsorption isotherms for lead by HAP in 0.01 M NaNO₃ and in ammonium acetate buffer. As can be seen, HAP has a high maximum adsorption capacity and adsorption equilibrium constant in NaNO₃ solution.

To study the distribution patterns of lead forms, depending on soil type, experiments were conducted. The results were presented in the form of diagrams showing the distribution of lead under conditions of maximum soil saturation (Figure 5). For Luvic Anthrosols, results from experiments A and B were presented.

Figure 6 shows the results of the distribution of water-soluble and mobile forms of lead depending on the HAP/soil ratio in semi-logarithmic coordinates. The experimental data are fairly well described (R² > 0.9) by a linear regression equation that can be used to predict the effect of reducing bioavailable lead forms when HAP is added.

Table 4. Parameters of the linear equation describing the dependence of logarithmic fraction of lead on the content of HAP in soil.

The Form of Lead in the Soil	Experiment	Parameters of the Linear Equation
		B	A	R2
Water-soluble	Luvic Anthrosols, «Experiment A»	−1.6 ± 0.1	−15 ± 1	0.995
	Luvic Anthrosols, «Experiment B»	−1.6 ± 0.2	−18 ± 3	0.981
	Gleyic Cambisols	−3.1 ± 0.1	−10 ± 1	0.997
Mobile	Luvic Anthrosols, «Experiment A»	−0.10 ± 0.02	−3.5 ± 0.3	0.991
	Luvic Anthrosols, «Experiment B»	−0.14 ± 0.01	−3.6 ± 0.2	0.996
	Gleyic Cambisols	−0.17 ± 0.01	−3.6 ± 0.1	0.998

shows the calculated parameters for the linear equations. B is the logarithm of water-soluble or mobile Pb in soil without HAP. A is the slope of the line that characterizes the increase in efficiency of HAP with increasing content.

The adsorption capacity of HAP to lead during long-term exposure to Luvic Anthrosols was estimated. Figure 7 shows the relationship between the SEC value and the amount of added HAP. The experiment lasted for 14 days.

Laboratory studies were conducted to evaluate the effectiveness of HAP as an additive for reducing lead uptake into plant tissue using oats (Avena sativa) as an example.

Table 5. Results of evaluation of the effects of lead and HAP supplementation on oat growth after 14 days.

Experiment	Germination Rate (%)	Weight of Roots(g)	Weight of Sprouts (g)	Length of Sprouts (mm)	Lead Content in Tissues (mg/g)
Control	49 ± 16	6.1 ± 2.6	3.8 ± 2.8	101.3 ± 10.9	0
H(0)	26 ± 17	3.7 ± 2.8	0.2 ± 0.2	12.1 ± 1.8	5.04 ± 1.06
H(10)	39 ± 21	4.3 ± 2.0	0.4 ± 0.7	19.6 ± 4.3	1.71 ± 0.33
H(20)	44 ± 11	4.7 ± 0.5	1.5 ± 2.0	27.8 ± 6.1	0.82 ± 0.26

shows the results of the effect of adding lead and HAP on oat growth, formation, and lead accumulation in tissue.

Figure 8 shows the results of statistical analysis of the data obtained during the laboratory experiment to investigate the effect of the application of HAP and dose on the germination, growth, and weight of sprouts, as well as the lead content in oat tissues (Avena Sativa).

4. Discussion

The presence of increased clay content in Luvic Anthrosols (Table 2) is explained by plowing, during which the upper horizon, enriched in organic matter, is mixed with the lower horizon, which is enriched in aluminosilicate minerals. The increase in humus in Gleyic Cambisols (Table 3) is associated with the enrichment of the upper horizons with plant litter due to the small biological cycle. A genetic feature of this type of soil in Primorsky Krai is the high content of fulvic acids in humus. Fulvic acids have active adsorption centers in the form of carboxyl and hydroxyl functional groups [88,89], which explains their low hydrogen index compared to Luvic Anthrosols.

The adsorption isotherms of lead by soils are close to the Sips model, which indicates complete saturation of the monomolecular layer in the concentration range. However, significant heterogeneity of adsorption centers is observed, which is associated with the variety of adsorption-active components (oxides, silicates, aluminosilicates, and high-molecular organic compounds) in the soil. Gleyic Cambisols have a higher maximum capacity (G_max) than Luvic Anthrosols, which is explained by the increased content of organic substances, in particular fulvic acids. Lee et al. [90] showed that soils with a high content of organic matter adsorb lead better. The main active centers are associated with organic matter, and metal oxides do not play a significant role in lead adsorption by soils [90]. Zhao et al. [91] showed that solid organic matter plays a significant role in lead adsorption through interaction with phenolic and carboxylate functional groups. However, minerals reduce the affinity of lead compared to humic acid [91]. Coutinho et al. [92] showed, using Histosols as an example, that lead has a greater affinity for soil compared to copper. Lead adsorbs non-specifically to fulvic acid, which is enriched in aliphatic structures. Compared with Gleyic Cambisols, Luvic Anthrosols are characterized by a higher adsorption equilibrium constant (K), which may indicate that adsorption occurs on an inorganic part with a higher affinity for Pb compared to organic components. Lead can adsorb on feldspar, which is described in the literature as a material for extracting heavy metals [93,94,95].

The process of lead absorption by HAP involves the dissolution of Ca₁₀(PO₄)₆(OH)₂ and the precipitation of a hydroxypyromorphite with the composition of Ca_8.51Pb_1.49(PO₄)₆(OH)₂. This is caused by an ion exchange reaction between Ca and Pb ions. The molar ratio of Ca to Pb in the solid phase decreases until a pure HAP structure is reached [96,97]. However, the final product in our experiment was lead phosphate (Figure 3, curve 2). The adsorption equilibrium constant (K) for HAP is 28 and 207 times higher than the values for Gleyic Cambisols and Luvic Anthrosols, respectively (Figure 4a). This suggests that HAP will significantly reduce the bioavailable fractions of Pb by irreversibly binding to form lead phosphate. However, in ammonium acetate buffer, the efficiency of Pb extraction decreases and the experimental values are described by Freundlich’s equation (Figure 4b). This is due to the increase in the ionic strength of the solution and the competing effects of K⁺ and NH₄⁺ ions. The results indicate that HAP is effective in reducing the fraction of water-soluble lead forms, but less effective for mobile forms.

The diagrams of Pb forms’ distribution in soil samples at maximum saturation (Figure 5) showed some features. For Luvic Anthrosols, the proportion of the water-soluble form is 2.8% and 5.4% for “Experiment A” and “Experiment B”, respectively. For Gleyic Cambisols, the proportion of water-soluble form does not exceed 0.15%, which is consistent with the adsorption isotherm data (Figure 4). To analyze the results in detail, the distribution of water-soluble and mobile forms of Pb depending on the HAP/soil ratio was plotted in semi-logarithmic coordinates (Figure 6) with subsequent linear regression. It was found that, for Luvic Anthrosols, a decrease in the proportion of the water-soluble form of Pb with an increase in the HAP content can reach three orders of magnitude in both experiments. For Gleyic Cambisols, the effect of HAP application on reducing the water-soluble fraction is less pronounced and does not exceed two orders of magnitude. The value of A is more than 1.5 times lower than that of Luvic Anthrosols (Table 4). Thus, Gleyic Cambisols are characterized by a higher degree of Pb binding, which is reflected in a decrease in the content of the water-soluble fraction compared to Luvic Anthrosols. Therefore, the effect of applying HAP to reduce the content of the water-soluble form of Pb will be more pronounced in the case of Luvic Anthrosols and less pronounced in the case of Gleyic Cambisols. In comparison with Luvic Anthrosols, the proportion of the mobile fraction is reduced in Gleyic Cambisols by more than 5%. This indicates their high ability to adsorb Pb. The values for parameter A are similar for both soil types, indicating the presence of sites that can adsorb Pb through ion exchange.

An unexpected result was the decrease in the capacity of HAP for Pb when added to Luvic Anthrosols (Figure 7). The absorption capacity of HAP decreased more than 30 times compared to the calculated maximum capacity (Figure 4). This is probably due to the negative impact of the soil components (iron, manganese, etc.) blocking the adsorption centers of HAP.

The results of the experiment show that adding Pb to irrigation water decreases seed germination and sprout length, as well as biomass (Figure 8). The introduction of Pb has a negative effect on plant growth, decreasing the average length of sprouts by 110 cm with a phytoeffect of −88.3. Preliminary application of HAP to the soil generally has a stimulating effect, affecting plant growth through strong binding of lead and reducing its availability to the root system. By adding HAP to soil and increasing its content, the negative effect of Pb on plant growth decreases. However, only the results of seed germination and sprout length in experiment H(20) are reliable (p > 0.05). When HAP was added at a concentration of 20%, the germination of oat seeds increased by 18.7%, sprout length increased by 7 mm, and the phytoeffect was 131.7 compared to experiment H(0).

There is no statistical significance in estimating the mass of root and above-ground parts of plants. Although the magnitude of effect indicates an increase in biomass when HAP is added (Figure 8b), the addition of lead in irrigation water has a more negative effect on above-ground plant parts (p < 0.01) than on root systems.

The greatest positive effect of the HAP application was observed when assessing the uptake of Pb into plant tissue. So, compared to H(0), the decrease in the Pb content of plant tissue was 66% for H(10) and 83% for H(20) (p < 0.05).

5. Conclusions

The physicochemical properties of Luvic Anthrosols and Gleyic Cambisols, including the content of physical clay, humus and its type, and the adsorption characteristics towards lead, were evaluated. Gleyic Cambisols are characterized by a fulvic humus type and a higher maximum capacity (1.8 times), possibly due to Pb adsorption on fulvic acids. Luvic Anthrosols contain more physical clay and have a higher affinity for lead, which is reflected in a higher value of the adsorption equilibrium constant (7.2 times higher than that of Gleyic Cambisols).

The adsorption properties of the synthesized hydroxyapatite were characterized in a 0.01 M NaNO₃ solution, and the maximum adsorption capacity for Pb was 1500 ± 100 mg/g. The Langmuir adsorption equilibrium constant was calculated to be 110 ± 10. When hydroxyapatite was introduced into soil, its capacity decreased by more than a factor of 30, probably due to mineral soil components, such as iron and manganese. The forms of Pb in the studied soil samples were evaluated under conditions of maximum saturation. Gleyic Cambisols have a reduced content of the water-soluble fraction of Pb of more than 20 times. The mass fraction of the mobile fraction of Pb for both samples is comparable.

The introduction of hydroxyapatite into the soil reduces the proportion of the water-soluble fraction of lead; for Luvic Anthrosols, the positive effect can be up to three orders of magnitude. The decrease in mobile Pb does not exceed one order of magnitude when the mass ratio of HAP to soil is 0.2.

The toxic effect of lead at a mass ratio of 0.2 hydroxyapatite to soil is reduced, manifesting in an increase in germination of 18.7% and an increase in sprout length of 7 mm. The proportion of Pb in oat tissue is reduced by 83% and the phytoeffect effect is 131.7 compared to a control sample without HAP.