Biosorption of Pb²⁺, Cd²⁺ and Zn²⁺ from aqueous solutions by Agrobacterium tumefaciens S12 isolated from acid mine drainage

Abstract

Heavy metal pollution is a global environmental issue, and microorganisms play a crucial role in the bioremediation of heavy metal-contaminated wastewater. The study isolated heavy metal-resistant bacterium and observed their absorption ability toward Pb²⁺, Cd²⁺ and Zn²⁺. We isolated Agrobacterium tumefaciens S12 from acid mine drainage. The various factors influencing its adsorption performance, including pH, biomass dosage, initial metal ion concentration, and adsorption temperature, were investigated in detail. Chemisorption controls the adsorption rate due to the results better fitted by pseudo-second order kinetics. The maximum adsorption capacities of Pb²⁺, Cd²⁺ and Zn²⁺ on A. tumefaciens S12 were 234, 58 and 51 mg g⁻¹ at 30 °C from Langmuir isotherm, respectively. The adsorption processes for the three heavy metal ions were spontaneous and exothermic in nature. In bimetallic systems, biosorption of Pb²⁺ ions was preferential to that of Cd²⁺ and Zn²⁺. Furthermore, scanning electron microscopy coupled to energy dispersive spectroscopy, Fourier-transform infrared spectra and X-ray photoelectron spectroscopy analysis demonstrated that the adsorption mechanisms include ion-exchange, complexation interaction between the heavy metal ions and the functional groups on the surface of biomass. The obtained results indicated that A. tumefaciens S12 can be applied as an efficient biosorbent in bioremediation technology to sequestrate heavy metal ions from aqueous solution.

1 Introduction

The metal productions are main processes from the industries of mining, mineral processing and extractive metallurgy, and these industrial activities may cause environmental pollution [1, 2]. Acid mine drainage (AMD) comes from non-valuable sulfide minerals reacting with water and oxygen [3]. ADM possesses the features of low pH containing large amount of hazardous heavy metals and metalloids [4,5,6]. Heavy metals tend to accumulate in living organisms as they are not biodegradable that can cause ecological pollution and pose a threat to human health. Various traditional physicochemical methods including chemical precipitation, ion exchange resins, membrane filtration, electrodialysis, coagulation and flocculation have already been exploited to treat wastewater containing heavy metal [7, 8]. Each method has its own merits and drawbacks for removal of heavy metals from wastewater [9,10,11]. For example, chemical precipitation is easy to operate and cost-effective for treatment of heavy metals. Nevertheless, it has the following disadvantages, such as producing a large amount of sludge, increasing cost for sludge disposal, and having low efficiency for low heavy metal concentration [12]. Whilst ion exchange resins own the advantages of high removal efficiency and short processing periods, it is high cost for the synthesis of resin and can cause secondary pollution due to regeneration of the resins. For membrane filtration, it possesses the merits of high selectivity, small area occupancy and low-pressure requirement with these drawbacks containing high operational cost, complicated operation and low flux limiting its application [13]. Electrodialysis with high selectivity requires high cost because of energy consumption and membrane fouling. Although the process of coagulation and flocculation has the ability of removing multiple pollutants, it can generate sludge and is ineffective for heavy metals removal without combination with other treatment techniques.

Biosorption is one method of bioremediation, and it can be defined as the passive uptake of pollutants by using bio-derived products or waste materials as biosorbents to separate heavy metals from aqueous solutions [7]. Bioadsorption has gained enormous attention for its superiority in many aspects. Compared with traditional methods, bioadsorption is easy to conduct, environment-friendly, low cost and low energy-consumption. Additionally, bioadsorption can effectively remove soluble and insoluble pollutants without generating hazardous by-products [14]. Bioadsorption is effective when treating low metal concentration drinking water and can be regenerated for circle reuse [15]. Biosorption is dependent on active substances owing to the ability of metal- binding on surface of biomass [16]. The surface of microbes containing bacteria, fungi and algae has many functional groups, for instance, carboxyl, phosphate, hydroxyl or amine provided by polysaccharide, protein and lipid with the adsorption capability for the removal of heavy metal ions [17,18,19]. Living and dead microbial cells as biological adsorbent have been extensively studied [20]. Living microorganisms can be used to investigate tolerance to heavy metals and adsorption specificity to different heavy metals [21]. Resistance mechanisms of microorganisms to some harmful heavy metals have been developed and therefore these special microorganisms are potential biological materials to remediate wastewater containing heavy metals [22,23,24].

Therefore, finding microorganisms with heavy metal resistance for environmental remediation is a key process for biosorption technology. In addition, for biosorption use in the remediation of heavy metals, the source, safety, cost, and adsorption capacity should be considered for the selection of any suitable biomaterial. Agrobacterium tumefaciens are Gram-negative bacterium, generally exists on the root surface of plants and lives on nutrients infiltrated from root tissues [25]. It shows better biosecurity based on its wide application in plant genetic engineering [26, 27]. Easy and cheap preparation of A. tumefaciens lowers its application cost and makes it competitive among various bio-sorbents. Besides, simultaneous removal of multiple metal ions from complex environments using bioadsorption remains a challenge. Herein, we studied the A. tumefaciens with heavy metal resistant isolated from AMD, then explored in detail its adsorption performance in single and bimetallic systems, isothermal behavior and adsorption kinetics for Pb²⁺, Cd²⁺ and Zn²⁺. Subsequently, the scanning electron microscopy coupled to energy dispersive spectroscopy (SEM), Fourier-transform infrared spectra (FTIR) and X-ray photoelectron spectroscopy (XPS) characterization were carried out to clarify adsorption mechanisms of A. tumefaciens S12.

2 Materials and methods

2.1 Materials

This study used analytic grade chemical reagents to conduct all experiments. These chemicals included beef extract, peptone (Beijing Aoboxing Biotechnology Co., China), yeast extract (Shanghai Swan Beer Co., China), glucose (Tianjin Shengao Chemical Reagent Co., China), NaCl, Pb(NO₃)₂ (Tianjin Fengchuan Chemical Reagent Technology Co., China), Cd(NO₃)₂ (Tianjin Damao Chemical Reagent Factory, China) and Zn(NO₃)₂ (Chengdu Kelong Chemical Reagent factory, China). All reagents were prepared with ultrapure water (GWB-1E/2E water purification system, Beijing General Instrument Co., China).

2.2 Sampling sites and isolation of bacterial strains

The used bacteria in this study were isolated from AMD with pH value of 6.12 in the Yunnan province, China in 2015. In this study, five AMD samples were collected, with the average levels of Pb, Cd, and Zn in these samples being 0.43, 0.015, and 1.12 mg L⁻¹, respectively. An atomic absorption spectrophotometer (AAS, TAS-990AFG, China) was used to analyze the concentration of metal ions in the sample. Water sample was collected and stored in sterile glass bottle. Dilution technique was applied to isolate microorganisms from obtained water samples [21]. Fully mix with vortex 10 mL water sample and sterilized 90 mL water to attain 10^–1 dilute solution. We applied 100 µL of 10^–4–10^–6 dilute solution spreading on Nutrient Agar (NA) and incubated at a constant temperature incubator (MJPS-150, China) for 3–5 d at 30 ℃. After incubation periods, strains were purified through repeated plate streaking. The purified isolates were observed under light microscopy.

2.3 Minimum inhibitory concentrations (MIC) of Pb²⁺, Cd²⁺ and Zn²⁺

MIC represents the minimum heavy metal concentrations at which the strains cannot grow normally. We used Pb(NO₃)₂, Cd(NO₃)₂ and Zn(NO₃)₂ as source of metal ions to conduct our experiments. The heavy metal ions solutions were added to sterilized and cooled NA (about 50 ℃). The number of microbial cells in each culture flask is approximate 1.5 × 10⁵ CFU mL⁻¹ and culture temperature is 30 ℃. After 2 d, aliquots of 100 µL bacterial suspension was spread onto NA and observed visually the growth of strains regularly. The concentration of Pb²⁺, Cd²⁺ and Zn²⁺ in the nutrient broth was gradually increased while keeping other conditions unchanged until no bacteria grew on the NA. Tolerance of isolates to heavy metal was determined by comparing MIC values of isolates obtained from this study and Bacillus subtilis NRRL-B-209 strain [28]. Among the isolated bacteria, three strains, designated as S1, S5, and S12, demonstrated superior resistance to Pb²⁺, Cd²⁺, and Zn²⁺ at a concentration of 250 mg L⁻¹ compared to other strains. Therefore, they were selected for further confirmation through morphological and molecular characterization.

2.4 Identification of bacteria

Gram stain was used for determining G( +) or G(-) of the three strains. 16S rDNA sequencing was utilized to analyze molecular characterization of isolated bacteria. Extracting total DNA with Ezup column bacterial genomic DNA extraction kit (Sangon Biotech, China.) was the first step, and then 16S rDNA was amplified using bacterial universal primers. The amplification was performed in a specific mixture involving 0.5 µL template, 2.5 µL 10 X PCR buffer, 1 µL dNTP (2.5 mM), 0.2 µL Tap polymerase, 0.5 µL F(10 µM), 0.5 µL R(10 uM) and then added distilled water to 25 µL. The procedure of PCR reaction was conducted following these steps: the first step is denaturation for 4 min under 94 ℃, then performing 30 times denaturation (94 ℃, 45 s) followed by annealing (55 ℃, 45 s) and extension (72 ℃, 1 min). After finishing above steps, it needs to repair extension for 10 min and control the temperature at 72 ℃. Finally, the reaction was terminated at 4 ℃. The electrophoresis was used to extract the amplified fragments through 1.0% agarose gels, which was treated for 20 min under 150 V. Molecular weight marker can be completed by using specific DNA ladder ranging from100 bp to 10 kb. Sequencing of obtained gel purified amplicon was completed by Sangon Biotech (Shanghai), China. Obtaining DNA sequences were imported onto the web of National Center for Biotechnology Information to compare with reference strains. The phylogenetic tree of isolated strains was made by Clustal X 1.83 software and MEGA version 7.0.

2.5 Bacterial preparation as a biosorbent

The strain S12 was identified as A. tumefaciens, exhibiting a higher adsorption capacity than the other strains. Therefore, it was selected as biosorbent for removing heavy metals in wastewater. This strain S12 was cultivated in 250 mL volumetric flasks including 100 mL nutrient broth. Biosorbent used subsequently was prepared through the following steps: collecting bacterial cells through centrifugation (10,000 rpm, 4 ℃) by a centrifuge (GL-21 M, China) at late-exponential stage cultivating (28 ± 2 ℃, 150 rpm) in an incubator (SPH-1112D, China), and then washing these cells 3 times using deionized water that had been sterilized. Harvested cells from the last step were stored in refrigerator with 4 ℃ after centrifuging [29].

2.6 Batch biosorption studies

In single systems, various influencing factors were considered, including pH, S12 dosage, concentrations of heavy metal ions added to the reactor, adsorption time and temperatures. A certain amount of heavy metal solution with different concentrations and A. tumefaciens S12 (dry weight) were placed in conical flasks to measure the pH, which were cultivated in a rotary shaker with setting specified time, temperature and rotate speed. The mixtures were obtained by centrifuging about 10 min under the conditions of 10,000 rpm, 4 ℃ after the absorption process reached equilibrium. Then the supernatant was used to analyze residual concentration of studied metal ions through AAS. For this study, each experiment was repeated 3 times, and experimental results were analyzed using the average values.

R (%) and q_e (mg g⁻¹) were used to display the removal efficiency and adsorption capacity of S12 for heavy metal ions, respectively. They were obtained through Eqs. (1) and (2), respectively.

$$R=\frac{{C}_{0}-{C}_{e}}{{C}_{0}}\times 100\%$$

(1)

$${q}_{e}=\frac{\left({C}_{0}-{C}_{e}\right)\times V}{m}$$

(2)

where C₀ and C_e are the initial and equilibrium concentrations of heavy metal ions (mg L⁻¹), respectively. V (L) and m (g) are reaction volume and amount of biosorbent, respectively.

For bimetallic systems (Cd /Pb, Zn/ Pb and Cd/Zn), initial Pb²⁺, Cd²⁺ and Zn²⁺ concentrations were regulated from 0 to 100 mg L⁻¹. Changing another metal ion concentration in above bimetallic systems was done when the concentration of the first metal ion was fixed.

2.7 Characterization analysis

The dried S12 biosorbent (free and metal treated) was obtained about 48 h through vacuum freeze drying oven at -50 ℃. The SEM–EDS (JSM-7500F, Japan) showed the morphology characteristics and elemental distribution of the tested samples at 20 kV acceleration voltage. The FTIR was conducted by a Perkin Elmer spectrometer (Nicolet iS5N, USA), using KBr as one of the materials and controlling the mass ratio of KBr to tested sample as 700/1. XPS was also used to further illustrate the surface element situation of the samples by Kratos Axis Ultra DLD (Escalab 250Xi, USA) and room temperature could meet temperature requirements of this test. Al Kα X-rays (hν = 1486.6 eV) could provide monochromatic beam to obtain ejected photoelectrons, and the binding energy of the elemental peaks obtained from tested samples is based on the C1s peak.

3 Results and discussion

3.1 Isolation and heavy metal resistances of microorganisms

Forty-five bacterial strains were isolated from AMD, and their resistance to Pb²⁺, Cd²⁺ and Zn²⁺ was studied. The results indicated that eight strains exhibited stronger resistance to these heavy metal ions compared to the others (Fig. 1). The MIC of three isolates (designated as S1, S5, and S12) for Pb²⁺, Cd²⁺ and Zn²⁺ exceeded 250 mg L⁻¹, demonstrating higher resistance than other strains. Consequently, these strains were selected for further identification through morphological and molecular characterization.

Fig. 1

The resistance of isolated strains towards Pb²⁺, Cd²⁺ and Zn²⁺

3.2 Phenotypic and genotypic characterizations

Three heavy-metal-resistant S1, S5 and S12 were identified as Klebsiella sp., Bacillus megaterium and A. tumefaciens. The isolated A. tumefaciens exhibited a higher adsorption capacity than strains S1 and S5, making it the primary focus of this study. Gram staining (Fig. 2) revealed that strain S12 is a Gram-negative, rod-shaped bacterium.

Fig. 2

Gram staining of S12 strain

The identity of S12 was further confirmed through 16S rDNA sequencing. A comparison of the 16S rDNA sequence between S12 and known sequences indicated that it belongs to the Agrobacterium or Rhizobium genera, with 100% similarity to A. tumefaciens. Additionally, Fig. 3 presents a phylogenetic tree illustrating the evolutionary relationships between S12 and its closest bacterial relatives. These findings confirm that S12 can be identified as A. tumefaciens. This strain was subsequently used as a biosorbent to remove Pb²⁺, Cd²⁺, and Zn²⁺ from wastewater.

Fig. 3

Phylogenetic tree of S12

3.3 Batch biosorption studies

3.3.1 Effect of pH

Adsorption properties influenced by pH were assessed through experiments conducted at various pH levels. Within the pH range of 2.0 to 5.5, the adsorption capacity of A. tumefaciens S12 for Pb²⁺, Cd²⁺, and Zn²⁺ increased correspondingly, likely due to the deprotonation of functional groups [30]. The experimental results indicated that the maximum adsorption capacities for Pb²⁺, Cd²⁺, and Zn²⁺ were 182, 53, and 31 mg g⁻¹, respectively, at an optimal pH of 5.5 (Fig. 4a). A slight decrease in adsorption capacities was observed when the pH exceeded 5.5, possibly due to the formation of hydroxide precipitates, which could interfere with the adsorption process [31, 32]. Zeta potential measurements were performed to determine the optimal pH range for metal sorption. The Zeta potential represents the point where the sorbent surface charge is neutralized, known as the point of zero charge (pHzpc). When the pH is below the pHzpc, the adsorbent carries a positive charge, leading to electrostatic repulsion with metal ions and inhibiting adsorption. Conversely, at pH levels above the pHzpc, the adsorbent acquires a negative charge, enhancing its adsorption ability. The pHzpc value for A. tumefaciens S12 was estimated to be 3.25, indicating that metal sorption is favored at pH levels higher than 3.25. Consequently, a pH of 5.5 was selected for subsequent experiments, as this value promotes the dissociation of functional groups, enhancing metal ion adsorption while minimizing hydroxide precipitation.

Fig. 4

Effect of pH on the adsorption of Pb²⁺ (a), Cd²⁺ and Zn²⁺ by A. tumefaciens S12 and Zeta potential of A. tumefaciens S12 (b)

3.3.2 Effect of biomass dosage

Figure 5 shows that the concentration of A. tumefaciens S12 had a significant impact on the adsorption of heavy metal ions. Increasing the concentration of A. tumefaciens S12 enhanced the removal of Pb²⁺, Cd²⁺, and Zn²⁺, likely due to the increased availability of functional groups for metal ion binding as the adsorbent dosage increased. The removal efficiency (R value) of Pb²⁺ notably increased from 34 to 73% as the biomass dosage was raised from 0.01 g to 0.02 g (dry weight). Similarly, the removal efficiencies of Cd²⁺ and Zn²⁺ improved from 9% and 1.0% to 51% and 40%, respectively, when the biomass dosage was increased from 0.01 g to 0.05 g. However, when the concentration of A. tumefaciens S12 exceeded certain thresholds (Pb²⁺, 0.02 g; Cd²⁺ and Zn²⁺, 0.05 g), no significant further improvement in the removal rate was observed. Consequently, dosages of 0.02 g, 0.05 g, and 0.05 g of A. tumefaciens S12 were selected for follow-up research on the removal of Pb²⁺, Cd²⁺, and Zn²⁺, respectively.

Fig. 5

Effect of biosorbent dosage on the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ by A. tumefaciens S12

3.3.3 Adsorption kinetics

Figure 6 illustrates the impact of reaction time on the adsorption capacity of A. tumefaciens S12 for Pb²⁺, Cd²⁺, and Zn²⁺. According to the experimental results shown in Fig. 6, the adsorption process can be roughly divided into three stages, reaching equilibrium within 60 min. The rapid uptake phase, designated as the first stage, occurs within the first 30 min, during which the adsorption capacity increases significantly with time. In this initial stage, the adsorbent provides numerous available active sites, and mass transfer resistance is minimal [28]. Additionally, this adsorption process may be spontaneous and does not require energy consumption. The subsequent stage is characterized by slower adsorption. In the later period of adsorption, adsorption capacity appeared slight drop caused by insufficient active sites and difficulty in occupying empty sites. This is caused by repulsive forces between heavy metal ions in the solid and liquid phases [33]. Finally, the active sites become saturated with adsorbed heavy metal ions, and the adsorption process reaches equilibrium at 60 min. To further analyze the adsorption kinetics, two typical models, the pseudo-first-order (Eq. (3)) and pseudo-second-order (Eq. (4)), were applied to the experimental data.

$${q}_{\text{t}}={q}_{e}\left(1-{e}^{-{k}_{1}t}\right)$$

(3)

$${q}_{\text{t}}=\frac{{k}_{2}{q}_{e}^{2}t}{1+{k}_{2}{q}_{e}t}$$

(4)

where qt represents the adsorption amount at time t (min) in mg g⁻¹, and k₁ (min⁻¹) and k₂ (g mg⁻¹ min⁻¹) are the rate constants for the respective models. Figure 6 and Table 1 show the fitting results.

Fig. 6

Effect of contact time on adsorption of Pb²⁺ (a), Cd²⁺ (b) and Zn²⁺ (c) by A. tumefaciens and the fitting results of kinetic models

Table 1 Adsorption kinetic parameters of A. tumefaciens S12 for Pb²⁺, Cd²⁺ and Zn²⁺

The correlation coefficients (r²) in Table 1 fitting from Eq. (4) are better than Eq. (3). Moreover, the q_e values obtained from Eq. (4) at various heavy metal ions concentrations align more closely with the experimental values compared to those from Eq. (3). Luo et al. [34] proposed that the pseudo-second-order model effectively explains dynamic data. This classic model suggests that chemisorption controls the adsorption rate and predicts the entire adsorption process based on the sorption capacity of the sorbent.

3.3.4 Isothermal study

Figure 7a illustrates how the initial concentrations of Pb²⁺, Cd²⁺, and Zn²⁺ affect the adsorption capacity of A. tumefaciens S12. The change of adsorption capacity was consistent with the change of initial Pb²⁺, Cd²⁺ and Zn²⁺ concentration. This can be attributed to the increased collision frequency between A. tumefaciens S12 and the heavy metal ions as the concentration of metal ions increases in the solution. Higher Pb²⁺, Cd²⁺ and Zn²⁺ concentration leads to a corresponding increase in adsorption capacity, as the higher concentration gradient effectively reduces mass transfer resistance, allowing more ions to reach the surface of A. tumefaciens S12 [35]. The q_e remained nearly unchanged when the initial heavy metal ions concentrations exceeded certain value (the initial Pb²⁺, Cd²⁺ and Zn²⁺ concentration were 100, 50 and 50 mg L⁻¹, respectively, in this work) because of the saturation of active sites on A. tumefaciens S12.

Fig. 7

Effect of initial Pb²⁺, Cd²⁺ and Zn²⁺ concentration (a) and isotherms curve (b)

Two representative isotherm models, the Langmuir (Eq. (5)) and Freundlich (Eq. (6)) isotherm models, were applied to further fit the experimental data, as shown in Fig. 7b.

$${q}_{e}=\frac{{\text{q}}_{\textrm{m}}{\text{K}}_{\textrm{L}}{\text{C}}_{\textrm{e}}}{1+{\textrm{K}}_{\textrm{L}}{\textrm{C}}_{\textrm{e}}}$$

(5)

$${q}_{e}={\text{K}}_{F}{C}_{e}^{1/n}$$

(6)

where q_m (mg g⁻¹) and K_L (L mg⁻¹) from Eq. (5) represent the fitted maximum adsorption capacity and the Langmuir model constant, respectively, for the adsorption process of A. tumefaciens S12 for heavy metal ions. K_F [(mg g⁻¹) (L mg⁻¹)^1/n] and 1/n from Eq. (6) describe Freundlich model constant and the empirical parameter. The reaction intensity between heavy metal ions and A. tumefaciens S12 affects K_F, and 1/n can vary according to the adsorption intensity.

The results fitted by these two models and displayed in Fig. 7b and Table 2 illustrate the Eq. (5) provides a better fit than Eq. (6), based on the correlation coefficient (r²) and q_m values. This suggests that the surface of A. tumefaciens S12 is homogenous, and the adsorption process follows monolayer adsorption. The K_L values for the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ were 0.31, 0.26 and 0.14 L mg⁻¹, respectively, indicating that the affinity between heavy metals ions and A. tumefaciens S12 follows the order: Pb²⁺  > Cd²⁺  > Zn²⁺. Additionally, the n values between 1 and 10 suggest that A. tumefaciens S12 is a favorable adsorbent for the removal of Pb²⁺, Cd²⁺ and Zn²⁺.

Table 2 Fitting results of experimental data through two isotherms models

The q_m values for Pb²⁺, Cd²⁺ and Zn²⁺ determined by Langmuir isotherm are compared with those reported in the literature, as shown in Table 3. This comparison indicates that A. tumefaciens S12 has significant potential for adsorbing heavy metal ions, particularly Pb²⁺.

Table 3 Comparison of biosorption capacities of A. tumefaciens S12 for Pb²⁺, Cd²⁺ and Zn²⁺ with other biosorbents

3.3.5 Thermodynamic study

Temperature is an important influencing factor for adsorption process and was investigated. Figure 8a presents the results. As the temperature increased, the q_e values for Pb²⁺ initially rose, reaching a maximum at 30 °C, before subsequently decreasing. The increase in temperature enhances the kinetic energy of metal ions and increases their collision frequency with the adsorbent, leading to higher adsorption capacity [35]. However, excessive temperature (35 ℃ in this study) can damage binding bonds, resulting in a decrease in adsorption capacity [43]. This trend is consistent with findings from other researchers [44]. For Cd²⁺ and Zn²⁺, increasing the temperature improved their adsorption capacities. However, there was little effect on adsorption quantity when the temperature increased from 30 to 35 °C. Consequently, 30 °C is considered the most suitable temperature for this study.

Fig. 8

Effect of temperature (a) and the adsorption thermodynamics curve (b)

Equilibrium data obtained at 20, 25, 30 and 35 ℃ were used to calculate the thermodynamic parameters, including ∆G (kJ mol⁻¹), ∆H (kJ mol⁻¹) and ∆S (J (mol K) ⁻¹). These parameters were determined using the following Eqs. (7), (8), and (9).

$$\Delta G=-NTlnK$$

(7)

$$lnK=\frac{\Delta S}{N}-\frac{\Delta H}{NT}$$

(8)

$$\Delta G=\Delta H-\text{T}\Delta S$$

(9)

In these equations, N [8.314 J (mol K)⁻¹], T (K) and K represent the universal gas constant, the adsorption temperature, and the equilibrium constant at temperature T, respectively. The slope and intercept of the curve between ∆G and T were used to determine the values of ∆S and ∆H, as shown in Fig. 8b and Table 4.

Table 4 Thermodynamic result for the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ by A. tumefaciens S12

The negative ∆G at different experimental temperatures demonstrate that the adsorption process was feasible and spontaneous. The decrease in ∆G with increasing temperature (from 20 to 35 °C) indicates that the adsorption between S12 and heavy metal ions is favored within this temperature range. The positive values of both ∆H and ∆S suggest that the reaction was endothermic and that randomness at the solid/solution interface increased.

3.3.6 The adsorption behavior in bimetallic systems

Figure 9 shows the interaction effects between pairs of metal ions. In the Cd/Pb and Zn/Pb binary systems, the presence of Cd²⁺ (Fig. 9a) and Zn²⁺ (Fig. 9c) had negligible effects on the biosorption of Pb²⁺. The adsorption capacity of Pb²⁺ remained relatively unchanged under varying Pb²⁺ concentrations, even as Cd²⁺ and Zn²⁺ concentrations increased. In contrast, the presence of Pb²⁺ inhibited the ability of A. tumefaciens S12 to adsorb Cd²⁺ (Fig. 9b) and Zn²⁺ (Fig. 9d), with increasing Pb²⁺ concentrations leading to a noticeable decrease in the uptake capacity of Cd²⁺ and Zn²⁺. These findings are consistent with previous research. In the Zn/Cd binary system (Fig. 9e and f), the coexistence of one metal in the solution consistently reduced the removal of the other, indicating that competitive bioadsorption occurred between Cd²⁺ and Zn²⁺ for the available binding sites.

Fig. 9

Adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ by A. tumefaciens S12 in binary metal system ((a) and (b) are Cd/Pb binary system, (c) and (d) are Zn/Pb binary system, (e) and (f) are Zn/Cd binary system)

This adsorption behavior may be influenced by the different electronic configurations of the metal elements ([Xe] 4f¹⁴5d¹⁰6s²6p² for Pb, [Kr] 4d¹⁰5s² for Cd, [Ar] 3d¹⁰4s² for Zn) [45]. Metals with more orbitals and outer orbital electrons exhibit stronger affinities with functional groups. Additionally, differences in electronegativity and ionic radius contribute to these results. The electronegativity of Pb²⁺, Cd²⁺ and Zn²⁺ are 2.33, 1.39 and 1.65, respectively, and their ion radii are 1.32, 0.97 and 0.74 Å, respectively [46,47,48]. Metal ions with higher electronegativity and larger ionic radii are more easily adsorbed by adsorbents.

In bimetallic systems, Pb²⁺ showed a greater affinity for A. tumefaciens S12 than Cd²⁺ and Zn²⁺, based on the physical properties of the metals. This observation is consistent with the constants in Table 2 fitted by the Langmuir model. The above explanations demonstrate that the biosorption of Pb²⁺ ions is preferred over Cd²⁺ and Zn²⁺ in binary systems. For the Cd/Zn binary system, the competitive bioadsorption may be attributed to their similar affinities for the binding sites and the limited number of adsorption sites on the adsorbent.

3.4 Adsorption mechanism

3.4.1 SEM and EDS spectroscopy

The morphological changes in biomass before and after the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ can be observed in Fig. 10. The pristine A. tumefaciens S12 cells were rod-shaped, with smooth and regular surfaces (Fig. 10a). However, after exposure to heavy metal ions, some A. tumefaciens S12 cells showed signs of structural damage, which may be attributed to the restriction of microbial metabolic activity due to nutrient deficiency and the toxicity of heavy metals. Similar observations were reported by Huang et al. [49]. Additionally, the aggregation of A. tumefaciens S12 cells became more pronounced, and their surfaces appeared rougher after the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ (Fig. 10c, e and g). This roughness likely resulted from mechanical force damage and interactions between surfactants and metal ions [50].

Fig. 10

SEM and EDS results of before (a and b) and after the adsorption of Pb²⁺ (c and d), Cd²⁺ (e and f) and Zn²⁺ (g and h) by A. tumefaciens S12

The EDS analysis revealed significant changes in the elemental composition of A. tumefaciens S12 after the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ compared to the composition before adsorption (Fig. 10b, d, f and h). Distinct peaks corresponding to Pb, Cd, and Zn were observed in the post-adsorption samples, confirming that A. tumefaciens S12 successfully absorbed these metal ions. Besides, the disappearance of the Na peak after adsorption suggests that Na⁺ was released from the cells during the process, indicating that ion exchange was involved in the adsorption mechanism [51].

3.4.2 FTIR analysis

FTIR analysis can be used to identify the functional groups present on the surface of A. tumefaciens S12 cells. As shown in Fig. 11a, the surface composition of A. tumefaciens S12 is complex, with various functional groups identified through the FTIR spectra. The broad band observed at 3600–3200 cm⁻¹ corresponds to overlapping peaks from hydroxyl (-OH) and amine (-NH) groups. In addition, the peaks at 3000–2800 cm⁻¹ are attributed to C-H stretching vibrations in alkyl chains [52]. Strong bands at 1655, 1547 and 1239 cm⁻¹ are associated with amide bond formation [53, 54]. Similarly, absorption bands around 1453 and 1384 cm⁻¹ have been identified as the C-O stretching mode of -COOH groups and C-H stretching vibrations, respectively [49]. Furthermore, the peaks at 1083 and 1049 cm⁻¹ indicate the presence of organic phosphate groups [55]. The peaks between 701 and 881 cm⁻¹ are mainly due to P-O-C and P-O-P stretching, originating from phospholipids and ribose phosphate chain pyrophosphate [56]. This analysis reveals that hydroxyl, carboxyl, amino, and phosphate groups are present in A. tumefaciens S12, with the primary components being polysaccharides, proteins, and phospholipids.

Fig. 11

FTIR spectra of before and after adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ by A. tumefaciens S12 (a: before, b: Pb²⁺,c: Cd²⁺, d: Zn²⁺)

The shifts in peak positions and changes in intensity observed in the FTIR spectra of A. tumefaciens S12 with and without heavy metals (Fig. 11b–d) suggest that the functional groups represented by these peaks are involved in the adsorption process. After the adsorption of Pb²⁺, Cd²⁺ and Zn²⁺, the peak at 3297 cm⁻¹ shifted to 3284, 3295 and 3295 cm⁻¹, respectively, with a corresponding reduction in peak intensity. Similarly, the intensity of the peaks at 2970 and 2927 cm⁻¹ decreased after adsorption. Furthermore, both the intensity and position of the adsorption peaks for amide groups changed. Moreover, the peaks at 1083 and 1049 cm⁻¹ observed for pristine A. tumefaciens S12 cells merged into a single peak and shifted to 1071, 1060 and 1069 cm⁻¹ after adsorption. These changes conclusively indicate that multiple functional groups, including hydroxyl, amino, carboxyl, and phosphate, participate in the adsorption process and play a dominant role in the removal of metal ions.

3.4.3 XPS analysis

XPS is a powerful tool for elucidating the adsorption mechanism of A. tumefaciens S12 with heavy metal ions, as it provides detailed information on the chemical composition of the samples and chemical-state changes in the elements under study. By comparing the wide-scan results of A. tumefaciens S12 before and after adsorption (Fig. 12a), it is clear that Pb 4f, Cd 3d and Zn 2p peaks appear distinctly in the XPS spectra, confirming that A. tumefaciens S12 successfully captured these three ions. The spectrum analysis of Pb²⁺-bound A. tumefaciens S12 (Fig. 12b) reveals a characteristic binding energy for Pb²⁺ at 138.3 eV, corresponding to the Pb 4f7/2 peak. Additionally, the presence of a Pb 4f5/2 peak at 143.2 eV suggests that Pb²⁺ was chemisorbed to functional groups on the cell surface [35]. Figure 12c shows that the Cd peak is split into two components, 3d3/2 (411.8 eV) and Cd3d5/2 (405.1 eV). Figure 12d presents the Zn 2p photoelectron spectrum, featuring Zn 2p3/2 and 2p1/2 peaks at 1021.9 and 1044.9 eV, respectively, consistent with literature values [57].

Fig. 12

XPS spectra of A. tumefaciens S12 before and after adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ (a: broad scan, b: Pb 4f, c: Cd, 3d, d: Zn 2p,e: O 1 s, f: N 1 s, g: P 2p)

Interestingly, the wide-scan results also reveal a prominent Na 1 s peak in the pristine A. tumefaciens S12, which disappears after adsorption of Pb²⁺, Cd²⁺, and Zn²⁺. This suggests that Na⁺ was displaced during the adsorption process, as Pb²⁺, Cd²⁺, and Zn²⁺ have a stronger affinity for the functional groups on the biosorbent. This displacement hints at an ion-exchange mechanism being involved in the adsorption process.

Table 5 highlights a noticeable decrease in the proportion of N, O, and P elements in A. tumefaciens S12 after adsorption of the heavy metal ions. Figure 12 and Table 6 further show the high-resolution XPS spectra for O 1 s, N 1 s, and P 2p.

Table 5 The composition on the surface of A. tumefaciens S12 from XPS spectra

Table 6 The XPS spectra carves fitting results of A. tumefaciens S12 before and after adsorption of Pb²⁺, Cd²⁺ and Zn²⁺

Figure 12e shows that the O 1 s peaks in A. tumefaciens S12, both before and after metal ion adsorption, consist of C = O/C–OH and C-O groups. Prior to adsorption, the O 1 s spectrum reveals peaks for C = O/C–OH and C-O at 531.2 and 532.7 eV, respectively [58]. Post-adsorption of Pb²⁺, Cd²⁺, and Zn²⁺, the area of the C = O/C–OH peak decreases from 55% to 42, 37, and 38%, respectively, indicating that hydroxyl and carboxyl groups played a significant role in the adsorption process, aligning with FTIR results. Moreover, the binding energies of C = O/C–OH groups shift to 531.5 eV, and the C-O binding energy changes to 532.8 eV after metal ion binding. This suggests that these functional groups were actively involved in the adsorption process, likely through charge transfer interactions with Pb²⁺, Cd²⁺, and Zn²⁺.

Figure 12f presents the fitted N 1 s spectra. Before the adsorption of Pb²⁺, Cd²⁺, and Zn²⁺, a single N 1 s peak at 400.0 eV, corresponding to -NH₂ or –NH groups from proteins, is observed. After Pb²⁺ adsorption, the complexation reaction between -NH₂ or –NH groups and Pb²⁺ results in the protonation of amines, as evidenced by the emergence of a new weak N 1 s peak at approximately 401.4 eV. The reaction is illustrated by the following equations:

$$\mathrm{R}-{\text{NH}}_2\;+\;\text{Pb}^{2+}\;\rightarrow\;\text{R}-{\text{NH}}_2 \text{Pb}^{2+}$$

(10)

$$\mathrm{R}-\text{NH}\;+\;\text{Pb}^{2+}\rightarrow\;\text{R}-\text{NHPb}^{2+}$$

(11)

Additionally, the binding energy of -NH₂ or -NH groups increases to 400.1 eV due to Pb²⁺ sharing lone pair electrons provided by N atoms, which diminishes electron cloud density [59]. For Cd²⁺ and Zn²⁺ adsorption, the binding energies of -NH or -NH₂ groups decrease to 399.9 eV, indicating coordination bonds formed between N atoms and the metal ions. In these bonds, Cd²⁺ and Zn²⁺ share electrons with N atoms, resulting in a decrease in binding energy due to the increased electron cloud density around the N atoms.

Although the atomic contents (%) of P is lower compared to O and N, the change in P content after adsorption is significant, as shown in Table 5. The P2p spectrum of A. tumefaciens 12 (Fig. 12g) features two peaks assigned to P-O (133.1 eV) and P = O (133.9 eV) [60, 61]. This confirms that phospholipids, containing phosphate groups with P = O and P-O functionalities, are part of the surface composition of A. tumefaciens S12. The shifts in their binding energies after adsorption suggest that these phosphate groups were involved in the adsorption process through complexation and electrostatic attraction.

In summary, based on the XPS spectra analysis, A. tumefaciens S12 proves to be an excellent adsorbent for Pb²⁺, Cd²⁺, and Zn²⁺ because of the abundance of functional groups containing N, O, and P. These groups facilitate the formation of complexes with the heavy metal ions, with mechanisms such as ion exchange, complexation, and electrostatic attraction likely playing a dominant role in the adsorption process.

4 Conclusion and future prospect

A.tumefaciens 12 with heavy metal tolerance isolated from AMD has demonstrated exceptional tolerance to heavy metals, specifically Pb²⁺, Cd²⁺, and Zn²⁺. This bacterium not only exhibits high resistance to these metals with MIC of 400 mg L⁻¹ for Pb²⁺, 300 mg L⁻¹ for Cd²⁺, and 250 mg L⁻¹ for Zn²⁺, but also shows effective adsorption capabilities. The adsorption was obviously impacted by pH, biomass dosage, initial Pb²⁺, Cd²⁺ and Zn²⁺ concentration and adsorption temperature. The adsorption process is best described by a pseudo-second-order chemisorption model, and the Langmuir isotherm effectively fits the adsorption data for Pb²⁺, Cd²⁺ and Zn²⁺ with maximum adsorption capacities of 234, 58 and 51 mg g⁻¹, respectively. The adsorption of Pb²⁺, Cd²⁺ and Zn²⁺ by A. tumefaciens 12 was spontaneous and endothermic in nature through thermodynamic study. In bimetallic systems, A. tumefaciens 12 showed a preference for adsorbing Pb²⁺ over Cd²⁺ and Zn²⁺, with competition observed between Cd²⁺ and Zn²⁺ for binding sites. The high adsorption efficiency of A. tumefaciens 12 is attributed to the presence of functional groups such as hydroxyl, carboxyl, amino, and phosphate on its surface, which facilitate ion-exchange and complexation with the metal ions. This makes A. tumefaciens 12 an excellent candidate for bioremediation, offering an economical and environmentally friendly method for the removal of heavy metals from wastewater.

To enhance the applicability and feasibility of A. tumefaciens 12, further research is necessary. Conducting pilot-scale experiments using actual wastewater, such as mining effluents containing heavy metals, would help explore its practical application conditions and feasibility. Additionally, several limitations need to be addressed. For instance, it is crucial to develop effective methods for separating metal-laden cells from water bodies. Furthermore, lacking of specifificity for heavy metals, simple biosorbents could be hard to recycle the desired metals.

A. tumefaciens 12 is a novel and promising bacterium for the removal of heavy metals. Further research could provide a more comprehensive understanding of its metal absorption capabilities. Firstly, exploring the bacterium’s ability to absorb additional heavy metals, such as Cu²⁺, Ni²⁺, and Cr⁶⁺, would broaden its potential applications. Delving deeper into the specific mechanisms involved in each process could also offer valuable insights into how the bacterium interacts with heavy metal ions, enabling the optimization of these processes for more efficient biosorption. Additionally, evaluating the long-term effects of biosorption is essential. This includes assessing the potential risks associated with releasing genetically modified bacteria into the environment and understanding the impact of repeated heavy metal exposure on bacterial viability and performance over time. These areas of research are crucial for fully realizing the potential of A. tumefaciens 12 in practical applications.