Characterization of PM₁₀-bound metal speciation, bioavailability, and plant responses in Andrographis paniculata at a traffic junction

Abstract

Understanding the intricate fate of atmospheric particulate matter (PM₁₀)-associated heavy metals within medicinal plants is crucial for environmental and public health. This study rigorously investigated the speciation, bioavailability, and cellular-physiological impacts of PM₁₀-borne lead (Pb), cadmium (Cd), and arsenic (As) on Andrographis paniculata, tracing their dynamics across soil and plant tissues. The experiment was conducted near a busy traffic junction, comparing uncontaminated (T1) and metal-contaminated (T2) soils. PM₁₀-Pb was predominantly (76%) in the residual fraction, whereas Cd was largely (66.5%) in the oxidizable fraction. As concentrations were too low for detailed PM₁₀ fractionation. In A. paniculata, Pb exhibited higher accumulation but lower mobility than Cd. PM deposition induced significant leaf surface roughening, aggregate formation at trichomes, and altered stomatal morphology. The multivariate analysis delineated the dynamic relationship between PM₁₀-associated metal speciation and their subsequent subcellular sequestration. Also, the distinct metal sequestration profiles between T1 and T2 treatments were discriminated. Analysis results indicated direct interaction of water-soluble PM₁₀ metal forms with cellular components. In contrast, oxidizable and residual forms demonstrated intracellular biotransformation into more bioavailable states. In T1, a higher proportion of bioavailable metal forms (EtOH and H₂O-extractable) was observed, correlating with enhanced organelle accumulation, indicative of a comparatively less robust detoxification capacity. Conversely, T2 shifted to less bioavailable forms (HAc, HCl, and NaCl extractable), reducing organelle uptake. While antioxidant enzyme activity increased in both, growth and andrographolide content decreased. This research provides novel insights into plant adaptive strategies against atmospheric metal stress under varying soil conditions.

Supplementary Information

The online version contains supplementary material available at 10.1038/s41598-025-24536-0.

Introduction

Atmospheric particulate matter (PM) is a crucial vector for the transport and dispersion of potentially toxic metals (PTMs). The severity of these impacts is intrinsically linked to the chemical and physical characteristics of the metals, particularly their bioavailability and toxicity upon deposition or inhalation. The chemical fractionation of toxic metals suggested higher bioavailability and accessibility of water-soluble and exchangeable fractions of PM in plants via leaf surfaces or indirectly through root uptake from the soil^1,2. A substantial contribution of atmospheric deposition to metal accumulation in crops was demonstrated in different studies. The atmospheric deposition of copper (Cu), zinc (Zn), and lead (Pb) contributed 44–47%, 23–25%, and 35–37% to their accumulation in rice grains, respectively³. The deposition of Cd and Pb accounted for 18%-63.5% and 27.6–41.1% of their total accumulation in rice grains, respectively⁴. Significant accumulation of Cd and increased translocation from roots to aerial parts were also observed in Cyperus rotundus growing near highway⁵. It was observed that newly deposited Cu and Pb represented a small fraction (0.2%-14%) of their total soil pool. Still, it contributed a substantial amount (8%-77% for Cu and 14%-84% for Pb) of their accumulation in soybean plants⁶. Therefore, the contribution of atmospheric toxins is important factor for toxic metal contamination in edible crops.

The deposited atmospheric particles have negative impacts on plant physiological processes due to the presence of toxic compounds. The chemical forms and subcellular localization of PTMs in atmospheric deposition are the controlling factors for plant toxicity⁷. A significant reduction in biomass, surface area of trifoliate leaves (34%, 23%, and 37%, respectively), reducing sugar (71%), and total sugar content (63%) was observed in Vigna radiata after PM exposure⁸. Likewise, a significant reduction in biomass, protein, and chlorophyll content and an enhancement in antioxidant enzyme activity in A. paniculata were observed due to deposition of atmospheric Pb and Cd ⁹. The atmospheric dust deposition on four common medicinal plants, Ocimum sanctum, Andrographis paniculata, Catharanthus roseous, and Kaempferia galanga, affects their growth, levels of essential biochemical constituents, and heavy metal concentration¹⁰. The reduction in fresh and dry weights, as well as the activities of photosynthetic enzymes in wheat seedlings, was observed by atmospheric deposition carried by particles from the coal mine area¹¹. The atmospheric deposition also altered the uptake velocities of Pb, Cd, and As from soil and their subcellular distribution in plants⁹. Despite evidence of atmospheric metal toxicity in plants, current research offers limited insight into the mechanisms governing their subcellular distribution and chemical speciation within plant tissues.

This study, therefore, aims to elucidate the intricate interrelationship between the chemical fractions of toxic metals in PM₁₀ and their speciation and sequestration within plant tissues. Our approach involved comparing metal distribution in plants cultivated under two distinct conditions: uncontaminated soil (representing atmospheric uptake only) and metal-contaminated soil (representing combined atmospheric deposition and root uptake). The study employed various multivariate analysis techniques (correlation, cluster analysis, principal component analysis, and correspondence analysis) to establish the interrelationships and discriminate the mechanisms of subcellular metal distribution under two distinct environmental conditions. The deposition and translocation of Pb, Cd, and As in Andrographis paniculata (Burm. f.) Nees, a medicinal plant, was used for initial biomonitoring in a polluted area, providing information about the interrelationship between the chemical fractions of toxic metals in PM_10, their speciation, and sequestration within plant tissues. Also, their physiological and biochemical status in plants. This plant is widely employed in traditional medicine systems for various ailments, including skin infections, stomach pain, inflammation, fever, influenza, dyspepsia, diabetes, dysentery, malaria, cancer, and respiratory diseases. Its dried leaves are also key components in herbal materials and extracted herbal products. This study specifically focuses on the deposition and translocation of As, Pb, and Cd in plants because they are classified as the first, second, and eighth of the 20 priority hazardous substances, respectively. In addition, pharmacopeia guidelines in the U.S., India, and other countries report a limit of three PTMs for medicinal herbs¹². Despite the recognized threat of atmospheric heavy metals, the mechanistic link between PM-associated Pb, Cd, and As’s chemical fractionation and subsequent chemical speciation and subcellular distribution within a medicinally important species like Andrographis paniculata remains unreported. This study fills this knowledge gap and provides a mechanistic insight into the subject.

Materials and methods

Study area

The site was situated at a busy traffic junction on the Lucknow-Sitapur Highway (NH-24) (26.8° N, 80.9° E), an area heavily influenced by local traffic (Fig. S1). The site was characterized by a crowded environment with food stalls, wastewater drainage, and ongoing bridge construction activities during the sampling period. Also, this site was mutually affected by the nearby mixed sources of agricultural and industrial areas. An agricultural farm of CSIR-Central Institute of Medicinal and Aromatic Plants (CIMAP) (26.5° N, 80.5° E) with the dominant agricultural activities situated about 2 km from the traffic site. The Kukrail picnic spot forest, spanning over 5000 acres, is located northeast, about 5 km from the sampling site. Additionally, typical industries are present in the vicinity, such as cable manufacturing, rubber and plastic products manufacturers, plywood factories, metal and sewing companies, and small-scale units like battery, plastics, paint, detergent, and soap factories, which cause potential polluting activities. The average vehicular density was 4,330 ± 90 vehicular density/hour during the morning, 2,292 ± 55 vehicular density/hour during the afternoon, and 5165 ± 96 vehicular density/hour during the evening period. The ambient temperature varied from 21 °C to 33 °C at the sampling site during the study period. The predominant wind direction was north to south, nearly 46.42% during the study period, and the wind speed was 1 to 2 m/s. The relative humidity varied from 74% to 94%.

Materials

The uncontaminated soil was taken from the experimental farm of the CSIR-CIMAP at a depth of 0–15 cm. The Pb, Cd, and As content were below the detection limit. Heavy metals in natural soils in UP (Uttar Pradesh, India) are often below detection limits or at minimal concentrations. This soil was taken as a control. The metal-contaminated soil collected from the traffic junction had Pb: 32.09 ± 0.8 µg g^− 1, Cd: 1.91 ± 0.06 µg g^− 1, and As: 0.82 ± 0.08 µg g^− 1. The seeds of A. paniculata were taken from the National Gene Bank for Medicinal and Aromatic Plants (NGBMAP) of CSIR-CIMAP, having Accession number- CIMAP-948.

Experiment design

The experiment was performed using a random block design (RBD). The experiments were set up at two sites: (1) A traffic junction to examine the effect of airborne metal (2) Control the site in a closed environment to avoid any deposition of airborne particles. The 5 kg earthen pots were filled with uncontaminated and metal-contaminated soils. One-month-old plantlets were transferred to pots, allowing for acclimatization before placing them in different environments. At the traffic junction, six pots containing uncontaminated (T1) and metal-contaminated (T2) soils were positioned 30 cm apart in a field of 2 m × 3 m × 2 m. For the closed chamber study, three translucent cubes of 2 m ×3 m ×2 m covered with transparent plastic sheets were used to avoid atmospheric heavy metal deposition. The plastic shed environment had reduced lighting by 34% and increased humidity compared to outside. This study used a closed-control system, eliminating the need to consider environmental factors due to the absence of PTMs contamination. One pot with uncontaminated soil was placed in each cube. The closed chamber treatment was taken as a control (C), indicating conditions where neither atmospheric nor soil pollution contributes. Consequently, the plant exposure limited to atmospheric deposition is designated as T1. In contrast, plant exposure encompassing atmospheric deposition and metal-contaminated soil is termed T2. Plants were irrigated regularly with deionized water. The atmospheric dust deposition on the surface of plant leaves was taken at different time scales. Plants were harvested after 90 days of transplantation. Plant tissues were carefully cut at harvest using scissors to separate the leaves, stems, and roots. The tissues were washed with distilled water to remove dust from the leaves or stems and soil particles from the roots. The biomass and root length of the harvested plant were measured. The plants were blotted with paper towels to remove excess water and immediately weighed to determine fresh weight. After that, each part was placed in plastic bags (HiMedia) for further analysis. Leaf samples and rhizospheric soil were collected weekly for four sampling events over the 30 days leading up to harvest, allowing for subcellular fractionation and analysis of chemical forms (only for leaves). Leaf samples were divided into two parts; one was shade-dried for metal analysis. The other part was transferred to an ice box, and then samples were frozen immediately in liquid nitrogen and stored at -20 °C to prevent tissue degradation. This part is used for subcellular distribution and chemical forms of metals. The rhizospheric soil was collected and placed in plastic zip-lock bags according to treatments. All soil samples were transported to the laboratory for metal analysis and BCR (European Community Bureau of Reference) metal fractionation. The pH analysis of control and treatment soil was conducted using the Mettler Seven Go DuoTM pH/Conductivity meter SG23. The total organic carbon (OC) was analysed by dichromate oxidation and titration with ferrous ammonium sulfate.

Collection of PM₁₀ and leaf dust deposit

The PM₁₀ samples were collected at a traffic junction (N = 12) during the study period. A dual-channel ambient fine dust sampler model PEM-DCAFDS 2.5/10 (Polltech Instruments Pvt. Ltd., Mumbai, India) was used at a 15 LPM (liters per minute) flow rate with flow measure accuracy of ± 2%. During the sampling period, an inbuilt gas flow calibrator ensured consistent flow during sampling. Glass microfiber filters (Whatman, 46.2 mm diameter) were used for sample collection. Before sampling, filters were pre-heated for 4 to 6 h at 60 °C and kept in a desiccator. Filters were handled with tweezers to reduce contamination. Before sampling, blank and sample filters were weighed in a gravimetric cabin at 22 ± 1 °C temperature and 45 ± 5% relative humidity. A calibrated analytical balance (Mettler-Toledo, Switzerland) with one µg sensitivity was used for triple weighing of each filter. The sampling was conducted for 8 h. After sample collection, the filters were desiccated to eliminate moisture content introduced during the sampling and transportation process and then weighed to obtain PM mass. Filter blanks were also collected to assess potential contamination by passing air at 15 LPM for 30 s. During the measurement period, contamination was not detected in the sample handling and passive sedimentation of particles. The PM₁₀ samples were sealed in a petri dish and stored at 4 °C for metal and BCR fractionation analysis.

For the collection of dust deposition on the leaves of plants (N = 3 from each plant), the method of Gupta et al. (2004) was used with some modifications. The randomly selected mature leaves (N = 03) were plucked and weighed for dry deposition on the leaf surface. Then, the leaf was placed in a 50 mL centrifuge tube to recover foliar dust and shaken thoroughly for five minutes with 30 mL Milli Q water. After that, this dusty water was filtered using a syringe filter (0.22 μm, PVDF) and digested with HNO₃ before metal analysis. After washing, the leaves were wiped clean, and weight and surface area were measured using a leaf area meter (CI 202, CID Inc., USA). The difference in leaf weight before and after washing was recorded as the dry dust deposition on the leaf surface.

BCR fractionation of PM₁₀ filter and soil

A three-step sequential method (BCR) was used to determine the fractionation of metals in soil and PM₁₀ ¹³. Detailed procedures are given in Table S1. For this, 1 g of soil (< 2 mm particle size) samples were divided into four fractions: highly exchangeable or water-soluble (SF1), reducible (SF2), oxidizable (SF3), and residual (SF4) fractions¹⁴. Likewise, highly exchangeable or water-soluble (FF1), reducible (FF2), oxidizable (FF3), and residual (FF4) metal fractionations in PM₁₀ were obtained. The highly exchangeable and reducible forms are more bioavailable than the oxidizable and residual forms, which are more stable and exhibit lower bioavailability.

Subcellular distribution of metals in plant tissues

For determining metal distribution in the subcellular part of plants, the fresh leaves (500 mg) of each treatment were ground in prechilled 10 mL of Tris-HCl buffer solution and centrifuged (20 min at 300 rpm). The precipitate was taken as a cell wall fraction. Further, the supernatant was centrifuged for 15 min at 2,000 rpm, and the subsequent precipitate obtained was referred to as the chloroplast and nucleus fraction. Then, the third centrifugation of supernatant at 10,000 rpm for 20 min, and the resultant precipitate was designated the mitochondrial fraction. Further, the remaining supernatant solution was termed a soluble fraction, primarily consisting of cytosol, cell sap, inclusions, and vacuoles. All the experiments were performed in triplicate at 4 °C¹⁵.

Chemical forms of metals in plant tissues

The chemical forms of metals in plant tissues were obtained using a sequential extraction in the following order: (1) inorganic forms priority to nitrate/nitrite, chloride, and aminophenol (EtOH-extractable) were extracted with 80% ethanol, (2) water-soluble organic acid complexes (H₂O-extractable) were extracted with deionized water, (3) pectates and protein-integrated forms (NaCl-extractable) were extracted with 1 mol L^{− 1} sodium chloride, (4) insoluble phosphate complexes (HAc-extractable) were extracted with 2% acetic acid (5) oxalate-complexes (HCl-extractable) were extracted with 0.6 mol L^{− 1} hydrochloric acid, and (6) metals residual forms (RD)¹⁶. A 500 mg sample of fresh leaf material was homogenized using a mortar and pestle in 20 mL of prechilled extraction solution. The homogenate was shaken at 25 °C for 24 h, then centrifuged at 5000 rpm for 10 min. The first supernatant was collected in a conical beaker. The remaining pellet was resuspended in 10 mL of extraction solution, shaken for 1 h at 25 °C, and centrifuged at 5000 rpm for 10 min. The supernatant was then transferred to a conical flask, and the extraction process was repeated by adding another 10 mL of extraction solvent. After the third centrifugation, the three supernatants were combined. The pellets were further treated with successive solvent extractions, following the same process: resuspension, shaking, centrifugation, and supernatant collection for each subsequent solvent. After extracting all five solvents, the supernatants and the residual pellet were dried to a constant weight in an oven at 105 °C.

Analysis of plant parameters

The morphological analysis of leaves was done by scanning electron microscopy (SEM) [Model- FEI QUANTA™ 250, USA at 15 keV]. Using commercial razor blades, the transverse section was performed on the midrib of young leaves at the leaf base by freehand for anatomical studies. All sections were rinsed in distilled water. The sections were stained with safranin with 1% solution in 50% alcohol. The transverse cross sections were stained with methylene blue and safranin. The sections were mounted under coverslips with glycerol (50%) and subsequently analyzed and photomicrographed using a Qwin system and video camera (Leica ICC50 HD) coupled to an optical microscope (Leica DM 750) for capturing images. The protein and total chlorophyll in the leaves of A. paniculata were analyzed using previously reported standard methods¹⁷. For enzyme analysis, the fresh leaves were crushed in liquid nitrogen at 4 °C and then homogenized in pre-cooled phosphate buffer (pH = 7.0). This extract was used for glutathione reductase (GR), superoxide dismutase (SOD), and catalase (CAT) analysis using previously standardized protocols^18–20. High-performance liquid chromatography (HPLC) (Shimadzu model SPD-M20A, Japan) was employed to quantify the metabolites in the plant extract as per the previously reported method²¹.

Metal and elemental analysis

For the metal analysis (Pb, Cd, As, Ba, B, Ca, Co, Cr, Cu, Mg, Zn, Fe, and K), PM₁₀ filter paper, dried plant (250 mg), and soils (100 mg) were digested with HNO₃: HCl (5:2 v/v) by Anton Paar Multiwave pro-oven for a total run of 41 min at 120 °C. These digested samples were used to determine metals and elements using Inductively Coupled Plasma (Perkin Elmer USA, optical emission spectrometer- Optima, 5300 DV). The total As content was determined using hydride generation with a vapor generation accessory (HG-VGA) coupled with an inductively coupled plasma optical emission spectrometer (ICP-OES Perkin Elmer Optima 5300 V) ²¹. All analyses were conducted in triplicate. Blanks were prepared and processed similarly to the reported sample procedures.

Quality assurance and quality control

The quality assurance and control of Pb, Cd, and As were also done by estimating blanks and assessing the method’s accuracy, recovery, and precision. Standard solutions with different concentrations were analyzed, and a calibration curve was prepared. The values of R² for the calibration line were 0.9997 for Pb, 0.9994 for Cd, and 0.9992 for As. The Pb, Cd, and As detection limits were 0.006, 0.0054, and 0.001 mg L⁻¹, respectively. The accuracy and precision of the methods were 4.17% and 1.65% for Pb, 4.2% and 1.13% for Cd, and 5.1% and 1.38% for As, respectively. The Pb, Cd, and As recoveries in spiked samples of PM₁₀, plant, and soil were 96–97%, 96.5–98%, and 92.5–94.5% respectively. Procedure blank and calibration standards were run after every ten samples.

Statistical analysis

Experimental data are expressed as the mean ± standard deviation (n = 3). One-way ANOVA (analysis of variance) followed by Tukey’s HSD test at a 5% (p < 0.05) probability level was done to assess significant differences among the treatments using SPSS version 25.0 (IBM). The correlation, hierarchical cluster analysis (HCA), principal component analysis (PCA), PCA-biplot, and correspondence analysis (CA) were applied to the data using R (R-Version 4.2.0). Plant fractionation and PM₁₀ samples collected from similar periods were taken for multivariate analysis.

Results and discussion

Concentration and bioavailability of metals and elements in PM₁₀

The concentrations of PM₁₀ during the study period ranged from 284 to 432.5 µg m^{− 3}, with an average value of 364.4 ± 57.5 µg m^{− 3}, respectively. The order of elements present in PM₁₀ was Ca > Ba > Fe > K > Al > Mg > Mn > Zn > Cu > Cr (Table S2). The Pb, Cd, and As concentrations in PM₁₀ were 0.03–0.78 µg m^{− 3}, 0.01–0.5 µg m^{− 3}, and 0.001–0.01 µg m^{− 3}, respectively. The bioavailability of all elements present in PM₁₀ (Table 1 and Fig. S2) displayed that Al, Ba, Ca, Cr, Cu, Fe, K, Mg, and Mn in PM₁₀ were dominant in water-soluble (51.5%), followed by oxidizable (42.1%), residual (5.4%), and reducible forms (0.8%). This fractionation of elements suggested the dominance of water-soluble salts like nitrate, chloride, and sulfates in aerosol²². The dominance of Pb in the residual fraction (76%), followed by water-soluble (13%) and then oxidizable (11%), suggested the contribution of road dust and industrial emissions. The presence of Pb salt in exhaust and its transformation into PbCO₃, Pb oxides, and PbSO₄ in the aerosol and co-precipitation with Fe and Mn oxides was reported earlier²³. The dominance of Cd in the oxidizable fraction (66.5%), followed by water-soluble (33.2%), suggested that the contribution of the combustion process, biological processes, and crustal sources also significantly contributed to Cd. The prevalence of crustal origin Pb and Cd in the residual and oxidizable fractions of PM₁₀ was reported earlier²⁴. Since the As concentration was low in PM₁₀, its fractionation was not performed.

Table 1.

Soil pH and organic carbon and Pb, Cd and As content and bioavailable fractions and proportion in PM₁₀ and soil.

				Soil treatments
				C	T1		T2	C	T1	T2
				Initial				Final
pH				7.49 ± 0.04a	7.49 ± 0.04a		7.52 ± 0.01a	7.45 ± 0.03a	7.44 ± 0.02a	7.44 ± 0.09a
OC (%)				0.34 ± 0.02a	0.34 ± 0.02a		0.31 ± 0.03a	0.34 ± 0.02a	0.33 ± 0.01a	0.30 ± 0.01a
PM10				Soil
				C		T1	T2	C	T1	T2
	Concentration ( µg m− 3 )		Proportion (%)	Concentration (µg g − 1 )				Proportion (%)
Pb
F1	0.06 ± 0.005a		12.8	ND	ND		ND	ND	ND	ND
F2	ND		ND	ND	1.45 ± 0.01c		20.7 ± 0.1c	ND	64.4	57.1
F3	0.05 ± 0.005a		11.1	ND	0.747 ± 0.005b		10.9 ± 0.05b	ND	33.3	30.2
F4	0.38 ± 0.01b		76.0	ND	0.051 ± 0.002a		4.60 ± 0.01a	ND	2.2	12.6
Cd
F1	0.004 ± 0.001		33.2	ND	ND		0.90 ± 0.02b	ND	ND	54.5
F2	ND		ND	ND	0.30 ± 0.01b		0.247 ± 0.005a	ND	60	15.1
F3	0.008 ± 0.001		66.5	ND	0.10 ± 0.01a		0.246 ± 0.005a	ND	20	15.1
F4	ND		ND	ND	0.103 ± 0.005a		0.246 ± 0.005a	ND	20	15.1
As
F1	ND		ND	ND	ND		ND	ND	ND	ND
F2	ND		ND	ND	0.004 ± 0.001a		0.13 ± 0.01b	ND	15.3	16.1
F3	ND		ND	ND	0.002 ± 0.001a		0.071 ± 0.005a	ND	7.6	8.7
F4	ND		ND	ND	0.020 ± 0.001b		0.643 ± 0.002c	ND	76.9	75

(significance difference p < 0.05).

Concentration and bioavailability of Pb, Cd, and As in soil

The concentrations of Pb, Cd, and As in T1 were 2.24 ± 0.08 µg g^− 1, 0.5 ± 0.01 µg g^− 1, and 0.024 ± 0.004 µg g^− 1, while in T2, concentrations were 35.26 ± 0.03 µg g^− 1 for Pb, 1.82 ± 0.06 µg g^− 1 for Cd, and 0.856 ± 0.002 µg g^− 1 for As. The Pb, Cd, and As were below the detection limit in the control soil. The soil pH, organic carbon content, and concentration of bioavailable fractions of Pb, Cd, and As in control and treatments are given in Table 1.

In soil, the predominant form of Pb was reducible, followed by oxidizable in T1 and T2 treatments, suggesting that atmospheric deposition of Pb did not change its speciation in soil. However, Cd exhibited distinct fractionation patterns in the two treatments. In T1, the prevalence of reducible fractions (60%) could be due to precipitation of atmospheric water-soluble Cd (dominant fraction in aerosol) onto iron (Fe) and manganese (Mn) oxyhydroxides of the soil. Additionally, about 20% of the fractions were classified as oxidizable and residual, indicating a contribution from an oxidizable fraction of atmospheric deposition²⁵. A small but insignificant decrease in soil pH was observed with their initial concentration, possibly due to exudation release from plant²⁶. In T2 treatment, Cd was higher in water-soluble fractions (54.4%), while reducible, oxidizable, and residual fractions contributed a similar proportion of about 15%. The fractionation pattern of As was similar in both treatments, i.e., residual > reducible > oxidizable fractions. This aligns with a previous study²⁷.

Dust deposition and morphology of the leaf surface

The dust deposition on the leaf’s surface was 1.82–1.96 mg cm^{− 2} and alkaline (pH: 6.6–7.2). The order in which elements are deposited onto the leaf surface is as follows: Ca > Ba > K > Mg > Al > Fe > Pb > Mn > Cu > Zn > Cd > Cr (Table S2). This order differs slightly from that of PM₁₀ due to the influence of additional processes during deposition. For example, although As was present in PM₁₀, its absence in the dust deposition on leaf surfaces could be attributed to its high mobility and/or volatilization by phyllosphere bacteria²⁸. The deposition flux concentration of Pb and Cd was 0.31 ± 0.01 ng cm⁻² h^{− 1} and 0.016 ± 0.002 ng cm⁻² h^{− 1} in T1, respectively, while deposition flux in T2 was 0.280 ± 0.009 ng cm⁻² h^{− 1} for Pb and 0.020 ± 0.002 ng cm⁻² h^{− 1} for Cd. The higher deposition flux of Pb than that of Cd and As was associated with its high content in PM₁₀.

The morphological analysis of the leaf surface revealed a distribution of mineral particles, resulting in a visibly rough texture in T1 and T2 compared to the smooth epidermal surface of the leaves in the control (Fig. 1a–c). The formation of aggregates near multicellular glandular trichomes of the abaxial region of leaves in T1 and T2 treatments suggested the solubilisation of atmospheric dust and their entry into the plant through glandular trichomes. The transformation of heavy metals exudates released on the leaf surface facilitates their conversion into bioavailable ionic forms, which may diffuse via aqueous films over guard cells or aquaporins, enhancing metal uptake potential²⁹.

Fig. 1.

Scanning electron micrographs of particles deposited on the adaxial and abaxial leaf surfaces of A. paniculata with C (close control soil), T1 (open control soil), and T2 (open polluted soil (T2); (a–c) showing trichomes on leaf surface; (d–f) showing cystoliths; (g–i) showing stomatal length; (j–l) showing stomatal width.

Additionally, the occurrence of more rod-shaped cystoliths on the adaxial surface of the leaf surface T1 and T2 than in the control (Fig. 1d–f) suggested acceleration of the bio-mineralization at the leaf surface, an adaptation mechanism of detoxification associated with Wang and Ji (2021). The higher number of stomata and trichomes, reduced stomatal aperture length (Fig. 1g–i) and width (Fig. 1j–l) on the abaxial surface in T1 and T2 treatments than in the control indicated incorporation of particles through stomatal pores. This morphological adaptation likely minimizes the internal accumulation of particulate-bound heavy metals³⁰. The anatomical study of the midrib of leaves displayed thickening of the cell wall T1 and T2 compared to the control (Fig. 2a–c). This was due to the formation of cation complexes or metal present in dry deposition with negatively charged functional groups within the cell wall matrix.

Fig. 2.

T.S. of midrib with lamina showing intercellular space thickening in open environment treatments T1 and T2 as compared to close control (a) C, (b) T1 and (c) T2, and presence of cystoliths in T1 and T2 with control (d) C, (e) T1, and (f) T2.

Plant growth and oxidative stress

The plant growth, biomass, secondary metabolites, and antioxidant enzyme activities are shown in Fig. 3. A significant (p < 0.01) decrease was observed in biomass (44.3% and 52.4%), root length (15.8% and 32.4%), total protein content (68.2% and 70%), and chlorophyll content (39.3% and 58.2%) in T1 and T2, respectively, as compared to the control. A significant increase in plant height was observed in the treatments compared to the control.

Fig. 3.

Plant growth parameters (a) plant height, (b) root length, (c) biomass, (d) protein, (e) chlorophyll; secondary metabolite (f) andrographolides, and antioxidant enzymes (g) superoxide dismutase, (h) catalase, and (i) glutathione reductase in close control C, and open environments T1 and T2 treatments.

This indicated the physiological and biochemical stress induced by atmospheric and soil-borne PTMs. The biomass reduction in plants exposed to atmospheric dust is directly linked to the physical blocking of leaf stomata, which disrupts normal cellular physiological reactions and alters the uptake and subcellular distribution of these metals. The accumulation of Pb, Cd, and As in the subcellular part of the plant caused physiological disorders, leading to higher production of stress enzymes³¹. A significant increase in SOD, CAT, and GR in T1 (93.5%, 72.7%, and 56.2%, respectively) and T2 (95.8%, 76.9%, and 73.07%, respectively) compared to the control is an indicator of the plant responses to oxidative stress caused by atmospheric deposition and metal accumulation. The more toxic effect in T2 than T1 in A. paniculata was in line with previous reports demonstrating less toxicity in chilli by single foliar and root uptake of Cd than foliar-root uptake³². A significant decrease in the major secondary metabolite of A. paniculata in T1 (andrographolide content: 1.52%) and T2 (andrographolide content: 1.08%) compared to the control (andrographolide content: 2.04%) indicated that atmospheric deposition altered the secondary metabolite production pathways. The decrease in production of andrographolide by As accumulation via soil was reported earlier²¹.

Concentration of Pb, Cd, and As in plant tissues and subcellular parts

In leaves, the Pb content was 1.5 ± 0.001 µg g^{− 1} in T1 and 2.8 ± 0.004 µg g^{− 1} in T2, while Cd concentration was 0.9 ± 0.004 µg g^{− 1} in T1 and 1.10 ± 0.01 µg g^{− 1} in T2. However, the As content was 0.01 ± 0.005 µg g^{− 1} in T1 and 0.155 ± 0.003 µg g^{− 1} in T2. In the stem, the Pb, Cd, and As concentrations in T2 treatment were 0.80 ± 0.002 µg g^{− 1}, 0.600 ± 0.001 µg g^{− 1}, and 0.09 ± 0.002 µg g^{− 1}, respectively. While in root, the concentration of Pb, Cd, and As was 10.9 ± 0.01 µg g^{− 1}, 5.10 ± 0.001 µg g^{− 1}, and 1.49 ± 0.007 µg g^{− 1}, respectively. The concentration of Pb, Cd, and As was not observed in the control of all tissues. In the T1 treatment, the absence of Pb, Cd, and As in the stem and root suggested no translocation of metals in these tissues due to lower pollution load or total absorption of particulate trace metals on the leaves³³.

The concentration and fraction of Pb, Cd, and As in the subcellular fractions of leaf, stem, and root tissues are presented in Figs. 4a–c and Table S3. In leaves, Pb was bound mainly to the cell wall (45.5% in T1 and 44.2% in T2), followed by the nucleus and chloroplast (37.2% in T1 and 28.4% in T2), then the soluble fraction in T1 (8.6%) and mitochondria in T2 (17.4%). Meanwhile, the fraction of Pb was lower in the mitochondria (8.4%) of T1 and the soluble fraction (9.9%) of T2. In this study, Cd mainly contributed to the cell wall (37.6% in T1 and 38.5% in T2), followed by mitochondria (32.9% in T1 and 30.8% in T2), then nucleus and chloroplast (26.6% in T1 and 23.6% in T2), and soluble fraction (2.7% in T1 and 6.9% in T2). The Pb and Cd are primarily bound to the cell wall, followed by soluble fractions in the stem and root. The cell wall acts as a primary barrier, and binding metal ions with carboxylic groups of pectin and sugars, precipitation, and ion exchange sequestration is a plant’s detoxification strategy^34–37. In leaf tissues, a higher proportion of Pb was found in the nucleus and chloroplast, while Cd was present in the nucleus, chloroplast, and mitochondria in T1 treatment than in T2, suggesting that these metals present in atmospheric deposition had more bioavailability or were in a form that can enter the organelles and bind there.

Fig. 4.

(a) leaf (b) stem, and (c) root, showing proportion of Pb, Cd, and As in different subcellular fractions of T1 and T2 treatments (CW; cell wall, NC; nuclear and chloroplast, M; mitochondria, S; soluble fraction). The proportion of different Pb, Cd, and As chemical forms in the (d) leaf, (e) stem, and (f) root in T1 and T2 treatments (CF1, EtOH-extractable; CF2, H₂O-extractable; CF3, NaCl-extractable; CF4, HAc-extractable; CF5, HCl-extractable; CF6, residual). *(Pb, Cd, and As not detected in control (C) of leaves, stem and root; As in T1 of leaves, and Pb, Cd, and As not detected in T1 of stems and root).

In leaves, As was bound mainly to the soluble fraction (47.1%), followed by the cell wall (27.69%), then mitochondria (25.1%) in T2. However, As was mainly observed in the soluble fraction in the stem and root, followed by the cell wall. The higher accumulation of As in the soluble fraction, followed by the cell wall fractions, was reported earlier³⁸. However, the mitochondrial association of As in T2 could be related to its high mobility; arsenate can easily cross internal membranes by accumulating various Pi transporters³⁹.

Chemical forms of Pb, Cd, and As in different tissues

After entering the plant, cell activity, migration ability, and toxicity of PTMs are closely related to their chemical forms⁴⁰. In foliar uptake, these metals penetrate the plasma membrane of leaf epidermal cells and stomata or via the rims of drying droplets on the leaf surface. The concentration of Pb, Cd, and As in different chemical forms of leaf, stem, and root tissues is presented in Fig. 4d–f and Table S4.

In the leaf, Pb had predominantly HCl-extractable (T1: 65.9%, T2: 67.5%), followed by HAc-extractable (T1: 15.2%, T2: 24.19%) fraction. T1 showed higher proportions of EtOH-extractable, H₂O-extractable, and residual forms of Pb than T2. However, Cd primarily was observed in HCl-extractable (T1: 39.9%, T2: 41.4%) and HAc-extractable (T1: 39.41%, T2: 39%), followed by residual form (T1: 9.6%, T2: 11%) and EtOH-extractable (T1: 7.9%, T2: 7.79%). In other forms, T1 had more H₂O-extractable Cd than T2. The As was highest in H₂O-extractable form (24%), followed by EtOH-extractable (4.3%), residual form (2%), and NaCl-extractable (1.3%). The dominance of Pb and Cd in HCl-extractable fractions suggested their incorporation into plants as detoxification strategies⁴¹. The metal complex with oxalate and insoluble phosphate in the cell walls could be present in HCl-extractable and HAc-extractable^40,42. In leaves, the higher percentages of mobile forms of Pb and Cd in T1, specifically EtOH and H₂O-extractable, suggest that these metals can easily enter organelles due to their high-water solubility. The greater accumulation of Pb and Cd in the nucleus, chloroplast, and mitochondrial fractions of T1 compared to T2 supports this finding. The soluble forms of these metals are distributed within the cytosol. They are partitioned among various cytosolic ligands and intracellular organelles based on their redox-active nature, allowing them to bind readily with other plant components⁴³. In the stems, the predominant forms of Pb, Cd, and As in T2 were HCl-extractable and HAc-extractable due to increased sequestration of soil-derived PTMs at the cell wall, where they bind with polysaccharides and ions. Conversely, in the roots, the contents of Pb, Cd, and As in HCl-extractable, NaCl-extractable, HAc-extractable, and H₂O-extractable forms were consistent with their availability and distribution in the cell wall, mitochondria, and soluble fractions. While As was dominant in oxalate forms, followed by insoluble phosphate complexes. Previous studies have reported the prevalence of As in these two fractions during the vacuolar sequestration of S. nigrum L. ⁴⁰.

Interrelationship

In this study, multivariate tools (correlation, hierarchical cluster analysis (HCA), principal component analysis (PCA), and correspondence analysis (CA) were employed to examine the association between the fractions (PM₁₀, subcellular distribution, chemical forms, and soil fractions) and similarities/dissimilarities between the treatments. The As data in multivariate analysis were used only for the T2 treatment. Soil bioavailable fraction, water-soluble and reducible fraction of Cd and the reducible fraction of Pb in T2 analysis were taken as these forms are more bioavailable.

PCA and CA were examined to understand the similarity/dissimilarity between treatments. (Fig. 5a-d). Correlation analysis, hierarchical cluster analysis (HCA), and principal component analysis (PCA) were performed separately for each metal in T1 and T2 treatments to understand the relationships between speciation, bioavailability, and sequestration of specific PTMs in each treatment (Figs. 6 and 7, and 8).

Fig. 5.

(a) Pb, (b) Cd, and (c) As showing Principal component analysis and (d) Correspondence analysis among fractions of PM₁₀ (FF1, water-soluble fraction; FF3, oxidizable fraction; FF4, residual fraction), subcellular distribution (CW; cell wall, NC; nucleus and chloroplast, M; mitochondria, S; soluble fraction), chemical forms of leaves EtOH-extractable, H₂O-extractable, NaCl-extractable, HAc-extractable, HCl-extractable, RD; residual form and soil fractions (SF1; exchangeable-water soluble fraction, SF2; reducible fraction) with T1 and T2 treatments.

Fig. 6.

(a) Correlation analysis (b) Hierarchical cluster analysis (HCA) (c) PCA-biplot analysis of T1Pb, and (d) Correlation analysis (e) Hierarchical cluster analysis (HCA) (f) PCA-biplot analysis of T1Cd among fractions of PM₁₀ (FF1, water-soluble fraction; FF3, oxidizable fraction; FF4, residual fraction), subcellular distribution (CW; cell wall, NC; nucleus and chloroplast, M; mitochondria, S; soluble fraction), chemical forms of leaves EtOH-extractable, H₂O-extractable, NaCl-extractable, HAc-extractable, HCl-extractable, CF6 residual form of uncontaminated soil (T1).

Fig. 7.

(a) Correlation analysis (b) Hierarchical cluster analysis (HCA) (c) PCA-biplot analysis of T2Pb and (d) Correlation analysis (e) Hierarchical cluster analysis (HCA) (f) PCA-biplot analysis of T2Cd among fractions of PM₁₀ (FF1, water-soluble fraction; FF3, oxidizable fraction; FF4, residual fraction), subcellular distribution (CW; cell wall, NC; nucleus and chloroplast, M; mitochondria, S; soluble fraction), chemical forms of leaves EtOH-extractable, H₂O-extractable, NaCl-extractable, HAc-extractable, HCl-extractable, CF6 residual form and soil fractions (SF1; exchangeable-water soluble fraction, SF2; reducible fraction) of contaminated soil (T2).

Fig. 8.

(a) Correlation analysis (b) Hierarchical cluster analysis (HCA) (c) PCA-biplot analysis of T2As among fractions of PM₁₀ (FF1, water-soluble fraction; FF3, oxidizable fraction; FF4, residual fraction), subcellular distribution (CW; cell wall, NC; nucleus and chloroplast, M; mitochondria, S; soluble fraction), chemical forms of leaves EtOH-extractable, H₂O-extractable, NaCl-extractable, HAc-extractable, HCl-extractable, CF6 residual form and soil fractions (SF2; reducible fraction) of contaminated soil (T2).

Similarity/dissimilarity between treatments

The PCA analysis (Fig. 5a-d) assessed the variability in PM₁₀, subcellular distribution, chemical forms, and soil fractions of Pb, Cd, and As between T1 and T2 treatments. The total variance explained was high for all metals: 87% for Pb, 90% for Cd, and 92.7% for As. The first two principal components (PC1 and PC2) accounted for 71% and 16% of the Pb variance, 74.9% and 15.1% of the Cd variance, and 78.8% and 13.9% of the As variance, respectively. The PC1 vs. PC2 plots segregated T1 and T2 treatments for each metal, suggesting a differential pattern of metal sequestration and chemical forms.

Correspondence analysis (CA) was conducted to identify statistical relationships between T1 and T2 treatments of Pb, Cd, and As fractions across PM₁₀, soil fractions, subcellular distribution, and chemical forms in plants (Fig. 5d). The two factors accounted for 57% and 30.4% of the variance observed. Pb and Cd in the T1 treatment, along with As in the T2 treatment, clustered with water-soluble, oxidizable, and residual fractions of PM₁₀, leaf chemical forms H₂O-extractable, HCl-extractable, and RD forms subcellular fractions of cell wall, nucleus and chloroplast, and soluble fraction of cell. Conversely, Pb in the T2 treatment was associated with residual soil fraction and NaCl-extractable chemical form. Cd in the T2 treatment was linked to water-soluble soil fraction, mitochondrial fraction, EtOH, and HAc-extractable chemical forms. The CA further discriminates between the two treatments on speciation and sequestration patterns of metals.

Pb, and Cd speciation, bioavailability, and sequestration in uncontaminated soil (T1)

The correlation, HCA, and PCA results for Pb in T1 treatments (Fig. 6a–c) demonstrated that water-soluble and oxidizable fractions of Pb in PM₁₀ were associated with chemical forms H₂O-extractable, NaCl-extractable, HAc-extractable, and RD form, and the soluble fraction of the cellular part. This indicated that water-soluble and oxidizable fractions, i.e., organic acid complex of Pb in PM₁₀, are sequestered into vacuoles of plant cells. The ion exchange or weak binding on the organic surface of Pb was reported earlier⁴⁴. The association of the residual fraction of Pb in PM₁₀ with chemical forms EtOH-extractable and HAc-extractable, as well as the soluble fraction of the cellular part, suggested vacuole sequestration.

The oxidizable fraction of Cd in PM₁₀ (Fig. 6d–f) was associated with EtOH-extractable, H₂O-extractable, HAc-extractable, and HCl-extractable chemical forms, and the soluble fraction of plant cells. The water-soluble fraction of Cd in PM₁₀ was associated with the cell wall, nucleus, chloroplast, and mitochondrial fractions of the cell and chemical forms NaCl-extractable and RD forms. The HCl-extractable and HAc-extractable chemical forms of Cd were associated with soluble fractions in PCA. This indicates that water-soluble fractions of Cd were transported into the nucleus, chloroplast, and mitochondria through transporters or ion channels. However, it may be possible that oxidizable forms of Cd were transformed into bioavailable forms such as nitrate, chloride, aminophenol, and organic acid complex (citrate and malate) and enter the cell organelle. The redox-active metal forms might be oxidized with O₂ and H₂O₂ molecules and release ionic forms of PTMs. This oxidative transformation could contribute to changes in the solubility and mobility of PTMs within the cell⁴⁵.

Pb, Cd, and As speciation, bioavailability, and sequestration in metal-contaminated soil (T2)

In T2, PM₁₀ water-soluble and oxidizable fraction of Pb were paired with chemical form NaCl and HAc-extractable and nucleus, chloroplast, mitochondrial, and soluble cell fraction (Fig. 7a–c). The clustering of the residual fraction of Pb in PM₁₀ with chemical form EtOH-extractable, NaCl-extractable forms, RD forms, and nucleus, chloroplast, and soluble cell fractions was observed in T2. Generally, the RD chemical form is biologically inaccessible; however, its clustering with mobile forms suggested that Pb bio-transformed into soluble forms binds with nitrate, chloride, aminophenol or protein and pectate in the cell. Furthermore, the reducible soil fraction in T2 was associated with HCl-extractable chemical form, possibly due to the formation of Pb-oxalate complex in the cell wall or adsorption of Fe/Mn hydroxides⁴⁶.

Correlation, PCA, and HCA (Fig. 7d–f) results indicate that the oxidizable fraction of PM₁₀ and water-soluble soil fraction of Cd were positively associated with mitochondrial and soluble cell fractions and NaCl-extractable and HCl-extractable forms. These forms of Cd might be complexed with oxalate and sequestered in vacuoles or precipitated in the cytoplasm, which reduces Cd translocation. At the same time, Cd binds with proteins containing -SH groups, glutathione complex, and metallothioneins, which might be transported into mitochondria after increasing the oxidizable fraction of PTMs in soil due to atmospheric deposition. Unbound small molecules and inorganic Pb, Cd, and As ions can enter plant cell organelles through various protein ion channels and receptors⁴⁷. The association of the reducible soil fraction of Cd with a water-soluble fraction of PM₁₀ might confirm our hypothesis that atmospheric Cd precipitates on Fe/Mn oxides in soil. The association of the water-soluble fraction of soil with the cell wall and EtOH-extractable, HAc-extractable, and RD chemical forms suggests its binding and precipitation in the form of phosphate and carboxylates. It has been reported that the cell wall and soluble fraction may not always fully retain Cd due to its high mobility and difficulty in precipitating within the cell wall⁴⁸. It may be possible that protein or pectate-bound Cd might be trafficked to vacuoles or precipitated as Cd-phosphate/Cd-sulfide granules with time. The multivariate analysis showed the As concentration associated with soluble and mitochondrial, and chemical forms EtOH-extractable, H₂O-extractable, NaCl-extractable, HCl-extractable, and RD forms (Fig. 8a–c). The reducible soil fraction of As is positively associated with the cell wall, nucleus, and chloroplast of the cell, and the HAc-extractable form. It suggests that As might change their solubility and mobility inside the cytosol by transforming into more bioavailable forms such as nitrate, chloride, and aminophenol. The formation of organic complexes with citrate and malate could be a reason for their mitochondrial association⁴⁹. The reducible fraction of As forms a complex with insoluble phosphates in the cell wall, reducing As bioavailability⁵⁰.

Conclusion

Our findings highlight the complex interplay between atmospheric deposition and plant-soil interactions in dictating the fate and impact of PM₁₀-associated heavy metals. While Pb, Cd, and As in PM₁₀ originate from diverse crustal, biological, and secondary sources, their behavior within Andrographis paniculata shows remarkably dynamic. The atmospheric metal fractions, particularly water-soluble forms of Pb and Cd, readily enter plant cellular components and organelles, suggesting a direct pathway for toxicity. Notably, less bioavailable atmospheric forms (oxidizable and residual) might undergo significant biotransformation into more mobile fractions once inside the plant. Crucially, the soil environment profoundly modulates the plant’s response. In T1 treatment, higher bioavailability of these metals leads to greater cellular accumulation, with atmospheric deposition being the sole source. Conversely, T2 treatment prompts the plant to activate robust detoxification mechanisms, shifting metals towards less bioavailable forms and significantly reducing organelle uptake due to the dominance of the root uptake process. This research suggests that both atmospheric concentration and the specific chemical forms of deposited metals, alongside their subsequent transformations within the plant and soil, are pivotal in determining their bioavailability and toxic effects. Future research should further explore the intricate details of these metal bio-transformations within plant tissues.

Supplementary Information

Below is the link to the electronic supplementary material.

Author contributions

Sweta Yadav: Investigation, Methodology, Formal analysis, Writing-original draft, Data interpretation. Puja Khare: Funding acquisition, Conceptualization, Writing, review, and editing.

Data availability

The data that support the findings of this study are available in the supplementary material of this article.

Declarations

Competing interests

The authors declare no competing interests.

Planting material

The study complies with relevant institutional, national, and international guidelines and legislation for planting material used.