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Seagrasses are essential blue carbon ecosystems that are increasingly threatened by the accumulation of emerging global pollutants, such as microplastics (MPs). However, in regions like Sri Lanka, MP contamination within seagrass beds remains critically understudied. This study investigated MP abundance in seagrass blades ( Thalassia hemprichii ), sediments, surface water, and sea cucumbers ( Holothuria atra ) and characteristics of isolated MPs from six selected sites along the Southern and Eastern coastlines of Sri Lanka (Indian Ocean). The average abundances of MPs in seagrass blades, Holothuria atra , surface water, vegetated sediments, and non-vegetated sediments were 1.59 ± 0.11 items/blade, 177.19 ± 53.28 items/kg, 9.60 ± 0.97 items/m 3 , 69.72 ± 8.38 items/kg, and 33.61 ± 2.15 items/kg, respectively. In this study, fibers were the most prevalent morphological type of MPs, while blue was the most dominant color. Fourier Transform Infrared Spectroscopy (FTIR) analysis revealed the predominant presence of low-density polyethylene (LDPE) in all types of samples, followed by polyethylene terephthalate (PET), high-density polyethylene (HDPE), polypropylene (PP), and polyamide or commonly known as nylon (PA) in variable amounts. A strong positive correlation between MPs on seagrass blades and those in sediments and surface water suggests that MPs from both water column and seabed contribute to their accumulation on seagrass. Furthermore, the ingestion of MPs by sea cucumbers indicates the potential for fragmentation and trophic transfer within the seagrass ecosystem, enabling MPs to enter and move through the marine food chain.
Plastic particles less than 5 mm in size are known as microplastics 1 . Microplastics (MPs) are emerging global contaminants in the marine environment. Historically, MP studies have focused predominantly on open sea areas 2 . Benthic ecosystems, including seagrass beds, are increasingly being recognized as potential hotspots for MP pollution. However, the majority of previous studies in benthic ecosystems focus on MPs in surface water and sediments in a range of geographical areas, such as seagrass beds in Deerness Sound, Orkney 3 , South Japan 4 , East Asian surface water, the Bohai Sea, China 5 , Barranglompo Island in Indonesia 6 , Xincun Bay and Li’an Bay, Hainan, China 7 , Mu Ko Phetra National Park, Satun Province, Thailand 8 . Specific environmental conditions (e.g. decreased water flow, increased settling rates, and the ability of seagrass blades to trap particles) can promote the accumulation of MPs in seagrass ecosystems 7 , 9 , 10 . For example, a significant number of MPs accumulate in epibionts of Thalassia testudinum blades even in remote locations. Grazers consume these seagrasses with higher concentrations of epibionts 11 . Accordingly, a high potential remains to transfer MPs into higher trophic levels 12 , and various marine organisms such as fish 13 , 14 , crabs 15 , sea urchins 6 , 16 , sea cucumbers 17 , and mussels 18 ingest MPs due to their high bioavailability and small particle size. Therefore, plastic pollution in seagrass ecosystems reduces primary productivity and associated multifunctional ecosystem services, including acting as nursery and spawning grounds for marine organisms, facilitating water filtration, protecting coastlines, cycling nutrients, and sequestering and storing carbon 9 , 19 , 20 .
Similarly, MP research has received slight attention in the seagrass ecosystems of Sri Lanka. Thalassia hemprichii is the most widespread seagrass species on the Southern and Eastern coasts of Sri Lanka, and these ecosystems are highly diverse with associated organisms 21 . However, seagrass ecosystems are under threat due to coastal infrastructure development, harmful fishing practices, and pollution from agricultural and aquaculture activities 20 , but are not well documented for MP contamination. Therefore, estimation of abundance and composition of MPs in Sri Lankan seagrass beds is vital for effectively addressing the silent threats caused by plastic pollution. Based on the existing literature and identified knowledge gaps, we hypothesized that (i) seagrass beds in Sri Lanka function as accumulation areas for MPs, resulting in significantly higher MP concentrations in vegetated sediments compared to adjacent non-vegetated areas, and (ii) distribution and characteristics of MPs vary across different ecosystem components. The objectives of this study are (i) to evaluate the MP abundance in selected seagrass beds on the Southern and Eastern coasts of Sri Lanka by analyzing seagrass blades, sediments, surface water, and sea cucumber samples, (ii) to determine the physical and chemical characteristics of the isolated MPs and (iii) assess spatial contamination patterns using ecological pollution indices, including the Microplastic Load Index (MPLI), Polymer Hazard Index (PHI), and Microplastic Distribution and Impact Index (MDII), to provide an integrated evaluation of MP contamination across ecosystem components. Finally, this study provides a comprehensive understanding of the extent of MP prevalence in seagrass beds along the Southern and Eastern coasts of Sri Lanka and contributes valuable baseline data for future management and conservation efforts.
Six shallow seagrass sites were selected from the Southern ( n = 3, Ahangama, Dondra, and Dickwella) and Eastern ( n = 3, Adukkuparu, Salli Beach, and Kalkudah) coasts of Sri Lanka, considering the availability and distribution of Thalassia hemprichii , along the coastline (Fig. 1 ). Sampling was conducted considering the inter-monsoon periods of each coast, in April 2023 on the Southern coast, and from April to May 2023 on the Eastern coast. In this study, four types of samples were collected, including seagrass blades ( Thalassia hemprichii ), surface water and sediment samples, and sea cucumbers ( Holothuria atra ), from each site. A minimum of three replicates were collected for each sample type at each site.
Seagrass blades ( T. hemprichii ) were collected at each sampling site following the methods described in the literature 3 , 6 . Seagrass blades were cut via scissors and separately added to labeled pre-rinsed glass bottles with distilled water to avoid cross-contamination. The scissors were cleaned with filtered water before and after use. The collected samples were kept in an ice box and transported to the laboratory. Three replicates (3 blades × 3 = 9) for each sampling site were maintained following the same method.
Surface water samples were collected into sterilized glass bottles from the seagrass meadow, using a 300 μm mesh manta net 22 . After the cap was carefully removed from the cod end, all the concentrated particles were transferred to labeled, sterilized glass bottles and stored in an ice box until further analysis. Before and after each sampling, the net was rinsed with filtered water. Three sample replicates were collected from each sampling site following the same procedure. A total of 18 (6 × 3) samples were obtained from the six selected sampling sites.
Three sediment samples (vegetated sediment) were collected at each sampling site within the seagrass beds, using the methods described by the literature 3 , 23 . About 250 g of sediments were collected from 5 cm beneath the surface using a core sampler. Each sediment sample was stored in labeled aluminum foil bags and transported to the laboratory. Three non-vegetated samples were collected away from the seagrass beds and near the shoreline using the same method as control samples.
Three sea cucumber samples were hand-harvested randomly from the seagrass beds at each site. All samples were transported to the laboratory in sterilized glass bottles using ice packs, and the samples were stored in a − 40 °C freezer until laboratory analysis 24 . Species identification was based on external morphological features such as body coloration, presence of longitudinal muscle bands, following descriptions by 25 , 26 .
Map of Sri Lanka showing the sampling sites.
Microplastics were separated from seagrass blades, surface water, surface sediments, and sea cucumbers according to established methodologies. Microplastic isolation was performed for seagrass blades, according to the literature 3 , 6 . Seagrass blades were cleaned with distilled water in a clean 50 mL skirted centrifuge tube. After vigorous agitation on a shaker rotator for 30 s, the samples were transferred to a sterile tray. Each seagrass blade underwent scraping with the blunt edge of a scalpel blade to eliminate MPs that adhered to the seagrass or biofilms present. Any dislodged particles were subsequently rinsed into the centrifuge tube using a glass pipette filled with distilled water. Samples were subsequently filtered through 0.7 μm GF/C filter paper via a vacuum pump and stored in Petri dishes for subsequent microscopic analysis.
Microplastics were extracted from surface water using the wet peroxide oxidation (WPO) method 27 . The mixture was treated with hydrogen peroxide (30%) transferred to a density separator, and allowed to settle overnight. The resulting filtrate was then passed through GF/C Whatman filter paper (1.6 μm) to determine the MP count.
For surface sediment, the MP isolation process followed the methodology outlined by 23 . Sediment samples were initially oven-dried at 60 °C for 24 h until reaching a constant weight. Approximately 250 g dried sediment of each sample was floated in a 1.2 g/cm 3 saturated NaCl solution for density separation. Visible plastic particles floating on the surface were manually removed using forceps and transferred to a Petri dish. To remove residual organic matter, 20 mL of 30% H 2 O 2 was added to the separated material and allowed to react overnight at room temperature for oxidative digestion of organic components. The resulting solution was then filtered through GF/C filter paper (1.6 μm) with a Büchner funnel.
The MP extraction from sea cucumber samples followed the methodologies introduced by 24 , 28 . H. atra samples were thawed at a temperature range of 25–27 °C. Before the intestines were removed, measurements of wet weight (in grams) and length (in centimeters) were recorded. Each sample was subsequently dissected individually, and the gastrointestinal tract (GIT) was isolated from the top of the esophagus to the anus. The isolated GIT was then transferred to a clean glass beaker, and 20 mL of 10% KOH solution was added to digest the tissue. After 24 h of digestion at 40 °C, the resulting mixture was filtered through a 0.7 μm glass microfilter.
The filter papers were examined under a light microscope at 10 × 10 magnification using a 10-megapixel camera (model name: BOECO), and the MP length was measured using ImageJ bundled with 64-bit Java 8 software. During the visual sorting phase, a hot needle test was randomly applied solely as a preliminary screening tool to differentiate synthetic plastics from natural organic debris (observing for melting or curling behaviors). On the basis of the length along the longest axis, the MP particles were further categorized into large and small MPs with ranges of 0.06–0.99 mm and 1.00–4.56 mm, respectively. The MP particles collected were categorized according to size, shape, and color.
The isolated plastic particles were analyzed for chemical identification using the attenuated total reflectance mode of Fourier Transform Infrared Spectroscopy (ATR-FTIR) (ALPHA-T, South Africa). Approximately 50% of the randomly selected microplastic particles, primarily from the large size fraction (1.00–4.56 mm), were subjected to FTIR analysis to ensure reliable spectral acquisition. Smaller particles and fibers were analyzed whenever feasible, and those materials were not proportionally subsampled due to practical constraints. Spectra were collected over the range of 450–4000 cm −1 at a resolution of 4 cm −1 using 32 scans per sample. Bruker’s OPUS 23 spectroscopy software was used for spectral acquisition and polymer identification. These spectra were then compared with reference polymer libraries, and matches with a quality index above (e.g. 70% or 80%) were accepted for polymer identification 29 , 30 .
Prior to field sampling, all equipment used for sample collection and storage was thoroughly rinsed with distilled water to minimize the potential cross-contamination. Items of clothing prone to shedding fibers were avoided during all procedures. Before the laboratory analyses, benchtop working surfaces, laboratory equipment, and all necessary instruments were cleaned with alcohol, and investigators’ hands were thoroughly washed and scrubbed to prevent contamination. The glassware, metal, and dissecting equipment were rinsed three times with distilled water. The quality control samples ( n = 19) were incorporated into the sampling (field and transportation) and sample preparation, method development, and isolation protocols in the laboratory. Field ( n = 2), transportation ( n = 2), and laboratory control ( n = 15) samples were prepared with distilled water. Three spiked samples were prepared using polyethylene terephthalate (PET), polyvinyl chloride (PVC), and low-density polyethylene (LDPE) particles. Spiked samples ( n = 21) were prepared for each matrix using three polymer types, including polyethylene terephthalate (PET) and PVC, and processed in parallel with the corresponding sample batches. These spiked samples were subjected to the same extraction and analytical procedures as the environmental samples. Mean recovery rates were within the generally accepted range across all matrices, with values of 87 ± 5% for seagrass blades, 93 ± 3% for sediments, 94 ± 5% for surface water, and 89 ± 6% for sea cucumbers. The identified average suspected MP particles in the quality control blanks were subtracted from the original samples for background correction. Tissue digestion was performed in a fume hood, and control samples consisting of 10% KOH were filtered onto glass fiber filter paper for microscopic examination.
The recovery rate of spiked Polyvinyl Chloride (PVC), Low-Density Polyethylene (LDPE), and Polyethylene Terephthalate (PET) particles ranged from 94 to 97%. The procedural blanks ( n = 19) showed minimal background contamination, with an average of 4.1 ± 1.2 particles per blank consisting of blue and transparent fibers. This ensures that the reported abundances are representative of the actual environmental conditions.
Ecological risks associated with MP contamination were evaluated using quantitative indices including, the Microplastic Pollution Load Index (MPLI), Polymer Hazard Index (PHI), Microplastic Diversity Integrated Index (MDII), and Bioaccumulation Factor (BAF). These preliminary methods integrate MP abundance and polymer hazard characteristics to assess environmental risk levels 31 , 32 .
Microplastic Pollution Load Index (MPLI)
MPLI was calculated to compare relative pollution intensity among sampling sites:
where contamination factor (CF) is:
where \(\:{C}_{site}\) represents mean microplastic abundance at each site and \(\:{C}_{baseline}\) is the lowest observed concentration among sites.
Polymer-associated ecological risk was estimated as:
where \(\:{P}_{n}\) is the percentage of each polymer type and \(\:{S}_{n}\) is the hazard score derived from polymer risk rankings.
Microplastic Diversity Integrated Index (MDII)
Polymer diversity was assessed using the Shannon diversity index:
where \(\:{p}_{i}\) represents the proportional abundance of each polymer type.
Bioaccumulation potential in Holothuria atra was calculated as:
where the organism concentration represents MPs in sea cucumbers and the sediment concentration represents MPs in the surrounding vegetated sediment.
Grain size analysis was carried out for sediment samples at each site using a mechanical sieve shaker. Sediment samples (500 g) were oven-dried at 60 °C for 24 h before being sieved. Mechanical sieve analysis was performed for 20 min at quarter phi intervals (BS sieves), with mesh sizes of 3.15 mm, 2.00 mm, 1.00 mm, 0.5 mm, 0.25 mm, 0.125 mm, 0.063 mm, 0.050 mm, and a pan. The sieved materials were collected and weighed. GRADISTAT V9.1 software was used for the classification of sediments.
All the numerical data are displayed as the mean ± standard deviation (SD). The data were tested for normality (Shapiro‒Wilk test) and homogeneity of variance (Levene’s test). The differences in MP abundance among seagrass blades, surface water, surface sediments, and sea cucumbers were analyzed using a one-way analysis of variance (ANOVA) test. A post hoc Tukey (HSD) test was performed to determine the statistical significance of the differences. The Pearson correlation coefficient was calculated for MP abundance in surface water, sediments, seagrass blades, and sea cucumbers. For all the statistical analyses, the significance level was set at p < 0.05. Origin 2023b software was used to develop the correlation matrix.
Figure 2 shows average MP concentrations in seagrass blades (items/blade), surface water (items/m 3 ), surface sediment (items/kg), and sea cucumbers (items/kg) of the entire animal. Microplastics were detected in all seagrass blade samples collected across the six sampling sites on the Southern and Eastern coasts of Sri Lanka. In total, 145 MP items were found from T. hemprichii blades. Accordingly, the overall average MP abundance was 2.68 ± 0.11 items per blade, and the average MP abundance ranged from 1.07 ± 0.07 to 2.73 ± 0.18 items per blade (Fig. 2 a), with significant variation among the sites ( p < 0.05). On the Southern coast, the highest MP concentration was recorded at Dickwella (1.27 ± 0.13 items/blade), while Ahangama and Dondra had the lowest concentrations of 1.07 ± 0.07 items/blade. On the Eastern coast, Salli Beach exhibited the highest MP abundance of 1.93 ± 0.13 items/blade, whereas Kalkudah had the lowest abundance of 1.47 ± 0.07 items/blade. The elevated MP concentrations at Salli Beach may be linked to intensive fishing activities, inadequate waste management, and nutrient influx 33 , 34 . These variations can reflect site-specific influences such as anthropogenic activities, hydrodynamic conditions, and seagrass morphology.
The average MP concentrations in different environmental matrices, ( a ) seagrass blades (items/blade), ( b ) surface water (items/m 3 ), ( c ) surface sediment (items/kg), and ( d ) sea cucumbers (items/kg) of the entire animal, across six sampling sites. Error bars represent standard deviations (± SD) of the means.
The current study indicates a relatively lower MP abundance in seagrass ecosystems compared to previous literature (Table 1 ). For example, MP concentrations range from 7.7 to 13.7 items/blade in five seagrass species from the Puttalam Lagoon, Sri Lanka 28 . The comparatively lower MP values in the current study may be partly attributed to species-specific differences in blade morphology and remote environmental conditions. Seagrass beds, particularly those of T. hemprichii , reduce water flow velocity and enhance MP deposition and retention 11 . In addition, blade morphology, such as surface roughness and the presence of biofilms and epiphytes, can enhance MP retention 3 , 35 . Species with broader and more textured blades like E. acoroides and T. hemprichii , generally retain more MPs compared to finer-leave species like H. uninervis 6 . In contrast, Table 1 indicates that no standardized protocol or globally accepted manual exists for defining the relative abundance of MP units in different types of samples.
A total of 240 MP particles were identified in surface water samples collected from six seagrass beds along the Southern and Eastern coasts of Sri Lanka. The average MP concentration ranged from 6.43 ± 0.71 items/m 3 at Kalkudah to 15.71 ± 1.24 items/m 3 at Salli Beach, with an overall average of 9.60 ± 0.97 items/m 3 (Fig. 2 b). Statistically, significant differences were observed in MP concentrations among the six locations ( p < 0.05). Among the Southern sites, Dondra recorded the highest MP load (8.81 ± 1.04 items/m 3 ), while Ahangama had the lowest average value of 7.38 ± 0.24 items/m 3 . Interestingly, even relatively remote locations such as Kalkudah had measurable levels of MP contamination, likely due to wind-driven dispersion, ocean currents, and land runoff 37 . These findings could be linked with both anthropogenic activities (e.g. waste disposal practices, fisheries) and environmental factors (e.g. wind, currents) that influence MP distribution in coastal ecosystems. Since sampling was conducted using a 300 μm manta net, the reported concentrations represent only MPs larger than 300 μm, and smaller particles were not captured in this study. The observed MP concentrations in surface water align with previous studies conducted in Sri Lanka, such as 9.8 ± 7.6 items/m 3 in reef ecosystems 48 , 0–29 items/m 3 in Southern coastal waters 49 , and 17.45 ± 3.35 items/m 3 in Southern coastal waters 22 . However, some studies reported lower values like 0.85 items/m 3 in the northwestern coast, and 0.515 ± 0.054 items/m 3 in Hikkaduwa and Bundala areas 39 . Similarly, international studies also report a broader range, such as 0.33 ± 0.34 items/m 3 in the Bohai Sea, China 5 , 3.74 items/m 3 in South Japan 4 , and 7.5 ± 1.5 items/m 3 in Orkney, Scotland 3 . These regional differences can be attributed to a variety of factors, including waste management practices, marine activities, and hydrodynamic conditions such as current strength and water retention time. Notably, semi-enclosed or low-energy waters can retain MPs more effectively than open or turbulent coastlines. Furthermore, differences in methodologies, such as mesh sizes and sampling timing, can substantially influence measured concentrations. Therefore, comparisons among studies should be interpreted with caution.
A total of 251 MPs were detected in vegetated surface sediments across the six sampling sites. The MP abundance ranged from 46.67 ± 4.41 to 96.67 ± 9.28 items/kg, with a mean of 69.72 ± 8.38 items/kg (Fig. 2 c). Significant differences ( p < 0.05) in MP concentrations were observed among the sites. On the Southern coast, Dondra recorded the highest MP abundance (75.00 ± 13.23 items/kg), while Ahangama recorded the lowest value of 63.33 ± 8.33 items/kg. On the Eastern coast, Salli Beach exhibited the highest MP content (96.67 ± 9.28 items/kg), whereas Kalkudah recorded the lowest value of 46.67 ± 4.41 items/kg. In contrast, non-vegetated sediment samples contained a total of 121 MP particles, with abundances ranging from 21.67 ± 1.67 to 51.67 ± 1.67 items/kg and a mean of 33.61 ± 2.18 items/kg (Fig. 3 ). Therefore, MPs were significantly abundant in vegetated sediments than in non-vegetated sediments ( p < 0.05), emphasizing the role of seagrass beds in MP retention. Similarly, MP content is higher in vegetated sediments (62.8 ± 30.7 items/kg) than in non-vegetated sediments (29.8 ± 12.1 items/kg) of Puttalam Lagoon, Sri Lanka 34 . In addition, studies from Europe 3 , 19 have documented elevated MP levels in vegetated sediments (300 ± 30 MP/kg in Scotland and 215 ± 163 MP/kg in England), reinforcing the notion that seagrass meadows act as microplastic sinks. As shown in Table 1 , microplastic concentrations vary among countries, with vegetated sediments in Sri Lanka exhibiting significantly high levels, suggesting that seagrass beds enhance microplastic retention.
Microplastic abundance of sediments in the vegetated and non-vegetated sites. Data were analyzed using two-way ANOVA followed by Tukey’s post hoc test ( p < 0.05).
The enhanced retention capacity in vegetated sites can be attributed to the structural complexity of seagrass beds, which trap suspended particles, including MPs, through reduced flow velocity and increased surface roughness 12 . Accordingly, sheltered intertidal mudflats or seagrass-covered areas with minimal disturbance tend to accumulate more MPs than exposed or non-vegetated zones 42 . Sediment characteristics such as mud content and grain size also influence MP accumulation. In general, finer sediments can retain higher MP concentrations 50 , 51 . In the present study, no significant correlation was observed between sediment grain size and MP abundance. However, this lack of correlation likely reflects the restricted sediment variability across the selected sampling sites, which were predominantly composed of medium to coarse sand (Table 2 ). This narrow range of grain size distribution may have limited the statistical power to detect a true sediment-MP relationship. Therefore, MP retention can be influenced by local hydrodynamics and human activity. In addition, anthropogenic influences such as intensive fishing activities, recreational use, and domestic waste inputs are dominant at high MP sites like Salli Beach and Dondra 33 . Remote sites like Kalkudah had the lowest MP concentration. Consequently, site-specific characteristics provide valuable insight into MP distribution in coastal Sri Lankan sediments, and further investigation is required to clarify the influence of hydrodynamic conditions on MP retention.
A total of 209 MP particles were detected in Holothuria atra across the six seagrass sites. This study is one of the first to offer quantitative evidence of MPs in sea cucumbers from Sri Lanka, helping to fill a significant gap in existing data. The average MP abundance ranged from 111.18 ± 15.48 to 258.16 ± 25.00 items/kg, with a mean of 177.19 ± 53.28 items/kg (Fig. 2 d). Significant differences ( p < 0.05) in MP concentrations were observed among the six locations. On the Southern coast, Dondra, Dickwella, and Ahangama exhibited MP abundances of 232.23 ± 26.89, 210.28 ± 37.84, and 134.26 ± 22.86 items/kg, respectively (Fig. 2 d). On the eastern coast, Salli Beach had the highest MP load (258.16 ± 25.00 items/kg), while Kalkudah and Adukkuparu showed relatively lower levels of 111.18 ± 15.48 and 117.04 ± 22.36 items/kg, respectively. These findings highlight the susceptibility of deposit-feeding organisms like H. atra to MP ingestion in seagrass ecosystems. The high MP loads are linked to feeding behavior such as feeding rate, sediment MP load, and local habitat conditions. For example, H. atra ingests surface sediments using oral tentacles and inadvertently consumes MPs entangled in organic detritus 28 , 50 .
The comparatively higher MP concentrations in sea cucumbers from fishing-active areas such as Dondra and Salli Beach suggest a direct influence of anthropogenic inputs, likely from fishing gear and coastal littering. In contrast, relatively lower values in remote or recreational areas (e.g. Kalkudah and Adukkuparu) may reflect reduced exposure. Similarly, research in Malaysia found MP ingestion (approximately 60–70 particles in just 4–5 g) in Stichopus horrens gut content 28 . Furthermore, Table 1 summarizes microplastic concentrations in various sea cucumber species associated with seagrass beds across different countries. Beyond ecological concerns, the presence of MPs in sea cucumbers carries serious food safety implications. For example, H. atra is an economically significant species in Sri Lanka and the broader Indo-Pacific region, harvested for domestic use and export, especially to East Asian countries where it is prized as both a delicacy and a component of traditional medicine 52 . Therefore, MP ingestion raises potential risks not only to human consumers but also to higher trophic organisms that prey on sea cucumbers 53 . However, the ecotoxicological effects of MPs on internal organs such as the respiratory tree or coelomic fluid remain underexplored. Future research is thus required to examine MP translocation within tissues and physiological impacts, especially under varying environmental conditions and seagrass habitat structures 8 .
Figure 4 shows the characteristics (i.e. shape, color, size, and polymer composition) of retrieved MPs from seagrass blades, surface water, surface sediments, and sea cucumbers found from all sampling sites. The collected MPs exhibited considerable variability in terms of shape, such as fibers, fragments, and films. Fibers were the most dominant in all types of samples (Fig. 4 a). Fibers accounted for 80.69% of MPs in seagrass blades, 57.08% in surface water, 56.18% in surface sediments, and 53.11% in sea cucumbers (Fig. 4 a). This dominance is consistent with previous findings in marine environments, where fibrous MPs are commonly linked to sources like fishing gear, ropes, and synthetic textiles. This fine and thread-like morphology (Fig. 5 ) increases the likelihood of ingestion during sediment filtering 42 , 54 . Their lightweight and elongated structure allows them to be easily transported by water currents, contributing to their widespread distribution in coastal environments 8 . Microplastic color diversity encompassed nine distinct colors, with blue being the most prevalent, and accounting for over 40% of all MPs in each sample type (Fig. 4 b). The abundance of blue-colored MPs has also been reported elsewhere 3 , 28 , 55 , probably from ropes, nets, and containers used in fisheries and coastal activities. Microplastics were categorized into small (0.06–0.99 mm) and large (1.00–4.56 mm) particles. Small MPs dominated seagrass blades (69.66%) and sea cucumbers (58.85%), whereas larger MPs were more prevalent in surface water (60.00%) and sediments (61.75%) (Fig. 4 c). This size distribution may reflect differences in retention capacity and biological interaction. For example, biofilms on seagrass blades can preferentially capture smaller MPs, while sedimentation processes can trap larger particles. The predominance of small MPs is ecologically important due to their likelihood of ingestion by marine organisms, especially benthic invertebrates like sea cucumbers.
Distribution of MP categories according to ( a ) shape, ( b ) color, ( c ) size, and ( d ) polymer type.
Morphological variation of microplastics identified in the present study: fibers ( a – d ), fragments ( e – g ), and films ( h – i ). Microplastics were isolated from various environmental matrices, including vegetated sediments ( a , e , h ), seagrass blades ( c , i ), and sea cucumbers ( d , g ). Scale bars represent 0.5 mm.
The chemical composition of MPs, determined through FTIR analysis, revealed a predominance of low-density polyethylene (LDPE) in all sample types, particularly in seagrass blades (75%) and sea cucumbers (75%), followed by polyethylene terephthalate (PET), high-density polyethylene (HDPE), polypropylene (PP), and polyamide or commonly called nylon (PA) in varying proportions (Fig. 4 d). NaCl-based separation may preferentially recover low-density polymers. Therefore, sediment polymer composition may be biased toward lighter materials. The detection of PET in the present study may be attributed to particle weathering, fiber morphology, and suspension effects during agitation, which may allow partial recovery of higher-density particles. The identity of these polymers was verified by their characteristic spectral analysis (Fig. 6 ). LDPE and HDPE were identified by the prominent doublet of CH 2 stretching vibrations at 2914 cm −1 and 2847 cm −1 , while PET was confirmed by the sharp carbonyl (C=O) stretching at 1713 cm −1 .
Representative ATR-FTIR spectra of microplastic polymers ( a ) Low-density polyethylene (LDPE), ( b ) Polyethylene Terephthalate (PET), ( c ) Polypropylene (PP), ( d ) High-density polyethylene (HDPE), and ( e ) Polyamide (PA) identified across all sample sites. Characteristic absorption bands used for polymer verification are indicated by wavenumbers (cm −1 ).
Low-density polyethylene (LDPE) is frequently found in the environment because it is widely used in everyday items like plastic bags, bottle caps, and fishing gear. Polypropylene (PP) and polyamide (PA) were also detected, often linked to fishing ropes and nets 22 . The predominance of fibrous, blue-colored, and LDPE-based MPs is consistent with patterns observed in other tropical and subtropical coastal regions like Asia, such as Thailand, Malaysia, and China 8 , 23 , 42 highlighting shared pollution sources and mechanisms across these regions. However, in this study, we acknowledge that FTIR analysis was limited to a subset of larger particles, which may introduce some bias in the reported polymer distribution and may not fully reflect the composition of smaller particles and fibers present in the samples.
The correlation analysis also demonstrated important ecological connections among MP distributions in different sample types (Fig. 7 ). A strong positive correlation was found among seagrass blades and both sediments ( r = 0.829, p = 0.001) and surface water ( r = 0.793, p = 0.001), indicating that MPs in the water column and on the seabed are likely influencing MP loads on seagrass. Moderate correlations between surface water and sea cucumbers ( r = 0.502, p = 0.034) and weaker correlations between seagrass blades and sea cucumbers ( r = 0.433, p = 0.073) were not statistically significant. These patterns suggest possible trophic transfer and environmental overlap but also point to selective ingestion or retention differences in biota. Furthermore, these patterns can be explained by the persistence of plastics in marine environments due to their durability and resistance to degradation. Processes such as physical degradation, photodegradation, chemical degradation, and biodegradation break plastics into MP fragments in the environment 30 . For example, exposure to UV radiation leads to surface cracking, while wave action accelerates fragmentation, resulting in the prevalence of plastics in smaller size fractions like MPs and nano plastics. These degradation processes can alter surface chemistry, potentially affecting the detection and classification of MPs, and thereby may mislead quantitative assessments. Therefore, these results highlight that the complex dynamics of MP distribution within seagrass ecosystems may be shaped by physical transport, polymer properties, degradation, and biological interactions.
Correlation matrix across MP abundances in surface water, surface sediments, seagrass blades, and sea cucumbers.
Ecological risk indices revealed spatial variation in microplastic contamination across seagrass ecosystems (Table 3 ). Microplastic Pollution Load Index (MPLI) values indicated comparatively higher pollution levels at eastern coastal sites, particularly Salli, corresponding with elevated microplastic abundance observed in multiple environmental matrices. Lower MPLI values recorded at southern sites suggest relatively reduced contamination intensity. The Polymer Hazard Index (PHI) indicated moderate ecological risk associated with identified polymers, reflecting the dominance of relatively low-hazard polymers such as polyethylene and polypropylene, although the presence of higher-risk polymers, including PET, contributed to increased hazard scores. Similar polymer-risk patterns have been reported in coastal microplastic studies globally 31 .
Polymer diversity assessed using MDII depicts moderate diversity across sampling locations, suggesting multiple microplastic sources influencing seagrass ecosystems. Higher diversity values at eastern sites may reflect increased anthropogenic inputs and hydrodynamic transport processes.
Bioaccumulation Factor (BAF) values exceeding unity at several sites indicate effective uptake of microplastics by Holothuria atra , consistent with its deposit-feeding behavior and continuous interaction with sediment particles. However, variability among sites suggests that accumulation is influenced by local environmental availability rather than uniform trophic transfer efficiency. Therefore, the integration of abundance data with ecological risk indices provides a more comprehensive understanding of microplastic impacts within Sri Lankan seagrass ecosystems.
This study examined MP pollution in selected seagrass ecosystems from the Southern and Eastern coasts of Sri Lanka. All samples (seagrass blades, surface water, sediment, and sea cucumber samples) confirmed the accumulation of a significant number of MPs. Fibers of blue and LDPE polymers were the most common MPs found in most of the seagrass beds. The presence of MPs in seagrass blades and sea cucumbers highlights the potential for the transport of MPs through marine food webs. Higher MP values were recorded at sites subjected to anthropogenic activities, such as fishing and tourism. This study provides evidence for the ability of MPs to accumulate within seagrass beds, based on ecological pollution indices. Our findings suggest a potential risk to the health of seagrass ecosystems and highlight the importance of an integrated evaluation of contamination intensity amidst continuous plastic inputs. Therefore, Sri Lanka needs to develop better waste management practices, strengthen policies for alternatives to conventional plastics, and reduce the amount of plastic that enters the marine environment. As a recommendation, future research should focus on understanding the possible impacts of this MP accumulation on seagrass beds under controlled conditions and the transport pathways of MPs through seagrass food webs.
The datasets used and/or analyzed during the current study are available from the corresponding author on reasonable request.
Pilapitiya, P. G. C. N. T. & Ratnayake, A. S. The world of plastic waste: a review. Clean. Mater. 11 , 100220. https://doi.org/10.1016/j.clema.2024.100220 (2024).
Ratnayake, A. S., Gunawardhana, G. M. S. S. & Perera, U. L. H. P. The biography of microplastics (MPs): occurrences, sources, weathering/degradation, characterization, ecological/human risks, removal methods, policy development, and current trends and future perspectives. Anthr Coasts . 7 , 30. https://doi.org/10.1007/s44218-024-00064-8 (2024).
Jones, K. L., Hartl, M. G., Bell, M. C. & Capper, A. Microplastic accumulation in a Zostera marina L. bed at Deerness Sound, Orkney, Scotland. Mar. Pollut Bull. 152 , 110883. https://doi.org/10.1016/j.marpolbul.2020.110883 (2020).
Isobe, A., Uchida, K., Tokai, T. & Iwasaki, S. East Asian seas: a hot spot of pelagic microplastics. Mar. Pollut Bull. 101 , 618–623. https://doi.org/10.1016/j.marpolbul.2015.10.042 (2015).
Zhang, W. et al. Microplastic pollution in the surface waters of the Bohai Sea, China. Environ. Pollut . 231 , 541–548. https://doi.org/10.1016/j.envpol.2017.08.058 (2017).
Sawalman, R. et al. Microplastic abundance in sea urchins ( Diadema setosum ) from seagrass beds of Barranglompo Island, Makassar, Indonesia. IOP Conf. Ser. : Earth Environ. Sci. 763 , 012057. https://doi.org/10.1088/1755-1315/763/1/012057 (2021).
Huang, Y. et al. Seagrass beds acting as a trap of microplastics: emerging hotspot in the coastal region? Environ. Pollut . 257 , 113450. https://doi.org/10.1016/j.envpol.2019.113450 (2019).
Kongsap, V., Rattanachot, E., Prathep, A., Buaphol, W. & Mayakun, J. A high abundance of Holothuria (Halodeima) atra (Holothuroidea Aspidochirotida) in a Halimeda-dominated habitat. J. Mar. Sci. Eng. 11 , 451. https://doi.org/10.3390/jmse11020451 (2023).
Kreitsberg, R. et al. Seagrass beds reveal a high abundance of microplastic in sediments: a case study in the Baltic Sea. Mar. Pollut Bull. 168 , 112417. https://doi.org/10.1016/j.marpolbul.2021.112417 (2021).
Sanchez-Vidal, A., Canals, M., De Haan, W. P., Romero, J. & Veny, M. Seagrasses provide a novel ecosystem service by trapping marine plastics. Sci. Rep. 11 , 254. https://doi.org/10.1038/s41598-020-79370-3 (2021).
Article ADS CAS PubMed PubMed Central Google Scholar
Gross, H., Jaskiel, J. & Rotjan, R. Thalassia testudinum as a potential vector for incorporating microplastics into benthic marine food webs. Mar. Pollut Bull. 135 , 1085–1089. https://doi.org/10.1016/j.marpolbul.2018.08.024 (2018).
Ameen, H. M., Rohini, P., Madhu, M., Rajani, V. & Jayadev, A. Impact of microplastic (MP) pollution in seagrass ecosystem: a review. Curr. World Environ. 19 , 5. https://doi.org/10.12944/CWE.19.1.5 (2024).
Lin, X., Gowen, A. A., Pu, H. & Xu, J. L. Microplastic contamination in fish: critical review and assessment of data quality. Food Control . 153 , 109939. https://doi.org/10.1016/j.foodcont.2023.109939 (2023).
Alabhai, J. M., Pramath, H. P., Hanumanaika, R. N. & Sangeetha, J. Accumulation of microplastic in edible marine species from North Kerala, India. Environ. Monit. Assess. 197 , 607. https://doi.org/10.1007/s10661-025-14048-9 (2025).
Watts, A. J. et al. Uptake and retention of microplastics by the shore crab Carcinus maenas . Environ. Sci. Technol. 48 , 8823–8830. https://doi.org/10.1021/es501090e (2014).
Article ADS CAS PubMed Google Scholar
Feng, Z. et al. Microplastics in specific tissues of wild sea urchins along the coastal areas of northern China. Sci. Total Environ. 728 , 138660. https://doi.org/10.1016/j.scitotenv.2020.138660 (2020).
Bulleri, F., Ravaglioli, C., Anselmi, S. & Renzi, M. The sea cucumber Holothuria tubulosa does not reduce the size of microplastics but enhances their resuspension in the water column. Sci. Total Environ. 781 , 146650. https://doi.org/10.1016/j.scitotenv.2021.146650 (2021).
Sparks, C., Awe, A. & Maneveld, J. Abundance and characteristics of microplastics in retail mussels from Cape Town, South Africa. Mar. Pollut Bull. 166 , 112186. https://doi.org/10.1016/j.marpolbul.2021.112186 (2021).
Unsworth, R. K. Canopy accumulation: are seagrass meadows a sink of microplastics? Oceans 2 , 162–178. https://doi.org/10.3390/oceans2010010 (2021).
Veettil, B. K. et al. Blue carbon ecosystems in Sri Lanka: a review. Estuar. Coast Shelf Sci. 306 , 108907. https://doi.org/10.1016/j.ecss.2024.108907 (2024).
Udagedara, S. & Dahanayaka, D. D. G. L. Current status and checklist of seagrass in Sri Lanka. Int. J. Aquat. Biol. 8 , 317–326. https://doi.org/10.22034/ijab.v8i5.619 (2020).
Athapaththu, A. M. et al. Plastics in surface water of southern coastal belt of Sri Lanka (Northern Indian Ocean): distribution and characterization by FTIR. Mar. Pollut Bull. 161 , 111750. https://doi.org/10.1016/j.marpolbul.2020.111750 (2020).
Zhang, B. et al. Microplastic pollution in the surface sediments collected from Sishili Bay, North Yellow Sea, China. Mar. Pollut Bull. 141 , 9–15. https://doi.org/10.1016/j.marpolbul.2019.02.021 (2019).
Husin, M. J., Mazlan, N., Shalom, J., Saud, S. N. & Abdullah Sani, M. S. Evaluation of microplastics ingested by sea cucumber Stichopus horrens in Pulau Pangkor, Perak, Malaysia. Environ. Sci. Pollut Res. 28 , 61592–61600. https://doi.org/10.21203/rs.3.rs-207775/v1 (2021).
Dissanayake, D. C. T., Athukorala, S. & Amarasiri, C. Present status of the sea cucumber fishery in Sri Lanka. SPC beche-de-mer Inf. Bull. 30 , 14–20 (2010).
Purcell, S. W., Samyn, Y. & Conand, C. Commercially important sea cucumbers of the world. FAO Species Catalogue for Fishery Purposes 6, 150 pp; ISBN 978-92-5-106719-2 (2012).
Masura, J. et al. Laboratory methods for the analysis of microplastics in the marine environment: recommendations for quantifying synthetic particles in waters and sediments. (2015). https://repository.library.noaa.gov/view/noaa/10296
Mohsen, M. et al. Microplastic ingestion by the farmed sea cucumber Apostichopus japonicus in China. Environ. Pollut . 245 , 1071–1078. https://doi.org/10.1016/j.envpol.2018.11.083 (2019).
Perera, U. L. H. P., Subasinghe, H. C. S., Ratnayake, A. S., Weerasingha, W. A. D. B. & Wijewardhana, T. D. U. Maritime pollution in the Indian Ocean after the MV X-Press Pearl accident. Mar. Pollut Bull. 182 , 114301. https://doi.org/10.1016/j.marpolbul.2022.114301 (2022).
Gunawardhana, G. M. S. S., Perera, U. L. H. P., Ratnayake, A. S., Weerasingha, W. A. D. B. & Subasinghe, H. C. Formation of secondary microplastics during degradation of plastics originating from the MV X–Press Pearl maritime disaster. Discov Environ. 2 , 22. https://doi.org/10.1007/s44274-024-00044-2 (2024).
Lithner, D., Larsson, Å. & Dave, G. Environmental and health hazard ranking and assessment of plastic polymers based on chemical composition. Sci. Total Environ. 409 , 3309–3324. https://doi.org/10.1016/j.scitotenv.2011.04.038 (2011).
Li, J., Liu, H. & Chen, J. P. Microplastics in freshwater systems: a review on occurrence, environmental effects, and methods for microplastics detection. Water Res. 137 , 362–374. https://doi.org/10.1016/j.watres.2017.12.056 (2018).
Jayalath, G. N. et al. Status of plastic waste generation and propose a circular economy-based policy framework to manage plastic waste in Sri Lanka. J. Environ. Sci. Toxicol. Food Technol. 15 , 24–38 (2021).
Bhathiya, M. A. M. et al. Seagrass beds may act as a trap of microplastics: a case study at Puttalam Lagoon, Sri Lanka. Proc Int For Environ Symp. 28: 77 (2023). https://doi.org/10.31357/fesympo.v28.6956
Rummel, C. D., Jahnke, A., Gorokhova, E., Kühnel, D. & Schmitt-Jansen, M. Impacts of biofilm formation on the fate and potential effects of microplastic in the aquatic environment. Environ. Sci. Technol. Lett. 4 , 258–267. https://doi.org/10.1021/acs.estlett.7b00164 (2017).
Datu, S. S., Supriadi, S. & Tahir, A. Microplastic in Cymodocea rotundata seagrass blades. Int. J. Agric. Environ. Biotechnol. 4 , 1758–1761. https://doi.org/10.22161/ijeab.46.35 (2019).
Priscilla, V., Sedayu, A. & Patria, M. P. Microplastic abundance in the water, seagrass, and sea hare Dolabella auricularia in Pramuka Island, Seribu Islands, Jakarta Bay, Indonesia. J. Phys. : Conf. Ser. 1402 , 033073. https://doi.org/10.1088/1742-6596/1402/3/033073 (2019).
Seng, N. et al. Early evidence of microplastics on seagrass and macroalgae. Mar. Freshw. Res. 71 , 922–928. https://doi.org/10.1071/MF19292 (2020).
Dharmadasa, W. S., Andrady, A. L., Kumara, P. T. P., Maes, T. & Gangabadage, C. S. Microplastic pollution in marine protected areas of Southern Sri Lanka. Mar. Pollut Bull. 168 , 112462. https://doi.org/10.1016/j.marpolbul.2021.112462 (2021).
Jeyasanta, K. I., Patterson, J., Grimsditch, G. & Edward, J. P. Occurrence and characteristics of microplastics in the coral reef, seagrass and nearshore habitats of Rameswaram Island, India. Mar. Pollut Bull. 160 , 111674. https://doi.org/10.1016/j.marpolbul.2020.111674 (2020).
Mahagamage, M. G. Y. L. et al. Water quality and floating microplastics pollution along Southern to Northwestern Coast, Sri Lanka. Sri Lanka J. Aquat. Sci. 30 , 37–52. https://doi.org/10.4038/sljas.v30i1.128 (2025).
Rasyid, N., Andryan, D., Bengen, D. G. & Subhan, B. Assessment of marine debris in seagrass beds of Pramuka Island, Kepulauan Seribu. IOP Conf. Ser. : Earth Environ. Sci. 967 , 012014. https://doi.org/10.1088/1755-1315/967/1/012014 (2022).
Esmeralda, G. & Patterson, J. Accumulation of microplastics in relation to the sediment textural characters of seagrass meadows in the coastal areas of Tuticorin, Southeast Coast of India. Mar. Pollut Bull. 212 , 117557. https://doi.org/10.1016/j.marpolbul.2025.117557 (2025).
Riani, E. & Cordova, M. R. Microplastic ingestion by the sandfish Holothuria scabra in Lampung and Sumbawa, Indonesia. Mar. Pollut Bull. 175 , 113134. https://doi.org/10.1016/j.marpolbul.2021.113134 (2022).
Idris, F., Febrianto, T., Hidayati, J. R. & Nugraha, A. H. Microplastic abundance in sea cucumber at seagrass ecosystem of Bintan Island and surrounding area, Indonesia. IOP Conf. Ser. Earth Environ. Sci. 967 , 012009. https://doi.org/10.1088/1755-1315/967/1/012009 (2022).
Umasangaji, H. & Ramili, Y. The occurrence of microplastics in the gastrointestinal tract of sea cucumbers on the conservation area of Mare Island, North Maluku, Indonesia. Aquacult. Aquarium Conserv. Legis . 15 , 2822–2839 (2022).
Fong, S. et al. Widespread microplastic contamination across Singapore’s coral reef, mangrove and seagrass ecosystems. Sci. Total Environ. 939 , 289717. https://doi.org/10.1016/j.scitotenv.2025.289717 (2025).
Hansani, K. U. D. N. et al. Contamination of microplastics in tropical coral reef ecosystems of Sri Lanka. Mar. Pollut Bull. 194 , 115299. https://doi.org/10.1016/j.marpolbul.2023.115299 (2023).
Koongolla, J. B., Andrady, A. L., Kumara, P. T. P. & Gangabadage, C. S. Evidence of microplastics pollution in coastal beaches and waters in southern Sri Lanka. Mar. Pollut Bull. 137 , 277–284. https://doi.org/10.1016/j.marpolbul.2018.10.031 (2018).
Hidalgo-Ruz, V., Gutow, L., Thompson, R. C. & Thiel, M. Microplastics in the marine environment: a review of the methods used for identification and quantification. Environ. Sci. Technol. 46 , 3060–3075. https://doi.org/10.1021/es2031505 (2012).
Boshoff, B. J., Robinson, T. B. & von der Heyden The role of seagrass meadows in the accumulation of microplastics: insights from a South African estuary. Mar. Pollut Bull. 186 , 114403. https://doi.org/10.1016/j.marpolbul.2022.114403 (2023).
Purcell, S. W., Williamson, D. H. & Ngaluafe, P. Chinese market prices of beche-de-mer: implications for fisheries and aquaculture. Mar. Policy . 91 , 58–65. https://doi.org/10.1016/j.marpol.2018.02.005 (2018).
Francour, P. Predation on holothurians: a literature review. Invertebr Biol. 116 , 52–60. https://doi.org/10.2307/3226924 (1997).
Zhu, Z. et al. Structural features and digestive behavior of fucosylated chondroitin sulfate from sea cucumbers Stichopus japonicus . J. Agric. Food Chem. 67 , 10534–10542. https://doi.org/10.1021/acs.jafc.9b04996 (2019).
Linders, T. et al. Particle sources and transport in stratified Nordic coastal seas in the Anthropocene. Elem. Sci. Anth . 6 , 29. https://doi.org/10.1525/elementa.149 (2018).
The authors would like to acknowledge the Research Committee, Uva Wellassa University of Sri Lanka.
The Authors received no funding for this work.
Department of Animal Science, Faculty of Animal Science and Export Agriculture, Uva Wellassa University, Passara Road, Badulla, 90000, Sri Lanka
H. M. Madushani, E. P. D. N. Thilakarathne, Anjula Sachintha Mahaliyana, Sakuni Dhananja Samarakoon, Chathuni Sulochana Divyanjalee & Madushi Anuththara Jayalath
School of Chemistry and Physics, Faculty of Science, Queensland University of Technology (QUT), GPO 2434, Brisbane, QLD, 4001, Australia
Department of Applied Earth Sciences, Faculty of Applied Sciences, Uva Wellassa University, Passara Road, Badulla, 90000, Sri Lanka
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H. M. M.: conceptualization, microplastic analysis, writing original draft, visualization, statistical analysis. D. N. T.: supervision, conceptualization, project conceptualization, project administration, reviewing, editing, and funds acquisition. J. B. K.: writing, reviewing, editing, and methodology. A. S. M.: visualization, data curation. S. D. S.: investigation, methodology, writing-review and editing. C. S. D.: visualization, data curation. M. A. J.: methodology, data curation. A. S. R.: reviewing, editing, visualization, data curation, supervision.
Correspondence to Amila Sandaruwan Ratnayake .
Approval was obtained from the Research Ethics Committee of Uva Wellassa University of Sri Lanka.
The authors declare no competing interests.
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Madushani, H.M., Thilakarathne, E.P.D.N., Koongolla, J.B. et al. Distribution and characteristics of microplastics in the seagrass ecosystem of Sri Lanka (Indian Ocean). Sci Rep 16 , 22723 (2026). https://doi.org/10.1038/s41598-026-53757-0
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