Effects of Biochars Derived from Sewage Sludge and Olive Tree Prunings on Cu Fractionation and Mobility in Vineyard Soils over Time

Abstract

Copper-contained products that are widely employed yearly in viticulture for vine disease management, lead to Cu accumulation in topsoil creating an increased risk for land workers and for leaching and/or uptake of Cu by plants, especially in acidic soils where Cu mobility is higher. In this study, the impact of two biochar types on Cu distribution and redistribution in fractions was evaluated in four acidic vineyard soils in relation to incubation time. The two biochars were derived from sewage sludge (SG) and olive tree prunings (OL). Soils (control) and biochar-amended soils with application rate of 20 % (w/w) were spiked with CuCl₂ (160 mg kg⁻¹) and incubated in the laboratory at ambient temperature 22 ± 5 °C. After 1, 3, 7, 15, 36, and 90 days of incubation, modified BCR sequential extraction procedure was used to determine Cu distribution in the four soil chemical phases and to examine potential Cu redistribution between these fractions both in soils and in amended soils with biochars. Results show that biochar amendment affects Cu distribution in different soil fractions. In SG treatment, from the 1st and up to 36th incubation day, both exchangeable and reducible Cu fractions decreased, while oxidizable Cu increased, in relation to control soils. At 90th incubation day, a redistribution of Cu was observed, mainly from the oxidizable to the residual fraction. In OL treatment, during the first 36 incubation days exchangeable and oxidizable Cu slightly increased, while reducible Cu decreased. At the 90th incubation day the higher percentage of Cu was extracted from the residual fraction, but exchangeable Cu was present in remarkable quantities in the three of the four studies soils. SG application in the studied soils highly restricted the availability of added Cu promoting Cu-stable forms thus reducing the environmental risk while OL did not substantially reduce Cu available fraction over the experimental incubation period. Fourier transformation infrared analysis (FTIR) provided convincing explanations for the different behavior of the two biochar types.

1. Introduction

The contamination of soils with potentially toxic elements (PTEs) is one of the main risks to planetary natural and environmental resources [1,2]. Although the origin of PTEs in soils may be either lithogenic (naturally occurring in the soil environment as a result of the weathering of parent material) or anthropogenic, high PTE concentrations are typically the result of human activities [1,2]. A typical example of a PTE that has received multiple reports in the literature on the contamination of agricultural soils with its presence is Cu [3,4,5].

Copper has an average abundance of 60 mg Cu kg⁻¹ in the Earth’s crust, but typical soil concentrations range from 2 to 50 mg kg⁻¹ [6]. Cu is a trace element that is necessary for the functioning of all lifeforms (humans, plants, animals, and micro-organisms). It is found in a wide variety of enzymes and proteins, including cytochrome C oxidase and particular superoxide dismutases (SODs) [6,7]. Soils with low bioavailability of Cu can cause crop yield losses and deficiency symptoms in livestock, particularly in intensive farming systems. The daily recommended intake of Cu for humans is 1–2 mg, while for plants, the optimum Cu concentration levels depend on the individual plant species and its needs [6,7,8]. Plastocyanin, which is involved in the photosynthetic electron transport in the thylakoid lumen of chloroplasts, is the most abundant Cu protein in higher plants. Cu/Zn SOD (superoxide dismutase), another major Cu protein, is found in the cytoplasm, chloroplast stroma, and peroxisomes and is involved in the scavenging of reactive oxygen species. Despite its physiological importance, excess Cu is toxic for plants because of its potential participation in the Fenton reaction [7,8].

Elevated soil Cu concentrations can be toxic to soil organisms (plants, invertebrates, and microorganisms) and disrupt the soil ecosystem’s functioning [6,7]. Cu toxicity to terrestrial organisms strongly depends on its bioavailability in soil and the sensitivity of the organisms [6,7,8]. The cupric ion often attaches to both inorganic and organic ligands in the soil when Cu is introduced into the environment. Cu is able to bind to dissolved organic matter in the pore water of the soil (e.g., humic or fulvic acids). The ion Cu in these organic acids forms stable complexes with the –NH₂ and –SH groups, and to a lesser extent, with the –OH groups. Additionally, cupric ions have the ability to attach to both inorganic and organic components of soils, but with varied degrees of affinity [6]. The following is a general sequence of the maximum adsorption of Cu by various components of soil: manganese oxides > organic matter > iron oxides > clay minerals [6]. However, in most cases, specific Cu adsorption in soil is dominated by organic matter in the soil, and organic matter is primarily responsible for sequestering Cu that has been adsorbed. The ability of Cu²⁺ to bind to inorganic and organic soil colloids is affected by the pH, the oxidation-reduction potential, and the presence of other ions that compete with it [6].

Its origin to soils is mainly due to the intensive and long-term use of copper-based fungicides, which are widely used in agriculture for the control of bacterial and fungal diseases [4,6]. Especially in vineyards following organic cultivation protocols, copper-based products have for decades played a primary role in disease control, with the most characteristic example being their use against downy mildew (Plasmopara viticola) [4]. Due to this fact and to provide a realistic simulation, the soil samples used in this study were taken from areas of vineyards cultivated with organic farming protocols.

Pyrolysis is a method that is widely utilized and effective for the production of biochar. This thermochemical conversion takes place in the presence of only a small amount of oxygen or in its complete absence, which results in the release of volatile gases (oxygen and hydrogen), hydrocarbons (bio-oils), and a biochar that is rich in carbon [9,10]. It is possible to produce biochar from a wide variety of raw materials, some of which are products derived from the industrial sector, bio-materials derived from agricultural and forestry practices, in addition to numerous waste products obtained from human activities, such as sewage-sludge [9,10,11]. As a result of the pyrolysis process, pathogenic bacteria species are eliminated, and the biochar that is produced as a result contains pathogenic free C residues along with other nutrients such as Ca, Mg, P, and K [9,10,11]. Biochars are considered ecologically friendly soil additives that can regulate the bioavailability of nutrients in soils and enhances plant growth and productivity. Depending on the initial biomass used to produce biochar, incorporating these carbonaceous materials into soils can improve water retention, nutrient availability, biological activity, and alkalinity management [9,10]. Biochars showed promising results when tested for their ability to inhibit the movement of PTEs in soil ecosystems [9,11,12]. The application of biochar has played a crucial role in recent years to remediate heavy metals-contaminated soils. Biochar has been used to immobilize heavy metals in contaminated soils by reducing their crop uptake through different mechanisms that can be grouped into four distinct categories electrostatic attraction, ion exchange, complexation, and precipitation. For a more detailed discussion on these mechanisms see Bilias et al. [10].

In this study, two biochars were used that differed in terms of the raw material used for their production, as one came from olive prunings and the other from sewage sludge.

Considering that continuous Cu application to agricultural soils, especially in vineyards, leads to the buildup of element’s concentration in these soils, posing a greater environmental risk, the purpose of the present study was to investigate (i) the distribution of Cu in the various chemical phases over time of four acidic soils, and (ii) whether the addition of olive tree- and sewage sludge-derived biochars to the studied soils affects the distribution of the element in the soils chemical phases. The selection of acidic soils for the performed experiments was initiated by the well-established effect of soil pH on Cu behavior in the soil system that leads to increased Cu mobility at lower pH values.

2. Materials and Methods

2.1. Soil Sampling and Characterization

Four surface soils (S1, S2, S3, and S4) with moderate to slightly acidic pH value were collected from four organically cultivated vineyards in the Peloponnese region, Greece. The selection of these soils was mainly intended to investigate the effect of biochar application to acidic soils on the immobilization and fractionation of Cu derived from copper-based phytosanitary treatments that are widely employed in viticulture. The soils were taken along the vine rows from a 0–20 cm depth using a stainless auger, air-dried and passed through a 2 mm sieve prior to analyze. Soil pH and electrical conductivity (EC) were measured electrometrically at a solid:distilled water suspension with a ratio of 1:1 (w/v) using a pH meter (Selecta 2000, J.P. Selecta S.A., Barcelona, Spain) and a conductivity meter (Selecta 2000, J.P. Selecta S.A., Barcelona, Spain) [13]. Soil organic matter content was estimated by the Walkley–Black wet oxidation with potassium dichromate method as described by Nelson and Sommers [14]. Particle size distribution was performed following the hydrometer method according to Bouyoucos [15]. Determination of cation exchange capacity (CEC) was performed using 1 mol L⁻¹ CH₃COONH₄ (pH 7.0) [13]. Free and amorphous Fe and Mn oxides content were determined after extraction by the dithionate–citrate–bicarbonate (DCB) method, and by the acid ammonium oxalate method in dark conditions, respectively [16,17]. Pseudo-total Cu concentration was determined using an atomic adsorption spectrometer (AA240FS, Varian, Middelburg, The Netherlands) after a microwave (Start D, Milestone, Bergamo, Italy) assisted digestion with aqua regia (HCl:HNO₃ 3:1 v/v) (EPA 3051A) [18].

2.2. Biochars

Sewage sludge (SG) biochar was received from Pyreg GmbH, (Dörth, Germany) and used without further treatment. For its preparation, dewatered sewage sludge (at 75% dry weight) was fed into a Pyreg P500 pilot scale, rotary pyrolysis kiln unit (9000 mm length × 3000 width × 5800 mm height) and pyrolyzed at 600 °C for 20 min. No inert gas was used as flush gas to drive off pyrolytic vapors. The biochar was allowed to cool down for 15 min and was quenched with water. SG complied with the standards for heavy metals as set in Directive 86/278/EEC (regarding the use of sewage sludge in agriculture). The production conditions for SG are described in detail in earlier published work [19]. SG contains 185 g kg⁻¹ organic carbon, 21.3 g kg⁻¹ total nitrogen, 60.5 g P kg⁻¹, and 176.51 mg Cu kg⁻¹, while pH was 7.09. The surface area of SG was 25.6 m²/g.

Olive tree prunings were collected locally (Chania, Crete, Greece), were washed to remove impurities, and air-dried in order to prepare OL biochar through flame curtain pyrolysis, as described below. The design, materials, and dimensions of the pyrolysis kiln used in this study were published in a previous study [20]. The olive tree prunings were randomly divided into three groups, and flame-curtain pyrolysis was performed in triplicate. Each batch was pyrolyzed for 1 h at 600 °C, utilizing thermocouples mounted to the kiln. There was no need for an external heating source, and the process was self-sustaining once the first prunings began to pyrolyze. After being quenched with water, the biochar was air-dried for 96 h before being stored for future use. OL contains 767 g kg⁻¹ organic carbon, 8.1 g kg⁻¹ total nitrogen, 4.4 g P kg⁻¹, and 104 mg Cu kg⁻¹, while pH was 8.28 (CaCl₂). The surface area of OL was 375 m²/g.

2.3. FTIR Spectra

Fourier transformation infrared analysis was performed using a JASCO 4200 FTIR spectrometer (Easton, MD, USA) using KBr pellets in the wavelength of 400–4000 cm⁻¹ to characterize the surface functional groups of the two biochars before and after Cu sorption.

2.4. Experimental Design

This study explores the influence of two types of biochar on Cu fractions in four acidic vineyards soils as a function of incubation time. For the incubation experiment, the four selected soils (S1, S2, S3, and S4) were amended with SG and OL biochar at a rate of 20 % (w/w) including controls (non-amended soils). In detail, soils alone or biochar-amended soils, totaling 1.0 g, were placed in 50 cm³ Teflon centrifuge tubes, and contaminated with 160 mg Cu kg⁻¹ in the form of CuCl₂ solution. The samples were incubated for 1, 3, 7, 15, 36, and 90 days. A quantity of distilled water equals to 30 % of the water holding capacity was added periodically. The tubes were incubated in the laboratory at 22 ± 5 °C. Weigh measurements were used to assess the moisture level on a regular basis, and distilled water was added if needed to maintain it throughout the incubation period. All incubation experiments were carried out in duplicates.

2.5. BCR Sequential Extraction

At the end of each incubation period, a sequential extraction of Cu was applied to the samples using the modified BCR sequential extraction procedure [21], aiming to determine four soil Cu fractions. Definitions of chemical reagents, extraction conditions, and associating fractions are summarized in

Table 1. BCR fractionation scheme: quantity of the reagents, and experimental procedure, based on the use of 1.0 g of air-dried sample.

Step	Fraction	Reagents	Procedure
1	F1 Exchangeable and weak acid-soluble	40 ml 0.11 mol L−1 NaHCO3	16 h shaking 200 rpm T = 22 ± 5 °C
2	F2 Reducible–bound to Fe/Mn oxides	40 ml 0.5 mol L−1 NH2OH.HCl pH = 2.0	16 h shaking 200 rpm T = 22 ± 5 °C
3	F3 Oxidizable—bound to organic matter and sulfides	10 ml 8.8 mol L−1 H2O2 10 ml 8.8 mol L−1 H2O2 50 ml 1 mol L−1 CH3COONH4 pH = 2.0	1 h digestion T = 22 ± 5 °C 1 h digestion T = 85 °C 1 h digestion T = 85 °C 16 h shaking 200 rpm T = 22 ± 5 °C
4	F4 Residual	9 ml conc. HCl + 3 ml conc. HNO3	Digested in a microwave 15 min until T = 200 °C 15 min at 200 °C

. After each extraction step, centrifugation for 10 min at 4000 rpm was conducted in a benchtop centrifuge (K241, Centurion Scientific, Chichester, UK) to separate the extractant from the residue and the supernatant was collected after filtration in a polyethylene container and stored at 4 °C for further analysis. Before and after each extraction step, the centrifuge tube was weighted to correct for the remaining amount of extractant in the residue. All the laboratory glassware and Teflon centrifuge tubes used for the extraction procedure were pre-cleaned with diluted HCl and rinsed thoroughly with deionized water. All the reagents were analytical grade and purchased from Merc Millipore (Darmstadt, Germany). Cu concentration in the extractants was determined using atomic adsorption spectrometer (AA240FS, Varian, Middelburg, Netherlands).

Cu quality control in soils was determined by measuring its content in the certified reference material (ERM-CC141 European Reference Material), which was digested and measured as the soil samples did and the obtained recovery of Cu was 103%. Every 10 samples, a control sample was analyzed, and 30% of the samples were reexamined to test for reproducibility. The correctness of the BCR sequential extraction results was confirmed by recovery rate which was estimated as the sum of the Cu concentration in the four BCR fractions divided by the total Cu concentration (all in units of mg kg⁻¹). For soil–biochar mixtures, the total Cu was the sum of Cu concentrations in the participating soil and biochar at a 0.8:0.2 ratio.

$$R{R}_{Cu}\%=\frac{F1+F2+F3+F4}{Cut}\%,$$

(1)

where $R{R}_{Cu}\%$: recovery rate of Cu (%); F1: Cu concentration in the acid-soluble fraction; F2: Cu concentration in the reducible fraction; F3: Cu concentration in the oxidizable fraction; F4: Cu concentration in the residual fraction; and Cut: pseudo-total Cu concentration.

2.6. Mobility Factor (MF) of Cu

Mobility Factor (MF) that is an index of the strength of metal binding, which is commonly used to assess heavy metal pollution in sediments and soils, was used in this study to estimate Cu availability in the studied soils [22]. MF is expressed as the percentage of Cu concentration (mg kg⁻¹) in the acid-soluble fraction (F1) to the sum of Cu concentrations (mg kg⁻¹) in all studied fractions (F1, F2, F3, and F4) according to the following equation:

$$MF=\frac{F1}{F1+F2+F3+F4}\%,$$

(2)

A proportion of heavy metals in F1 of less than 1% indicates no risk (NR), 1–10% indicates low risk (LR), 10–30% medium risk (MR), 30–50% high risk (HR), and >50% indicates very high risk (VHR) [23,24].

3. Results and Discussion

3.1. Soil Characteristics

Selected physicochemical properties of the studied soils are illustrated in

Table 2. Physicochemical properties of the studied soils.

Soil Sample	S1	S2	S3	S4
Clay (%)	21.1	27.4	16.1	14.2
Silt (%)	28.3	19.0	6.7	38.0
Sand (%)	50.6	53.6	77.2	47.8
Texture	SCL	SCL	LS	L
pH (1:1)	5.01	5.80	6.48	5.80
EC μS cm−1	690	1249	1800	940
Organic Matter (%)	1.48	2.10	2.46	4.68
CEC cmolc kg−1	14.9	20.5	24.3	21.7
Feo (%)	0.76	0.72	0.61	0.54
Fed (%)	0.80	0.88	0.63	1.10
Mno (%)	0.40	0.41	0.25	0.01
Mnd(%)	0.07	0.07	0.05	0.01
Cut mg kg−1	57.80	82.45	46.20	19.65

SCL: Sandy Clay Loam; LS: Loamy Sand; L: Loam; EC: electrical conductivity; CEC: cation exchange capacity; Fed Mnd: free oxides; Feo Mno: amorphous oxides; Cut: pseudo-total Cu concentration.

. Following the pH criterion soils S1, S2, and S4 are moderately acidic since the respective pH values are 5.01, 5.80, and 5.80, whereas soil S3 is slightly acidic with pH 6.48. Distinct differences in terms of other soil properties include texture, organic matter, and Fe and Mn oxides content that may have an impact on soil Cu mobility were observed. The texture of the soil samples varies markedly, and the clay and sand contents range from 14.2 to 21.1% and from 47.8 to 77.2%, respectively. Soil organic matter content is low to high for Mediterranean soils ranging from 1.48 to 4.68%. Cation Exchange Capacity (CEC) was varied from 14.9 to 24.3 cmol_c kg⁻¹ of soil. Soil S4 had the highest percentage of free iron oxides (Fed), while the highest amorphous Fe oxides (Feo) presence was observed in soil S1. Pseudo-total Cu concentration ranged from 19.65 to 82.45 mg kg⁻¹.

3.2. Soil Fractionation

Sequential extraction is widely employed to describe and predict the possible mobility and bioavailability of metals in soil environment [25]. The acid-soluble fraction (F1) is thought to have the higher mobility potential. When the redox potential of the soil environment changes, the reducible and oxidizable fractions (F2 and F3) of metals can change into more accessible forms. The residual fraction (F4) is the most stable form and least bioavailable. The amount of Cu determined in each fraction is expressed as percentage of the pseudo-total Cu amount. Cu recovery for the soils and the biochar-amended soils was in the range of 73.01–116.63%. The obtained results for the recovery support the BCR’s sequential extraction procedure suitability and reliability for identifying Cu speciation in the studied solids. Similar results for Cu recovery have been reported by various researchers as for example by Tytla [26] (41.1–124.4%), Gusiatin et al. [27] (81.2–130.9%), and Tytla et al. [28] (29.7–117.9%).

3.2.1. Control Soils

Figure 1 shows the percentage fractionation values of Cu in the control soils over time. After one day of incubation, the concentrations of Cu fractions follow the order: F1 > F2 > F3 > F4 in S1, F2 > F1 > F3 > F4 in S2, F2 > F4 > F3 > F1 in S3, and F3 > F2 > F1 > F4 in S4.

Acid-soluble is the predominant Cu fraction constituting the 47% of the pseudo-total Cu content in S1, followed by the reducible one which corresponds to the 31% of the pseudo-total. Acid-soluble and reducible fraction account for 78% of the pseudo-total Cu content, while the oxidizable and residual fractions accounts for only 22% of the pseudo-total content. In S2 and S3 high amounts of Cu were extracted from the reducible fraction constituting 39% and 43% of the pseudo-total Cu, respectively. Acid-soluble fraction was a remarkable amount in S2, and accounts for the 30% of the pseudo-total. In S3 acid-soluble fraction accounts for only 10% of the pseudo-total, while oxidizable and residual fraction corresponds to 20% and 27%, respectively. Oxidizable fraction was the predominant in S4, corresponds to 37%, followed by the reducible and acid-soluble constituting 34% and 20% of the pseudo-total amount, respectively.

The distinct variations among the four investigated soils observed on the first day of incubation may be attributable to differences in soil properties. Low pH values increase the mobility of Cu ions in the soil environment, as has been repeatedly reported in the literature [1,6], whereas the crucial role of organic matter in the element’s immobilization in the soil environment is also known [1,6]. Thus, it is possible that in S1, despite the fact that the clay content was quite high, the low pH value and the low percentage of organic matter led to an increased amount of the mobile acid-soluble fraction. Acidic soil conditions favor spiked metals to remain as soluble forms in soil solution or weakly adsorbed on soil particles, and restricts the metal ions to diffuse into the micropores of soil or complex with soil minerals or organic matter [1,6]. The pattern in which soil pH affects the distribution of Cu between the different fractions is consistent and largely explains the results obtained for S3 and S4 soils, where higher pH levels led to more negative Cu adsorption sites on the soil metal oxides [1,6], which made the reducible Cu fraction more common. Moreover, in S4 soil which contains comparatively the highest amount of organic matter, Cu was mainly extracted from oxidizable fraction constituting the 37% of the total Cu amount.

In general, the fractionation pattern showed a constant decline of F1 fraction and an increase in F2 fraction during incubation period without substantial variations among all soil types, indicating the aging process of Cu in the fractionation scheme [6].

3.2.2. SG Biochar-Amended Soils

Figure 1 shows the Cu distribution between chemical fractions (F1, F2, F3, and F4) in the studied soils amended with SG biochar, over incubation time. During the first incubation day a rapid process of metal redistribution among the four fractions occurred. It was obvious that, at the first incubation day, the acid-soluble fraction of Cu reduced in all amended with SG biochar soils compared with the control. This pattern was maintained throughout the incubation period.

More specifically it was observed that at the 1st incubation day acid-soluble Cu fraction in S1 was impressively reduced from 47% of the pseudo-total Cu content in the control to 26% in SG amendment. This was the most intense reduction in acid-soluble fraction among soils. The corresponding decrease in acid-soluble fraction for S2 was from 30% to 21%, for S3 from 10% to 8%, and for S4 from 20% to 17%, respectively. Our findings agreed with Méndez et al. [29] who reported that the biochar derived from sewage sludge significantly decreased the exchangeable fraction of metals in a Mediterranean agricultural soil. The reducible fraction of Cu was dramatically decreased in all amended soils with SG biochar and remained in very low amounts throughout the incubation period accounts only for 5–10% in S1, 6–13% in S2, 6–14% in S3, and 4–7% of the total amount in S4.

In contrast, the oxidizable Cu fraction was noticeably increased during 1–36 incubation days, compared with the controls which varied throughout 1–36 days of incubation from 56 to 59% in S1, 54 to 57% in S2, 53 to 69% in S3, and 62 to 72% in S4. Increasing soil organic matter with SG biochar enhanced the formation of stable Cu complexes with the functional groups of organic matter. Our findings were consistent with Lu et al. [30], who observed that the application of a modified biochar resulted in a significantly higher oxidizable Cu fraction in an acidic soil. On the 90th incubation day oxidizable Cu fraction was decreased and accounts for 37%, 31%, 32%, and 36% for S1, S2, S3, and S4, respectively, indicating that after the 36th incubation day oxidizable Cu was re-distributed to the residual fraction and therefore was increased.

The residual Cu fraction remained at the same levels and was not affected by the addition of SG biochar during 1–36 days of incubation. Instead, it was increased at the 90th incubation day in all studied soils, reaching 44%, 54%, 53%, and 49% of the pseudo-total Cu amount in S1, S2, S3, and S4, respectively.

In a laboratory experiment, Arias-Estévez et al. [31] observed aging-like effects in the distribution fractions of Cu applied to an acidic agricultural soil with increasing incubation duration (>500 days). As reported by many authors, biochar application increased the pH of acidic soils, and presented a mechanism effect of liming when applied to that type of soil [32]. Therefore, the decrease in acid-soluble fraction following the application of SG biochar may be partially ascribed to the increase in the soil pH, thereby enchasing Cu hydrolysis and favoring the specific adsorption and complexation of metal cations by soils constituents [33,34], as well as favoring the heavy metals adsorption on the biochar particles themselves [35], reducing Cu availability both ways. As reported by many authors [11,36,37], biochar application has the ability to increase soil CEC because during the pyrolysis process, the basic cations in biomass (Ca, Mg, K, Na) are transformed into alkaline compounds (for example oxides, hydroxides, or carbonates) [38].

3.2.3. OL Biochar-Amended Soils

Figure 1 shows the effect of OL biochar addition on Cu distribution in soil fractions (F1, F2, F3, and F4), over incubation time. At the first incubation day and throughout the incubation period, the acid-soluble fraction of Cu increased in all amended soils with OL biochar addition compared with the controls. At the first incubation day acid-soluble was the predominant Cu fraction in S1 and S2, reaching 55% and 41% of the pseudo-total Cu amount, respectively. The significant proportion of acid-soluble Cu fraction in OL biochar-amended soils suggested that ion exchange was important process in Cu availability. The incorporation of OL biochar had a great influence on reducible Cu fraction, mainly in S1, S2, and S3 that decreased compared with the controls, from 31% to 11%, from 39% to 5%, and from 43% to 20% at the first incubation day, respectively. In S4 the decrease that was observed was smaller namely from 34% to 21%. When OL biochar was added to the soils, the amount of oxidizable Cu increased noticeably accounts at the first incubation day for 17%, 28%, 25%, and 43% of the pseudo-total Cu amount in S1, S2, S3, and S4, respectively. When OL biochar was added to the soils, the amount of oxidizable Cu increased noticeably during 1–36th incubation day owing to the formation of Cu complexes with organic functional groups on the biochar [39]. The proportion of residual Cu fraction was increased when OL biochar was incorporated into soils varied at the 90^th incubation day from 27% to 45% among soils.

3.3. FTIR Spectra

To better understand the sorption mechanism of Cu, FTIR spectra of the two biochar, before and after contamination with Cu, were obtained. FTIR analysis of SG biochar (

Table 3. FTIR chemical bond vibrations in SG and OL biochar before and after Cu contamination.

Functional Group	SG Biochar	SG Biochar + Cu	OL Biochar	OL Biochar + Cu
O-H stretching	3420 cm−1	3460 cm−1	3421 cm−1	3434 cm−1
C=C stretching (aromatic)	1622 cm−1	1622 cm−1	1558 cm−1	1558 cm−1
-COO- stretching	-	-	1456 cm−1	1417 cm−1
C-O stretching	1052 cm−1	1050 cm−1	1127 cm−1	1108 cm−1
C-H bending (aromatic)			874 cm−1 800 cm−1	874 cm−1 800 cm−1

, Figure 2) revealed a broad absorption band at 3420 cm⁻¹ that can be attributed to O-H stretching of hydroxyl groups. Moving to lower wavenumbers, a strong band is observed at 1622 cm⁻¹ due to stretching vibration of C=C bonds, found in aromatic rings or in newly formed aromatic compounds obtained during pyrolysis, while the broad band at 1052 cm⁻¹ can be assigned to C-O stretching vibrations of C-OH groups or C-O-C ester groups. On the other hand, the FTIR spectra of OL biochar (Figure 3) has distinct differences regarding the number and the intensity of the appearing bands. More specifically, a broad band at 3421 cm⁻¹ is observed, owing to the O-H stretching present in cellulose that remain in the sample even after pyrolysis. Furthermore, the presence of aromatic compounds and carboxylic groups in OL biochar is confirmed via the weak bands at 1558 cm⁻¹ (aromatic C=C stretching), 874–800 cm⁻¹ (aromatic C-H bending), and 1456 cm⁻¹ (-COO- stretching). Similar to SG biochar, at 1127 cm⁻¹ a band appears in the spectra of OL biochar, attributed to C-O stretching vibrations; however, the intensity of the band in OL biochar is much weaker than SG biochar, suggesting the presence of less C-O groups.

The structural differences between the two types of biochar play a crucial role in their ability to act as soil amendment, since the different functional groups present in biochar can greatly affect the sorption mechanism of metal ions. This mechanism lies on the tendency of metal cations, such as Cu to form complexes with oxygen-containing organic functional groups [40,41,42]. When these complexes are formed, the chemical composition of biochar is altered, leading to a shift of the characteristic peaks of the functional groups participating in the complexes in FTIR spectra. Following contamination of SG biochar with Cu, the vibrational band of O-H stretching at 3420 cm⁻¹ and the vibrational band of C-O stretching at 1052 cm⁻¹ shifted at 3460 cm⁻¹ and 1050 cm⁻¹, respectively, suggesting that the hydroxyl groups in SG biochar function as complexing agents for the sorption of Cu. The spectra of OL biochar after enrichment with Cu demonstrated also chemical shifts. The peak attributed to O-H stretching at 3421 cm⁻¹ shifted to 3434 cm⁻¹, whereas the peak attributed to -COO- stretching of carboxylic groups at 1456 cm⁻¹ and the shift ascribed to C-O stretching at 1127 cm⁻¹ shifted to lower wavelengths, 1417 cm⁻¹ and 1108 cm⁻¹, respectively, revealing that both hydroxyl and carboxylic groups of OL biochar contribute to the sorption of Cu. It has been shown that the surface area of biochar plays a significant role in Cu sorption [43]. In our case, SG had a much lower surface area than OL; however, it showed a higher affinity for Cu adsorption. This is further proof of the critical role of the surface functional groups in the process.

3.4. Mobility Factor

Mobility factor (MF) values (Figure 4) were employed to assess the risk of Cu addition into soils with and without biochar application. Among control soils, and throughout the incubation period, MF ranged from 37 to 47%, 21 to 30%, 4 to 10%, and 13 to 20%, for S1, S2, S3, and S4, respectively, indicating that the added Cu may pose a high potential ecological risk in S1 and S2, low in S3 and medium in S4 soils. After SG biochar application to soils, at the 90th incubation day the corresponding decrease was 3.08, 2.62, 1.13, and 1.44 times the controls, resulting in lower ecological risk for all the studied soils. Among the possible Cu sequestration mechanisms, an increase in soil pH may enhance the Cu adsorption on biochar while the increase in CEC due to the oxygen-containing organic functional groups of biochar (-COOH and -OH) may contribute to the decrease in Cu-accessible content via ionic exchange [44]. Moreover, due to the high P content of SG-biochar, possible precipitation [45,46,47,48] and co-precipitation [49] may occur leading to surface sorption processes on biochar for Cu.

On the contrary, when OL was incorporated into soils, at the 1st incubation day, MF was increased 1.17, 1.37, 1.40, and 1.10 times in S1, S2, S3, and S4, respectively, compared with the controls. Even after 90 days of incubation increased MF values were determined (1.03, 1.10, 1.25, and 1.23 times, for S1, S2, S3, and S4, respectively).

4. Conclusions

In this study, the impact of two biochars, derived from sewage sludge (SG) and olive tree prunings (OL), on Cu distribution and redistribution in soil fractions was evaluated in four acidic vineyard soils (control treatment) in relation to incubation time. The main findings, regarding the control soils, were that the soil physicochemical properties (pH, organic matter content, clay content, and the Fe-Mn oxides content) largely determined the behavior of the added Cu over time. The addition of the two biochars to the soils remarkably affected the distribution of Cu between its fractions with different mechanisms. Regarding the presence of SG in the soil, the main conclusion was that, through indirect (alteration of soil properties) and direct (functional groups) mechanisms, it reduced the mobility of Cu remarkably. On the other hand, the presence of OL seemed to increase the Cu mobility, as the available fractions were higher compared with control soils and with amended soils with SG biochar. The current study’s findings support that the addition of biochars from different feedstocks have the potential to alter the geochemical behavior of Cu both immediately and over longer periods. Therefore, the application of biochars in phytoremediation schemes, biofortification techniques, and in management strategies that require either reducing or enhancing the Cu availability and mobility is emerging.