The distribution, characteristics and ecological risks of microplastics in the mangroves of Southern China

Journal Pre-proofs The distribution, characteristics and ecological risks of microplastics in the mangroves of Southern China Ruili Li, Lingyun Yu, Minwei Chai, Hailun Wu, Xiaoshan Zhu PII: DOI: Reference:

S0048-9697(19)35017-X https://doi.org/10.1016/j.scitotenv.2019.135025 STOTEN 135025

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Science of the Total Environment

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13 August 2019 11 October 2019 15 October 2019

Please cite this article as: R. Li, L. Yu, M. Chai, H. Wu, X. Zhu, The distribution, characteristics and ecological risks of microplastics in the mangroves of Southern China, Science of the Total Environment (2019), doi: https:// doi.org/10.1016/j.scitotenv.2019.135025

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The distribution, characteristics and ecological risks of microplastics in the mangroves of Southern China Ruili Lia, b*, Lingyun Yua, Minwei Chaia, Hailun Wua, Xiaoshan Zhuc

a

School of Environment and Energy, Peking University Shenzhen Graduate School, Shenzhen 518055, Guangdong, China.

b

ZTE Instruments Co.,Ltd.,Shenzhen 518000, Guangdong.

c

Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, P.R. China.

*Corresponding author. E-mail address: [email protected] (Ruili Li).

Abstract: During the production, use and disposal of plastic products, microplastics (MPs) are dispersed into the surrounding environment and have inevitable impacts on mangrove ecosystems in estuaries and offshore areas. In the mangroves of Southern China, the systematic evaluation of the distribution, characteristics and ecological risks of MPs is lacking. In this study, surface sediments (0-5 cm depth) were collected from six representative mangroves in China to explore MP contamination and its associated ecological risk. Based on the results, MP concentrations of MPs in mangrove sediments were as follows : FT (2249±747 items/kg), ZJ (736±269 items/kg), DF (649±443 items/kg), DZG (431±170 items/kg), YX (424±127 items/kg), and FCG (227±173 items/kg). The higher MP concentration in the Futian mangrove was mainly related to inputs from the Pearl River, the third largest river in China. The predominant shape, colour, and size of MPs were fibrous, white-transparent, and 500 -5,000 μm, respectively. The main MP polymer types were polypropylene, polyethylene, and polystyrene. Degradation artefacts were present on surface of MPs as well as metallic and non-metallic elements. MPs concentration in mangrove sediments increased with increasing social-economic development of surrounding districts, which indicated the clear influence of anthropogenic activity on MP pollution in these mangroves. Furthermore, total organic carbon (TOC) and silt content were positively associated with MPs (P < 0.01), indicating a facilitatory role in deposition of MPs in mangroves. Based on a comprehensive evaluation using the potential ecological risk factor (𝐸𝑖 ), potential ecological risk (RI), polymer 1

risk index (H) and pollution load index (PLI), MPs were found to present ecological risks in these mangroves, with the highest risk occurring in the Futian mangrove. Key words: mangroves; microplastics; ecological risk; impact factor; China 1 Introduction Plastic products are widely used in almost every aspect of daily life due to their durable and lightweight characteristics (Andrady and Meal, 2009). Globally, the production of plastics reached 350 million tonnes in 2017 and is expected to continue to increase in the future (Plastics Europe, 2018). Microplastic (MP) particles, with diameters of less than 5 mm, mainly originate from direct manufacturing and the fragmentation of larger plastics discarded in the environment (Fendall and Sewell, 2009; Andrady, 2011; Wright et al., 2013). MPs have been found in different aquatic environments and can be taken up and bioaccumulated by a range of marine organisms occupying different tropic levels (Lusher et al., 2015; Ding et al., 2019; Erni-Cassola et al., 2019). The marine environment has been the main focus of MP research in many countries, including Belgium (Claessens et al., 2011), China (Peng et al., 2017), Germany (Dekiff et al., 2014), Italy (Vianello et al., 2013), Singapore (Ng and Obbard, 2006), and the UK (Thompson et al., 2004). MPs can adversely affect organisms via physical damage that causes inflammation and bioaccumulation, and thereby limits metabolism (Von Moos et al., 2012; Rochman et al., 2014). MPs can also adsorb various pollutants from the surrounding environment, such as persistent organic pollutants and heavy metals, which enhances their potential toxicity (Chua et al., 2014; Brennecke et al., 2016). In addition, MPs can release internal toxic substances that are added during the manufacturing process, thereby posing additional threats to the surrounding environment (Gallo et al., 2018). Estuaries and other coastal ecosystems are productive for wildlife and provide various ecosystem services for humans (Barbier et al., 2011). Meanwhile, large quantities of domestic sewage produced by surrounding populations may be discharged into these systems, which act as an important source of MPs in marine 2

environments (Browne et al., 2011; Cesa et al., 2017). Generally, mangroves, seagrass beds and coral reefs, are distinct from other coastal ecosystems with respect to the transformation of pollutants because of their high productivity and biomass (Fourqurean et al., 2012; Li et al., 2016; Booth and Sear, 2018). In particular, mangroves can retain land-based pollutants and litter, and are thus receiving increasing research attention (e.g. Martin et al., 2019). The first concentrations of MPs in mangrove were reported as 63 items/kg in Singapore (Nor and Obbard, 2014). Since then, many studies have reported higher concentrations in other mangroves and salt marsh wetlands, with wetland vegetation appearing to be particularly effective at retaining MPs (Sutton et al., 2016; Weinstein et al., 2016; Li et al., 2018). China is the largest producer of plastics globally, accounting for approximately 26% of global production, and, consequently, is one of the largest producers of plastic litter entering the environment (Jambeck et al., 2015; Plastics Europe, 2018; Wang et al., 2019a). In the offshore oceans of China, MPs have mainly been documented in marine sediments, seawater and fishes, without considering the sink function of offshore ecosystems (Zhao et al., 2018; Zheng et al., 2019; Su et al., 2019). On the other hand, several models have been developed to evaluate the ecological risks posed by MPs in sediments (Peng et al., 2018), water columns (Xu et al., 2018) and the atmosphere (Liu et al., 2019). Yet, studies on ecological risk and the associated impact factors in mangrove systems have not been systematically conducted (Peng et al., 2018; Xu et al., 2018; Li et al., 2019). In this study, we selected six representative mangroves along the coast of China to study the characteristics of MP pollution. To avoid inconsistencies between different mangrove communities, Avicennia marina was selected as a pioneer mangrove species that is widely distributed along the coast of China (Pi et al., 2009; Shi et al., 2019). Stereo microscopy, scanning electron microscopy (SEM) and micro-Raman spectrometry were used to identify MPs. The main aims of the study were to: (1) investigate the concentrations and characteristics of MPs in the mangroves of China; (2) assess the ecological risks posed by MPs; and (3) explore the factors affecting the distributions of MPs in these mangrove systems. 3

2 Materials and methods 2.1 Study areas and sample collection The following representative mangroves were selected with A. marina as the dominant species: Dongzhaigang (DZG) and Dongfang (DF) in Hainan; Fangchenggang (FCG) in Guangxi; Futian (FT) and Zhanjiang (ZJ) in Guangdong; and Yunxiao (YX) in Fujian (Figure 1). General information for these mangroves including location, area, climate, and pollution characteristics are provided in Table S1 (Chai et al., 2019; Shi et al., 2019). In August, 2017, surface sediments (0-5 cm) were collected. In each case, two habitats were selected, namely mudflat (MF) and A. marina forest (AM). In both the MF and AM habitat types, three sampling sites were selected with a spacing of approximately 100 m along the shoreline and at distance of approximately 20 m from the forest edge. At each sampling site, a 5 m × 5 m area was established and three sediment samples (0-5 cm depth) were randomly collected during low tide. Each sample was then divided into two parts: one of which was used for the extraction and determination of MPs, and the other was completely air-dried to determine selected physico-chemical parameters including pH, total organic carbon (TOC) content, and grain size distribution. 2.2 Determination of physico-chemical parameters The methods followed for determining physic-chemical parameters of sediment samples were based on Chai et al. (2019). Briefly, the air-dried sediments were pre-treated by being ground, homogenised, and sifting through a 0.5 mm sieve. Sediment pH was measured using a pH meter (Sarturis PB-10, Germany) with a sediment:water ratio of 1:5 (w/v). Sediment TOC content was determined by combustion at 550 °C (Cambardella et al., 2001). Grain size was determined using a particle size analyser (Microtrac S3500, USA). Grain-size fractions were classified according to Gao and Chen (2012), as follows: sand (> 63 μm); silt (4-63 μm); and clay (< 4 μm). 4

2.3 Extraction and characterization of MPs MPs in the sediment were separated using zinc chloride density separation following the steps of density separation, oxidation digestion, and visual inspection (Townsend et al., 2019; Zheng et al., 2019). The modified procedure was as follows: (1) 5 g of air-dried sediment was put into a 50-ml centrifuge tube with 40 ml of ZnCl2 (1.5 g ml-1) to obtain a total volume of approximately 45 ml, which was shaken for 3 h. The centrifuge tube was then placed in an ultrasonic cleaner for 10 min, and centrifuged at 3,000 r/min for 10 min; (2) after filtration through a 0.45-μm filter membrane, the retained material was transferred to a 500-ml beaker with 200 ml of H2O2 and heated in water-bath at 100 °C for 3 h. When the temperature had dropped to room temperature (25 °C), the solution was filtered (0.45 μm) and transferred back into the beaker to continue digestion with the addition of more H2O2. After digestion was completed, the sample was placed into the drying box at 60 °C for 24 h; (3) the numbers of particles on the filters were detected using a stereo microscope equipped with an electronic eyepiece (Nikon SMZ 745, Japan). The observed MPs were classified into four classes (<50 μm, 50-100 μm, 100-500 μm and >500 μm) and four shape categories (fibre, granule, film and foam). During the sediment sampling process, pre-cleaned stainless steel spades and containers were used to avoid possible MP contamination. In the laboratory analysis, to avoid contamination, all instruments and vessels were rinsed three times with distilled water. All researchers wore laboratory coats and gloves during the analysis. Blank tests were conducted in which no MPs were found. A micro-Raman spectrometer was used for identifying the MP polymer types (Li et al., 2019; Sobhani et al., 2019).For this, all suspected and representative particles extracted from the sediments were placed on quartz glass slides and identified using a micro-Raman spectrometer (iHR320, Horiba, Japan). Laser energy, laser wavelength, exposure time and the emission wave number were 15 mW (5%), 785 nm, 2.0 s, and 1304000 cm-1, respectively. For each analysis, three microzones were randomly selected and averaged to obtain 5

a final estimate. The morphological structure and chemical composition of typical MPs (n = 10 per sample) extracted from the sediments were also analysed using scanning electron microscopy (SEM) and energy-dispersive X-ray spectroscopy (DES) (JEOL JSM-7800F, Japan). For this, the MPs were first coated with a thin film of evaporated gold prior to analysis. 2.4 Ecological risk assessment Several preliminary assessment methods were applied to assess the ecological risks posed by MPs. A modified assessment model for potential ecological risk was calculated as follows: 𝐶

𝐸𝑖 =𝑇𝑖 ×(𝐶 𝑖 ), RI=∑𝑛𝑖=1 𝐸𝑖

(1)

0

where 𝐸𝑖 and RI are the potential ecological risk factor and the potential ecological risk, respectively; 𝑇𝑖 is the chemical toxicity coefficient for the constituent polymer (Lithner et al., 2011); and 𝐶𝑖 /𝐶0 is quotient for the observed MP concentration versus the background level. Due to a lack of available background data, the lowest MP concentration measured in this study was adopted as the background value. Obtaining values of 𝐸𝑖 and RI were not the main focus of this study but instead were used for further assessment. The different categories of 𝐸𝑖 and RI are shown in Table S2. The hazard scores of plastic polymers and polymer types were used as indices to assess the risks of MPs (Lithner et al., 2011), as follows: H=∑ 𝑃𝑛 × 𝑆𝑛

(2)

where H is the calculated polymer risk index; 𝑃𝑛 is the percentage of each MP polymer type at each sampling site; and 𝑆𝑛 is the score for the polymers comprising the MPs. The 𝑆𝑛 values of polypropylene (PP), polyethylene (PE), and polystyrene (PS) were 1, 11 and 30, respectively (Lithner et al., 2011). The pollution load index (PLI) has been widely employed for assessing pollution levels in estuaries (Tomlinson et al., 1980), and is calculated using the following formula:

6

𝐶𝐹𝑖 = 𝐶𝑖 /𝐶𝑜𝑖 , PLI=√𝐶𝐹𝑖

(3)

where the PLI of MPs in each mangrove was calculated based on the respective MP concentration factors (𝐶𝐹𝑖 ), which are the quotients of the MP concentration in each mangrove and the minimal MP concentration (𝐶𝑜𝑖 ). In this study, the lowest MP concentration was taken to be 𝐶𝑜𝑖 . The criteria for the risk levels based on the risk index and pollution load index are shown in Table S2. 2.5 Statistical analysis MP concentration were expressed in terms of items per unit mass of dry sediment (items/kg). Mapping of the distribution of MPs was conducted using ArcGIS10.2 (ESRI, Redlands, CA). Linear regression analysis among MP concentration and selected physico-chemical parameters was also performed. Furthermore, principal component analysis (PCA) and hierarchical cluster analysis (HCA) were carried out to explore the contamination characteristics of MPs. All statistical tests were implemented using SPSS 18.0 software (SPSS Inc, Chicago, IL, USA). 3 Results and discussion 3.1 Microplastics in mangrove sediments Marine sediments have been found to function as important environmental media and major sinks of plastics in MP pollution surveys (Cózar et al., 2014; Woodall et al., 2014). In this study, MP concentrations were explored in surface sediments (0-5 cm depth) collected from six mangroves in China. Figure 2 showed photographs of the MPs typically found in all of the sediments, being consistent with previous reports of the widespread distribution of MPs in the marine environment (Browne et al., 2011; Eriksen et al., 2014). As shown in Figure 3A, MP concentrations were FT (2249±747 items/kg), ZJ (736±269 items/kg), DF (649±443 items/kg), DZG (431±170 items/kg), YX (424±127 items/kg) and FCG (227±173 items/kg). The MP concentration in the FT mangrove was one order of magnitude higher than that of the other mangroves. This 7

mangrove is located in the Pearl River Estuary (PRE), which received significant riverine inputs of pollutants, with estimated midpoint mass inputs of plastic waste higher than 1.1 × 105 tons/year (Lebreton et al., 2017). Furthermore, the Pearl River is the third largest river in China and is an important source of various organic and heavy metal pollutants from the surrounding areas that are home to an estimated 60 million people (Lo et al., 2018). Vianello et al. (2013) also found that anthropogenic micro-pollutant sources are associated with the distribution of MPs in Venice lagoon, Italy. In the Beibu Gulf of China, MP concentrations in mangroveadjacent zones (170-990 items/kg) were found to be lower than in the mangrove sediments (1,780-2,310 items/kg), which indicated the effective retention ability of mangrove plants with respect to MPs (Li et al, 2019). In the present study, with the exception of the FT mangrove, MP concentrations in the AM habitat type were higher or at least comparable to the MF habitat type (Figure 3A). Indeed, the FT mangrove is located in the semi-closed Shenzhen Bay on the east side of the PRE with a low water exchange capacity (Figure S1). We suggest that the MPs in the FT mangrove were mainly marine derived from the PRE and are related to the particular hydrographic characteristics of Shenzhen Bay, China. As shown in Table 1, the ranges of MP concentrations in mangroves of FT, ZJ, and DF were lower than that of Qinzhou Bay (15-12852 items/kg, Li et al., 2018), similar to that of South Yellow Sea (560-4205 items/kg, Wang et al., 2019b) and coastal areas of Hong Kong (0.58-2116 items/kg, Lo et al., 2018), and higher than some other areas in China and elsewhere, with MPs in DZG, YX, and FCG mangroves to be relatively lower. 3.2 Microplastic characteristics MP colour is an essential factor affecting its ingestion by marine organisms, probably due to colour-driven predator capturing behaviours (Hoarau et al, 2014; Trevail et al., 2015; Abayomi et al, 2017). As shown in Figure 3B, the proportions of MPs with different colours, in ranked order, were white-transparent (80.60%) > black (10.52%) > blue (4.58%) > red (2.27%) > green (2.03%). The higher proportion of white-transparent MP may, at least partially, be related to degradation, which can bleach out colourings and increase the relative 8

proportion of white-transparent MPs in the environment (Fan et al., 2019). Coloured MPs might originate from packaging and clothing from surrounding residential areas. As shown in Figure 3C, the higher percentage occurrence of fibre-shaped MPs in the FT mangrove may be the result of the disposal of municipal wastewater given its location in the PRE. Here, human activity is intense, with fishing activities presenting a potential source of fiber type MPs in the other mangroves (Figure S1). The film-shaped MPs may mainly derive from the fragmentation of plastic wrapping, bags and agricultural films, although restrictions have been imposed on their use in China. Foam-shaped MPs originate from the packaging and fisheries industries (Wang et al., 2019c), with granules mainly derived from cleaning and cosmetic products as well as the breakdown of larger degradable plastics (Fendall and Sewell, 2009; Cole et al., 2011). The potential threats of MPs in soil/sediment to biota include their adsorption of persistent organic pollutants (POPs) and their subsequent transportation into biota tissues (Batel et al., 2018). Figure 3D showed that the proportions of MPs with different sizes were as follows: 500 μm-5,000 μm (61.1%); 100-500 μm (19.8%); 50-100 μm (9.0%); and < 50 μm (10.1%), respectively. Li et al. (2019) suggested that small MPs (<1 mm) are degraded more effectively and as a result affect mangrove benthos negatively-although this has not be verified (Li et al., 2019). In this study, more than half of the recovered MPs were larger than 500 μm, indicating that the decomposition of plastic pollutants in these mangrove sediments is limited. It is speculated that mangroves act as a ‘sieve’ to hold back larger plastic particles whilst the interception of smaller MPs from anthropogenic activities is combatively limited. In the FT mangrove, although the MP concentration in the AM habitat type was lower than that in the MF habitat type (Figure 3A), higher concentrations of fine MPs in the AM samples (<100 μm) may have been related to the more efficient degradation of plastic wastes by marine bacteria (Auta et al., 2019). As shown in Figure 3E and Figure S2, MP polymer types were mainly composed by polypropylene (PP, 67.47%), polyethylene (PE, 13.05%) and polystyrene (PS, 10.45%), being similar to some other places in 9

China and elsewhere (Table 1). PP is widely used in microwave-proof containers and food packaging, with PE being common in homeware, bottles, pipes and toys, and PS is common in packaging, plastic cups and spectacle frames (Plastics Europe, 2016). In Europe, PP, PE and PS account for 80% of plastic production (Plastics Europe, 2016). These belong to the buoyant group of plastics, and their high accumulation in sediments may be related to ageing processes that improve formation of biofouling on MPs surface, reduce their hydrophobicity, and increase their deposition (Cole et al., 2011; Lobelle and Cunliffe, 2011; Cózar et al., 2014; Chae et al., 2015). PCA and HCA were used to explore the distribution characteristics of MPs based on shape, colour, size, and polymer type (Figure S3). The few differences between the MPs recovered from the sampled mangroves and the lack of clear clustering (except for DF-AM and DZG-AM) indicated similar characteristics. Although the concentration of MPs was highest in the FT mangrove, the similar pollution characteristics among all mangroves based on the PCA and HCA indicated that shape, size, colour, and polymer type should not negligible in evaluating MP pollution. On the other hand, the distinction of the DF-AM and DZG-AM samples may have been related to their lower proportion of fibres (Figure 3C). The reason why the proportions of MP fibres were higher in the DF and DZG samples is not clear and deserves further investigation. 3.3 Surface morphology of microplastics In the coastal environment, mechanical, chemical and biological processes affect the surface morphology of MPs, with a rough or cracked texture typical evidence of degradation (Veerasingam et al., 2016; Zhou et al., 2018). The SEM observations showed that the surfaces of MP fibres were relatively even, indicating a lower degree of degradation (Figure 4A). In comparison, the edges of MP films were irregularly fractured (Figure 4B). The MP foams had many fine cracks on their surfaces (Figure 4C), and granule had an uneven surface to which some amorphous substances were adhered (Figure 4D). In Figure S4, the surface of MP fibre (constituting the highest proportion of the samples) were further investigated using energy-dispersive X-ray 10

analysis (EDX). Silica (Si) was detected on surfaces of the MPs, which can affect surface morphology and other properties as has been found in coastal sediments of the Bohai Sea and the Yellow Sea, China (Zhou et al., 2018). EDX also showed that carbon, oxygen and some metals (Na, Al, K and Fe) had been adsorbed onto surfaces of the MP fibres. This may be related to the degradation of fibres, which leads to the occurrence of curled and cracked morphologies that enhance their ability to adsorb various organic and inorganic pollutants (Guo et al., 2013, 2018). In the past, the six studied mangroves have all suffered pollutions by polybrominated diphenyl ethers (PBDEs) and heavy metals to varying extents (Chai et al., 2019; Shi et al., 2019). Compared with the mean concentrations of MPs in the AM and MF habitat types (Table S4), there were similar variations in total heavy metals in the FT, ZJ, FCG and DF mangroves; similar patterns of decabromodiphenyl ether (BDE-209) were detected in the YX, DZG and DF mangrove sediments, with that of ∑PBDEs in ZJ, FCG and DF mangroves. Thus, the geochemical characteristics of MPs among the AM and MF habitat types were, at least partially, similar to those of heavy metals and PBDEs. In marine environments, the degradation of plastic is apparently different from non-biodegradable heavy metals and biodegradable organic pollutant PBDEs. Based on the observed variations in MPs between the AM and MF habitat types, the functional role of mangroves with respect to MPs cannot adhere to a simple “sink” or “source” model, and the impacting factors require some further systematic evaluation. 3.4 Factors impacting microplastics in mangrove sediments Generally speaking, the concentration and distribution of MPs in marine regions are impacted by anthropogenic activities (Frère et al., 2017), ocean currents (Xu et al., 2018) and meteorological conditions (Wang et al., 2019b). Anthropogenic factors including urbanisation, densely populated and industrial areas, and economic development are also related to MP pollution in coastal areas (Andrady, 2011; Jambeck et al., 2015; Yu et al., 2018; Zheng et al., 2019). In this study, selected socio-economic parameters were

11

investigated to quantify the links between MP pollution in mangroves and the surrounding anthropogenic activities (Figure 5). Based on the mean sedimentation rate in mangroves in China (4.1-57 mm a-1; Tan and Zhang, 1997) and given the limited data on socio-economic factors for the surrounding districts, partial data for 2015 were employed in this study (Figure 5). Population density (PD), urban land area (ULA) and gross domestic product (GDP) were chosen as simple measures of the intensity of anthropogenic activity. The variations in PD, ULA and GDP among the six districts surrounding the sampled mangroves were similar to that of MPs found in their sediments, indicating that MPs in mangroves may be originated from anthropogenic activities. Previous studies have also reported close links between population density, urban development and MP pollution (Yonkos et al., 2014; Townsend et al., 2019). For example, the larger the urban area and the larger the population, the more waste plastics are produced (Tang et al., 2018). Plastic products could directly reflect anthropogenic loading on the environment with respect to the plastic industry, which showed similar variation with MP pollution in the six studied mangroves (Figure 5D). Based on these observation, significant effort is required to establish appropriate management for discarded plastic products (such as reduced use and improved recycling of plastic products) to limit plastic input to rivers and other land run-off pathway. Generally, the fishery industry is an important source of MPs in coastal areas (Chen et al., 2018; Zhu et al., 2018; Wang et al., 2019d). With the general exception of the FT mangrove, variations in the seawater cultivated area (SCA), freshwater cultivated area (FCA), seawater aquatic products (SAP) and the production of freshwater aquatic products (FAP) were similar to the observed patterns of MPs (Figure 5). The declining fishing industry in the FT area relative to the other five areas may reflect broader on-going changes of land use that have significantly reduced the extent of aquaculture. This likely reflects a reduced impact from the fishing industry on MP pollution in the FT mangrove relative to the other mangroves. 12

In offshore oceans, ocean currents can influence the distribution of MPs (Yu et al., 2018; Pan et al., 2019). At lower latitudes, sea surface transfers may affect MPs in the Southern Ocean (Waller et al., 2017). Pent et al. (2017) also found that the southerly monsoon and the plume front in the Yangtze River Estuary may, at least partially, determine the fate of MPs in seawater and sediments in the East China Sea. The role of ocean currents in the distribution of MPs in offshore of the China coast was considered (Figure S5). MPs in the studied mangroves were assumed to be mainly affected by the Nanhai Sea Coast Current, which flows from north-to-south in winter and south-to-north in summer. However, the concentrations of MPs in the FT mangrove were higher than those in other mangroves (Figure 3A). In view of the important impacts of anthropogenic activities on MP pollution (Figure 5), the effects of ocean currents on the distribution of MP appeared to be limited in these mangroves. MPs had positive exponential relationship with sediment TOC (R2=0.2772, P < 0.01) and silt (R2=0.0844, P < 0.01), and increased with increasing TOC concentration (Figure 6A, C), indicating that finer sediment with a higher TOC content favours the deposition of MPs. This may be related to the tendency of mangroves to be distributed in offshore and lower wave-energy environments where finer particle and organic matter tend to accumulate (Lo et al., 2018). Furthermore, MP concentrations reduced with increasing percentages of sand fraction (Figure 6D). This may have resulted from the relatively low specific surface area-and hence adsorption capacity-of sand relative to silt, especially in the FCG mangrove, which was found to have a higher proportion of sand and lower MP concentrations (Figure 6, Table S3). These observations indicate that depositional environments with higher percentages of silt might affect the distribution of MPs. This is in contrast to previous research reporting no correlation among grain sizes and the occurrence of MPs (e.g. Blašković et al., 2017). Organic matter is important for determining the distribution and characteristics of MPs in the semi-enclosed Maowei Sea in the South China Sea (Li et al., 2019). Similarly, the structure, composition and stability of organic matter could affect the distribution of MPs in mangrove sediments (Li et al., 2018). 13

Thus, the positive relationship between MPs and silt observed in this study may be related to the fact that siltsized sediment can facilitate the deposition of organic matter and, thereby, enhance the deposition of MPs. More evidence is now required to verify these linkages. 3.5 Ecological risk posed by microplastics in mangroves A number of studies on the ecological risks posed by MPs have focussed on the harm they cause to organisms, their bioavailability and the additives embedded within them (Pedà et al., 2016; Hahladakis et al., 2018; Zhu et al., 2018) without evaluating the degree of pollution over entire areas. Currently, there are no systematic and standardised models to assess the ecological risks posed by MPs, and especially for mangroves. Here, 𝐸𝑖 , RI, H and PLI were applied to explore the ecological risks associated with the MP pollution of mangroves (Table 2). Essentially, the risks posed by PE were higher than PP and PS due to its higher 𝐸𝑖 values. The risks caused by PP were minor in all of the studied mangroves, regardless of habitat type. Overall, RI values in the FT mangrove were relatively higher, falling within the danger/extreme danger risk categories. The FT mangrove acts as a MP source and, therefore, poses a possible risk for MPs exposure, deserving further attention. There remain problems in risk assessment of MPs, however, including a lack of unified quantification models and well-defined background values. Furthermore, MPs can adsorb various pollutants (including heavy metals and POPs) from the surrounding environment (Zhang et al., 2015; Brennecke et al., 2016), which inevitably enhance their toxicity. How to evaluate the complex toxicity of MPs in combination with other pollutants remains unresolved and deserves further investigation. Some plastic polymers are biologically inert and have less of an impact on the aquatic environment (Matlack, 2001). In this study, the risk index (H) values for MPs in mangroves were lower than 10, indicating low levels of chemical risk (Table 2). However, a high concentration of MPs would result in a certain degree of environmental risk, as is the case in the FT mangrove. Importantly, H may underestimate the toxicity of MPs in mangrove sediments owing to a lack of polymer toxicity coefficients for other polymer types, which were 14

not identified in this study. The PLI values for MPs in the Futian mangrove fell within category II (AM) and category III (MF), while the other mangroves fell into category I. It was worth noting that the mangroves with a high PLI values had higher concentration of MPs but did not have the most hazardous polymers (Figure 3, Table 2). MP toxicity was mainly related to the low hazard scores of PP (1), PE (11) and PS (30) compared to much higher scores for, for example, polyvinyl chloride (10551), polyurethane (7384) and styreneacrylonitrile (6788) (Lithner et al., 2011). Moreover, the quantitative analysis of health risks associated with exposure to MPs is still lacking. Human may be exposed to MPs through food, drink and air pathways (Rist et al., 2018; Cox et al., 2019; Vianello et al., 2019). Therefore, further work is required to explore the potential exposure pathways of MPs in mangroves to the human body to clarify their potential risk to humans. Conclusion Along the coast of Southern China, MP concentrations in mangroves were found to be highest in Futian (2249±747 items/kg), followed by Zhanjiang (736±269 items/kg), Dongfang (649±443 items/kg), Dongzhaigang (431±170 items/kg), Yunxiao (424±127 items/kg), and Fangchenggang (227±173 items/kg). In terms of shape, colour and size, MPs were mainly fibrous, white-transparent and 500 μm-5,000 μm, respectively. The dominant polymer types of the recovered MPs mainly were polypropylene (67.5%), polyethylene (13.1%), and polystyrene (10.5%). Degradation of the recovered MPs was reflected in their surface morphology, and their gradual breakdown may enhance the adsorption of heavy metal pollutants onto their surfaces. MP pollution in mangrove was significantly linked to surrounding socio-economic development. The TOC and silt content of mangrove sediments also appeared to significantly affect the deposition of MPs. 𝐸𝑖 , RI, H and PLI risk indices showed that the ecological risks associated with the MPs in the studied mangroves are highest in Futian, which deserves further attention. Overall, the results present in this study provided fundamental data on MPs occurring in the mangroves of Southern China that can support further studies of the ecological consequences of MPs on mangrove macrofauna, shrimp, fish and even human. 15

Acknowledgement We thank the senior engineer Zhigang Qiu (ZTE Instruments Co., Ltd., Shenzhen) for a lot of help in Field investigation and lab operation of this research. This work was financially supported by the Program “ZeroWaste Agricultural Mulch Films for Crop in China ” (2017YFE0121900), the Program of Science and Technology of Shenzhen (JSGG20170413103811649), the Program of Science and Technology of Guangdong Province (Study on Ecological Investigation and Protection Patterns of Typical Coastal Mangroves in Guangdong Province), Special fund of State Key Joint Laboratory of Environment Simulation and Pollution Control (18K05ESPCP), and Shenzhen Municipal Development and Reform Commission (Discipline construction of watershed ecological engineering) References Abayomi, O.A., Range, P., Al-Ghouti, M.A., Obbard, J.P., Almeer, S.H., Ben-Hamadou, R., 2017. Microplastics in coastal environments of the Arabian Gulf. Mar. Pollut. Bull. 124, 181-188. Andrady, A.L., 2011. Microplastics in the marine environment. Mar. Pollut. Bull. 62, 1596-1605. Andrady, A.L., Neal, M.A., 2009. Applications and societal benefits of plastics. Phil. Trans. R. Soc. B. 364, 1977-1984. Auta, H.S., Emenike, C.U., Fauziah, S.H., 2019. Screening of Bacillus stains isolated from mangrove ecosystems in Peninsular Malaysia for microplastic degradation. Environ. Pollut. 231, 1552-1559. Barbier, E.B., Hacker, S.D., Kennedy, C., Koch, E.W., Stier, A.C., Silliman, B.R., 2011. The value of estuarine and coastal ecosystem services. Ecol. Monogr. 81, 169-193. Batel, A., Borchert, F., Reinwald, H., Erdinger, L., Braunbeck, T., 2018. Microplastic accumulation patterns and transfer to benzo[a] pyrene to adult zebrafish (Danio rerio) gills and zebrafish embryos. Environ. Pollut. 235, 918-930. Blašković, A., Fastelli, P., Čižmek, H., Guerranti, C., Renzi, M., 2017. Plastic litter in sediments from the Croatian marine protected area of the natural park of Telaščica bay (Adriatic Sea). Mar. Pollut. Bull. 114, 583-586. Booth, D.J., Sear, J., 2018. Coral expansion in Sydney and associated coral-reef fishes. Coral Reefs 37, 995. 16

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Figure captions Figure 1 Geographic location and sampling sites in mangroves of China (YX = Yunxiao; FT = Futian; ZJ = Zhanjiang; FCG = Fangchenggang; DZG = Dongzhaigang; DF = Dongfang). Figure 2 Photographs of typical microplastics in mangrove sediments of China Figure 3 Abundance (A), colour (B), shape (B), size (D), and polymer type (E) of microplastics in mangroves of China (YX = Yunxiao; FT = Futian; ZJ = Zhanjiang; FCG = Fangchenggang; DZG = Dongzhaigang; DF = Dongfang; AM = Avicennia marina; MF = mudflat; PP = polypropylene; PE = polyethylene; PS = polystyrene). Figure 4 Examples of microplastics found in mangrove sediments under a scanning electron microscope. Figure 5 Microplastics in mangrove sediments and indicators of socio-economic developments in the surrounding districts. (YX = Yunxiao; FT = Futian; ZJ = Zhanjiang; FCG = Fangchenggang; DZG = Dongzhaigang; DF = Dongfang). The surrounding cities for mangroves are as follows: YX = Zhangzhou; FT = Shenzhen; ZJ = Zhanjiang; FCG = Fangchenggang; DZG = Haikou; DF = Dongfang. For socio-economic 24

indicators, PD = people density; GDP = gross domestic product; PP = plastic products; SAP = seawater aquatic products; FAP = freshwater aquatic products; SCA = seawater cultivated area; FCA = freshwater cultivated area. Socio-economic data were obtained from the Statistics Bureau of local government for 2015. Figure 6 Pearson correlation analysis between microplastic concentrations and TOC (A), clay (B), silt (C) and sand (D).

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0.03

0.06 0.09 TOC

(C) 8000

0.12

0

0.15

0.1

0.2

Clay

0.3

(D)8000

R2 = 0.0844 P < 0.01

6000

MPs (items/kg dw)

MPs (items/kg dw)

R2 = 0.0217 P = 0.1278

4000 2000 0

0.4

0.5

R2 = 0.0682 P < 0.01

6000 4000 2000 0

0

0.2

0.4

Silt

0.6

0.8

1

0

0.2

0.4

Sand

0.6

0.8

1

Table 1 Microplastics in sediment environment in China and elsewhere. Location Sample type (depth) Concentration (dw) Yunxiao mangrove, China Sediment (top 5 cm) 260-580 items/kg Futian mangrove, China Sediment (top 5 cm) 980-3100 items/kg Zhanjing mangrove, China Sediment (top 5 cm) 460-1280 items/kg Fangchenggang mangrove, China Sediment (top 5 cm) 120-640 items/kg Dongzhaigang mangrove, China Sediment (top 5 cm) 260-700 items/kg Dongfang mangrove, China Sediment (top 5 cm) 160-2600 items/kg Mangrove in Qinzhou Bay, China Sediment (top 2 cm) 15-12852 items/kg South Yellow Sea, China Sediment (top 3 cm) 560-4205 items/kg Coast of Hong Kong, China Surface sand/sediment 0.58-2116 items/kg Mangrove of Maowei Sea, China Sediment (3-4 cm) 520-940 items/kg Pearl River Estuary, China Surface sediment 258±133 items/kg Bohai Sea, Yellow Sea, China Surface sediment 72-171.18 items/kg East China Sea, China Sediment (top 10 cm) 142±38 items/kg Yangtze Estuary, China Sediment (5-10 cm) 121±9 items/kg North Yellow Sea, China Sediment (top 5 cm) 37.1±42.7 items/kg Nakdong River, South Korea Sediment (top 2cm) 1970±62 particles/kg South east Coast, United States Surface sand 43-443 pieces/kg Mangrove, Singapore Sediment (3-4 cm) 12-62 items/kg

Main polymer types PP, 69.1-88.6%; PE, 0.9-8.9%; PS, 7.3-11% PP,75.2-84.7%; PE, 7.4-7.5%; PS, 7.5-4.0% PP, 69.8-75.4%; PE, 13.9-18.0%; PS, 7.7-11.3% PP, 65.7-77.2%; PE, 1.3-3.9%; PS, 14.7-16.5% PP, 39.8-70.1%; PE, 16.0-45.1%; PS, 12.4-13.9% PP, 39.4-54.7%; PE, 12.7-20.9%; PS, 9.0-10.3% PS, >98% PP, 31%; PE, 24%; Nylon, 19%; PS, 15% PE, 46.9%; PP, 13.8%; PET, 13.5% PE, 47.5-79.2%; PP, 2.5-42.5%; PS, 1.8%-12.8% PP, PE Rayon, 61.24%; PE,16.29%; PET, 12.36% Celophane, 37.2%; PET, 21.6%; PE, 17.6% Rayon, 63.1%; Polyester, 18.5; Acrylic, 13.9% PP, 44.5%; PE, 33.3%; Nylon, 11.1% PP, 24.8%; PE, 24.5% PET, 24% PE, PP, nylon, PVC

PP, polypropylene; PE, polyethylene; PS, polystyrene; PET, polyethylene terephthalate; KF, potassium formate; LMT, lithium metatungstate.

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Extraction (g ml-1) ZnCl2 solution, 1.5 ZnCl2 solution, 1.5 ZnCl2 solution, 1.5 ZnCl2 solution, 1.5 ZnCl2 solution, 1.5 ZnCl2 solution, 1.5 CaCl2 solution, 1.38 NaCl solution, 1.8 ZnCl2 solution, 1.6-1.7 KF solution, 1.5 KF solution, 1.5 NaCl solution, 1.2 NaI solution, 1.6 NaCl solution, 1.2 NaCl solution, 1.2 LMT solution, 1.6 NaCl solution, 1.27 Saline solution, 1.18

References This study This study This study This study This study This study Li et al., 2018 Wang et al., 2019d Lo et al., 2018 Li et al., 2019 Fan et al., 2019 Zhao et al., 2018 Zhang et al., 2019 Peng et al., 2017 Zhu et al., 2018 Eo et al., 2019 Yu et al., 2018 Nor and Obbard, 2014

Table 2 Potential ecological risk evaluation for microplastics in six representative mangroves of Southern China. 𝐸𝑖 (Risk level) RI H PLI PP PE PS YX-AM 2.2(Mi) 187.2(Da) 48.5(Me) 237.8(Me) 5.6(I) 3.1(I) YX-MF 3.2(Mi) 22.0(Mi) 36.9(Mi) 62.1(Mi) 3.3(I) 3.6(I) FT-AM 12.5(Mi) 869.9(Ex) 175.4(Da) 1057.7(Da) 4.3(I) 16.3(II) FT-MF 18.7(Mi) 1090.1(Ex) 124.6(Hi) 1233.4(Ex) 3.0(I) 23.5(III) ZJ-AM 3.8(Mi) 649.6(Ex) 85.4(Hi) 738.8(Da) 6.1(I) 5.2(I) ZJ-MF 1.6(Mi) 198.2(Da) 23.1(Mi) 222.9(Me) 4.7(I) 2.1(I) FCG-AM 1.1(Mi) 44.0(Me) 34.6(Mi) 79.8(Mi) 6.5(I) 1.6(I) FCG-MF 1.0(Mi) 11.0(Mi) 30.0(Mi) 42.0(Mi) 6.2(I) 1.3(I) DZG-AM 3.0(Mi) 2257.3(Ex) 145.4(Hi) 2405.6(Ex) 9.6(I) 7.3(I) DZG-MF 2.2(Mi) 341.3(Ex) 55.4(Me) 399.0(Hi) 6.3(I) 3.1(I) DF-AM 1.9(Mi) 407.4(Ex) 69.2(Me) 478.5(Hi) 7.8(I) 3.7(I) DF-MF 2.8(Mi) 715.7(Ex) 64.6(Me) 783.1(Da) 6.6(I) 5.1(I) 𝐸𝑖 , potential ecological hazardous single index; RI = potential ecological risk; H = polymer risk index; PLI = pollution load index; YX = Yunxiao; FT = Futian; ZJ = Zhanjiang; FCG = Fangchenggang; DZG = Dongzhaigang; DF = Dongfang. AM = Avicennia marina; MF = mudflat. Mi = minor, ; Me = medium, ; Hi = high, ; Da = danger, ; Ex = extreme danger, . Mangroves

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YX mangrove FT mangrove ZJ mangrove FCG mangrove DZG mangrove DF mangrove

China

Highlights (1) Mean concentration of MPs was the highest in Futian mangrove with 2249 items/kg dw. (2) MPs in mangroves were mainly characterized as fiber, white-transparent, and 500 μm-5 mm. (3) MPs in mangroves were mainly composed of polypropylene, polyethylene, and polystyrene. (4) Surrounding social-economic development affected distribution of MPs in mangrove. (5) MPs in mangroves had ecological risks, with that to be the highest in Futian mangrove.

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Conflict of interest The authors declared that they have no conflicts of interest to this work.

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